Watershed Sensitivity Measurement
Strategy for Identifying Resources at
Risk from Acidic Deposition
Institute of Ecology, Indianapolis, IN
Prepared for
Environmental Research Lab.-Duluth, MN
Jan 84
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EPA-600/3-84-011
January 1984
WATERSHED SENSITIVITY MEASUREMENT STRATEGY FOR
IDENTIFYING RESOURCES AT RISK FROM ACIDIC DEPOSITION
by
Orie L. Loucks
Roland W. Usher
The Institute of Ecology
Indianapolis, Indiana 46208
David Rapport William Swanson
University of Toronto Miami University
Toronto, Ontario, Canada Oxford, Ohio 45056
Richard W. Miller
The Institute of Ecology
and Butler University
Indianapolis, Indiana 46208
Cooperative Agreement CR809328
Project Coordinator
Gary E. Glass
Environmental Research laboratory-Duluth
Duluth, Minnesota 55804
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
DULUTH, MINNESOTA 55804
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. HEPOHTNO.
EPA-600/3-84-011
3. RECIPIENT'S ACCESSION NO.~
pap A 1*1 209
4. TITLE AND SUdTITLE
Watershed Sensitivity Measurement Stragety for Identify-
ing Resources at Risk from Acidic Deposition
5. REPORT DATE
January 1984
6. PERFORMING ORGANIZATION CODE
I. AUTHORIS)
0. L. Loucks
8, PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
The Institute of Ecology
Indianapolis, Indiana 46200
10. PROGRAM ELEMENT NO.
11. CONTRACT/GRANT NO.
CR809328
12. SPONSORING AGENCY NAME AND ADDRESS
Environmental Research Laboratory
Office of Research and Development
U.S. Environmental Prot-ction Agency
Duluth, MN 55804
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
13. SUPPLEMENTARY NOTES
15. ABSTRACT
--Several lines of research on the effects of acidic deposition have been supported by EPA through
Its Environmental Research Laboratories and through a Cooperative Agreement with North Carolina State
University. The study reported here was carried out as ,a subcontract to The Institute of Ecology from N.C.
State University for research supporting programs at the Environmental Research Laboratory-Duluth. The
objectives Included review of existing literature on the use of Indices for quantifying resource. *tatus and
predicting long-term trends In relation to'acidic deposition, review of options as to the form of a "sensitivity
Index" or loading tolerance model for use In determining resources at risk, and Identification of validation . .
steps needed to complete testing of the measure or model, arid to begin Its application.
One section of the report describes the suite of measures which, when taken together, best Identify
areas potentially sensitive to acidic Inputs. Each of the component measures, when viewed separately, has
certain limitations which prevent It from being an adequate measure of sensitivity; when considered together,
however, as an Integratlve measure, the limitations are less significant.
For non-agricultural systems, forest site Index appears to be a wy|l-establIshed Integratlve
measure capable of responding to altered soil and water chemistry. The extent to which site Index Is related to
changes In cation nutrient storage (due to cation stripping by acid precipitation) or other pollutant Impacts
still Is Incompletely documented, however. Measurements of aquatic sensitivity have been developed more fully,
and a number of experimental and field data-based approaches exist. These Include the Calclte Saturation Index,
the Henrlckson nomograph, the Almer/Dlckson relation and an additional measure proposed here based on pH shock
effects during acid flushing events. This Integratlve response property appears to describe a complex
environment leading to.specles and population effects associated with periodic but physiologically Important
exposures to H+ and AI \.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
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RELEASE TO PUBLIC
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'"A Form 2220-1 {Re*. 4-77) PREVIOUS EDITION is OB>OL£TE 1
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NOTICE
This document has been reviewed in accordance with
U.S. Environmental Protection Agency policy and
approved for publication. Mention of trade names
or commercial products does not constitute endorse-
ment or recommendation for use.
11
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TABLE OF CONTENTS
Page
LIST OF TABLES 1V
LIST OF FIGURES . v
SECTION I - INTRODUCTION 1
Concepts of Measurements for Ecosystem Response to Acid Deposition 3
Study Objectives 4
SECTION II - SUMMARY AND FINDINGS 7
Approach 7
Results 8
SECTION III - SUB-COMPONENTS OF LAKE/WATERSHED SENSITIVITY 11
Water Flow Pattern 11
Nitrogen Cycling 13
Sulfur Cycling 16
Alkalinity 18
Organic Interactions 19
Cation Nutrient Stripping' 20
Hydrogen Ion Toxicity 26
Aluminum Mobilization and Toxicity 35
Heavy Metal Mobilization and Toxicity 46
SECTION IV - METHODOLOGIES FOR QUANTIFYING THE SENSITIVITY OF
TERRESTRIAL AND AQUATIC AREAS 53
Measures of Sensitivity to Forest Productivity Response? . 54
Measures of Aquatic Ecosystem Sensitivity 56
SECTION V - RESEARCH AND DATA REQUIREMENTS FOR APPLICATION
AND VALIDATION 71
An Approach to Assembling Sensitivity Measurements 71
Other Information/Research Needs for Sensitivity Measure
Validation 75
REFERENCES 77
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LIST OF TABLES
TABLE III-l: Mean and median concentrations of dissolved
ionic materials in bulk precipitation, canopy
throughfall, forest floor, and A2 horizon per-
colate, and springwater during the 1975 and 1976
growing seasons at the Mt. Moosilauke subalpine
sampling sites, New Hampshire
Paoe
24
TABLE III-2: Net export of major ions for calibrated water-
sheds in Canada
TABLE III-3: Literature survey on pH toxicity to various fish
species
TABLE III-4: Summary of the effect of pH values on fish
29
34
TABLE III-5: Summary of literature concerning aluminum toxi-
city to fish
44
TABLE III-6: Times to 50% mortality for fish in trials A-F
48
TABLE V-l: Listing of data needs and computation procedures
applying sensitivity measures and pollutant
loading tolerance models in northeastern North
America
72
IV
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LIST OF FIGURES
Figure 1-1: Flow diagram showing system linkages for acid
precipitation formation, deposition and effects as a conse-
quence of nitrogen and sulfur oxide emissions from fossil
fuel combustion. 2
Figure III-l: Detailed diagram illustrating acid precipitation
formation, deposition and effects on terrestrial and aqua-
ti'- ecosystems. 12
Figure III-2: Simplified nitrogen cycle showing acid-forming
and acid-consuming reactions. 15
Figure III-3: Simplified sulfur cycle illustrating acid-forming
and acid-consuming processes. 17
Figure III-4: Nutrient budgets for podzol-brown earth lysi-
meters treated with artificial rain of pHs 3, 4, and 5.6:
21
Figure III-5: Nutrient budgets for podzol lysimeters treated
with simulated rain of pHs 2, 3, 4 and 6. 22
Figure III-6: Solubility of aluminum as affected by pH. 36
Figure III-7: Aluminum concentrations found in Norwegian and
Swedish clear water lakes. 37
Figure III-8: Nutrient release from snowpack during 1978 spring
melt as determined by monitoring runoff from mini catchments
in Norway.
Figure III-9: Relationship between H+ loading rate and observed
aluminum concentrations in lysimeter effluent, minicatch-
ments or freshwater lakes. ' 42
Figure 111-10: Effect of altering pH on cumulative mortality of
brook trout exposed to three concentrations of aluminum. 47
Figure IV-1: Site index classes determined by plottir.n age
against average height of dominant tree. 55
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Figure IV-2: Theoretical height/age curves for three soils. 57
Figure IV-3: Results of a mortality test at various tempera-
tures plotted on logarithmic-probit paper.. 60
Figure IV-4: Effects of various sulfate loading rates on lake
pH for lakes in very sensitive and somewhat less sensitive
surroundings in Sweden. 62
Figure IV-5: A nomograph to predict the pH of lakes. 64
Figure IV-6: River pH, rainfall pH, rainfall accumulation and
discharge rate for the Shavers Fork River, W. Va. 67
Figure IV-7: pH depression in Little Moose Lake, N.Y., occur-
ring during spring snowmelt. 68
Figure IV-8: . Relationship between sulfate loading rate and
potential pH changes occurring during spring melt for
sensitive areas and slightly less sensitive areas. 69
vi
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SECTION I
INTRODUCTION
Orie L. Loucks, David Rapport, and Richard Miller
Atmospheric deposition of acidic pollutants appears to have increased
in both area and intensity since the middle 1950's (Likens et a_L 1979)
and possibly over a longer time (Davis 1980). This increase has been cor-
related with increasing use of fossil fuel for automobiles as well as
electricity, and the associated releases of oxides of nitrogen and sulfur
to the atmosphere. Utility emissions increased most rapidly, during the
1960's (EPA 1978), and, increasingly over this period, were emitted from
elevated stack heights where dispersion in the lower atmosphere (rather
than local scavenging) becomes more effective. Once in the atmosphere,
the emissions can be converted to nitric and sulfuric acids. Processes of
rainfall and dry deposition then lead to the accumulation of these pro-
ducts once again on land and surface water resources. Despite questions
as to rt;tes of reactions and the control mechanisms, enough is known about
the acidic deposition phenomenon, however, to describe the principal types
of impact, the mechanisms by which effects are expressed, and the nature
and variability in resource response times.
Figure 1-1 is a conceptual model showing reactions, transport path-
ways and the principal ecological consequences resulting from atmospheric
deposition of elevated ozone, hydrogen, nitrate and sulfate ions. The
materials and processes producing the ultimate effects of this deposition
must be viewed as a coupled system which, for simplicity, is shown here as
a flow diagram. The effects of interest are the direct effects on vege-
tation, the alteration of groundwater and associated stream chemistry, the
alteration of soil chemistry and associated terrestrial productivity, and
the effects on fish and other biota in aquatic systems. These effects are
largely mediated through the watershed, alt-hough direct flux of acidic
materials to surface waters can be important in some situations. The
magnitude of the effects will vary depending on characteristics of the
watershed; only those ecosystems with little or no carbonate buffering are
likely to show the listed effects from acidic inputs. The actual
mechanisms of the chemical and biological transformations implied in
Figure 1-1 are complex and will be discussed throughout the report.
Effects from oxidants and acid precipitation are both .partly
transient and partly irreversible in nature. Transient effects include
alteration of soil solution chemistry and adverse impacts on crops,
forests, fish and other biota. Irreversible effects include long-term
changes in species composition, stripping of cation nutrients from soil
systems, modification of groundwater. and relatively permanent reduction
of buffering capacity for entire catchment areas (see later sections for
references). Nutrients leached from the soil column are permanently lost,
and because of the rates of natural leaching, geological replacement may
be on the order of thousands of years. Shallow groundwater can be des-
cribed as a sink for sulfate and hydrogen ions, heavy metals and aluminum
leached from the soil column, but these ions also reach surface waters.
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ATMOSPHI
SO,
HO,
H,0
HXCX°X ,
ERIC REA(
:TIONS
sof
H*
o,
FLUX TO
SURFACE WATERS
H*. SO, N03
FLUX TO
VEGETATION
SO,, NO,. 03
CELLULAR
METABOLIC
TOXICITIES
CATION DISPLACEMENT
METAL MOBILIZATION
EFFECTS ON 1
FISH AND
OTHER BIOTAJ
Figure 1-1. Flow diagram showing system linkages for acid precipitation
formation, deposition and effects as a consequence of
nitrogen and sulfur oxide emissions from fossil fuel com-
bustion.
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Once in the groundwater system, these elements can 'influence drinking
water. Acid-altered groundwter is released slowly as seepage to springs
and other surface waters.
Government, industry, and the public are all concerned with the
long-term responses to deposition of ozone and acidic pollutants. One
possible response is the control of atmospheric inputs through reductions
of S02 and NO emissions from fossil fuel combustion, a responsibility of
regulatory agencies at state and federal levels of government. Glass e_t
aj_. (1980) reported on a series of workshops held in late 1979 with state
regulatory personnel of Wisconsin, Minnesota and Michigan regarding the:r
concerns with and role in emissions control and environmental protection.
These three states contain large areas characteristic of the strsitive
resources being affected elsewhare, although acidification responses in
lakes have not yet been demonstrated.
There was considerable interest among the 1979 workshop participants
in controlling the precursors of acid deposition, while recognizing that
there must be a clear indication of the prospective benefits in resource
protection to be gained from such action. This interest resulted in
specific suggestions for development of a "damage susceptibility index"
which could be used to inventory susceptible resources in regions that are
not yet seriously affected. Specific criteria suggested for such an index
included (1) identification of buffering capacity of the lake/watershed
system, and (2) consideration of potential effects on all biological
resources, including timber and fish.
CONCEPTS OF MEASUREMENT FOR ECOSYSTEM RESPONSES TO ACID DEPOSITION
The loading rate of pollutants that is just below a rate producing
demonstrable effects can be defined as a loading tolerance for a system
with a defined sensitivity. This tolerance and sensitivity must be
related to specific effects observed and to measurable lake/ watershed
characteristics - chemical, geological, and biological. The pathways by
which effects of acidification are expressed in ecosystems are indicated
in Figure 1-1 as sequential events: Fossil fuel emissions, followed by
atmospheric reactions, deposition, chemical changes and, finally,
biological effects. While this sequence is an oversimplification, given
the state of existing knowledge, one could expect that early symptoms of
acidification need not lie solely in altered biota, nor in the chemical
properties of the soils and water. Various biological measures might be
responsive solely to che periodic or.3et of acid stress. For example,
changes in the establishment and recovery patterns of algal species at the
mouth of a stream receiving acidic pulses from spring snowmelt could be a
very sensitive measure of irregularly expressed stresses.
In developing indicators of pollutant impacts on both terrestrial and
aquatic ecosystems, three classes of measures can be used to reflect
system transformations (Rapport and Regier 1980). Two of these, termed
here "indicators" and "integrators," are well :"ited to representing the
effects of acidification. The third, the so-called "environmental quality
index," is basically unsound because it attempts to compress the many
dimensions of environmental transformation into a single dimensionless
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index number, often without -benefit of a scientifically valid model or
logical basis for such a representation. It appears to be unwise to
attempt construction of environmental quality indices, or agglomerative
.measures of environmental quality to reflect environmental transformation,
whether due to acid rain or any other factors. Inhaber (1974) proposed
such an index for Canada, and the results have led to a general disen-
chantment with the environmental index approach.
The alternative, namely to suggest a suite of direct or relatively
direct measures which reflect various dimensions of the impact process,
and which are based on scientifically sound models, appears mere promising
(Rapport and Regier 1980). These two types of measures can be defined as
fo11ows:
Indicators - measures of ecosystem alteration which
are t?latively specific or diagnostic of particular types
of stresses, and which are likely to be useful as early
warning of stressed environments. Some examples in the
acid rain context include reduction in the populations of
shcrt-lived sensitive species and specific chemical changes
such as the mobilization of aluminum during acid flushing
events. Indicators must be suited to picking up transient
events which may otherwise be difficult to measure.
Integrators - measures of ecosystem alteration which
are capable of reflecting a number of stress impacts
through time, and which reflect these impacts retrospec-
tively and synergistically. An example is the forest site
index measure which integrates a variety of environmental
transformations as they affect forest responses (growth and.
productivity) over the lifespan of the trees being
measured.
To a large extent these two types of measures complement each other.
The "indicators" are diagnostic of particular types of stresses, and
would, due to their greater specificity, function in evaluating both
causative and curative strategies. On the other hand, the integrators,
although retrospective and often non-specific, overcome to a large degree
the difficulty of capturing short-term events and synerTistic effects in
the ecosystem. Tc use a medical analogy, the integrators are somewhat
akin to the "vital signs" of an ecosystem, while the indicators function
as the clinical symptoms of distress (Rapport e_t a_L 1980).
STUDY OBJECTIVES
The initial sections of this report have as their objective a review
of the various measurements and relationships being used to quantify
resource status in oxidant and acid-stressed systems, and to identify
integrative measures of lake/watershed response to acidic pollutants.
Considering the voluminous literature on the subject and the many possible
effects, the review will be limited to those topics most closely related
to the effects described in Figure 1-1. '
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A second objective will be to examine a group of measures and/or
models selected to represent ecosystem responses, such as resource
nutrient cycling and biotic composition and quality. These measures
require a compromise between generality, reality and precision because not
all of these qualities can be achieved simultaneously (Levins 1968). The
emphasis in the report will ba on generality and reality for the time
being. The measures used must enhance the likelihood of prediction of
future lake/watershed responses in lightly impacted systems, given current
acid loading rates over the period of a few decades.
Finally, the data needed to evaluate the usefulness of a group of
measures or predictive models will be identified. The entire group of
measures should be tested and verified with appropriate studies before
effective regulatory use can be expected. The final test, application,
can be met best if all others are satisfied, and if the measures relate to
obvious economic benefits from acidic pollutant abatement.
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SECTION II
SUMMARY AND FINDINGS
Orie L. Loucks
In August, 1979, President Jimmy Carter established a ten-year feder-
ally funded acid rain assessment program. At about the same time, the
Environmental Protection Agency awarded a Cooperative Agreement to North
Carolina State University to conduct a program of subcontracted studies on
biological effects of acid precipitation. The Institute of Ecology (TIE)
was awarded a subcontract entitled "Assessment £f the Sensitivity Index
Concept for Evaluating Resources at Risk from Atmospheric Pollutant
Deposition (Acid Rain)," to be carried out in support of studies at the
ERL-Ouluth, U.S. Environmental Protection Agency.
The study was concerned principally with developing sensitivity
measures for evaluating terrestrial and aquatic resources at risk from
atmospheric pollutant deposition (oxidants as well as acid rain). The
main objectives have been:
(1) To review existing literature on the use of indices for quan-
* tifying resource status, predicting long-term trends in eco-
system and resource responses to acid deposition, and for
assessing overall risks from atmospheric pollutant deposition in
relation to air emissions management;
(2) To consider several options as to the form of a "sensitivity
index" or pollutant loading tolerance model for use in deter-
mining resources at risk from energy development, and outline
how such a measure would function in regional inventory of risk
from pollutant deposition or in the assessment of benefit from
acid precursor control;
(3) To identify validation steps needed, data required (existing
data or new measurements), and the steps required to complete
testing and begin application of the sensitivity measures or
loading tolerance model in regional and national energy develop-
ment decisions.
APPROACH
Early in the study, attempts were made to compress various .acid rain
indices (i.e., the McFee. soil sensitivity measure, the Calcite Saturation
Index, etc.) into one sensitivity index. As this concept was examined, it
was recognized as an unsound approach because it attempts to express too
many dimensions of environmental transformation into a single dimension-
less index number. Instead, a suite of integrative sensitivity measures
was considered which, when viewed as a wh^le, would be more likely to be
suitable for quantifying sensitive systems and their response to pollutant
inputs. To provide a technical basis for considering such measures, an
Preceding page blank
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extensive review is presented on the mechanisms by which pollutant inputs
alter the biological productivity of terrestrial and aquatic systems.
RESULTS
One major section of the report, entitled Sub-Components of Lake/
Watershed Sensitivity, contains the background material on pertinent
characteristics of acid precipitation effects on soils, lakes and water-
sheds, and lays a foundation for both the recommended sensitivity measures
and the outline of research and data needs in later sections. This
section also provides an understanding of the various processes occurring
within the system, information that is required for developing and evalu-
ating measures of watershed/lake sensitivity. Topics of interest include
hydrologic flows, the nitrogen and sulfur cycles, alkalinity relation-
ships, interactions of acids with organic material, nutrient stripping, H+
toxicity, the mobilization and toxicity of aluminum and heavy metals, and
synergisms between H+, aluminum and heavy metals.
The section principally focused on new results, entitled Methodolo-
gies for Identifying Sensitive Terrestrial and Aquatic Areas, describes
the various options for measures that best identify potentially sensitive
areas. Three separate measures are employed for the terrestrial compo-
nent: McFee's (1980) soil sensitivity classification based on cation
exchange capacity; the soil sensitivity classification based on base
saturation (Coote et al. 1980); and the forest site index. The site index
(SI) concept has been accepted as a measure of forest productivity for
many decades and is examined here as a methodology for measuring changes
in potential forest growth du,: to long-term acidic inputs. The magnitude
of site index changes due to a combination of oxidants, changes in cation
nutrient storage (resulting from cation stripping by acidic precipita-
tion), and aluminum toxicity effects is still incompletely quantified,
however. Studies will be required using available data bases on oxidant
exposures, apparent changes in total nutrient stocks, and aluminum
mobilization in relation to acidic inputs.
Integrative measures for expressing aquatic sensitivity have been
developed more fully, and a number of experimental and field data-based
methods exist. These include the Calcite Saturation Index, the Henricksen
nomograph and the Almer/Oickson relation. An additional measure, based on
pH shock effects during acid flushing events, attempts to identify
species/population impacts associated with short-term, physiologically
important exposures of critical life stages to H+ and A13+. This approach
can be quantified for the pH depression levels already being observed in
various sensitive regions. As with the other models, its applicability is
relatively untested for regions where acidic inputs are moderate and pH
shock effects are intermediate in significance.
The final section, entitled Concluding Comments and Research Needs,
is a brief statement on the data needed to achieve an effective validation
of the sensitivity measurement approaches. Insufficient data presently
exist for fully quantifying hydrogen ion or sulfate fluxes through a wide
variety of watersheds, or aluminum mobilization during peak H+ concentra-
tions. Nutrient stripping effects on forest productivity, effects on
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mammals and predatory birds from metal mobilization and food chain altera-
tions, and the role of organic matter in mediating acid deposition effects
are all too poorly known to have a fully reliable, locally applicable
sensitivity measure at this time.
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SECTION III
SUB-COMPONENTS OF LAKE/WATERSHED SENSITIVITY
Richard Miller, Roland Usher, Orie Loucks and William Swanson
Devising an adequate measure of forest, watershed and lake sensiti-
vity to combinations of pollutants requires an understanding not only of
the response mechanisms, but also of various processes occurring within
the system which might mediate or interact with the effects of acidic
deposition. Within the lake/watershed system these processes include
water flow, nitrogen and sulfur cycling, alkalinity production, inter-
actions with organic material, nutrient stripping (leaching), H+ toxicity,
mobilization and toxicity of aluminum and heavy metals, and synergisms
between H+, aluminum and heavy metals.
Figure III-l is an expanded version of Figure 1-1 and is meant to put
all the sub-components of forest, lake and watershed sensitivity into
perspective. The scheme is obviously complex and highly linked; changes
in one place have effacts elsewhere. Systems of differing overall sensi-
tivity are indicated, and fluxes from the different systems will vary from
zero in the highly buffered areas to generally high fluxes in systems
undergoing alteration (Overrein 1980). The events occurring following
watershed deposition are discussed below.
WATER FLOW PATTERN
The pathways by which watar flows through a watershed are implicit in
Figure III-l and are of crucial importance to acidification by determining
the contact time between acid inputs and p-tt .ially neutralizing soil
constituents. A considerable amount of surface runoff will greatly in-
crease the through-put of acid directly to the aquatic components of the
system. The exact flow patterns depend on a variety of factors, including
topography and slope, iioil type, rate of input and vegetative cover.
Two types of water flow exist: surface flow and sub-surfaca flow.
Surface flow is water move.-?'/., across the top of the soil to an arbitrary
depth of 2 cm. This flow .an be interpreted in two ways: dynamic and
continuous. Dooge (1973) haj developed mathematical equations describing
surface flow that take into account flow velocity, slope, friction of
slope, depth of flow, flow velocity/unit area, length of slope, length of
flow, outflow, outflow equilibrium, time, characteristic time dependent on
flow intensity, and duration of uniform flow. Predicting the flow rate of
water through the surface soil is important because this layer has the
most biological activity and is therefore one of the most influential
components of the system.
Sub-surface flow is the movement of water through the soil profile
below the surface. This aspect of hydrology is perhaps the most difficult
to accurately simulate in mathematical form. The .movement of water can
proceed in two directions: (1) percolation (downward movement to the
11
Preceding page blank
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NATURAL SOURCES
FOSSIL FUEL EMISSIONS
CO,.SO..NO.,H.C.O.
FLUX TO WATERSHEDS
1
ATMOSPHERIC REACTIONS
SO, ) f SOI'
NO. _ NO',
H,0 *" H*
H.C.O. O,
FLUX TO LAKES
HIGHLY BUFFERED i
( SOILS j
_._i—
{ POORLY Ni
I BUFFERED |
\ SOILS /
ACID
! SOILS !
SOLUBLE NUTRIENT
IONS Ca^Mg7tNa*,K:
PO:;SO];HCOI
I
SOLUBLE TOXIC
IONS HtAI^PbT
CuTHgTZn'*
ELEVATED H*. METALS,
AND NO', LEVELS IN
GROUNDWATER
TO FISH AND
AOUATiC BIOTA
ALTERED
PHYSIOL.
PROCESS
..I.. ,-J-_
,' ACID N| ,' POORLY ")
.1.
LAKES !
i BUFFERED
1 LAKES ;
i HIGHLY BUFFERED
yCARBONATEl LAKES*
DISRUPTION OF
ORGANIC MATTER A
NITROGEN CYCLE
N01,NH4,ORG. N
SOLUBLE TOXIC
IONS H'.AI ,Pb ,
SOLUBLE NUTRIENT
IONS Ca2*,Ma2*. NO;
FROM
WATERSHED
SOILS / /
ALTERED
PHYSIOL.
PROCESS
EFFECTS ON CROPS
AND FORESTS
EFFECTS ON
HUMAN HEALTH
EFFECTS ON FISH AND
OTHER AC'.'ATIC BIOTA
Figure III-l. Flow diagram showing system linkages for acid precipitation formation, deposition and effects as a
consequence of nitrogen and sulfur oxide emissions from fossil fuel combustion.
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groundwater), and (2) capillary n' ~e (upward movement from the ground-
water). Implicit in both direction, of flow is equilibrium between the
water and the soil. Mathematical equations describing percolation have
been developed that take into account such factors as soil suction; hy-
draulic conductivity, hydraulic diffusivity, soil moisture content and
elevation above water table.
The above discussion has dealt with the liquid form of precipitation
only. The solid form, snow, does not contribute to v.aterflow until snow-
melt occurs. Light (1941) has proposed a complex bjt realistic approach
for predicting snowmelt that takes into account many of the major micro-
climatic variables that can influence snowmelt.
Once the snow is in liquid form it can be placed into the surface and
sub-surface flow equations. Wright and Dovland (1977) have discussed the
problems associated with snowmelt, snowpacking and major ion concentra-
tions in Norway. Fahey (1979) and Wright and Dovland (1977) imply that
slow input of acid, metals and ions from the snow into the watershed
occurs as a result of snowmelt. One question that has not been properly
addressed in the literature is the importance of snowmelt and soil contact
with the water before it enters the lake and/or river, as it pertains to
acid and heavy metal inputs to the watershed.
Finally, Glass e_t a_L (1981) note that because, in northern soils,
the sulfate ion is a relatively conservative substance (Harvey et al_.
1981), high rates of evaporation can leave the precipitation sulfate
concentrated in the soil solution (and lake water) by a factor controlled
by evaporative losses. The equations for lake sulfate concentrations
developed by Henriksen (1980, see later sections) show this factor, plus
the dry deposition, to be 1.9 for central Norway. Regions of propor-
tionately higher evaporative losses have higher observed sulfate concen-
trations in lake watar than are predicted by the Henriksen equations.
These processes vary with precipitation and temperature patterns between
regions and, in areas of strong topography, from one watershed to the
next.
NITROGEN CYCLING
Becaus* large quantities of amironium and nitrate can be present in
precipitation, it is of interest to Jiscuss their relationship to acidi-
fication. First, a brief review of ".he natural reactions of the nitrogen
cycle will help in understanding the possible fates of added ammonium and
nitric acid.
Atmospheric nitrogen (N2) is reduced by the soil microflora, collec-
tively termed nitrogen fixerr>, to ammonia (NH3). The NH3 is hydrogenated
in the soil system to the ammonium ion, NH4-f-, resulting in a loss of 1H+
from the soil system. The NH4+ either may be assimilated by plants (re-
sulting in an input of 1H+, thus balancing previous output), or it may be
oxidized to nitrate (N03~) forming 2H+. The ammonium ion assimilated by
plants will eventually be returned to the soil system as NH3. The fol-
lowing reactions simplify the oxidation of NH4+, a process termed
nitrification:
13
-------
NH4 + r-sOj, •* 2H + H20 + N02 (eq. 1)
N02" + 5s02 -> N03~ (eq. 2)
Nitrate is generally taken up by plants, resulting in an input of 1 OH-
ion to the system, but under anaerobic conditions it may be reduced to
nitrite (N02-), summarized in equation 3:
N03" + 2H+ + 2e- -> N02" + H20 (eq. 3)
The nitrite (under anaerobic conditions) may be further reduced to nitrous
oxide (N20) or atmospheric nitrogen (N2). Both of these reactions consume
H+ ions from the soil system. .The magnitude of H+ consumption by the
process of nitrate conversion to N2 (denitrification) is largely unknown
for most terrestrial ecosystems. These reactions probably occur more
readily in aquatic systems, where anaerobic conditions more frequently
occur.
The molecular forms of nitrogen deposited in acid rain can result in
either acidification or neutralization of surface waters. Nitrogen as
nitrate ions (N04-) can be incorporated directly by plants, and the
resulting reactions with water release hydroxyl ions (OH-) into the en-
vironment (Figure III-2). The hydroxyl ions can raise the pH of the soil
and water or neutralize the hydrogen ions of the nitric acid. Natural
decomposition of nitrogenous plant materials also releases hydrogen ions,
but net accumulation of plant tissue dominates in most ecosystems; hence
net production of neutralizing capacity from nitrate addition is usually
dominant.
Several scenarios operating in terrestrial and/or aquatic systems
involving ammonium and nitrate inputs and acidification are possible:
(1) If all nitrogen in precipitation is in the form of HN03, or if
the concentration of HN03 is much greater than the concentration
of NH4+, and if all nitrogen is assimilated, then a decrease in
the acidity of the system is possible. (The H+ associated with
N03- may be neutralized by the OH- released when nitrate is
assimilated).
(2) If the HN03 concentration equals the NH4+ concentration and if
there is no leaching, then there may be a slight increase in the
acidity of the system. (The H+ from HN03 may be neutralized
when N03- is assimilated; the NH4+ adds 1H+ to the system).
(3) If ell the nitrogen in rain is in the form of NH4+, or if the
NH4+ concentration is much greater than the N03- concentration,
and if there is no leaching, then an increase in the acidity of
the system will occur.
(4) If HN03 predominates in precipitation, and if leaching occurs,
then an input of H+ to the system will occur.
(5) If NH4+ predominates in precipitation and if assimilation does
not occur, three routes are possible:
14
-------
PLANT
NH2
r\ — v
F
SOIL
\
Nh
R-(
_, — pj -^a i\j n, •"•« —^p-
r
1 organic nitrogen H
+J
i ii
c
RtBm. M1 ' ^fel>- l~i.
\ft~- | VJ ( | «•>— J»«-
IM
i
N
4 ^ +
-J
^v
OH
1 1 "*" -^ &»- r
"4 °^ f
SJ(
>
~ —
R organic nitrogen
Figure III-2. Simplified nitrogen cycle showing acid-forming and acid-consuming reactions. Assumes all N
utilized as N03-. From Reuss 1976.
-------
(a) Leaching of NH4+ intact (acidifies receiving water);
(b) conversion to N03- and then leaching, adding 2H+ to the
jystem; or
(c) exchange reactions on soil colloids between NH4+ and soil
cations, resulting in cation stripping aM loss of
nutrients from the soil system.
What route incoming nitrate or ammonium will take depends on a
variety of factors, including: soil type (base saturation, cation ex-
change capacity and mineralogy); vegetation type, density and nutrient
condition; and water flow patterns.
One further impact of acid deposition on the nitrogen cycle is a
possible disruption of the rate (and quantity) of nitrogen cycled. Many
steps of the nitrogen cycle (i.e., nitrogen fixation, ammonific-'-ion,
nitrification and denitrification) are mediated by microorganisms known to
grow only in conditions above a pH of approximately 5.0. Acid conditions
are known to inhibit modulation of legumes (Andrew 1978; Evans et a_T.
1980), resulting in a decreased nitrogen fixation rate. Besides direct
injury to organism? involved, acid deposition could lead to a decreased
availability of molybdenum, which is an essential constituent of the
enzynes involved in nitrogen fixation, nitrification and denitrification.
Because nitrogen is such a critical plant nutrient, any disruption of the
nitrogen cycle could have severe consequences in terms of plant
productivity.
SULFUR CYCLING
Large quantities of sulfur are being added to terrestrial and aquatic
ecosystems via precipitation. It is important to determine what is the
fate of this added sulfur in terms of acidification. A brief review of
the sulfur cycle is in order.
The predominant form of sulfur in the soil system is organic sulfur,
generally in the S2- state or as sulfide, elemental sulfur, thiosulfate,
tetratliionate and sulfite (Reuss 1975). This sulfur is oxidized to
sulfate by autotrophic and heterotrophic microorganisms. Autotrophic
bacteria of the genus Thiobacillus are recognized as the most important
group of sulfur-oxidizing microorganisms. Five species predominate: T.
thioparus, T. denitrificans, T. thiooxidans, T. ferroxidans, and T.
novellus. T. ferrodoxins and T. novi?1lus are found in acid soils, even
dovn to pH 2.0. Regardless of the pathway or microorganism involved, the
oxidation of sulfides or sulfur results in the formation of H+:
S > 3/2 02 H20 - S0|" 2H* («q. 4)
H2S + 202 -" 2H* + SO^" (eq. 5)
Under anaerobic conditions ^'primarily deltas and flooded systems), sulfur
may be reduced to sulfides by members of the genera Desulfovibrio and
Desulfotomaculum. Iron sulfides often accumulate in such areas.
16
-------
PLANT
SOIL
anaerobic
Figure III-3. Simplified sulfur cycle illustrating acid-forming and acid-consuming processes. (From Reuss
1976.)
-------
A simplified sulfur cycle in Figure III-3 shows the acid-forming and
acid-consuming reactions (Reuss 1976). As shown, the sulfur cycle is
balanced and no net change in acidity sho-jld occur. (The 2H+ ions pro-
duced when organic S is oxidized to S0|- are neutralized by the 2 OH- ions
released when the plant incorporates the SO^-. The 2 OH- ions lost are
balanced when 2H+ ions are used to reduce S0|- to organic S in the plant
system.) However, because of lags between sulfate formation and plant
uptake, the H+ ions produced will displace basic cations from exchange
sites, and the cations plus S0|- will leach through the system. Sulfate.
reduction by anaerobic heterotrophs occurring in flooded areas results in
neutral or basic soils as H+ is consumed. Such soils will become acid if
they become aerobic .due to the oxidation of sulfides to sulfates. Sulfur
entering the soil system as S0,-S02, H2S03 or H2S04 will lead to the same
net increase in acidity (2H+) only if the sulfate produced is not imme-
diately absorbed by the plant system (i.e., if sulfur is in abundant
supply). If sulfur is in short supply, no acidification will occur
because the 2 OH- ions released will balance the 2H+ ions proJuced.
In spite of many possible reactions, much of the sulfate deposited in
acid precipitation is not retained in the Precambrian Shield watersheds,
thus functioning freely as an anion balancing the transport of H+ in
surface water and shallow groundwater. The amount of sulfate in runoff
from the Shield areas is very close to the amount deposited. At the
Experimental Lakes Area in Ontario, Schindler et al_. (1976) found
virtually 100% of the atmospheric S0|- input in the runoff. Likens e_t a_L
(1977) found 67% of the total input in runoff at Hubbard Brook, New
Hampshire, and Harvey et a 1. (1981) found 25% more sulfate in runoff from
four watersheds than was measured in the bulk deposition. This "excess"
sulfate may be dua to unmeasured inputs by dry and gaseous deposition.
Ultimately, excess sulfate can be leached through the poorly-buffered
-soils of these regions in association with cations released from exchange
sites. Sulfur taken up by plants (or S02 absorbed by foliage) is subse-
quently reduced to the sulfide state (primarily as amino acids) during
decomposition, thus returning to an initial state of the sulfur cycle
(Figure III-3).
ALKALINITY
One of the most important factors influencing the sensitivity of
lakes to acid input is alkalinity, or the capacity of the water to
neutralize acid. It is normally measured by titrating a water sample with
acid to a fixed end-point, depending on total inorganic carbon concentra-
tion. Carbonate and bicarbonate are the major contributors to alkalinity
in most natural waters, but other anions of weak acids (e.g., organic
anions) can also accept protons.
Lake alkalinity is largely a function of the type of rock and
biological activity (soil layr C02 production) in the drainage basin. If
carbonates are present, weathfring by acid will release bicarbonate into
the water as the major buffeiing agent. In this case alkalinity can be
high, and the lake will be insensitive to acid input. In the absence of
carbonate minerals, bicarbonate can be released by weathering of other
18
-------
minerals, but the quantities are slight. On granitic substrates, very
little alkalinity will be made available by weathering, and the main
sources are from carbonic acid, as controlled by C02 concentrations in the
soil (e.g., see Johnson et aj. 1977). Lakes in such areas will often be
very poorly buffered (Wright and Henriksen 1978, Glass and Louc.ks 1980).
There may be some alkalinity produced in the anaerobic hyoolimnion of
small lakes. Cook and Schindler (1980) report incrt'asintj bicarbonate and
ferrous iron during summer stratification due to the oxidation of organic
matter to C02 and the reduction of ferric iron. Hutchinson (1957) note* a
similar phenomenon in a number of lakes and explains that ferrous and
manganous bicarbonat.es may diffuse from the sediments and add noticeably
to alkalinity in softwater lakes.
Kramer (1976) has reviewed the relationship between pH and alkalinity
in lakes. Water in equilibrium with CaC03 has pH 8 and about 2 meq/£ of
alkalinity. The pH and alkalinity levels drop as this solution .is
diluted, but between pH 4 to 6, alkalinity remains nearly constant at
0.1-0.2 meq/£. Kramer attributes this buffering to alumino-silicates,
FeOOH and soluble organic acids. He concludes that low alkalinity lakes
in non-calcareous terrain are likely to be altered by acid rain, but that
detailed analysis of soil minerology at all depths is necessary to
accurately assess the risk.
ORGANIC INTERACTIONS
The role of organic materials in soils and water as they affect the
impact of acid deposition has been the subject of some speculation but
little quantitative study. This is a major oversight since organics may
influence almost all the reactions relating to acidification. Of course,
the formation of hurmc acid substances is one of the factors leading to
natural soil acidification. Acid deposition can be seen as an audition to
this process, so the role of organics deserves consideration.
One of the major characteristics of humic substances in soils is
their behavior in cation exchange. McFee e_t a_[. (1976) list the average
cation exchange capacity of humus as 200 meq/lOOg. This can serve as a
significant buffer for added H+ in precipitation. If base saturation is
high, the added H+ will exchange for cations absorbed to the organic
matter, such as Ca2+ and A13+. Organic soils with lower base saturation
are. already acid, and H+ in rainfall will have little additional effect
(Petersen 1980).
Another characteristic of organic materials is their ability to
chelate various ions. These ions rr.ay then either be retained in the soil
or leached out, depending on the behavior of the organic molecula.
Schnitzer (1980) notes that moderate acidity will decrease the solubility
of humic acids but increase the solubility of fulvic acids, which may then
be lost from the soil system with their absorbed ions. Organic chelation
and mobilization could be a significant factor in ion mobilization from
watershed soils, but the quantitative effects are unknown.
19
-------
Once into the aquatic system, dissociated organic acids may add--feb
the alkalinity of the s- ..tern by their ability to take up H+ (Chen et ai.
1978):
R-COOH = RCOO" + H+
This reaction may be especially important in waters where bicarbonate
buffering is low, and organic concentrations are high. The buffering by
this reaction will probably occur at pi! less than 5.
One final effect of organics relative to acidification problems is
the chelation of toxic metals. Baker and Schofield (1980) have observed
that aluminum chelated with organics is not toxic to fish. Similar
detoxification is possible with other metals mobilized by acid deposition,
e.g., Hg, Cd, Cu, etc.
Quantitative determination of the role of organics in lake/watershed
sensitivity to acid deposition will depend on the amount of organics
present, and the quantitative effect on the processes discussed above.
Various methods are available to measure organic content depending on the
medium: soil or water. Total organic matter in soil can be determined by
combustion of dried samples. Various organic fractions can be extracted
by treatment with acid or alkali as described by Schnitzer (1980). Both
humic and fulvic acids dissolve in alkali, then humics can be precipitated
by subsequent acid treatment. This allows separate determination of the
quantity of organic material in each fraction. Specific.organic molecules
or classes of molecules (e.g., emino acids) may be separated arid quanti-
fied by various techniques, including GC, GLC and spectrophotometry.
Organic matter dissolved in water can be determined by oxidation to
C02, followed by infrared absorption measurement of C02. This quantifies
total organic carbon (TOC). Natural waters can also be analyzed by UV
absorption as an indication of relative organic content, but this would
not be as accurate as TOC analysis. Other methods include color compari-
sons with standards and methods for specific compounds as indicated above.
Each method has advantages and disadvantages: TOC analysis is accurate
but time-consuming and non-specific, whil'i color comparisons are quick but
difficult to relate to actual organic content. The choice of a suitable
method depends on the relationship between quantity and effect, e.g., how
much organic matter and what type is necessary to cKlate and detoxify
aluminum at various concentrdtions.
CATION NUTRIENT STRIPPING
Nutrient stripping is one of the most serious consequences of acid-
deposition because loss of critical nutrients from the soil system may be
expressed in lowered growth rates and reduced productivity. Abrahamsen et
aj_. (1976), working with podsol and podsol-brown earth lysimeters, docu-
mented the effects of hydrogen ic concentration on nutrient leaching.
For the. podzol-brown earth lysimeters, net losses of Ca, Mg, Al and
sulfate occurred; whereas in the podsol lysimeters, net losses occurred
for Ca, Mg, Al, and K, (Figures III-4 and III-5). For all elements except
Al in tiie podsol-brown earth lysi.neters, increasing hydrogen ion concen-
tration resulted in increased leaching of nutrients.
20
-------
pHS.6
pH4
pH5.6
pH4
pH 3
PM5.6
pH4
pH3
C* Tiq/m'/vear Mg mq/m'/ysaf
500 1000 1500 2000 '00 200 300 400
111.! 1111
610
960 j
1410
K
50 100 150
1 1 1
2
25
70
I
NH4 +N0i) - N
100 200 300
1 1 1
350
345
355
400
|
100 j
1S5J
200
Al
10 20 30
1 1 1
19
,
17 )
1
" 21 1
SO,) (1975 oni»l
1000 2000 3000 9500
i 1 1 , 1 K t ' I
550 N
2440 j
— '
'335 » |
1
pH 5.6
pH 4
pH 3
26
50
1
37
100
1
150
1
142
Input
Output
Net
£0
300
Figure III-4. Nutrient budgets for pedzol-brown earth lysimaters treated
with artificial rain of pHs 3, 4, and 5.6. (From Abrahamsen
et al. 1976.)
21
-------
Si.
—I
SXG
I—I 1 1 i_J 1 L_/s/-J
- N
Al
J0i» I
Reproduced from
best available copy.
Figure III-5. Nutrient budgets for podzol lysimeters treated with simu-
lated rain of pHs 2, 3, 4 and 6. N.W. = not watered;
received incident'rain of Norway, pH 4.4. (From Abrahaiisen
et a_L 1976.)
22
-------
At Mt. Moosilauke, New Hampshire, Cronan (1980) found sulfate to be
the major anion in precipitation, throughfall, percolate and springwater;
aluminum the major cation in springwater; and hydrogen the major cation in
precipitation, throughfall and percolate. Potassium is readily leached
from the forest canopy (as is calcium) but is rapidly taken up by the
vegetation. Harvey et aj_. (1981) also note potassium retention by vege-
tation in Canadian watersheds. Calcium is also retained to a certain
extent in the soil system, although more than half of the throughfall
concentration is found in percolate and springwater (Table III-l).
Magnesium concentrations were highest in the springwater and lowest in
precipitation.
Harvey et ah (1981) have summarized nutrient budgets for several
Canadian watersheds (Table III-2). All watersheds studied have a net
output of major cations (ICa + Mg + Na + K) except for Clear Lake. The
ELA and Harp Lake results are of comparative interest (in terms of effects
of H ion input) as they both have similar bedrock geology and soil types.
The increased H inputs at Harp Lake have resulted in increased leaching
rates of Ca + Mg. Harvey et aj. note that on the basis of the limited
evidence available, it appears that the increased H+ input to southern
Ontario has resulted in a 2- to 4-fold increase in .net output of cations.
There are numerous problems associated with quantifying nutrient
leaching from the soil. These include (1) organic composition, content
and percentage in the soil (see Goring ?nd Jamaker 1972); (2) pH-dependent
CEC sites (Clark et aK 1956); (3) the buffering capacity of ions other
than Ca2+ and Mg2+ (DeVilliers and Jackson 1967, Turner and Clark 1965,
Clark and Turner 1965, Clark 1966); and (4) the toxicities of ions (Clark
1966, McCorroick and Steiner 1978, Cronon and Schofield 1979). The two
areas that have the most influence with regards to acid precipitation are
(1) organics and (2) pH-dependent CEC sites.
Currently there are only a few general equations that deal with the
complexity of nutrient leaching. The parameters for these equations are
molar ion concentrations denoted by "[ ]" and ion activities denoted by
"( )". Addiscott (1977) and Reuss (1978) have proposed simplified models
describing the dynamics of nutrient leaching. Reuss1 model will be out-
lined and discussed briefly.
Reuss utilizes (1) sulfate (SO?,-) absorption, (2) H+ and HC03 equi-
libria, (3) Ca2+ and H+ equilibria, (4) A13+ equilibrium, and (5) elec-
trical neutrality. This proposed model assumes that there is a low Ca2+
base saturation and a low exchangeable Ca2+/unit soil. The level of
SO2.-is an indicator of H+ input and can be expressed as:
SOl- (sorbed) = Km [SO;-] / K^ + [SO2,-] (eq. 6)
where K is the number of moles of S042-/unit soil at saturation and K, is
equal tS h K . The total S042- present can be expressed as: "*
SOf- (total) = SO2.- (sorbed) + S01--9-D (eq. 7)
where 6 is the volumetric moisture content and D is soil depth. Equations
(6) and (7) express the balance of SO2,- between the soil and the soil
23
-------
IABIE lll'l. Mean and median concentrations of dissolved ionic aaterials In bulk precipitation, canopy throtighfal I, forest floor anil A2 torizon perco-
late, and springwater during the 1975 and 1976 growing seasons at the Ht. Moosilauke subalpfne sampling sites. New Hampshire. Springs
also include 19/7 data. (From Cronan 1980.)
Sample pit
Bulk precipitation 4.08
Ihroughfal 1
mean 4.02
summed equivalents
ro
r*^
ned i an
Percolate
Tiean 4.04
sunned equivalents
median
Springs
mean 4.66
summed equivalents
median
C o n c e n
II* Ca'*
83 9
95 36
4
20
91 25
Z
22
22 26
0.5
25
t r a
Mg*'
3
14
1
8
15
1
15
19
0.6
18
t i 0
K'
3
37
3
28
16
2
10
10
0.3
11
n ( o i c r
Na NH^
4 13
3 6
0.2 0.5
3 S
7 5
0.3 0.7
5 3
13 3
0.7 0.7
13 3
o - e
Al
2
S
2
5
54
4
51
67
3
67
q u i v
Fe*'
1
1
0.4
1
4
0.3
4
2
1
2
a 1
Mn*'
0
4
0.7
4
2
0.2
2
1
0
1
e n t s per
Total
* Cations SOj'
118 75
201 143
11
117
219 137
3
132
163 132
2
131
trie
Cl"
7
13
1
11
16
1
13
7
0.3
7
r )
Total Aniun
NO, 110)3 Anions Deficit*
21 0 103 15
12 0 168 33
2 -
8 -
8 " 0 161 5S
1 - -
5 -
15 0 154 , 9
3 -
14 - - -
* The anion deficit is presumed to represent organic ligands; this has been confirmed on the basis of GLC and ultra-violet irradiation experiaents.
-------
TABLE III-2. Net export of major ions (meq nr2 yr-1) for calibrated watersheds in Canada (From Harvey et
al. 1981).
Watershed
Net Export (meq tn-2 yr-1)
Ca
^~r"- " v"""i '" ^—^ Input of H+
Mg++ Na* K+ ZM NH4+-N N03"-N HC03" S04-S (meq m-2 yr-1)
Carnation Creek,
Vancouver Is.
Jamieson Creek, B.C.
Haney, S.W. B.C.
12 watersheds, ELA,
Western Ontario
Rawson Lake Watershed,
Western Ontario
269 61.2 118.5 5.8 454.5 -7.7 -4.3 270.0 131.4
1-3
171.6 54.5 53.9 4.3 284
-0.2 -2.2 48.7
4.9
70.4 22.2 29.4 2.0 124 -8.8 -13.4 approx. 18.0 -v
200 (range 3.1-47)
*36.6 24.0 18.0 3.4 82.0 -0.9 -0.8
217.7 16.4 11.1 0.5 45.7
11.2 12.1 9.0 0.4 32.7 Total N reported
38.8 est. 7-10
0.9
•v.7-10
4 subwatersheds of Harp • 52.7 33.0 6.6 2.3 94.6 -30.2 -32.3 31.7 43.9 67
Lake; Hdi iburton-Muskoka,
Ontario
* Clear Lake, Haliburton -36.8 -25.3 1.3 -0.4 -61.2 NO
retention = 62. 1
1 Gross output
2 Estimated net output using input in precipitation from Rawson Lake watershed studies.
-------
solution. This balance can be interpreted as SO2,- retention or holding
capacity of the soil system which has obvious implications for buffering
capacity related to the sulfur cycle.
Distilled water in equilibrium with atmospheric C02 should have a pH
of 5.6._ Therefore, it is necessary to express the equilibria between H+
and HC03. This can be expressed as:
(H+) + (HCOa) = (C02)g • 10-7-81 • (eq. 8)
The balance between Ca2+ and H+ is an indicator of the buffering
capacity. Reuss assumes that the input of Mg2+ is insignificant in rela-
tion to the Ca2+ input. Therefore, the equilibria is expressed as
pH - \ pCa = KL (eq. 9)
which is simplified from Schofield and Taylor's (1955) "Lime Potential"
{ K. = pH - ^(Ca + Mg) }. Equation (9) can be modified to include the
exchangeable Ca2+, Ca2+ in solution and the CEC.
The A13+ equilibrium can be expressed as
(A13+) • (10-14 /(H+})3 = KA1 (eq. 10)
and is pH dependent
(DeVilliers and Jackson 1967).
There must also be an electrical balance between the soil and soil
solution. This balance is a function of molar ion concentration and ion
activity. The electrical balance can be expressed as
[H+] + 2[Ca+2] 3[A1+3] = 2[SO;2] + [HCOg] + [Cl-]. (Eq. 11)
It appears from the literature that Reuss' simulation model may be
applicable to acid precipitation-sensitive systems only; however, as Reuss
states, the model needs verification in the field and in the laboratory.
This model deals superficially with (1) the input/output and neutraliza-
tion of H+ and (2) the variability of the organics.
HYDROGEN ION TOXICITY
General Reactions of H+
Exchanges of H+ throuah the watershed depend largely on soil charac-
teristics, the balancing anions, and water flow patterns. So:Is with high
pH usually are well buffered by carbonates or have Ir'gh base exchange
capacity to balance anions, and have little flux of H+ or mobile anions to
surface waters. As soil pH, base saturation and/or cation exchange
capacity decrease, the likelihood that H+ and a mobile anion will pass
through in the soil water greatly increases (Wiklander 1980). In very
acid soils, almost no H+ will be retained by the soil. In general, pod-
zolic soils should be most susceptible to much alteration by acid inputs,
26
-------
(Petersen 1980), but exchangeable cations and water flow can modify the
results (Johnsen and Freedman 1980). Some podzols may contain appreciable
cations to exchange for H+, depending on inputs. In addition, water flow
through the system may occur too quickly at times for complete exchange to
occur. Bache (1980) points out that soil texture and intensity of rain-
fall are important factors determining the opportunity for H+ uptake by
soils. One further factor affecting H+ movement through watershed soils
is sulfate retention (Johnson 1980). Cation movement requires the move-
ment of balancing anions, so that if sulfate is retained ind other anions
are not mobilized, then H+ cannot be leached. Soils with a high per-
centage of sesquioxides are likely to be resistant to suIfuric acid
leaching.
The toxic effects of elevated H+ ion concentrations have been known
over a considerable period, for both terrestrial and aquatic ecosystems.
Acid soils are known to be infertile (Hewitt 1952), mainly because of
adverse concentrations of nutrients and elements (Andrew 1978), and the
known lower production rates of acid lakes (EIFAC 1969).
Because, at low pH, the H+ ions also may displace aluminum and heavy
metals from exchange sites, it is reasonable to assume that all three can
occur simultaneously, a*, least during peak mobilization events. Syner-
gistic effects of these three pollutants are thus a potential result of
acid deposition. Some research has been conducted in the area of
synergisms, but more is needed, especially with terrestrial organisms.
Terrestrial Effects.
Hewitt (1952) has summarized the factors affecting crop productivity
in acid soils:
(1) Direct injury to below-ground organs by hydrogen ions.
(2) Indirect effects of low pH:
(a) Physiologically impaired absorption of calcium, magnesium,
and phosphorous.
(b) Increased solubility, to a toxic extent, of aluminum,
manganese, heavy metals and possibly iron.
(c) Reduced availability of phosphorus partly by interaction
with aluminum or iron, possibly after absorption.
(d) Reduced availability of molybdenum.
(3) Effects from the low base status:
(a) Calcium deficiency.
(b) Deficiencies of magnesium, potassium or possibly sodium.
27
-------
(4) Induced abnormal biotic factors:
(a) Impaired nitrogen cycle.
(b) Impaired mycorrhizal activity.
(c) Increased attack by certain soil pathogens, e.g., "club
root".
(5) Accumulation of soil organic acids or other toxic compounds.
These same factors could affect forest and lake productivity, and many of
them could be induced by acid precipitation, depending on the buffering
capacity of the system.
Aquaeic Effects
The elevated H ion concentration in lakes where bicarbonate buf-
fering is negligible, or has been eliminated by acid inputs, is now being
linked with adverse effects on numerous fish species, and on phytoplankton
and zooplankton species. Hall and Likens (1980) have presented data sug-
gesting that the emergence of adult mayflies, stoneflies, caddis flies and
some true flies decrease in part due to lower pH, although elevated levels
of A13+, Ca2+, Mg2+, K, Mn2+, Fe, ard Ca2+ were present. Although many
parts of aquatic ecosystems are detrimentally affected by acidic condi-
tions, our discussion will focus on the effects on fish.
Laboratory experiments and field data are i.n general agreement con-
cerning effects of pH on fish (Table III-3), but only in cases where
toxicity is not complicated by the presence of ferric salts (EIFAC 1969).
Although factors ether than hydrogen ions often make it difficult to
identify the cause of the observed response, the development of a pH below
5 (or above 10) is considered to be unsafe for fish (Table III-4). Fromm
(1980) notes that a pH around 6.5 is the 'no effect1 level of pH effects
on fish reproduction. A pH between 5 and 9 is generally agreed as being
safe for fish. However, some species of fish can tolerate pH values lower
than those reported in the laboratory to be lethal, which suggests that
long-term acclimation is possible (EIFAC 1969).
A generally agreed observation in the literature is thnt early stages
of fish development are much more sensitive to acid stress than adult
stages. Thus, acid deposition in sensitive lakes characteristically
results in a failure to replace the older age rlasses; eventually the fish
population is eliminated.
Most researchers also agree that sodium imbalance is the dominant
cause of death, although Packer and Ounson (1970) contend that death is
attributable to lowered blood pH. The ameliorating effect of high levels
of Ca on toxicity at low pH is probably a result of reduced sodium loss
from fish (Wright and Snekvik 1978). Fromm (1980) has reviewed the
physiological and toxicological responses of freshwater fish to acid
stress. 1.v'er& is some question whether H+ ions alone are toxic at the
observed concentrations, but the development of low pH is a major factor
in lake sensitivity.
28
-------
TABLE III-3. LITERATURE SURVEY ON pH TOXICITY TO VARIOUS FISH SPECIES.
ro
10
fish species
Salmon and Trout
Salmon and I rout
Salmon
Salmon
Salmon
Salmon
S.i I mon •
Atldiillc Salmon
Atlantic Salmon
Atlantic Salmon
Atlantic Salmon
Atlantic Salmon
Trout
Trout
Trout
Trout
Splake
(brook x Take trout)
Lake trout
(SalveIinus namaycush)
pll level
5.5
5.3 & lower
5.0
5.0-5.5
4.5
4.59
4.34
4.3
4.49
5.8-6.2
4.0-5.0
4.0-5.2
>5.0
4.5
4.7-5.4
5.1-5.7
4.49
5.5-5.2
Effect and Comments
"critical to young"
"highly lethal"
Lethal to SOX of eyed eggs
Critical range for hatching
LOf,-, in 12 days to yolk sac fry
96% Hatching success
4BX Hatching IUCCOK*
lUfto °r embryos and alevlns
5G~ rortalHy of yolk sac fry in i9.0 days
Lethal pll In 2-day test; possible
CO, interaction
Delayed hatching
Matching prevented
High and probably total lethjlHy
Critical range for hatching
80* yolk sac died in 20 days
10% yolk sac died in 20 days
50% mortality of yolk sac in 16.0 days
Approximate pll when reproduction ceased
Reference
Sunde 1926 as cited llagen & Langeland 19/3
Sur.de 1926 as cited llagen & Ljngeland 1973
Johansson et aj. 1977
Bua & Snekvik 1972 as cited Hagen & Langeland 1973
Oahl 1927 as cited EIFAC 1969
M. Grande personal comm. as cltod EIFAC 1%9
Oaye and Cars(de 19/9
Grande et aj. 1976
Bishai 1960 as cited EIFAC 1969
Peterson et a_l. 1900
Peterson et a]. 1980
Dahl 1926 ar cited llagen & Langeland 1973
Oua and Snekvik 1972 as cited llagen & Langeland 1973
Oahl 1927 as cited EIFAC 1969
Dahl 1927 as cited EIFAC 1969
Grande et aj. 1978
Beamish 1976
-------
FABLE 111-3. (continued)
Fish species
Lake trout
Rainbow trout
Rainbow trout
Rainbow trout
Ralnb'-i. \rout
Rainbow t -out
Rainbow trout
Rainbow trout
Brown trout
(Salrco trutla)
Brown trout
Brown trout
Brown trout
Brown trout
Brown trout
Brown trout
(Avaa strain)
Brown trout
(Avaa strain)
Brown trout
(R. Dalai ven strain)
Brown trout
pll level
4.49
4.49
1 age x
1 age x
1* age x
5.5
4.5
4.2
4.4-5.2
5.0
4.49
1 age x
1 age x
1 age x
4.4
4.0
4.5
5.0
= 6.10
= 5.59
= 4.98
= 6.20
= 5.50
= 4.77
(R. QJ la I veil strain)
Effect and Comments
50% mortality of yolk sac In ->>1.3 days
50% mortality of yolk sac fry In •>•]..3 days
3.5 no. exposure; 3% mortality
3.5 00. exposure; 4% mortal Ity
3.5 ma. exposure; 7% mortality
Approx. lower Halts of tolerance
LDSO to finger)Ings in 15 days
8 day LD60
100% mortality of yolk sac fry due to ron-
bined low pll and low salt concentration
Approx. lower Molts of tclerance
50% mortality of yolk sac fry In MO.O days
3.5 no. exposure; 5% mortality
3.5 mo. exposure; 2% mortality
3.5 mo. exposure; 6% mortality
66% of eyed eggs survived yolk-sac stage
Lethal to 90% of eyed eggs
Lethal to 100% of eyed eggs
Lethal to 20% of eyed eggs
Reference
Grande et al. i<"'
Grande et al. 1978
Edwards A fljeldnes 1977
Edwards & Hjeldnes 1977
Edwards & Hjeldnes 1977
Berlins I960 as cited Johansson et aj. 1977
Lloyd & Jordan 1964 as cited EIFAC 1969
Lloyd & Jordan 1964 as cited EITAC 1969
Grande & Anderson 1979
Berlins 1960 as cited Johansson et a_l. 1977
Grande et al. 1978
Edwards & Hjeldnes 1977
Edwards & Hjeldnes 1977
Edwards & Hjeldnes 1977
Johansson et aj. 1977
Johansson et aj. 1977
Johansson et al. 1977
Johansson et al. 1977
-------
CO
1ADIE II1-3. (continued)
f ish species
Brown trout
Brown trout 4.77
Brown truut <5.0
Adult Brook Trout 4.5
Adult Brook (rout S.08
3rook Trout 5.09
Brook Trout 5.57
Brook Trout • 6.13
Brook Trout 6.55
Brook Trout (Control) 7.04
Brook trout 4.8
Brook trout 4.5
Brook trout 4.49
Sea trout 5.8-6.2
Perch 4.5 and 5.0
Perch 4.4-4.9
Yellow perch 4.7-4.5
(Pcrea fiavescens)
flavescens
-------
IABIE 111-3. (coullmied)
F ish species
Umbra llmi
Walleye
(Slizostedion vitreum)
Iroulperch
(Percopsis oroiscomaycus)
Arctic char
Arctic char
Arctic char
to Arctic ctidr
ro
Arctic char
Char
pjl leyej
4.2
6.0*-5.5
5.5-5.2
1 age x -
\ age x =
1 ay
-------
FABLE 1II-3. (continued)
fish species
pll level
Smal 1 mouth bass
(Hicrupterus dolumieui)
Ruck bass
(Ainhloplf tes r.upestrls)
Centrarchids
lake Chub
(Coues Ins plumbeus)
Cyprinids
Burbot
'Ljt.a lota)
Bur-bo t
Ourbot
Lake herring
(Coregonus arredli)
Brown bul Ihead
(Ictalurus nebulostis)
Roach
Roach
Pike
Cisco
Eel
6.0 -5.5
5.2-4. 7
4.G-5.0
4.7-4.5
>5.0
6.0*-5.5
6.0
5.0
4.7-4.5
5.2-4.7
4.2
S5.5
4.4-4.9
<5.0
•v4.5
effect and Comments
Approx. pll when reproduction ceased
Approx. pll when reproduction ceased
Healthy populations found in Wisconsin
Lake Survey
Approx. pll when reproduction ceased
Only found when pll 5.0
Approx. pH when reproduction ceased
Critical lower level for embryo
segmentation
Critical lower level for post-embryo
development
Approximate pll when reproduction ceased
Approx. pll when reproduction ceased
8 day LDSO
Critical pll for reproduction
Critical pll for reproduction
Critical pll for reproduction
Critical pll for reproduction
Reference
Beamish 1976
Beamish ?976
Rahel A Haijnuson 1980
Beamish J976
Rahel & Haymtson 1980
Beamish 1976
Volodin 1960 as cited E1FAC 1969
Votodln 1960 as cited EIFAC 1969
Beanish '"-6
Beamish 1975
Lloyd & Jordan 1964 as cited flfAC 1969
Alner et aj. 1978
Aimer et al. 1978
Aimer et aj. 1978
Alner el al. 1978
-------
TABLE III-4. SUMMARY OF THE EFFECT OF pH VALUES ON FISH (From EIFAC 1969).
Range Iff feet
3.0 - 3.5 Unlikely that any fish can survive for more than a few hours in this range
although some plants and invertebrates can be found at pH values lower than
this.
3.5 - 4.0 This range is lethal to salmonids. There is evidence that roach, tench,
perch and pike can survive in this range, presumably after a period of
acclimation to slightly higher, non-lat-hal levels, but the lower end of
this range may still be lethal for roach.
4.0 - 4.5 Likely to be harmful to salmonids, tench, bream, roach, goldfish and common
carp which have not previously been acclimated to low pH values, although
the resistance to this pH range increases with the size and age of the
fish. Fish can become acclimated to these levels, -but of perch, bream,
roach and pike, only pike may bs able to breed.
4.5 - 5.0 Likely to be harmful to the eggs and fry of salmonids and, in the lor.g
term, persistence of these values will be detrimental to such fisheries.
Can be harmful to common carp.
5.0 - 6.0 Unlikely to be harmful to any species unless either the concentration of
free carbon dioxide is greater than 20 ppm or the water contains iron salts
which are precipitated as ferric hydroxide, the toxicity of which is not
known.
6.0 - 6.5 Unlikely to be harmful to fish unless free carbon dioxide is present in
excess of 100 ppm.
6.5 - 9.0 Harmless to fish, although the toxicity of other poisons may be affected by
changes within this range.
9.0 - 9.5 Likely to be harmful to salmonids and perch if present for a considerable
length of time.
9.5 - 10.0 Lethal to salmonids over a prolonged period of time, but can be withstood
for short periods. May be harmful to development stages of some species.
10.0 - 10.5 Can be withstood by roach and salmonids for short periods but lethal ove- a
prolonged period.
10.5 - 11.0 Rapidly lethal to salmonids. Prolonged exposure to the upper limit of this
range is lethal .:o carp, tench, goldfish and pike.
11.0 - 11.5 Rapidly lethal to all species of fish.
Reference is made to different species on the basis of information known to us; the
absence of a reference indicates only that insufficient data exist.
34
-------
ALUMINUM MOBILIZATION AND TOXICITY
General Reactions of Aluminum
One of the indirect effects of watershed acidification may be the
mobilization of aluminum in soils, causing toxic effects on terrestrial
organisms and the biota of streams and lakes. Aluminum solubility is pH
dependent, and increase;' with increasing acidity (Figure III-6). Several
reports have documented elevated aluminum concentrations in acid lakes and
streams in areas known to be impacted by acid inputs (Figure II1-7, Davis
1980; Cronan and Schofield 1979; Schofield 1980), and in effluents from
lysimeters treated with acid solution (Dickson 1978; Abrahamsen et a1.
1976). While aluminum is typically leached from the upper soil horizon of
podsol soils by carbonic acid and organic chelation, it is usually
deposited in lower horizons. Under the influence of -'strong acids in
precipitation, .however, the aluminum may be transported through the soil
and leached into lakes and streams (Hall and Likens 1980; Herrmann and
Baron 1980). Peak aluminum concentrations generally occur during spring
melt of the sr.owpack (Figure III-8) when large quantities of H+ ions,
accumulated over winter, are released and flush aluminum from the soil
system (Schofield 1980). As illustrated in Figure III-8, the large
majority of impurities are released during the first stages of snowmelt.
Seip et a_L 1380 note that interactions of meltwater with soil and vege-
tation are important in determining the amounts of aluminum leached to
receiving waters.
Current precipitation in the northeast United States, Norway, and
Sweden has pH levels between 3.0 - 4:0 and lower (Likens et aj. 1979,
Wright et aj_. 1980, Wright and Dovland 1977). Magistad's (1925) data and
more recent data (Baker and Schofield 1980, Clark 1966, Cronan and Scho-
field 1979, Dalai 1975, Driscoll 1980) show that the pH of incoming
precipitation is well within the range of aluminum mobilization resulting
in toxic levels (as identified in the laboratory).
The mechanism supplying A13+ is the decomposition of alumino-
silicates and gibbsite (Norton 1976, Reuss 1976)
A12 Si2 05 (OH)4)<1 + 6H* = 2A13+ + 2H4Si02+ H20
A1(OH)3 + H+ = A1(OH)2'1' + H20
AKOH),"1" + H* = Al(OH)2"1" + H20
Al(OH)2* + H+ = Al3* + H20
This is likely to occur in watersheds where there are no carbonates to
consume H+, and the above reactions become a primary buffering mechanism
(Johnson iD79, Kramer 1976). The pH at which this buffering occurs is
around 4.5-5.0 as indicated by the Al solubility diagram of Dalai (1975),
which shows that A1(OH)3 begins to decline rapidly at pH 5 as Al3-*- begins
to increase. The other soluble Al species, A1(OH)2+, -.A1(OH)2+ and
Al6(OH)153-t-, never exceed 10% of total aluminum.
35
-------
4.0
.5.0
CD
O
6.0
2.0 3.0 4.0 5.0 6.0 7.0 8.0 9.0 10.0 11.0 12.0
pH
Figure III-6. Solubility of aluminum as affected by pH. Note increasing
solubility with increasing acidity. (From Andren et al.
1980.)
36
-------
1000
_ 500
^
at
a
z
200
g 100
z
o
u
i so
20
10
4.0
5.0
6.0
PH
7.0
8.0
Figura III-7. Aluminum concentrations found in Norwegian and Swedish
clear water lakes. (From Davis 1980.)
37
-------
150
5 so
o
o
Mini-catchment 5
ISO
100
SO
5 (
10 IS
%
.'i i
; \ ,
. « '
' i
i i
i \^
; N> '<•
•
/ u
20 25 X 5 X) 15 20
t
i
t
i
20 •
10
20 25
April
JO 5
59
D IS
May
20
Figure III-8. Nutrient release from snowpack during 1978 spring melt as
determined by monitoring runoff from mini catchments in
Norway. Note majority of contaminants released simul-
taneously at initial melt. Shaded area in lower graph
corresponds to 30% of the runoff during the period. Dashed
and solid lines to right of top figure represent average
concentrations of S04 and H+, respectively, during summer
and autumn 1973. (From Seip e_t al_. 1980.) . .-.•-••*.
-------
In aquatic systems, aluminum forms a variety of complexes with water,
hydroxide, fluoride, silicate, organic matter and sulfate (Everhart and
Freeman 1973, Oriscoll 1980, Baker and Schofield 1980). It is rarely
found as the free aluminum ion. According to Johannessen (1980), these
complexes act as a buffer in the pH range 4..5-5.0, but above and below
these levels buffering cannot be ascribed to aluminum complexes. Henrik-
sen (1980) shows that lakes with pH 4.6 - 4.8 are less acid than expected
from a theoretical "titration" curve based on bicarbonate buffering, and
that the extra buffering can be explained by aluminum. He postulates that
mineral acid addition to lakes will cause them to change from a carbonic
acid-bicarbonate buffered system to one buffered by strong acid-aluminum.
In surface waters of the Adirondack Region of New York, Driscoll
(1980) found aluminum-organic complexes as the predominant monomeric form
(Avg. = 44%), which increased linearly with total organic carbon content.
Aluminum-flouride complexes were the most abundant inorganic forn. 'avg. =
29% of the total monomeric Al), with their concentration increasing with
decreasing pH, although their formation was generally limited by fluoride
concentration. Aluminum sulfate species were of relative unimportance,
although they did increase with decreasing pH. Everhart and Freeman
(1973) note that the solubility of aluminum is a direct function cf pH in
the vase majority of natural waters (i.e., under conditions where hy-
droxide complexes dominate). In acid environments the soluble forms are
cationic and polymeric. Overall, pH controls complexation, polymeriza-
tion, hydrolysis and solubility.
Aluminum also indirectly affects availability of heavy metals and
phosphorus. Heavy metal availability increases because humic substances,
to which heavy metals are normally bound, are instead complexed and preci-
pitated by aluminum (Davis 1980). Aluminum also precipitates phosphorus,
as aluminum phosphate, thus decreasing its availability to aquatic biota.
Terrestrial Effects
The aluminum mobilized by strong acid deposition is of most concern
due to its toxic effect on terrestrial and aquatic organisms. Much re-
search on aluminum toxicity has been conducted on crops and legume species
(Black 1968, Andrew 1978, Hewitt 1952, Foy et aj_. 1978). From these
reports the following statements can be made:
(11 Excess aluminum affects cell division in roots,
causing inhibition of root growth leading to stubby
and brittle roots. The best indicator of aluminum
toxicity in the field is abnormal root development.
The tips and lateral roots become thickened and turn
brown, and no fine branching roots develop. Al is
often found in root cortex cells, being especially
concentrated in nuclei. It generally accumulates in
roots and is only found in above-ground plant parts of
a few plant species. Germinating plants and young
seedlings are generally more susceptible than older
, plants.
39
-------
(2) Excess aluminum fixes phosphorous in less available
forms in the soil and in or on plant roots. Conse-
quently, phosphorous deficiency is often associated
with toxic aluminum concentrations and is generally.
the form of symptom expression in above-ground plant
parts. Al has also been shown *.o interfere with
uptake, transport and use of other nutrients (Ca, Mg
and K) and water. Thus, Al toxicity can appear as Ca,
Mg or K deficiency. In cotton, excess aluminum also
decreased uptake of Mn, Fe, Ni and B, some of which
must be due to a decreased absorptive surface.
(3) Excess Al also decreases root respiration, interferes
with certain enzymes governing the deposition of
polysaccharides in cell walls, and increases cell-wall
rigidity (by cross-linking pectins). All of these
effects of aluminum lead to decreased growth and occur
below a pH of 5.0, although effects have been noted at
pH 5.5. Andrew (1978) also notes that excess Al is
1 detrimental to nodule initiation in legumes, the
efficiency of the nitrogen-fixing symbiosis and plant
growth. He also notes that low Al concentrations for
short periods of time can increase growth of some
plant species.
Various plant species have quite different tolerance levels of Al.
Many factors affect tolerance/susceptibility, including: pH around the
root zone (tolerant plants can produce a higher pH around root zones);
NH4+ vs. N03- nutrition; aluminum exclusion processes; calcium nutrition;
phosphorous nutrition; and organic-aluminum complexes. All of these
factors are discussed in detail by Foy et a_L 1978.
Species known to be sensitive to aluminum include: barley, sugar
beet, corn and alfalfa. Decreased growth of barley occurred at < 1 ppm;
for corn, 4-17 ppm Al was toxic; and for alfalfa 2.7 ppm was toxic (Black
1968). In experiments with alfalfa grown in nutrient solutions, Munns
(1965) found decreases (from controls containing 0.008 ppm Al) in shoot
dry weight of 70.3% and 83% at aluminum concentrations of 1.08 ppm and 5.4
ppm, respectively. The first observed symptom was inhibition of root
elongation and lateral formation, and plants with high aluminum concen-
tration showed characteristic phosphate deficiency. Tolerant crop species
include oats, most brassicas (i.e., cauliflower, narrowstem Kale, and
swede), rye, rice, soybeans, azaleas, cranberry, and triticale.
For forestry applications, McCormick and Steiner (1978) tested the
effect cf aluminum on root elongation in six tree genera. A hybrid poplar
was most sensitive, with complete inhibition at less than 10 ppm Al.
Autumn olive was sensitive at 10-40 ppm, while the other genera (Quercus,
Pinus, AljTus, Betula) did not show effects until the concentration was
80-100 p;m. However, concentrations greater than 4 ppm Al are rare in
soil solutions (Andrew 1978).
Manganese toxicity also is associated with aluminum toxicity (both
occur simultaneously, i.e., under acid conditions). Low concentrations of
40
-------
1-4 rng/2 affect sensitive plants such as lespedeza, soybeans and barley,
although corn tolerates >15 mg/£ and Oeschampsia flexuosa-' tolerates >60
mg/£. (Black 1968). Tolerance is usually attributed to reduced absorption,
less translocation of excess Mn to plant tops, and/or greater tolerance to
high Mn levels within plant tissues.
Aquatic Effects
Graphs such as those in Figures III-6 to III-8 are important in
establishing the relationship between pH and aluminum solubility, but they
do not adequately describe peak aluminum mobilization events, nor do they
describe the effects of continual input of H+ ions. Peak concentrations
in lakes are of utmost importance biologically because they may occur in
conjunction with spawning and hatching and are thus most likely to cause
adverse effects on fish recruitment (Gjessing et al. 1976). Peak aluminum
concentrations also will occur in conjunccion with peak hydrogen ion
concentrations (Figure III-9) which may increase the overall effect on
fish and other biota (see Schofielc! 1980; and pH and synergisms, this
report).
Figure III-9 graphs the relationship between hydrogen ion loading
rate and aluminum concentration, and indicates the effacts large inputs of
hydrogen ions would have on aluminum concentrations in lakes. The points
labeled N are from lysimeter studies in which complexation with organic
matter increased aluminum solubility. Caution should be used when inter-
preting the data point 0, as the full data set was not available for its
calculation. It was assumed that 100 kg-H2S04/hectare'year resulted in 1
mg Al/£ effluent (see Oickson 1978). Data used in Figure III-9 are from
Abrahamsen and Stuanes (unpublished data), Crisman and Brezonik (1980),
Dickson (1978), Cronan and Schofield (1979), Likens et aJL (1977), and
Wright et aj_. (1978).
The scatter in Figure III-9 indicates that other variables influence
Al mobilization. Factors such as soil type (clays retain aluminum),
amount of organic matter present, base saturation, cation exchange
capacity and absolute aluminum concentration will affect a watershed's
response to acid inputs. These factors emphasize the need for more data
from various watershed types before conclusions can be made regarding the
hydiogen ion loading rate that will result in toxic aluminum concentra-
tions.
Graphs such as Figure III-9 could be used to determine the hydrogen
ion input resulting in potentially toxic aluminum concentrations to lake
and stream biota, but aluminum toxicity is not that simple, as it is
affected by a variety of chemical and biological factors. For example,
the points labeled N would suggest that elevated (and hence potentially
toxic) aluminum concentrations can occur with relatively small inputs of
hydrogen ions. However, these points are a result of increased r.olubility
due to complexation with organic matter, and aluminum is less toxic when
complexed to organic matter. Other substances act as aluminum ligands and
have the same effect (see previous discussion).
Wright and Snekvik (1978) observed that aluminum caused no deleter-
ious effects at concentrations ranging from 0.05 - 0.3 mg/2, found in 90%
41
-------
40
35
- 30
^v.
o>
"o
I 25
e"
i 20
u
§ 15
o
10
N
N
N
Las2FN
S.4
H3"
I
B,
M
B
?3
HB
B
o
50 100
H+Loading meq/m2/yr
150
200
Figure III-9. Relationship between H+ loading rate and observed aluminum
concentrations in lysimeter effluent, minicatchments or
freshwater lakes. Data from a variety of sources cited in
the text.
42
-------
of the 700 lakes which they surveyed. Their results indicate that fish
st-^'is was correlated most with. Ca and pH, with barren lakes having low pH
levels and low Ca concentrations. They note that acid stress to fish is
apparently more acute in waters with very low ionic strength. They also
found that Ca ameliorates the effects of low pH, with more fish surviving
low pH with high Ca-concentration than low pH with low.Ca concentrations.
Nutrient stripping of soils leads to increases in Ca2+ concentrations in
lakes which could ameliorate the toxic effects of high aluminum and hydro-
gen ion concentrations.
Some research has been completed on the physiological effects of
elevated aluminum concentrations on fisn. Muniz and Leivestad. (1980)
noted a rapid loss of plasma Na and Cl at toxic Al levels, and also large
amounts of mucus clogging gills, resulting in hyperventilation and-
coughing. Venous oxygen tension was decreased in fish with clogged gills.
The toxic effect was thus a combination of impaired ion exchange and
respiratory distress brought on by mucus clogging of gills. Cronan and
Schofield (1979) observed that mortality from acute exposure to aluminum
appeared to result from severe necrosis of the gill epithelium.
Estimating peak aluminum concentrations and toxicity to fish with
sufficient accuracy is not possible from current data. The relationship
between H+ ion input and aluminum concentration graphed in Figure III-9
requires further data for validation. Data on aluminum toxicity (both
chronic and shock effects) to various fish species, -*nd under various
physical and chemical situations, also are needed. Numerous factors --
pH, ionic strength, calcium concentration, phosphates, organic matter,
nitrate, fluoride, sulfate and silicates -- all affect aluminum toxicity.
Relationships between these factors and aluminum toxicity need to be
identified before "safe" aluminum concentrations can be set. Based on the
data presented in Tables III-4 and III-5, highly tentative ranges are
proposed as follows: non-toxic concentrations would be below 0.05 mg/£,
potentially toxic concentrations would be between 0.05 and 0.3 mg/£, and
toxic concentrations would be greater than 0.3 mg/£. These ranges should
only be applied to species already tested, trout and salmon, and should
not be applied to other fish species.
H+ and Synergisms with A13+
Thtre has been some work on responses of fish to pH in conjunrtion
with other variables. McLeay et al. (1979) have investigated the toxic
effects on salmonid fish of paper pulp effluent as a function of pH.
Their results suggest that as pH decreases, mortality increases. They
found 50% to 67% mortality at a pH below 7.0, which was the normal pH of
the receiving water. Baird e_t a_L (1979) have investigated the toxicity
of ammonia as a function of pH.
Aluminum has been found to lower the toxicity of pH to fish, i.e.,
low pH and some aluminum are more toxic than low pH and no aluminum (Davis
1980, Dickson 1978, and Baker and Schofield 1980). Davis (1980) notes
that for brook trout, pH 4.9 was not lethal but that pH 4.9 and 1.0 mg
Al/2 were toxic to 50% of the population. At pH 4.0, 1.0 mg Al/2 was less
toxic, although pH 4.0 and 4.4 were toxic by themselves (Figure 111-10).
Schofield (1980) also found less aluminum toxicity at pH 4.0 than at 4.4
43
-------
TABLE III-5. SUMMARY OF LITERATURE CONCERNING ALUMINUM TOXICITY TO FISH.
Species p_H
Stickleback
Sa 1 mo salar 5.0
Al
0.07 mg/£
0.2 mg/S.
Commt its
Lethal
Toxic within
6 days
Reference
Dickson 1978
Dickson 1978
Brook trout
4.4-5.9
0.2 mg/S. Toxic response Cronan & Schofiel
1979
Brook trout
0.1-0.3 mg/S. Growth reduction Cronan & Schofiel
1979
Brook trout 5.2
(Salvelinus fontinalis M.)
0.42 mg/S.
28% survival
after 2 weeks
(mean of 4
replicates)
Driscol 1 et a_L
1980
Brook trout
4.4
0.48 mg/je
42% survival
after 2 weeks
(mean of 4
replicates)
Driscol 1 et aj_.
1980
Brook trout
4.9
Brown trout 4.3, 4.5,
(Salmo trutta) 5.0 & 5.3
1.0 mg/£
0.2 mg/2
LDso
Mortality
occurred
Davis 1980 (work
of Schofield?)
Muniz & Leivestad
1980
Brown trout 4.0 & 6.0
(Salmo trutta)
0.2 mg/i.
No effects
Muniz & Leivestad
1980
Salmon
near 5.0
0.2 mg/2
Very toxic
Grahn 1980
Ciscoe 5.0-5.5
(Coregonus albula)
max 0.5 mg/S, Large-scale fish Grahn 1980
kills in 2 lakes
(majority <5. 0)
4.0-7.0
,05-. 3 mg/2.
Not deleterious
in 700 Norwe-
gian lakes
Wright &
Snekvik 1978
44
-------
I I I I I I I I I
4 6 8 10
DAYS OF EXPOSURE
12
14
Figure 111-10. Effect of altering pH on cumulative mortality of brook
trout exposed to three concentrations of aluminum. (From
Davis 1980.)
45
-------
and higher. His results (Table III-6) conclusively show that pH and
aluminum do interact in producing fish toxicity. Cronan and Schofield
(1979) found that brook trout showed a toxic response to 0.2 mg Al/2. (and
above) in the pH range of 4.4-5.9. Everhart and Freeman (1973) .-and
Oriscoll et a_K (1980) also note that aluminum toxicity is pH dependent.
Grahn (1980) reports that 0.2 mg Al/2 in water near pH 5.0 is very toxic
to salmon. He also notes that two fish kills of Ciscoe, Coregonus albula,
were associated with low pH (5.0-5.5) and maximum aluminum concentrations
of 0.5 mg Al/2. Oickson (1978) relates that 0.2 to 0.6 mg Al/2 in asso-
ciation with a low pH (4-5) is toxic to many fish species. Muniz and
Leivestad (1980), working with brown trout, Salmo trutta, and aluminum
concentrations of 0.2-0.8 mg/2, recorded mortality at pH 4.3, 4.5, 5.0 and
5.3, but no effects at pH 4.0 and 6.0. Maximum toxicity was at pH 5.0
where mortality occurred in 18 hours.
No work has been done en the combined effects of acid pH's and heavy
metals, or of aluminum and heavy metals, on terrestrial and aquatic or-
ganisms. Synergisms between zinc and copper on fish have been reported by
Bandt (1946) and Doudoroff (1952) (both cited by Lloyd 1961), and
synergism between nickel and copper on decomposing organisms has been
reported by Freedman and Hutchinson (1980)- Lloyd noted a possible
synergistic effect of zinc and copper sulfates on rainbow trout in soft-
water. Czuba and Ormrod (1974) noted synergistic response in lettuce and
cress when ozone (03) was added in conjunction with Cd or Zn, and similar
interactions were noted with S02 and Ni or Cu.
HEAVY METAL MOBILIZATION AND TOXICITY
General Soil Reactions
The toxicity of heavy metals to biota was", at one time, assumed to be
determined by the total metal concentration in the system (Mancy and Allen
1977). It is now realized that this assumption does not hold, and that
toxicity of heavy metals is determined by their particular chemical
species. Research must center on determining what species of heavy metals
are toxic and at what concentrations; what physical, chemical and
biological factors determine the speciation of heavy metals; and the flux
rates of heavy metals from the atmosphere to soil to ground water.
Sources of heavy metals fall into two categories: natural and man-
made. Natural sources include weathering of mineral bedrock, deposition
of marine salts, erosion, volcanism, and forest fires. Man-made sources
include industrial and mining wastes, fossil fuel combustion, atomic
testing, fertilizers, pesticides, manures, and sewage-'s-tudge. To deter-
mine'transport rates between ecosystem compartments; accurate determina-
tions of the quantity of heavy metals originating from these sources
should be calculated, something which has not beer, done for source inputs
to the atmosphere (Goldberg 1975). Goldberg recommends that more research
be conducted on the vapor phase of heavy metals, and on vertical and
horizontal concentrations of heavy metals, before calculating transport
rates from the atmosphere.
46
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Table III-6. TIMES TO 50% MORTALITY FOR FISH IN TRIALS A-F. PERCENT SURVIVAL IS
GIVEN IN THOSE TRIALS IN WHICH 50% MORTALITY WAS NOT REACHED AFTER 14
TO 23 DAYS EXPOSURE. NOTE LACK OF EFFECT OF INCREASING ALUMINUM CON-
CENTRATIONS ON MORTALITY AT pH 4.0. (From Schofield 1980.)
Time (days)
Trial
A
B
C
D
E
F
Nominal Aluminum
pH (mq/2)
4.0 0.0
0.1
0.5
1.0
4.4 0.0
0.1
0.5
1.0 •
4.9 0.0
0.1
0.5
1.0
4.9 0.25
0.5
5.2 0.0
0.1
0.25
0.5
4.9 0.0
0.32
0.65
0.63**
Repl icate
A
4.1
5.4
1.9
4.8
9.7
11.2
1.5
3.7
100%
95%
2.1
1.7
90% (80%)*
3.1
90%
100%
65%
1.6
100%
95%
4.7
11.9
Repl icate
B
4.4
4.8
2.6
5.4
10.0
11.2
1.8
3.3
100%
95%
2.1
1.6
90% (70%)*
3.1
90%
95%
70%
1.7
90%
75%
4.1
5.4
Replicate
C
5.1
5.4
4.0
4.9
11.4
12.0
2.9
3.9
100%
95%
2.8
2.1
90% (75%)*
3.8
100%
90%
55%
1.5
100%
100%
5.3
4.5
Mean
4.5
5.2
2.8
5.0
10.4
11.5
2.1
3.6
100%
95%
2.3
1.8
90% (75%)*
3.3
93%
95%
63%
1.6
97%
90%
4.7
7.3
* days survival
** Brook water
47
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Our knowledge of chemistry of metals in soils is incomplete and often
speculative (Goldberg 1S75). However, numerous factors are known to
affect availability, toxicity, ar* mobility of heavy netals (Ratsch 1974).
These factors include (1) relative abundance, (2) form, (3) soil pH,
(4) interactions with other element1;, (5) physical conditions of the soil,
(6) temperature, and (7) soil moisture. Generally, heavy metals do not
react to the same conditions in the same manner, so generalizations from
one metal to another are not recommenced. However, an increase in H+ ions
in precipitation may decrease the ab'iity of the -i"il system to retain
heavy metal ions (Tyler 1972, Abrahamsen an^ Dollard 1979). In the soil
system, heavy metals are found as (1) minerals of soil structures, (2) in-
organic precipitates, (3) soluble and insoluble organic complexes or che-
lates, and (6) absorbed ions on charged surfaces of clays, precipitates
and organic matter (Goldberg 1975).
Heavy metals tend to accumulate in soils (VanHook et al. 1977, and
Tyler 1972) and generally do so by exchange reactions, whereby organic
matter of the soil complex binds heavy metals forming stable complexes
(Abrahamsen and Dollard 1979, Tyler 1972). Accumulation occurs in/on dead
organic matter, litter and humus, with concentrations increasing with age
and extent of decomposition (Tyler 1972, Van Hook et aj_. 1977). Humic and
fulvic acids are believed to play a prominent role in accumulation and
cycling of trace metals in soils and sediments (Nriagu and Coker 1980).
In aquatic systems, heavy-metal speciation is subject to change by
biological, physical and chemical factors (Wollast et a_[. 1975, Elder
1978, Chakoumakos et aj_. 1979). These factors include: pH, water hard-
ness, alkalinity, organic matter contant, nutrient content, and oxidation-
reduction potential (Goldberg 1975, Gambrell et a_L 1980, Elder 1978,
Chakoumakos et a_L 1979). In freshwater systems'"treavy metals can exist
in three forms: readily available (biologically), potentially available,
and essentially unavailable. Processes affecting the availability of
toxic metals include: (1) precipitation as insoluble sulfides under
highly reduced conditions; (2) formation of metal cxiaes and hydroxides of
less solubility; (3) absorption to colloidal hydrc-js oxides of iron and
manganese (primarily under aerobic, neutral or alkaline conditions); and
(4) complex formation with soluble and insoluble organic matter under all
conditions of pH and oxidation potential (Gambrell et a_L 1980). Accord-
ing to Allen et al_. (1980), most studies suggest that free metal ions are
the most toxic form, and that the stronger the complex, tne lower the
•;oxicity.
The following summaries illustrate how various factors interact with
specific heavy metals.
Mercury: Increased levels of soluble mercurv were found by
Gambrel et a_l_. (1980) in moderately acid, reduced conditions (pH
5.0, -15CmV) and weakly alkaline, oxidized conditions (pH 8.0,
+500mV). Wollast et aj. (1975) note that mercury is highly soluble
only in well-oxygenated water, and that under moderately oxidizing
conditions the .predominant species of mercury is undissociated.
They report that under more reducing conditions extremely insoluble
cinnabar (HgS) precipitates, but that under very reducing conditions
mercury may increase in solubility. Dissolved mineral mercury can
48
-------
be present as Hg(OH)2 in the aerobic zone and in elemental form (or
as organically complexed oxidized species) in oxidizing conditions.
Mercury also exists in several organic forms, e.g., methyl mercury,
which are products of bacterial activity.
Zinc: In freshwaters zinc has been found to occur in both
ionic and colloidal inorganic forms (Allen et a]_. 1980). pH was
found to affect the levels of total dissolved and exchangeable zinc,
with highest concentrations found at the most acid pH employed, 5.0
(Gambrell et a_L 1980). At that pH it is most likely to be found in
the free cationic form. Soluble zinc decreases under reducing
conditions.
Lead and Cadmium: Florence (1977), as cited in Allen et a_l_.
(1980), has shown that in freshwaters, lead is present to an equal
degree in both inorganic and organic forms, while cadmium is present
as the free metal ion. Gambrell et a_L (1980) found little dis-
solved lead at any pH-redox combination (pH 5.0 to 8.0; redox -150mV
to +500 mV). They also found that exchangeable lead (readily
available) was more influenced by pH than redox potential (higher
concentrations found at pH 5.0 than 6.5 and 8.0), whereas reducible
lead (metal oxides and hydroxides) was more influenced by oxidation
strength than pH (higher levels found at +500 mV than -150, +50 or
+250mV).
Copper: In fresh waters copper is predominantly associated
with organic colloidal matter (Allen et al_. 1980), and is often
bound to humic acids (Nriagu and Coker 1980). Elder (1978) studied
copper speciation in two alkaline freshwater lakes in California, pH
7.7 - 9.0. Based on the copper concentration and pH of a lake, he
constructed a model of expected speciation of copper in an aqueous
solution. Below pH 5, all of the copper is free Cu2+. He also
discusses the fact that all copper in the lake eventual'iy reaches
the sediments. Under anoxic conditions, there is a low redox poten-
tial and a release of hydrogen sulfide from sediments. Copper com-
plexes with sulfur to form CuS, which permanently sequesters copper
in sediments.
Terrestrial Effects
As indicated above, the soil system is capable of accumulating heavy
metals in potentially toxic concentrations, depending on the level of
inputs and mobilization reactions. The most common symptoms of heavy
metal toxicity on crops and other plants are chlorosis and stunting.
Chlorosis is generally caused by direct or indirect heavy metal inter-
actions with foliar Fe (Foy et a_L 1978). Oats have often been used as
indicator plants for Ni toxicity because of their unique chlorotic and
necrotic reaction. Stunting can be caused by specific metal toxicity,
antagonism with other nutrients, or inhibition of root penetration.
Toxicity is first expressed in root tips, and lateral root development can
be severely restricted. If root development is altered, this will also
affect normal nutrient uptake, which could lead to decreased productivity.
49
-------
Aquatic,. Effects
Once in the aquatic environment, heavy-metal chemistry becomes very
complex, as many factors interact to determine the chemical species found.
Because of this, extrapolating experimental results from one aquatic
system to another can lead to error; i.e., "safe" concentrations in one
system may not be "safe" in another. Even so, water quality criteria aro
set, with a single standard for all freshwater systems. The following
brief review of the literature shows the variability in heavy metal
toxicity, but also gives an idea of the concentrations which are "safe" to
freshwater fish. Unfortunately, most of the experiments were conducted
under neutral to alkaline conditions and thus do not reflect toxicities
under acid conditions.
Copper: Factors known to affect copper toxicity include pH,
hardness, alkalinity, and inorganic and organic complexes. Cha-
koumakos et aj. (1979) studied the effects of alkalinity, water
hardness, and pH on copper toxicity to Cutthroat trout. There was
an inverse relationship between acute toxicity and water hardness
and alkalinity. Copper was more toxic in soft water than hard
.water. Toxic species, of copper were Cu2+, CuOH+, Cu(HO)2° and
Cu(OH2)2+ while CuHC03, CuC03° and Cu(C03)|- were not toxic. Lett
et ajL (1976) studied copper toxicity on rainbow trout. They deter-
mined a 96-h LCSO of 0.25-0.68 mg Cu/£ for hard water (365 mg
CaC03/£) and pH 7.8-8.2. Beamish (1976) notes that 17 ug Cu/£ did
not affect survival, growth, or reproduction of adult rainbow trout
in soft water. Sauter et al_. (1976) recommend that the safe concen-
tration proposed in 1973 by the National Academy of Sciences (not to
exceed 0.1 of the 96 h-LC50 for the species of interest) should be
changed, based on the results of their study.
Zinc: Spehar (1976) determined a 96-h LCSO of 1500 pg Zn/£ to
juvenile (4 to 5-week-old) flagfish and an estimated MATC (maximum
acceptable toxicant concentration) of 26-51 ug/£ in Lake Superior
water. He found survival of larvae and growth of females to be the
most sensitive measures of toxicity. Lloyd (1960) notes that
rainbow trout are capable of acclimation to lethal concentrations of
zinc if they are first exposed to sublethal concentrations. Ball
(1967a) studied zinc toxicity to four freshwater species and found
five-day LCso's for each: rainbow trout 4.6.mg/£; bream 14.3 mg/£;
perch 16.0 mg/£; and roach (two strains) 17.3 mg/£. Cairns and
Scheier (1957) as .cited by Mount (1966) found that bluegills were
killed in soft water by 1.93-3.78 ppm Zn, whereas in hard water tne
-Values were 10.13-12.15 ppm.
Cadmium: For juvenile (4 to 5-week-old) flagfish in Lake
Superior water, a concentration of 2500 ug Cu/2 was toxic to 50% of
the population in 96 hours (Spehar 1976). Spehar determined an MATC
of 4.1-8.1 ug Cu/£ with the latter concentration inhibiting repro-
.duction on a chronic basis. Spawning and embryo production were the
most sensitive measures of cadmium effect. Ball (1967b) estimated a
7-day LC50 of 0.4 mg Cd/£ for rainbow trout, but he cites
Schweiger's (1957) 4 mg Cd/£ as lethal to rainbow trout in seven
days, and 3 mg/2 as a safe dose. Sauter et al. (1976) conclude that
50
-------
the National Academy of Sciences' recommended safe concentration of
0.004 mg Cd/£ for water with hardness 100 mg CaC03/£ is barely
adequate; and that for water with hardness 100 mg CaC03/£, a recom-
mended safe conce.itration of 0.03 mg Cd/£ is totally inadequate to
protect aquatic life.
Lead: Davies and Everhart (1973) found the following toxici-
ties for rainbow trout: in hard water (243 |jg CaC03/£) 96-h LC50
for total lead was 471 mg/£, but for dissolved lead it was 1.32
mg/£; in soft water (26.4 mg CaC03/£) the 96-h LC50 for dissolved
lead was 140 ug/£. They estimated MATC's for lead: hard water and
total lead (0.12-0.36 mg/£); hard water and dissolved lead (0.018 to
0.032 mg/£); and for softwater dissolved or total lead (6.0-11.9
mg/£). Pickering and Henderson (1965) as cited by Davies and
Everhart (1973) working with hard water (300-360 mg CaC03/£) and
s,oft water (18-20 mg/£), determined the following 96-h LC50's for
lead: fathead minnows in soft water - 5.58 mg/£; in hard water -
482 mg/£, for bluegills in soft water - 25.8 mg/£, in hard water -
442 mg/£. Sauter et aj. (1976) conclude that the National Academy
of Sciences' recommended maximum allowable concentration of 0.03
(jg Pb/£ appears adequate to protect most fish species.
Heavy-metal toxicity in soils and freshwater ecosystems is very
complex, with numerous factors (physical, chemical and biological) inter-
acting in complex ways to determine the chemical speciation, and hence
mobility and toxicity, of heavy metals. All heavy metals do not react in
the same manner to the same set of conditions, so extrapolation from one
heavy metal to another is not reliable. More research is needed to deter-
mine the residence times of heavy metals in various ecosystem components,
and to find the extent to which heavy metals are recycled in terrestrial
ecosystems. Finally, *he chemical species of heavy metals which are
toxic, and those that -»re not, need to be determined. In determining
toxicities, experiment should be conducted using soil and organisms from
the terrestrial site in question.
51
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SECTION IV
METHODOLOGIES FOR QUANTIFYING THE SENSITIVITY
OF TERRESTRIAL AND AQUATIC AREAS
Orie Loucks, Richard Miller, Roland Usher and William Swanson
The exposure of forests to oxidants and the addition of acidic pol-
lutants to a lake/watershed ecosystem both alter the producing system, but
in very different (although potentially interactive) ways. The oxidant
effects (mainly 03) on forest growth have been reviewed by the authors in
the recently completed ORBES study (Loucks et al_. 1980) and in a recent
symposium (Miller 1980). No additional review will be added here. The
aggregate effect of the acidic inputs to the ecosystem, however, may be
thought of as a large-scale titration (Henriksen 1980), but it is a
complex, uneven process only superficially similar to a laboratory titra-
tion. Gorham and McFee (1980) and Last et a_l_. (1980) discuss the
variables involved and note that the hydrogen ions deposited with precipi-
tation (or generated by other components of precipitation) may have
several fates in the lake/watershed system, as reviewed in the previous
section.
Various consequences follow. First, the mobilization of elements by
cation displacement can cause a long-term loss of soil fertility and lower
productivity. Even if there is an initial fertilization effect from the
added nitrogen in precipitation, continued acid input may result in future
nutrient limitation through the loss of essential cations. Secondly, the
mobilization of aluminum and heavy metals by cation displacement and
weathering can result in toxic effects for both the aquatic and terres-
trial components of the system. Finally, unneutralized H+ may itself be
toxic and a stress to the system.
The combined effects of these above-ground and below-ground
processes, as they influence the more highly integrative measures of
response, are discussed in the following sections from the viewpoint of
data needs in measuring forest, lake or watershed sensitivity.
MEASURES OF SENSITIVITY TO FOREST PRODUCTIVITY RESPONSES
The yields of forests usually are expressed as wood volumes, and
growth rates are expressed as cubic feet (or meters) per unit area, per
year. These volumes are affected by the annual diameter growth rate, but,
since total volume is greatly infuenced by the height of the cone (and
diameter increment is intercorrelated with annual height growth), height
growth is a very frequent measure of forest productivity (Forbes 1955).
Oxidants and acid rain, however, are relatively recent incursions on
forest processes (Miller 1980), and, therefore, most studies have reported
changes in foliate condition and Jiameter growth of mature trees -- not in
young, developing stands. Many other factors, including insects and
disease problems, interact with the mature tree, however, suggesting that
the height response of younger stands would have more general application.
53 Preceding oaee blank
-------
Various measures of soil texture and soil nutrient status have been
correlated with forest growth rates, and (as reviewed in the previous
section) these suggest various measurements linking a forest response to
the agents of chemical alteration in the soil, i.e., acidic deposition.
However, height and diameter growth of the fores't represents an "inte-
grator" of soil conditions, including, potentially, the alteration of
nutrient status due to cation stripping or aluminum toxicity. These
growth measures, however, are also the "integrative" measures that would
respond to toxic gaseous pollutant inputs to the forest canopy — S02 as
well as 03.
Of the height and diameter measures used in forestry, the most
generalizable is Site Index, a basic measure expressing rate of height
growth. The Site Index (SI) concept has been accepted routinely in
forestry as the best measure of current or potential forest productivity
because rates of height growth are readily converted to cubic volumes of
wood. The SI is most commonly expressed as the average height of dominant
and co-dominant trees at the age of 50 years. It is used in forestry to
express the rzce of productivity for a single species in a specified site,
or mapped aggregate of sites. Although SI usually is specified by height
and age measurements, it can be estimated indirectly by using soil and
topographical factors (Hannah 1968, Mader 1963), height and age (Spurr
1952), available nutrients (Ellerbe and Smith 1963, Gordon 1964, Heilnian
1968) and available soil water and nutrients (Fralish and Loucks 1975,
Mogren and Dolph 1972).
A review of the literature since 1960 shov~. that the primary vari-
ables associated with SI are the available soil water and topography
(Mogren and Dolph 1972, Fralish and Loucks 1975, Hannah 1968). The
importance of soil type and structure has been discussed (Hannah 1968,
Richards and Stone 1964, Linnartz 1963) and, although the literature
appears to be divided (Broadfoot 1969, Hannah 1968), one of the most
important underlying factors affected by site quality is the rooting
pattern of the trees under investigation. This directly correlates with
the available soil water and nutrients in solution, which, in turn, are a
function of precipitation and water use.
To determine site index directly the height and age of the dominant
and co-dominant trees must be known (Curtis £t a_[. 1974). Plotting height
versus age (Figure IV-1) of dominant stems in a series of stands results
in a collection of points which can be broken down into a family of
curves, each representing a separate site class. Thus, simply stated, the
site index is the height of the stand projected through time, either
forward or backward. This projection can only be accomplished through the
series of harmonic curves illustrated in Figure IV-1. In the United
States, 50 and 100 years have been chosen as reference points for the
expression of site index.
There have been a number of investigations relating the role of
nutrients to the SI (Gordon 1964, Heilman 1968). Mogren and Dolph (1972)
observed the influence of pH on the SI and concluded that there was no
effect. It should be noted that they did not report any pH values in the
report. A more recent preliminary literature search (1965 to date) has
shown that chronic and acute pH changes as they affect the SI have not
been investigated.
54
-------
120
UJ
LU
cc
100
z
<
z
o
o
u.
o
t-
z
g
UJ
z
UJ
O
<
CC
LU
20
10
20
30 40
AGE,YEARS
60
Figure IV-1. Site index classes determined by plotting age against
average height of dominant tree.
55
-------
Use of the forest site index as a measure of terrestrial ecosystem
stress from acidic <.1eposition is an attempt to integrate the effects of
nutrient additions (such as N and S), nutrient losses such as Ca, Mg, K
and Na, and the toxic effects of aluminum, H+ and heavy metals (Drablos
and Toll an 1980).
There are several liabilities to this approach. The first relates to
age classes. The site index can only be applied if the average site
quality is the same for each age class. This liability is important in
managed forest sites (Spurr 1956) due to many practices that are used to
increase yields. A second liability stems from the assumption that the
shape of the height/age curve is consistent between sites. This generali-
zation is capable of reliable results; however, it does not hold for all
soils (Figure IV-2). A third liability, related to the second, is that
site differences are apparent at an early age. In other words, the har-
monized curves that predict a taller tree at 50 years assume thac this
tree has always, been taller. To alleviate the influences of these lia-
bilities, the standard site index curve should be based on specific soil
types and field measurements of species growth on that site.
The site index is typically used in forest management to determine
harvest schedules for even-aged, monocultural stands. This relationship
severely limits the uses of the site index from an ecological viewpoint.
A mixed-species and uneven-aged stand presents many fundamental problems
in determining a site index curve. To alleviate these problems, Curtis
and Post (1964) have proposed a "Composite Site Index" to express site
quality in mixed hardwood stands. Their concept uses the- measured site-
index of each species, each of which is converted by means of equations
(Curtis and Post 1962) to an estimated maple site index. All the esti-
mated maple site indices are added to the measured maple site index and
then averaged, which results in the composite site index. This provides a
more reliable measure of site quality for an unmanaged, even-aged site.
Curtis £t a_L (1974) have presented a detailed review article on the
importance of the site index and how best to interpret the series of
curves obtained.
The effects of acid precipitation on a watershed/lake system are many
and intricately balanced, especially in a mature system. As has been
outlined above, the site index may be an appropriate tool to measure the
effects on, or changes in, site quality due to acid precipitation. Use of
the site index as an integrator of acid deposition effects (via nutrient
stripping and Al toxicity) has the advantage of being readily convertible
to economic terms for cost-benefit analysis. At this time, however, there
is not enough data to validate its use.
MEASURES OF AQUATIC ECOSYSTEM SENSITIVITY
At the present time there is no single, fully validated methodology
for estimating the sensitivity of aquatic systems to acid deposition.
Thus, measures of aquatic sensitivity to acid inputs should be thought of
as hypotheses constructed with a reasonable amount of quantitative inputs,
but evaluation should not be based on any one model or measurement proce-
dure. All of the methodologies given belr.w are limited in scope and/or
56
-------
en
-C
O)
Age, years
Figure IV-2. Theoretical height/age curves for'three soils: (1) homo-
geneous soil, (b)' shallow soil, (c) poor surface soil with
richer lower horizons (redrawn from Spurr 1964).
57
-------
availability of data, but taken together they present a comprehensive view
of our ability to quantify current resource status and, therefore, some
prediction of prospective future change.
Predicting Toxicity to Aquatic Organ'!sms
Determining toxic effects of acid precipitation on aquatic organisms
requires that standard methods be identified. Such standard methods exist
for determining general toxicity to aquatic organisms, and can be applied
to the toxicity of acid precipitation. Bioassay tests, conducted under
natural conditions and employing standard procedures, are a recommended
method.
Toy/!city is determined by dosage - the amount (in relation to body
weight or growth medium) of toxicant and the frequency and duration of
exposure (Gough et al. 1979). The term is used by a wide variety of
professionals, including doctors, pharmacists, microbiologists, and regu-
latory personnel.
Although chemical analysis and standardized tables would be the
easiest way to determine toxicities, they have not proven to be feasible
for many reason:-: (1) the specific organism in question often has not
been tested; (2) different bodies of wat^r have different characteristics
which greatly affect toxicity of many substances, and (3) there are dif-
ferences in temperature, fish size, seasonal lolerance of various fish
species, testing procedures, physiology, previous history, nutritional
state, and certain laboratory conditions which may have affec;ed the
results (Katz 1971; American Public Health Association 1965).
It is now recognized that bioassay with organisms of the indigenous
population under natural conditions is the most satisfactory means of
determining toxicities of substances (Peltier 1978, Geckler et al_. 1976,
American Public Health Association 1965, Katz 1971). Ideally, the tests
should employ a range of concentrations of toxicants, and should be con-
ducted at different seasons of the year to reflect differing water
quality, which has been shown to greatly affect the toxicity of sub-
stances, especially heavy metals (Geckler et al. 1976, U.S. EPA 1976, Katz
1971).
The earliest recommendations for standard methods were published by
Hart et al_. (1945) and Doudoroff et al_. (1951), and have largely been
adopted by the American Public Health Association (1965). Other documents
are available which outline standard procedures for determining toxicity
of elements to aquatic organisms: U.S. EPA (1972); Peltier (1978); and
The Committee on Methods for Toxicity Tests with Aquatic Organisms (1975).
Numerous review articles have been published on the measurement of pollu-
tant toxicity to fish (Sprague 1969, 1970 and 1971; Herbert 1965; Edwards
and Brown 1967; Alderdice 1967; Burdick 1967; and Warner 1967).
Acute .toxicity to fish is most frequently measured by the concentra-
tion of toxicant that is lethal to 50% of the population within a speci-
fied time period (usually 96 hours). Various terms are used, e.g., Median
Lethal Dose (LD50), Median Lethal Concentration (LC50), Median Tolerance
Limit (TLm), Median Effective Dose (ED50), and Median Effective Concer.tra-
58
-------
tion (EC5o), although LC50 and EC50 are preferred. The latter also can
refer to lethal or sublethal responses. When used to determine sublethal
responses, EC5(j is defined as the concentration causing an "adverse
effect" in 50 percent of the population within a specified time period,
whert "adverse effect" is generally defined as immobilization.
Three procedures for determining LC5o's and EC50's are discussed by
Sprague (1969) from which the following is largely taken. For all three,
tests are conducted using a series of concentrations of the toxicant, with
mortality being recorded at various times. The results are plotted on
log-probability paper, as shown in Figure IV-3 (from Sprague 1969). From
such a graph, one can determine median survival times (when acute mortal-
ity ceases), and the presence of differing modes of action. Changes in
slope or grouping of lines are clues to differing modes of action. The
next step in determining the LC50 is to plot all the median lethal times
(obtained from graphs similar to Figure IV-1) against concentration to
find a lethal threshold concentration.
A second procedure for determining LCso's is described in Standard
Methods (American Public Health Association 1965) and by the Doudoroff
Committee (1951). It yields LCSo's for specified time periods, usually 1,
2 or 4 days. For each fixed time, percentage mortalities are plotted
against test concentrations, and the concentration lethal to 50% of the
population is then estimated by interpolation. Sprague (1969) suggests
that more frequent observations be made and that investigators should
construct time-concentration toxicity curves as described above.
No reports specifically outline standard procedures for determining
chronic toxicity. This underscores the fact that chronic toxicity tests
have only recently gained the interest of toxicologists (Cough et al.
1979). Long-term studies using a wide range of concentrations of a parti-
cular toxicant on large numbers of aquatic organisms, which determine the
overall community response, are needed by regulatory agencies (U.S. EPA
1976). Geckler et aJL (1976) suggest that tests other than those on
survival, growth, and reproduction be conducted; specifically, that tests
based on behavioral responses may better predict effects of toxicants on
natural systems. They concluded from their tests with copper, conducted
over a 33-month period, that organisms avoided the highest" concentrations
of copper. Results of chronic tests can be expressed either in terms of
the lifetime of one organism or the time span of more than one generation.
Chronic effects often influence the species population and not just
individuals.
Sprague (1970, 1971) and Mount and Stephan (1967) discuss various
methods of determining sublethal or "safe" concentrations, including the
use of toxic units and application factors. The latter procedure involves
taking the LC50 and applying some factor (generally .1) to determine the
"safe" concentration.
In conclusion, the following recommendations can be made for mea-
suring effects of acid precipitation on lake/watershed ecosystems:
(1) That additional standard bioassay tests determining LC5o's
and EC50's' be used when toxicity tests are conducted;
59
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90
3 70
Q
~ 50
4)
Q.
10
30
26.5
10 50 100 500 1000
Time Of Observation, min.
5000
Figure IV-3. Results of a mortality test at various temperatures plotted
on logarithmic-probit paper. The results indicate two
modes of 'lethal action at 27.5°C. (From Sprague 1959.)
60
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(2) That the results of these toxicity tests be presented as a
range, indicating the variable effect that water quality
can have on the toxicity of many substances, particularly
heavy metals;
(3) That.extrapolation f results of toxicity tests from one
body of water to another, and from one species to another,
be avoided except where absolutely necessary. It should be
noted that extreme variability can exist from one aquatic
system to another, from one species to another, and from
one season of the year to another.
The Calcite Saturation Index (CSI)
Kramer (1976) has summarized the relationship between pH and alka-
linity in the calcite saturation index (CSI) which expresses the relative
saturation of water with CaC03.
CSI = p(Ca2+) + p(Alk) - p(H+) * pK,
where p(X) = -log10X, pK = 2, (Ca2+) is given as mol/£ and (Alk) and (H+)
are given as eq/2. When CSI < 1, the lake is nearly saturated with CaC03
and not susceptible to acidification. Increasing CSI indicates increasing
susceptibility, and, at CSI > 4, the lake is very likely to be affected by
acid (Glass and Loucks 1980).
The CSI has been used to classify lakes sensitive to acidification
(Glass and Loucks 1980, and others) and is a relatively simple way to
Account for lake buffering capacity. However, it is not comprehensive
'enough to be used as a sensitivity index as defined earlier for watersheds
because buffering capacity is only one aspect of lake/watershed
sensitivity. It is necessary to account explicitly for other possible
consequences of the acid input, including the mobilization if aluminum and
the loss of soil fertility by cation stripping. Moreover, the CSI
reflects only present conditions and does not relate past or future
changes to loading rates. There is no indication in CSI of how a lake
came to have a particular value, or of any explicit relationship between
the CSI and acidic inputs.
The Dickson Relation
One way to summarize the relationship between lake pH and atmospheric
loading of acidic pollutants is to plot lake pH in relation to measured
local sulfate loadings, as attributed to Dickson (Aimer e_t a_[. 1978),
Figure IV-4. Various combinations of watershed characteristics are
reflected in the family of curves connecting the different sets of lake
data. A larger number of curves may be envisioned, each representing
lake/watsrshed systems of comparable sensitivities as they apparently
respond to increases in atmospheric acid inputs. The curves are super-
ficially similar to titration curves, as might be expected, and each shows
buffering at high pH, then a rapid drop (as this-buffering apparently is
depleted) to lakes with somewhat lower pH and minimal buffering. The pH
in the more sensitive lakes drops sooner and more rapidly, relative co the
loading of SO^-.
51
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X
a
0 30 60 90
Sulfate Loading to Lake Water [Kg/ha-yrl
Figure IV-4. Effects of various sulfate loading rates on lake pH for
lakes in very sensitive (1) and somewhat less sensitive (2)
surroundings in Sweden. From Glass and Loucks 1980. Added
points are for: • Florida (Crisman and Brezonik 1980);
8 Como Creek (Lewis and Grant 1979); A Hubbard Brook
(Likens e_t al. 1977); and X Norway (Wright and Snekvik
1978).
62
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A graph of sulfate deposition versus lake pH could be used to express
lake/watershed sensitivity, and therefore acid loading tolerance, by
evaluating the position of the lake/watershed system on the response
curves. For example, the lakes of southern Norway are underlain by
granites and felsic gneisses (Wright and Snekvik 1978), and almost all
prove to be somewhat sensitive. From a knowledge of the response of very
sensitive geochemical/biological systems in moderately impacted systems,
it appears possible to anticipate similar responses in the future on
similarly sensitive systems if acidic inputs increase or are sustained.
Examination of Figure IV-4 suggests that annual sulfate loadings of
less than 15 tc 17 kg/ha would be unlikely to degrade lakes of the type
represented in curve (1). However, if the lower envelope of the data
distribution is viewed as a potential "family" of the most sensitive lakes
and streams, these appear to be just barely free of potential acid load-
ings effects at an annual rate of 9 to 12 kg/ha«yr. Thus, two "toler-
ances" can be defined, one associated with a possible protection of nearly
all sensitive aquatic resources, and the other with protection of some-
thing less than all sensitive lakes, e.g., only the half of the "sensi-
tive" resources that lie above curve (1) in Figure IV-4.
Since these curves were developed on the basis of Swedish data, and
represent a north-south geographic gradient in S0|- loadings, and are not
actual observations of acidification responses over time, these estimates
of acid loading tolerances must be viewed very cautiously for application
in a North American context.
The Henriksen Nomograph
Henriksen (1980) presents a model based on the concepts implicit in
titration of a bicarbonate-buffered lake with strong acid (principally
H2S04) from the atmosphere. In the process, bicarbonate is depleted and
lake pH can fall below 5 with consequent effects on aluminum mobilization
and fish. These relations are basically the same as those in the Dickson
work, and are summarized in the nomograph (Figure IV-5) using two key
measurements:
(i) ambient concentrations of in-lake calcium (or Ca+Mg) as an
estimator of the pre-acidification alkalinity; and
(ii) lake sulfate concentrations (in excess of marine input) as an
estimator of H+ added to the system.
The resulting semi-predictive nomograph is divided into three
sections:
(i) bicarbonate lakes, where original alkalinity was high and/or
added H+ is low, uo that the lakes remain bicarbonate buffered;
(ii) acid lakes, where original alkalinity was low relative to acid
irputs, and all bicarbonate appears to have been depleted by
the acid addition; and
63
-------
300-
5 200
(Q
o
(ft
O
X
UJ
100-r
250 •-
--150
--100
0°
O
z
--50
25 50 100 150 200
Excess Sulphate in Lakewater, ueq/l
4.7
—4 j 1_
4.4 4.3 4.1
pH of Precipitation
4.0
Figure IV-5. A nomograph to predict the pH of lakes given the-sum of
non-marine calcium and magnesium concentrations (or non-
marine calcium concentration alone) and the non-marine
sal fate concentration in lake water (or the weighted-
average hydrogen ion concentration in precipitation).
Nomograph also can be used to predict future changes in
lake pH when precipitation acidity increases. (From
Henriksen 1980. )
64
-------
(iii) transition lakes, in which bicarbonate appears to be undergoing
reduction (or is almost depleted) end large pH fluctuations
occur during runoff events.
The transition phase, in which the lake is shifting from a
bicarbonate-buffered equilibrium at moderate pH to an aluminum-buffered
equilibrium at low pH, represents the key process requiring prediction.
This shift apparently is forced by H+ and S05j- inputs, and Henriksen
presents regression equations based on Norwegian lake and precipitation
data for representing sulfate in lake water (S04*) in terms of sulfate in
precipitation [S04(p)] and H+ in precipitation [H-t-(p)]:
S04* = -19 f 1.9 S04(p)
S04(p) = -2.7 + 1.37 H+(p)
All concentrations are ueq/£, and the sulfate is excess over that from
marine origin. These equations have not yet been validated for North-
America.
The methodological component that can be applied to determine a
threshold loadings tolerance' must u^e assumptions as to the processes
maintaining steady state lake pH in various types of aquatic resources,
ranging from very sensitive to only moderately sensitive. Maintaining a
pH of 5.3 would, presumably, prevent lakes from entering the "transition
phase". This method., however, does not consider the depression of lake pH
due to spring snowmelt, although the results (based on summer data)
suggest the spring depression is being taken into consideration.
Episode Receptor Dose/Response Relation During Shock Events
A possible alternative model for describing and projecting aquatic
resource response to acid loading emphasizes the non-equilibrium nature of
the acidification process, equivalent to Henriksen's "transition phase".
During this phase, aquatic ecosystems can show some buffering and a rela-
tively unaltered pH during much of the year, but they also may experience
sudden pH drops during hydrologic flushing events, i.e., at times when
accumulated acidic material in the watershed is flushed relatively
suddenly into streams and lakes. The rapid increases in H+ concentrations
can also mobilize A13+ for a brief period, and the shock event can become
a multi-agent toxic stress affecting the survival of fish and other
organisms at various life stages, depending on the timing and magnitude of
the pH depression.
Research on brook trout and Atlantic salmon by Daye (1980) and Daye
and Garside (1975, 1976, 1977, 1980a, 1980b), and related research by
Beamish (1974, 1976) and Harvey (1975, 1979, 1980a, 1980b) have provided a
broad understanding of the response of several pH-sensitive fish species
to both long-term and short-term elevated H+ exposures. Mortalities of
fish eggs, sac fry and adult fish have been viewed as a response to
continuing chronic pH depression. At the seme time, effects on egg via-
bility, hatching1 success, and adult survival are known to occur as a
response to short-interval acute H+ and A13+ exposures. The experimental
data b-,se supporting these relationships can be generalized to support the
following statements (Andrews e_t aj_. 1980; see previous section).
65
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(1) The short-term acute exposure, or shock effect, can be expected
when pH drops by a range of 0.5 to 1.5 units of the pH scale in
a background of pH 5.5. to 6.5; and
(2) These shock exposures may be significant at a pH above the level
at which chronic effects ordinarily would be produced.
Taken together, these data suggest that for waters normally in a
range of pH 5.5 to 6.5, a pH depression (A pH) of 0.5 to 1.0 can cause
substantial, and physiologically significant, acid-induced alteration of
water chemistry. This relatively large increase in H+ during shock events
may be occurring only in regions where acidic deposition already has
produced a downward time-trend in stream pH and alkalinity, as shown in
several monitoring programs. Data from very sensitive regions relatively
free of long-term acid inputs (or in the first to second decade of moder-
ate deposition) indicate it is unlikely that a short-term pH depression of
these magnitudes can occur with an unusual watershed stimulus.
Given this dose/response relationship., a procedure for defining an
acid-loading tolerance, or loading threshold, also can be suggested. The
annual S0|- loading which, when subjected to a defined flushing event
(e.g., snowmelt or first m-jor rainfall following drought), leads to the
minimal biologically-significant short-term H+ and A13+ exposure, as
determined by controlled environment studies.
The data available on pH depression during flushing events indicate a
range in responses from very little to as much as 1.0 unit of the pH scale
during snowmelt in northern MinnecJta, a region of recent and only moder-
ate acid deposition (Glass et a_L 1981a). Depression of pH is shown to be
in the order of 1.0 unit on the Shaver's Fork River in West Virginia
(Dunshie 1979, Figure IV-6), to more than 2.0 units in the Adirondacks
(Schofield 1980, Figure IV-7). Values in this range also have been re-
ported for Hubbard Brook (Likens et aj. 1977) and Plastic Lake, Ontario
(Zimmerman and Harvey 1979). Jeffries et aj_. (1979) report spring pH
depression in several Ontario lakes and streams of 0.3-1.2.
The data presently available on the general relationship between
deposition ana pH depression (A pH) are presented in Figure IV-8. The
dashed lines indicate areas of differing geological sensitivity (i.e.,
buffering capacity). It is also important to recognize that two types of
streams will show little effect: "acidified" streams, those with pH in
the range of 4.7 to 5.0 during much of the year, and which have gone
through the "transition phase"; and well-buffered streams. Those likely
to show appreciable pH depression are those believed to be experiencing
the "transition phase" of Henriksen.
Pending further testing, a "significant" shock event response, de-
fined as a A pH of 0.5 to 1.0 unit, mav be a useful estimate of the annual
S0|- threshold loading required to produce an unacceptable level of pH
depression (this level to be based on the controlled environment effects
observed on fish populations). These estimates must be defined for
watersheds within a specified sensitivity range, a limited range of hydro-
logic dilution (i.e., stream size), and with a defined return interval for
the episode. Recurrence an average of once a year during critical life-
66
-------
cr>
7.0
6.5
II ICtll
6.0
(icnui
5.5
5.0
1.5
1.0
5.5
rl,;.n;irA:i DAILY PII ret us sii/v/rns FORX nivrn AF DUIIS. i/.v.
PI![Ciri!A(IG.-| LVtiir rl! All!) ACCUIIULAIIttl At AliBOT/AlC. I
A RIVER HI
O BAIill-AU Pll
j] nAi:ir/\iL ACCIESJLAI 101
— OISCIU1CC
10 o °
o
o
ilnIl_Ln t'l.nnll'illL
10 15 70 25
1//7
._
JO
Reproduced Irom ||pl
bes! available copy. %yj
2.0
1.75
1.5
1.25
1.0 -;
.75
0.5
.25
Figure IV-6. River pH, rainfall pH, rainfall accumulation and discharge rate for the Shavers Fork River,
West Virginia, illustrating river pH depressions during snowmelt, when river acidity is
generally greatest. Precipitation data were collected at Arborvale, West Virginia. (From
Dunshie 1979.)
-------
PH
7 -
7 I
Figure IV-7. pH depression in Little Moose Lake, New York, occurring
during spring snowmelt as measured in lake outlet and water
entering the laboratory from a 3-m intake from the lake.
Stream pH was more severely affected than lake pH. Little
Moose Lake aluminum concentrations increased from pre-thaw
leve'is of less than 20 ueq/£ to 320 ueq/£ during early
thaw. (From Schofield 1980.)
68
-------
3.0 r
2.5
2.0
pH 1.5
1.0
0.7
0.5
.J.
I3I
11.
10 20 30 40 50 60
SULFATE LOADING [Kg/ha/yrl
Figure IV-8. Relationship between sulfate loading rate and potential pH
changes occurring during spring melt for sensitive areas
(top dashed line) and slightly less sensitive areas (bottom
dashed line). Rapid pK changes of 0.5 to 1.0 unit are
considered detrimental to most fis;. species. Data shown
are from: (1) Minnesota (Glass 1980); (2) Ontario (Harvey
et
-------
cycle stages would be consistent with the physiological data base des-
cribed above, and implicit in the model. Flushing events, however, appear
to be possible at almost any time of the year and, therefore, could
threaten other life stages and other species. From Figure IV-8, a sulfate
loading of 5 to 7 kg/ha-yr is seen to produce a critical episode response
(A pH) in the range of 0.5 to 1.0 for the most sensitive streams.
The major advantage of the above model is that it relates acid
loading directly to biological effects of interest, namely, a pH depres-
sion sufficient to cause observable mortality of organisms at sensitive
life stages. It also emphasizes non-equilibrium transition events,
thereby predicting when resource impacts will tend to begin, not when they
will be past and the lake or stream already acidified. The model also
appears likely to apply equally to lakes as well as streams.
Although the episode receptor/dose relation incorporates physio-
logical data on organism sensitivity to H+, further work is required to
investigate fully the assumptions and relationships implicit in this
approach to measuring water quality alteration and. defining acid-loading
tolerances. The model needs to be evaluated fully in relation to types of
geological materials in the watershed, the pa-v, duration of acid loadings,
the range in flow rates of the streams or seasonal rivulets, and the
interval for deposition accumulation. It also needs to be evaluated for
applicability in southern watersheds where snow does not accumulate and
where flushing events are related to significant rainfall following
drought (Ounshie 1979). Finally, the role for fish mortality of short
periods of elevated aluminum concentrations must be investigated,
particularly the apparent interactions with H+ (Schofield, personal
communication).
70
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SECTION V
RESEARCH AND DATA REQUIREMENTS FOR APPLICATION AND VALIDATION
This report has considered various measures of ecosystem alteration
with respect to the effects of acid rain on terrestrial and aquatic eco-
systems. The details of the measurement proposals, however, beg the
question of what constitutes an appropriate or minimally sufficient des-
cription of ecosystem response. One should be particularly concerned that
the "indicator" measures not be selected from a restrictive set that
documents only a limited aspect of total ecosystem function. For example,
concerns over eutrophication seldom take into account the larger changes
in fauna and flora that accompany such processes—but more often restrict
documentation to a few "water quality" parameters, reflecting the
state-of-the-water chemistry.
Recognizing the need for a minimally sufficient set of well chosen
measures to reflect the various dimensions of ecosystem transformation,
Statistics Canada has developed a framework for environmental statistics
which seeks to establish taxonomies, both for human stress factors and
environmental responses, that are reflective of the major aspects of the
environmental transformation process (Rapport and Friend 1979; Rapport and
Regier 1980; Rapport 1981). With respect to ecosystem response, the
provisional taxonomy for response measures includes indicators and
integrators pertaining to the following ecosystem attributes: produc-
tivity, nutrient concentration and cycling, composition of the biota,
quality of biota, and quality of the environment.
The measures discussed in this report cover a number of aspects of
the overall response taxonomy. The forest site index reflects forest
ecosystem productivity; the calcite saturation index deals with aspects of
nutrient cycling; the toxicity measures refer to the quality of biota —
both with respect to mortality and morbidity; the composition of biota is
best documented in the sequential decline in reproduction and ultimately
extinction of various fish species in acidifying lakes. This latter work
could be extended to identifying the more sensitive "function:;1! groups" of
species that are displaced as well as new groups of dominants that may
become established. With respect to quality of biota, referred to above,
measures of changes in size distributions (particularly stunting of
trees), disease incidence on trees in acid-stressed areas, and failures to
reproduce might also be included — in addition to the toxic aspects
already mentioned.
AN APPROACH TO ASSEMBLING SENSITIVITY MEASUREMENTS
A preliminary listing of the data needed to document lake and water-
shed sensitivity is given in Table V-l, effectively summarizing the
various methodologies discussed previously. The table is meant only to
suggest the range of factors involved in lake and watershed sensitivity,
not to be comprehensive or complete at this time. However, one need not
have all the listed data to estimate some of the integrative measures. A
71
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TABLE V-l. LISTING OF DATA NEEDS AND COMPUTATION PROCEDURES FOR APPLYING
SENSITIVITY MEASURES AND POLLUTANT LOADING TOLERANCE MODELS IN
NORTHEASTERN NORTH AMERICA.
I.
GENERAL WATERSHED DATA
Location:
Area (ha):
% Urban
% Forest
Land
Mean Drainage Basin: Slope
Water Flow Volumes: Input
Water Flow Rates: Input
Retention Time:
Avg. Precipitation (mm)
Water
% Agricultural
% Other
Length
Output
Output
Mean Annual pH
II. GENERAL AIRSHED DATA .
Regional Atmospheric Bulk Loading (kg/ha-yr): S02 NO
SO^- NH4 Pb Hg C.3
Regional Cumulative Growing Season Exposure (ug/m3-yr above background): 0;
III. SOIL SENSITIVITY DATA
Major Soil Types Parent material
Horizon Depth (cm)
CEC (meq/100g)
Base Saturation (%)
Soil pH (water)
Soil texture
Extractable (ug/2)
Al
0?
A1
A?
BI
B?
Soil Sensitivity:
McFee
Coote et al.
Non-sensitive Slightly sensitive
Sensitive
IV. FOREST SITE INDEX SENSITIVITY DATA
Major Plant Communities:
Dominant/Codominant Tree spp.
Age of Dominants
Height of Dominants
Understory Spp.
Field Measurement of SI (ft.)
Regression Estimate of SI (ft.)
Calculated Nutrient Loss (Gain) Due to Acid Deposition (kg/ha-yr):
Ca +Mg K Other
Calculated peak AL3> (ug/2) +
Estimated SI with Nutrient Change and Al3 (ft.)
Estimated Change in Annual Fiber Production (kg/ha)
72
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TABLE V-l. (continued)
IV. AQUATIC SENSITIVITY DATA
pH Alkalinity (meg/2)
Specific Conductance (|jnihos/cm)
Secchi Disk Transparence (m) +
Concentrations (mq/2): DOC POC Caz
K+ . Na+ NO,- NH^-f
P02~ HCO.,- Al Mn
ZN Cu Hg Cd
Calcite Saturation Index
Mq2+
sor
Fe
CSI = p(Ca2+) + p(Alk) - p(H+) + 2
1. Ca2+ (mol/2) ; p (Ca2*)
2. Alk (eq/2) ; p (Alk)
3. H+ (eq/£) ; p (H+)
4. CSI =
(<1 . . Non-sensitive; 1-3 . . Potentially Sensitive; >3 . . Sensitive)
Henriksen Nomograph (See Figure IV-5)
1. Ca2+ cone. (ueq/2) 2. Mg2 cone. (ueq/2)
3. Excess S0|- (peq/2)
4. Lake Status: Acidified
Transition
Bicarbonate
(Based on Henrikson, 198C)
Dickson Relation
SO2, loading (kg/ha-yr)
Lake pH
Sensitivity category (see Figure IV-4)
V.. FISH SENSITIVITY DATA
Chronic effects: Al (mg/2) pH
(See Tables III-3 and III-4) '
Shock effects: SOj loading
Sensitivity category (from Dickson)
Predicted A pH
Predicted peak Al
73
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viable sensitivity assessment strategy would fill in theso data needs as
studies or existing resource surveys (e.g., soi'ls are assembled for as-
sessment purposes).
General Watershed Data
Coarse estimates of lake/watershed sensitivity could be made through
evaluation of the data collected under this heading. For instance, a
watershed with a large percentage of urban and agricultural lanJ (and
associated fine particulate runoff) is unlikely to be as sensitive as an
area with a large proportion of forest and water. Knowledge of water flow
rates, volumes and soil retention time also is important in assessing the
ability of soils to neutralize acidic inputs.
General Airshed Data
Application of the Site Index measure requires the use of several
dose/response relationships for sensitive tree species. Thus, ozone and
other gaseous pollutants should be available from regional monitoring. In
addition, use of Almer/Dickson relation, Henriksen Nomograph and pH shock
effects models requires measuring the atmospheric loading of acidifying
substances.
Soil Sensitivity Data
The data needs listed here are derived from McFee (1980) and Coote et
aj. (1980). Extractable aluminum should be measured to allow preliminary
estimation of potential toxicity problems, should sulfates accumulation in
soils reach that point. Together, these data provide a broad-scale
assessment of sensitivity.
Site Index Sensitivity Data
Forest site index is proposed as an integrative measure for terres-
trial sensitivity because oxidants, aluminum toxicity and nutrient losses
can all affect the rate of height growth, a term that translates readily
into terrestrial productivity. The magnitude of site index Decreases,
therefore, represents a measure of the resulting loss in wood fibre pro-
duction due to 03 and acid deposition.
Aquatic Sensitivity Data
The methodologies for evaluating aquatic sensitivity have been
developed to a greater extent than terrestrial methodologies, but field
validation is still required. The data needs listed here are based or. the
methodologies of Kramer (1976), Henrikson (1980) and Aimer et a]_. (1978),
as discussed in previous sections. Monitoring heavy-metal concentrations
would help to define potential toxicity problems.
Fish Sensitivity Data and Watershed Tolerance
The chronic effects of pH and aluminum on fish reproduction have been
described as fully as possible, although more work on aluminum toxicity
will be important. The pH and aluminum shock effect on various life-
74
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stages of fish is a dose/response relationship now being developed and in
need of validation. Development of this methodology is critical, as it is
the only measure that bears directly on the biological effects of acid
precipitation.
OTHER INFORMATION/RESEARCH NEEDS FOR SENSITIVITY MEASURE VALIDATION
A considerable amount of research will be necessary to show that the
measures given above are adequate and appropriate to determining forest,
lake/watershed sensitivity. While a variety of data needs is implicit
throughout the previous discussion, this section attempts to summarize
necessary research according to major subject areas.
Effects on Forest Productivity
Since forest site index is a highly integrative measure, it reflects
a variety of effects mechanisms, including oxidant damage to foliage,
nutrient addition and depletion, and the toxic effects of H+, aluminum and
heavy metals. The magnitude of the effect on height growth in relation to
the combined dose must ultimately be measured in.the field, not simply
calculated from a variety of separate studies. In general, much more data
are needed on the response of various forest species and entire communi-
ties to the various toxic agents.
Sulfate and H+ Transpo. t
The movement of H+ within the sytem is, along with movement of the
anion (SO^-), fundamental to predicting all other effects. Water flow
patterns influence the pathways of S0|- and H+ transport and the inter-
action with other system components. Simple quantitative methods are
necessary to predict water flow patterns in relation to soil type, topog-
raphy, rate of flow and other relevant variables. Of particular impor-
tance is study of flushing event processes as they affect S0|- and H+
contact with the soil, and the regulation of the nitrogen and sulfur
cycles in soil. The theoretical considerations of these questions given
in this paper need to be supplemented with experimental work and watershed
mass balance studies.
Nutrient Cation Stripping
While theory and some field studies indicate acid precipitation leads
to Teaching of cations such as calcium and magnesium, other field studies
indicate such losses may be small. Quantitative studies of nutrient
stripping and the consequential effects on forest productivity are needed,
therefore, for a variety of soils with different ratis of acidic inputs.
This is especially important in soils which are not now acid, but have
relatively low base saturation.
Aluminum Mobilization
The relationship between SOij and H+ input to a watershed and the
subsequent output of soluble aluminum must be studied on the watershed as
well as the soil column level. Of critical importance are the effects of
75
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variation in slopes, soil type and water, flow (flushing event)! patterns.
In order to predict the occurrence of spring or .fall aluminum flushes, one
must understand the dynamics of water flow during snowmelt or other
flushing events as water penetrates various soil horizons. Research
requirements include comparative watershed studies as well as regionally
placed lysimeter studies.
Role of Organic Substances in Water
Superimposed on many of the H+ and metal mobilization processes -and
effects are the modifying influences of organic molecules within the
system. These range from buffering of H+ input, to detoxification of
aluminum, to mobilization and transport in chelated form of various
elements and compounds. At the present time, many of these processes are
poorly understood, particularly in aqueous media. Further research snould
identify sources of organics in a variety of soils and determine quanti-
tatively the impact of the amounts and kinds of organics being observed.
Other Aquatic Effects
Most of the questions concerning'aquatic effects centc^ around the
dose/response relationships of H+, aluminum and heavy metals for fish and
other components of the aquatic community. Of particular concern are
potential synergisms between the various toxicants and overall community
and ecosystem responses. Effects on fish-eating birds and mammals as a
lc«:e/stream fishery declines, and certain species are c-1 inrinated, must be
investigated fully. It is possible that research in this area may
identify additional broadly based measures of aquatic ecosystem stress
comparable to the site index for forests.
76
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