Biological Services Program
FWS/OBS-80/40.3
JUNE 1982
Air Pollution and Acid Rain,
Report No. 3
THE EFFECTS OF AIR POLLUTION AND ACID RAIN
ON FISH, WILDLIFE, AND THEIR HABITATS
INTRODUCTION
Office of Research and Development
U.S. Environmental Protection Agency jBhF
Fish and Wildlife Service
U.S. Department of the Interior
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The Biological Services Program was established within the U.S. Fish and
Wildlife Service to supply scientific information and methodologies on key
environmental issues that impact fish and wildlife resources and their supporting
ecosystems.
Projects have been initiated in the following areas; coal extraction and
conversion; power plants; mineral development; water resource analysis, including
stream alterations and western water allocation; coastal ecosystems and Outer
Continental Shelf development; environmental contaminants; National Wetland
Inventory; habitat classification and evaluation; inventory and data management
systems; and information management.
The Biological Services Program consists of the Office of Biological Services in
Washington, D C., which is responsible for overall planning and management;
National Teams, which provide the Program's central scientific and technical
expertise and arrange for development of information and technology by contracting
with States, universities, consulting firms, and others; Regional Teams, which
provide local expertise and are an important link between the National Teams and
the problems at the operating level; and staff at certain Fish and Wildlife Service
research facilities, who conduct inhouse research studies.
For w»te by the Sii|><>rliiloii
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UNITED STATES
DEPARTMENT OF THE INTERIOR
FISH AND WILDLIFE SERVICE
EASTERN ENERGY AND LAND USE TEAM
Route 3, Box 44
Keameysville, West Virginia 25430
Dear Colleague:
The Eastern Energy and Land Use Team (EELUT) is pleased to present
you the results and synthesis of available information on the effects
of air pollution and acid rain or precipitation on fish, wildlife,
and their habitats. This Introduction is the first in a series of
nine reports relating acid rain and air pollution to forests, grass-
lands, lakes, rivers and streams, deserts and steppes, arctic tundra
and alpine meadows, urban ecosystems, and critical habitats of threat-
ened and endangered species. The series, "The Effects of Air Pollu-
tion and Acid Rain on Fish, Wildlife, and their Habitats" (FWS/OBS -
80/40.3, 40.4, 40.5, 40.6, 40.7, 40.8, 40.9, 40.10, 40.11) was pre-
pared under the direction of David Adler at Dynamac Corporation,
Rockville, Maryland.
This Introduction presents an overall view of the major air pollutants
including photochemical oxidants, particulates, and acidifying air
pollutants. The remaining series relate the effects of these pollu-
tants to several major ecosystems in the United States.
Please feel free to send suggestions or comments to EELUT so that
we may continually strive to improve our future products.
Sincerely
Edgar A. Pash
Team Leader, EELUT
Enclosure
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FWS/0BS-80/40.3
June 1982
AIR POLLUTION AND ACID RAIN, REPORT 3
THE EFFECTS OF AIR POLLUTION AND ACID RAIN
ON FISH, WILDLIFE, AND THEIR HABITATS
INTRODUCTION
by
M. A. Peterson
David Adler, Program Manager
Dynamac Corporation
Dynamac Building
11140 Rockville Pike
Rockville, MD 20852
FWS Contract Number 14-16-0009-80-085
Project Officer
R. Kent Schreiber
Eastern Energy and Land Use Team
Route 3, Box 44
Kearneysville, WV 25430
Conducted as part of the
Federal Interagency Energy Environment Research and Development Program
U. S. Environmental Protection Agency
Performed for:
Eastern Energy and Land Use Team
Office of Biological Services
Fish and Wildlife Service
U. S. Department of the Interior
Washington, D. C.
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DISCLAIMER
The opinions and recommendations expressed in this series are those
of the authors and do not necessarily reflect the views of the U.S. Fish
and Wildlife Service or the U.S. Environmental Protection Agency, nor
does the mention of trade names consitute endorsement or recommendation
for use by the Federal Government. Although the research described in
this report has been funded wholly or in part by the U.S. Environmental
Protection Agency through Interagency Agreement No. EPA-31-D-X0581 to
the U.S. Fish and Wildlife Service it has not been subjected to the
Agency's peer and policy review.
The correct citation for this report is:
Peterson, M.A. 1982. The effects of air pollution and acid rain on fish,
wildlife, and their habitats - introduction. U.S. Fish and Wildlife Service,
Biological Services Program, Eastern Energy and Land Use Team,
FWS/OBS-80/40.3. 181 pp.
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ABSTRACT
Air pollution arid acid rain impacts on living resources are a major
source of concern to the U.S. Fish and Wildlife Service and other govern-
mental agencies charged with the protection of natural resources and the
environment. This introductory volume synthesizes the results of scien-
tific research related to air pollution effects on fish and wildlife
resources. It is intended for use as a general reference and to provide
background information for the eight ecosystem specific reports in this
series: Deserts, Forests, Grasslands, Lakes, Rivers and Streams, Tundra
and Alpine Meadows, Urban Ecosystems, and Critical Habitats of Threatened
and Endangered Species.
Air pollutants related to effects on fish, wildlife and their habi-
tats are classified into three categories: photochemical oxidants, par-
ticulates, and acidifying air pollutants. A general summary of pollutant
origins, atmospheric transport, transformation and deposition is presented
in this volume. The bulk of this report describes plant, animal, and
ecosystem responses to air pollution as well as factors affecting the
sensitivity of receptive ecosystems. The Introduction also briefly
describes relevant features of air quality legislation. A computerized
bibliography and special reference library have been prepared in support
of this series.
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CONTENTS
Page
ABSTRACT iii
FIGURES vi
TABLES ix
1.0 INTRODUCTION 1
2.0 PRINCIPAL CATEGORIES AND ORIGINS OF MAJOR
AIR POLLUTANTS 5
2.1 Photochemical Oxidants 5
2.2 Particulates 9
2.3 Acidifying Air Pollutants 18
3.0 ECOSYSTEM EXPOSURE TO AIR POLLUTION 28
3.1 Atmospheric Transport and Transformation 28
3.2 Atmospheric Deposition Processes 33
3.3 Pathways of Air Pollution Exposure to
Fish, Wildlife, and Their Habitats 37
4.0 FACTORS AFFECTING ECOSYSTEM SENSITIVITY 40
4.1 Meteorology 41
4.2 Geology 42
4.3 Pedology 43
4.4 Hydrology 50
4.5 Hydrochemi stry 51
4.6 Topography 57
4.7 Biota 57
4.8 Human Activity 59
4.9 Summary 61
5.0 RESPONSES OF FISH, WILDLIFE, AND HABITAT
TO AIR POLLUTION AND ACID RAIN 64
5.1 Response to Photochemical Oxidants 66
5.1.1 Plant Response 67
5.1.2 Animal Response 70
5.1.3 Ecosystem Response 72
iv
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CONTENTS (continued)
Page
5.2 Response to Particulates 74
5.2.1 Terrestrial Plant Response . 75
5.2.2 Terrestrial Animal Response 77
5.2.3 Terrestrial Ecosystem Response 82
5.2.4 Aquatic Plant Response 85
5.2.5 Aquatic Animal Response 87
5.2.6 Aquatic Ecosystem Response 90
5.3 Response to Acidifying Air Pollutants 92
5.3.1 Terrestrial Plant Response 93
5.3.2 Terrestrial Animal Response 96
5.3.3 Terrestrial Ecosystem Response 98
5.3.4 Aquatic Plant Response 100
5.3.5 Aquatic Animal Response 105
5.3.6 Aquatic Ecosystem Response 116
6.0 AIR QUALITY LEGISLATION 120
6.1 The Clear Air Act 120
6.1.1 Federal Air Quality Standards 120
6.1.2 Federal Emission Standards 124
6.1.3 State Responsibilities Under
the Clean Air Act 124
6.1.4 Discussion 129
6.2 The Acid Precipitation Act 129
6.3 International Cooperation 131
REFERENCES 132
GLOSSARY 376
v
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FIGURES
Number Page
1 Principal components of the air pollution/
acid rain system in relation to the ecosystems
addressed by this series 2
2 Estimated trends in anthropogenic nitrogen oxide
emissions, by source (high growth scenario) 8
3 Estimated trends in anthropogenic hydrocarbon
emissions, by source (high growth scenario) 10
4 Estimated trends in net anthropogenic particulate
emissions, by source (high growth scenario) 17
5 The pH scale with comparisons of acid rain to
common acid and alkaline substances 19
6 Typical northeastern U.S. acid rain components
averaged annually 20
7 Weighted mean pH of precipitation in the
continental United States (1976-1979) 22
8 Estimated trends in anthropogenic sulfur oxide
emissions, by source (high growth scenario) 25
9 Monthly mean concentrations of sulfate as a
function of time at Cornell University (Ithaca),
Pennsylvania State University and the University
of Virginia 27
10 Particulate transformation and deposition
processes in relation to particle size 30
11 Mean annual hydrogen ion (H+) deposition in
precipitation over the continental United
States, 1976-1979 (Kg/ha) 36
12 Areas of differing acid-neutralizing capacities
in New York State, according to the geological
buffering classification of Table 11 44
13 Vulnerability map for the state of Georgia,
by county 45
vi
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FIGURES (continued)
Number Page
14 Geologically sensitive regions of North
America with lakes susceptible to acid
precipitation 45
15 An example of soil sensitivity mapping based
on the soil sensitivity classification of
Table 12 48
16 Henriksen's nomograph for predicting the acid
status of lakes 52
17 A Schofield diagram for Adirondack lakes, with
sulfate levels of 100-120 ueq/1, superimposed
on Henriksen's nomograph 54
18 Frequency histograms for fish status in 684
Norwegian lakes separated according to their
position on the nomograph of Figure 16 55
19 Total dissolved aluminum vs. pH in lakes of
the Adirondack Mountains, New York 56
20 Data from lakes in Sweden showing the relationship
between anthropogenic sulfate loadings and pH
change for (1) very sensitive and (2) somewhat
sensitive surroundings 60
21 Areas of the United States with soils sensitive
to atmospheric deposition overlain with 1978-
1979 pH isopleths from the National Atmospheric
Deposition Program 63
22 Lower limits of pH tolerance among the
phytoplankton 101
23 Number of phytoplankton species observed in
Adirondack Mountain lakes of different pH 103
24 Comparative pH tolerances of four groups of
fish-food organisms 107
25 Fish habitat selection in response to freshwater
acidification 114
vii
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FIGURES (continued)
Number Page
26 Extinction of brown trout populations in lakes
of southern Norway during the period 1940-1970 115
27 A schematic representation of the hypothesis
of auto-oligotrophication in acid lakes 118
28 An overview of the Clean Air Act 122
vii i
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TABLES
Number Page
1 Major pollutants of the photochemical oxidant
complex 6
2 Estimated national anthropogenic emissions of
photochemical oxidant precursors in 1977 7
3 Chemical constituents of particulate pollution 12
4 Estimated national anthropogenic emissions of
particulates in 1977 13
5 Estimated global emissions of atmospheric metals
from natural and anthropogenic sources 15
6 Average ambient trace metal concentrations in
remote, rural, and urban atmospheres 16
7 Primary and secondary forms of acidifying air
pollutants 18
8 Estimated national anthropogenic emissions of
acid rain precursors in 1977 23
9 Regional emissions of sulfur and nitrogen oxides
compared to population (percent of U.S. totals) 24
10 Average concentrations of metals in wet deposition. ... 35
11 Classification of rock types used to distinguish
geological sensitivity 43
12 Classification of soil sensitivity to acid
precipitation based on cation input and the
chemical characteristics of the top 25 cm 47
13 Soil sensitivity to acid precipitation based on
buffering capacity and hydrogen ion retention 47
14 Lake classification and fish population status
of 214 Adirondack moutain lakes based on the
nomograph of Figure 16 56
15 Factors indicative of potential ecosystem
sensitivity to acidifying air pollutants 62
ix
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TABLES (continued)
Number Page
16 A hierarchical classification of biotic and
ecosystem-level responses to air pollutant
uptake or deposition 65
17 Available reviews of the biotic and ecosystem-
level effects of air pollution and acid rain 66
18 The effects of photochemical oxidants on animals 72
19 Selected references on the use of lichens and
mosses in monitoring the deposition of atmospheric
trace metals 76
20 Incidents involving the adverse effects of
atmopsheric particulates on vertebrate wildlife 78
21 Selected references on the bioaccumulation of
trace metals in wild animal populations 79
22 The major biological systems of animals affected
by atmospheric particulates 80
23 Documented responses of wildlife, domestic and
laboratory animals to acute and chronic exposures
of atmospheric particulates 83
24 Studies of the use of lichens and tree bark in
the monitoring of acidifying air pollution 94
25 Processes and structural characteristics that
reduce plant sensitivity to acid deposition 95
26 Birds and mammals susceptible to indirect effects
of acid deposition 98
27 Amphibians susceptible to reduced reproductive
success from breeding-habitat acidification 109
28 A summary of effects of freshwater acidification
on fish 110
29 Acronyms for principal components of the Clean
Air Act 121
x
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TABLES (continued)
Number Page
30 National primary and secondary ambient air
quality standards 123
31 Mobile source exhaust emission factors for
1979, 1980 and 1985-1990 125
32 Prevention of significant deterioration
regulations 128
33 Federal departments and agencies participating
in the Interagency Task Force on Acid Precipitation . . . 130
xi
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1.0 INTRODUCTION
This introductory volume and accompanying reports taken together
constitute a review of the state of current knowledge concerning the ef-
fects of air pollution and acid rain on fish, wildlife, and their habi-
tats. The U.S. Fish and Wildlife Service (FWS), as a source of ecological
expertise within the Department of Interior, acts as a focal point for
research, methodological development, and dissemination of information to
improve the effectiveness of decisionmaking in areas affecting natural
resources. Particularly important are decisions which must weigh the ad-
vantages of development projects against potential long-term harm to the
environment from resulting pollution.
The FWS has statutory responsibility to protect the nation's fish and
wildlife resources and specifically to review and comment on the environ-
mental impacts of proposed developments. This set of reports on air
polution and related acid precipitation is intended to strengthen the
ability of the FWS to meet this responsibility by providing a usable
summary of what is known about the impacts and what is yet to be
learned. The focus is on the effects of air pollution, the long-range
transport of air pollution (LRTAP), and atmospheric deposition on fish
and wildlife. The reports are general in nature with extensive refer-
ence to the technical literature provided for those requiring additional
information beyond the scope of these documents.
The reports are the result of efforts to review and consolidate
information from current research and the available literature into a
format appropriate to FWS needs. They are primarily oriented toward the
requirements of biologists and Federal and state land managers called
upon to assess the ecological impacts of air pollution. However, as a
synthesis of information it is expected that the reports will be of
interest to a diverse audience including policymakers, public interest
groups, consultants, and others concerned specifically with the anticipa-
tion and mitigation of detrimental effects of air pollution on wildlife
resources.
This series consists of nine individual survey reports; the principal
topics covered are depicted in Figure 1. This volume provides an intro-
duction to the nature of air pollution, its sources, atmospheric trans-
port, transformation, deposition and fate in terrestrial and aquatic eco-
systems. The biological and ecological effects of air pollution, as well
as factors affecting ecosystem sensitivity, are discussed. A brief intro-
duction to air quality legislation is included. Also, a glossary of
important terms can be found at the end of this report. The remaining
reports in this series provide detailed information specific to effects
within the eight ecosystems represented schematically in Figure 1. The
series is supported by a computerized bibliography and a library of the
scientific literature employed.
To keep the discussion within manageable bounds, a simplified classi-
fication of air pollution has been used which is consistent with the focus
of the series on effects as observed in the field. Pollutants affecting
wildlife resources are classified into one of three categories:
1
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ro
i i
POLLUTANTS
Key: ~ 1 - Air Emissions
2 - Atmospheric Pollutants
3 - Pollutant Transport and
Transformation
4 - Dry Deposition
5 - Wet Deposition
ECOSYSTEMS
O A - LAKES
B - RIVERS AND STREAMS
C - FORESTS
0 - GRASSLANDS
E - URBAN ECOSYSTEMS
F - TUNDRA AND ALPINE MEADOWS
G - DESERTS
H - CRITICAL HABITATS OF THREATENED
AND ENDANGERED SPECIES
Figure 1. Principal components of the air pollution/acid rain system
in relation to the ecosystems addressed by this series.
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• Photochemical oxidants. These pollutants constitute a large vari-
ety of substances, or "secondary" pollutants, which are produced
by chemical reactions in the atmosphere as a result of sunlight
acting on nitrogen oxides and hydrocarbons. These "primary" pol-
lutants in turn may be the result of either natural or man-made
processes. The most important impacts of photochemical oxidants
are on terrestrial plants.
• Particulates and other toxic pollutants. Included in this cate-
gory are the heavy metals, fluorides, and organic micropollutants.
These can affect plants, animals, and the quality of their habi-
tats in many different ways.
• Acidifying air pollutants. These pollutants share the character-
istic that their effects on plants and animals are due to acidity.
They may be acids such as those found in acid deposition or they
may be gases, such as SO?, which are converted to acids when
incorporated into living systems.
Because the series addresses the needs of the FWS, the emphasis of
the reports is on the effects of air pollution as observed in the field.
In general, there are three types of effects:
• abiotic effects on air, land, and water such as leaching phenomena
resulting in the movement of nutrients or changes in the acidity
of surface water;
• aquatic and terrestrial biotic effects such as increased suscep-
tibility of plants to disease and predation or impacts on flora
and fauna resulting from changes in water or soil chemistry; and
• ecosystem effects including complex interactions between biotic
and abiotic components exemplified by forest ecosystems in which
decreased productivity of trees and increased nutrient mobility
caused by air pollution can cause changes in evapotranspiration,
microclimate and other factors, ultimately leading to large losses
of nutrients from the ecosystem and subsequent effects on down-
stream aquatic ecosystems.
These effects may be acute, episodic and short-term, or they may be chron-
ic , cumulative and long-term. A great deal of further study is needed to
assess their potential reversibility.
Each of the ecosystem-specific reports in this series contains a
brief background discussion of biotic, abiotic, and functional ecosystem
aspects relevant to an assessment of air pollution impacts. The current
state-of-knowledge concerning biological and ecological impacts of air
pollution is introduced for each ecosystem and related socio-economic
considerations are briefly outlined where appropriate.
3
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The final chapter of each ecosystem-spec ific report discusses topics
for further research related to the specific ecosystems. Some research
areas closely complement FWS experience and capability while others are
beyond the scope of the agency. As a general rule, these sections point
out that an effective evaluation of air pollution effects will require
further investigations in established disciplines of plant and animal
physiology, plant pathology and the biomedical sciences. The complexity
of effects stemming from phenomena such as LRTAP, the long-range trans-
port of air pollutants (especially acids, oxidants, trace elements, and
the organic micropol1utants), poses significant challenges to research
efforts in these fields. In view of the expense and specialization
required to develop knowledge of these subjects, FWS activities in the
following areas appear especially useful:
¦ a continuing synthesis and dissemination of the literature on
plant and animal effects in specific ecosystems as it relates to
the needs of FWS personnel;
• the creation of a working relationship between key researchers or
research institutions and the FWS personnel who rely on the devel-
opment of specific areas of knowledge or expertise; and
• efforts to expand the use of wildlife species and selected plants
on which they depend in field experimentation related to the
biotic effects of air pollutants and LRTAP.
Several reports point to the potential utility of indicator plants
and animals in pollution-related research applications. In view of the
FWS role in wildlife protection, habitat preservation and the development
of baseline ecological knowledge to support these activities, biomonitor-
ing research would appear to be a feasible and potentially productive
activity complementing other agency responsibilities. Furthermore, as the
monitoring of ambient air pollution and precipitation chemistry is ex-
panded to remote areas of the country, biomonitoring undertaken by field
specialists may be one of the more practical methods of discerning poten-
tial biotic effects in these remote areas.
Increased understanding of the complex impacts at the ecosystem level
will not be easily or quickly achieved. Again, a continuing synthesis
and dissemination of relevant ecosystem-specific literature is important.
The ecological research should be complemented by research related to the
social, economic, and intergenerational aspects of large-scale alterations
in ecosystem functioning. Although such studies have a high degree of
uncertainty associated with them, they are necessary for a complete pic-
ture of the impacts of air pollution and acid deposition on fish, wild-
life, and their habitats.
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2.0 PRINCIPAL CATEGORIES AND ORIGINS
OF MAJOR AIR POLLUTANTS
For the purposes of this series of reports the air pollutants of
greatest significance for potential impact to fish, wildlife and their
habitats will be classed into three functional categories:
• photochemical oxidants;
• particulates; and
• acidifying air pollutants.
This classification is useful to a review of biological effects since
the individual pollutants in each category possess similarities in physi-
cal and chemical properties and cause similar effects. The pollutants of
a class also share similar atmospheric dispersion and deposition process-
es, follow common exposure pathways to wildlife and habitat, and elicit
like responses among the different components of the ecosystem. Some
pollutants of less significance to fish and wildlife effects may fit only
loosely within this functional categorization.
Air pollutants may be emitted from point sources (fixed identifiable
sources) or area sources (numerous point or mobile sources such as chim-
neys or automobiles). Pollutants from point and area sources are refer-
red to as primary pollutants. Interactions among two or more primary
pollutants and normal atmospheric constituents create secondary pollu-
tants. Many of these chemical reactions require photoactivation. The
rates, reaction routes, and intermediate steps involved in the process are
influenced by many factors such as relative concentration of reactants,
degree of photoactivation, variable meteorological dispersive forces,
influences of local topography, ambient temperature and relative humidity.
Although air pollution is usuaUy associated with human activities,
many of the major gaseous pollutants also have natural sources. Taken on
a worldwide basis, natural emissions of some pollutants, such as nitrogen
oxides from volcanic eruptions, often exceed anthropogenic, or man-made,
emissions by several orders of magnitude (Urone 1976). The reason man-
made emissions so noticeably affect the quality of the environment is that
they tend to concentrate locally or regionally in air masses downwind of
urban areas or large point sources.
2.1 PHOTOCHEMICAL OXIDANTS
The group of pollutants known as photochemical oxidants are second-
ary pollutants, that is, they are produced by complex chemical reactions
between primary pollutants, notably nitrogen oxides and hydrocarbons.
These reactions are common in the atmosphere above urban centers where
there are hundreds of different hydrocarbons. Because of the multitude
and variety of photochemical oxidants, very few individual oxidants have
5
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been observed. Table 1 lists the major oxidants which are of primary
concern as a pollution problem (National Research Council 1977). These
are generally found in association with other photochemical products,
including aldehydes, formic acid, nitrous acid, gaseous and particulate
nitrates, and sulfates.
Table 1. Major pollutants of the photochemical oxidant complex.
Pollutant
Formula
Ozone
03
Peroxyacetylnitrate (PAN)
CH3COO2NO2
Ozone (O3) is by far the most abundant of the photochemical oxid-
ants. It is a product of many different reactions, and also occurs na-
turally in large quantities in the stratosphere (Singh et al. 1980). Peak
concentrations in urban and suburban areas may range from TT73 to 0.6 ppm
while peak ozone levels in rural areas seldom exceed 0.2 ppm (U.S. EPA
1978a). Average concentrations in rural areas may often equal or surpass
urban averages due to pollutant transport and the regional scale of oxid-
ant pollution. Averaged ambient ozone levels in urban areas showed no
long-term trend over the period from 1974 to 1979, however national aver-
ages were observed to decline by 3 percent between 1978 and 1979 (USEPA
1980a).
Peroxyacetylnitrate (PAN) is actually the most abundant member of a
series of similar compounds, but it appears to be the only one found to
occur in photochemical smog at concentrations high enough to pose a pollu-
tion problem. Rural daily maximums rarely exceed 3.0 ppb, however maximum
urban valves can reach levels of 0.05 to 0.2 ppm (USEPA 1978a). PAN is
known to be of greater phytotoxicity than ozone; however elevated levels
of ozone are more common, thus ozone is considered to be the most injuri-
ous of the photochemical oxidants.
The precursors of photochemical oxidants are numerous, but generally
fit into the categories of nitrogen oxides (N0X) or hydrocarbons (HC).
These pollutants are emitted in large quantities as a result of fossil
fuel combustion. Estimated national emissions of these precursors are
presented in Table 2 by source.
The nitrogen oxide precursors of the photochemical oxidant complex
figure prominently among the acidifying air pollutants and will be des-
cribed in detail in Section 2.3. Major types of hydrocarbon precursors
found in urban atmospheres include the alkenes, alkynes, cycloalkanes and
aromatics. Hydrocarbons, as a group, constituted the second largest form
of air pollution by mass in the U.S. in 1977 (USEPA 1978b).
6
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Table 2. Estimated national anthropogenic emissions of
photochemical oxidant precursors in 1977.
Pollutant Source Annual Production (106 metric tons/year)
NQx HC
Transportation 9.2 11.5
Fuel Combustion 13.0 1.5
(Stationary Sources)
Industrial Processes 0.7 10.1
Solid Waste Disposal 0.1 0.7
Miscellaneous 0.1 4.5
Totals 23.1 28.3
[From USE PA lS7Bb)
Transportation and stationary source fuel combustion (e.g., power
plants) are the principal sources of the oxides of nitrogen (USEPA 1978b).
Solid waste disposal, industrial process losses, and miscellaneous re-
leases account for the remaining N0^ emissions. Almost all anthropo-
genic N0X is produced by the oxidation of atmospheric nitrogen (N^) at
high temperature during combustion. Hence, the formation of N0X is of-
ten promoted under those operating conditions where the combustion of fuel
is optimized (Stoker and Seager 1976). Figure 2 presents estimated trends
in gross and net N0X emissions, by source, through the year 2000.
Natural sources of N0X contribute significantly more than human-
related sources (Stoker and Seager 1976, Soderlund and Svensson 1976).
Bacterial action in soils, resulting in the decomposition of nitrogen-
containing compounds, represents the major natural source of nitrous
oxide (N2O), some 592 million tons (Stoker arid Seager 1976). Bacterial
activity is also responsible for an estimated annual production of 430
million tons of nitric oxide (NO). Lightning also produces N0X but the
amount is negligible.
The highest average concentrations of nitrogen oxides are found in
heavily populated, industrialized urban areas. Average ambient levels
greater than 1.0 ppm N0X are seldom encountered in cities, while rural
air usually averages a few parts per billion (USEPA 1978a). One-hour max-
imum concentrations of 2.5 and 4.5 ppm nitrogen dioxide {NO2) have been
7
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40
35
30
25
20
15
10
8
6
k
2
Gross
~ Captured
HI Net
197 5 1985 2000
PETROLEUM
REFINING
1975 1985 2000
INDUSTRIAL
COMBUSTION
1975 1985 2000
OTHER
197 5 198 5 2000
ELECTRIC
UTILITIES
19 75 198 5 2000
TRANSPOR-
TATION
19 7 5 1985 2000
TOTAL
EMISSIONS
Figure 2. Estimated trends in anthropogenic nitrogen oxide
emissions, by source (high growth scenario). (From USEPA 1980b)
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recorded in Chicago and Philadelphia, respectively, whithout exceeding the
primary air quality standard, set by the U.S. Environmental Protection
Agency, of 0.05 ppm NO2 on an annual average; levels in excess of this
standard are consistently found in the Los Angeles area (National Research
Council 1977). Due to projected increases in transportation and station-
ary source emissions, observations of nitrogen dioxide concentrations
showed an increasing trend between 1974 and 1979 (USEPA 1980a).
Transportation and industrial processes, particularly refining, are
the principal emitting sources of hydrocarbons. Automobile combustion
alone accounts for half of the total anthropogenic emissions of hydrocar-
bons. Other large sources include solvent evaporation and hydrocarbon
emissions resulting from various chemical manufacturing processes (Clark
1980; Tilton and Bruce 1980). Trends in gross and net hydrocarbon emis-
sions, by source, are given through the year 2000 in Figure 3.
Natural sources of hydrocarbons are also important. In 1968, natural
sources of hydrocarbons contributed 85 percent of the total atmospheric
input (Stoker and Seager 1976). The two principal types of natural emis-
sions are methane (CH4), and a group of organics known as terpenes.
Copious amounts of methane are generated in swamps, marshes, and similar
water bodies as a result of the anaerobic bacterial decomposition of or-
ganic matter. Methane does not contribute to photochemical oxidant forma-
tion (Tilton and Bruce 1980). The terpenes, a family of high molecular
weight hydrocarbons, are produced by vegetation. About 43 percent of
these emissions occur in the summer months, and about 45 percent of the
annual emissions occur in the southern United States (Tilton and Bruce
1980).
On a nationwide scale, there is a direct correlation between popula-
tion density and ambient hydrocarbon levels. The areas with the highest
density of non-methane hydrocarbon emissions are: 1) the northeast corri-
dor; 2) certain industrial states (Ohio, New York, Pennsylvania, Illinois,
and Indiana); and 3) states along the gulf coast where petroleum refining
and storage activities contribute heavily to hydrocarbon emissions (Tilton
and Bruce 1980).
Non-methane hydrocarbon concentrations typically average 1.0 ppm in
urban air and less than 0.1 ppm in rural air, however levels as high as
10.0 ppm have been recorded in Los Angeles (USEPA 1978a). National hydro-
carbon emissions decreased by 4 percent in the period from 1970 to 1979,
primarily due to the achievement of lower emissions from transportation
sources (USEPA 1980a).
2.2 PARTICULATES
Particulate matter in the atmosphere exists in a wide variety of
forms, usually as smoke, fumes, mists, oils, and dusts (Anon. 1969). Vir-
tually all of these pollutants fall into the category of aerosols: solid
particles or liquid droplets which are dispersed or suspended in a gaseous
9
-------
Figure 3. Estimated trends in anthropogenic hydrocarbon
emissions, by source (high growth scenario). (From USEPA 1980b).
-------
medium. They may range in chemical composition from a single elemental
species to highly complex substances incorporating virtually any of the
other atmospheric pollutants. Common components of atmospheric particu-
late burdens are listed in Table 3.
The specific composition of particulates greatly depends on the type
of emission source, industrial operating conditions, characteristics of
fuels employed, and the nature of emission-control equipment. In view of
its chemical diversity, particulate matter is usually categorized accord-
ing to its size. Atmospheric particles generally range in size from mol-
ecular clusters (^. 0.005 um) to visible dust (>100 um). Within this
range, they are classified as coarse or fine particles, depending on their
diameter and mode of formation. The primary particulates are generally
coarse (>2.0 um) and are emitted directly from anthropogenic and natural
sources. The majority of fine particles (<1.0 um) in the atmosphere are
secondary particulates that result from chemical reactions and transforma-
tions among ambient air pollutant mixtures (Fennelly 1976).
In industrialized regions, sulfates (SO4 ~) are by far the most
abundant of the fine particulates (Fennelly 1976). Under natural condi-
tions, more sulfate aerosols are associated with ammonium ions (NH4 +)
than exist as sulfuric acid (H0SO4), however sulfuric acid and com-
pounds such as lead sulfate (P6SO4) predominate in urban air. Nitrate
ions (NO3 ~) are common in both gaseous and particulate form in rural
and urban air (Galloway et. aj_. 1981). These acidifying pollutants are
discussed in the following section.
Carbonaceous aerosols are also abundant in fine particulate matter.
These usually contain an elemental component, often graphite or soot, and
an organic component, usually a hydrocarbon (Miller et al. 1979). Trace
organics include the polychlorinated biphenyis (PCBs"J7 cFTlorinated hydro-
carbon pesticides (HCCls) and polycyclic aromatic hydrocarbons (PAHs)
(Galloway et £]_. 1981). Such substances may exist as solid entities or
they may adsorb to other particulates from a vapor phase.
Atmospheric metals and trace elements range widely in size, and are
usually present in the solid form. Those found in a vapor phase include
arsenic (As), cadmium (Cd), mercury (Hg), and selenium (Se) (Miller et al.
1979; Galloway et al. 1981). Most of the metals combine with oxygen, sul-
fur, sulfate, or nitrate following their release to the atmosphere, and
are virtually always found in association with various kinds of particu-
late matter (Linton et 1976).
Particulates of lesser abundance include a variety of radioactive
Rarticles, the chief among them being isotopes of uranium (^35jjs
"8U); radioactive gases such as radon (^^Rn) may adsorb to particu-
lates (Rivera-Cordero 1970). Nutrient elements such as nitrogen (N),
phosphorus (P), and potassium (K) also exist in a variety of forms as at-
mospheric particulates. Finally, there are many inorganic particles of
natural origin, including siliceous dusts, sea salts, and carbonates which
contribute to atmospheric loadings of particulate matter.
11
-------
Table 3. Chemical constituents of particulate pollution.
Non-metallic Ions
Sulfate (SO4 =)
Ammonium (NH4 +)
Nitrate (NO3 ")a
Chloride (Cl~)
Fluoride (F~)
Nutrient Elements
Nitrogen (N)
Potassium (K)
Phosphorus (P)
Radioactive Particles
Uranium (235u„ 238y)
Synthetic Organic Compounds5
Polycyclic Aromatic
Hydrocarbons (PAH)
Chlorinated Hydrocarbons (HCC1)
Polychlorinated Biphenyls (PCB)
Natural Organic Dust
Pollen, spores, fungal hyphae
and plant parts
Trace Elements
A1umi num (A1)
Antimony (Sb)
Arsenic (As)a
Beryllium (Be)
Cadmium (Cd)d
Chromium (Cr)
Cobalt (Co)
Copper (Cu)
Iron (Fe)
Lead (Pb)
Nickel (Ni)
Manganese (Mn)
Mercury (Hg)a
Molybdenum (Mo)
Selenium (Se)a
Silver (Ag)
Thallium (Tl)
Tin (Sn)
Titanium (Ti)
Vanadium (V)
Zinc (Zn)
aThese substances may also be present in a vapor phase
12
-------
The primary anthropogenic sources of atmospheric particles are sta-
tionary fuel combustion and industrial processes (USEPA 1978b). Indust-
rial sources include non-ferrous metal smelters, iron and steel mills,
petroleum refineries, cement quarries, pulp and paper plants, asphalt and
chemical works, and manufacturers of soap and synthetic detergents, glass
and glass fiber, and textiles (Anon. 1969). Other human activities gener-
ating particulate pollution include the cultivation and fertilization of
agricultural and forest land, solid waste disposal, commercial and resi-
dential heating, and the operation of motor vehicles. Table 4 quantifies
national particulate emissions according to various types of sources.
Table 4. Estimated national anthropogenic emissions
of particulates in 1977.
Pollutant Source
Annual Production (10 metric tons/year)
Particulates
Transportation
Fuel Combustion
(Stationary Sources)
Industrial Processes
Solid Waste Disposal
Miscellaneous
Total
1.1
4.8
5.4
0.4
0.7
12.4
(From USEPA 1978b)
The majority of fine ambient aerosols result from the incomplete com-
bustion of fossil fuels and from residuals produced during incineration,
manufacturing, photochemical, and condensation processes (Lee 1972).
Greater amounts of fine particles are known to be emitted from oil-fired
burners than from coal-fired units (Cheng j?t _a2« 1976). Moreover, the
finer particulates from fossil fuel combustion have been found to prefer-
entially concentrate toxic trace elements on the particle surface (Linton
et cU. 1976). Toxic vapor-phase contaminants, such as the polycyclic
aromatic hydrocarbons and volatile metals, adsorb preferentially to fine
particulates of respirable size (Mi 1 ler et^ aj_. 1979).
13
-------
The natural sources of gaseous pollutants often generate particulate
matter and trace metals as well. Volcanic activity is one of the impor-
tant geochemical origins of these substances; another is wind erosion
(Peirson ej; al. 1973). Sea spray is the major marine source and is an im-
portant locaTTzed contributor of atmospheric aerosols in coastal areas.
The principal sources of particulate matter related to biota include for-
est and grassland fires, wind-blown bacteria, and biological releases such
as plant pollen and the spores of fungi (National Atmospheric Deposition
Program 1978). Atmospheric metals also originate in vapor emissions from
land, the seas, and vegetation, including the process of low-temperature
volatilization from soils (Galloway et 1981). Historic base levels of
naturally-produced particulates are very often dwarfed by anthropogenic
emissions which dominate the air pollution situation in certain regional
and localized contexts. Table 5 illustrates this point by comparing
ratios of anthropogenic to natural emissions of the predominant metal par-
ticulates on a global scale. The ratios, or mobilization factors, may be
significantly greater than these averages in large portions of the United
States since emission sources are much more concentrated than in remote
portions of the globe (Galloway et 1981).
Regional concentrations of atmospheric particulates vary widely
throughout the United States. Variations in ambient levels for different
trace metals are shown in Table 6, which presents median levels of atmos-
pheric trace metals in remote, rural, and urban areas. Particulate con-
centrations also fluctuate both diurnally and seasonally. Levels of van-
adium and nickel, for example, correlate with the use of fuel oils for
heating and are thus present in air in greatest quantities during winter
(Galloway £t aK 1981). Submicron aerosols are found to predominate in
urban and suburban areas while larger particles comprise the majority of
rural and remote samples (Miller £t aj_. 1979; Van Vaeck a_l_. 1979).
Particulate emissions declined nationally by 50 percent between 1970
and 1979, largely due to controls on industrial and utility emissions, de-
creased burning of solid waste, and reduced coal burning by small sources
(USEPA 1980a). As a result, ambient levels have shown a consistent down-
ward trend over this period. Evidence suggests, however, that local con-
centrations of particulates in the fine or respirable range are generally
increasing due to the dispersal of point sources from urban areas. For
example, visibility degradation in the southwest and eastern United States
is observed to be spreading as a result of gradual shifts in particle size
distributions toward the finer, light-scattering sizes (USEPA 1980a, b).
Estimated trends in net particulate emissions, by source through the year
2000, are given in Figure 4. These projections also suggest that, despite
controls, total atmospheric particulate burdens may increase due to the
growing number of sources.
14
-------
Table 5. Estimated global emissions of atmospheric metals
from natural and anthropogenic sources. Ratios of anthro-
pogenic to natural emissions provide mobilization factors.
Mobilization Factors
Emissions (IQ^g/yr) (Ratio of Anthropogenic
Metals Natural Anthropogenic Natural Emissions)
Ag
0.6
50
83
As
29 [210]a
780
3.3
Cd
2.9
55
19
Co
70
50
0.71
Cr
580
940
1.6
Cu
190
2,600
13
Hg
0.4 [250]a
110
0.44
Mn
6,100
3,200
0.53
Mo
11
510
45
Ni
280
980
3.5
Pb
59
20,000
340
Sb
9.8
380
39
Se
•
CO
0
1 1
Oj
140
4.7
Sn
52
430
8.3
V
650
2,100
3.2
Zn
360
8,400
23
aVo1 atile-phase emissions
(From Galloway et.il* 1981)
15
-------
Table 6. Average ambient trace metal concentrations in remote,
rural, and urban atmospheres. Ratios indicate factors of con-
centration of urban and rural to remote ambient levels.
Average Ambient Concentrations (ug/m^) Ratios
Metal Remote Rural Urban Urban/Remote Rural/Remote
Ag
0.01
0.3
1.1
uo
30
As
0.2
6
25
125
30
Be
—
0.023
0.14
-
-
Cd
0.1
1 .0
2.0
20
10
Co
0.05
0.1
10,0
200
2
Cr
0.3
5.0
40.0
133
17
Cu
0.2
6.0
100
500
30
Hg
0.5
2.0
20
40
4
Mn
0.4
30.0
150
375
75
Mo
0.3
-
2
7
-
Ni
0.36
2
30
83
6
Pb
1.0
100
2000
2000
100
Sb
0.2
3
30
150
15
Se
0.1
1.5
4.7
47
15
V
1.0
5
50
50
5
Zn
0.5
100
1000
2000
200
(From Galloway et a[. 1981)
16
-------
on
z
o
lb
o
z
o
~-I
-1
z
o
15
10 -
1975 19B5 2000
1975 1985 2000
1975 1985 2000
1975 1985 2000
1975 1985 2000
1975 1985 2000
CONSTRUCTION
MATERIALS
ELECTRIC
UTILITIES
INDUSTRIAL
COMBUSTION
STEEL
PRODUCTION
OTHER
INDUSTRIES
TOTAL
EMISSIONS
Figure 4. Estimated trends in net anthropogenic particulate emissions, by source (high growth scenario).
(From USEPA 1980b)
-------
2.3 ACIDIFYING AIR POLLUTANTS
The acidifying air pollutants include primary gaseous emissions asso-
ciated with fossil fuel combustion and derivative acids resulting from the
long-range transport and atmospheric transformation of these precursors.
Table 7 presents the air pollutants most often implicated in the occur-
rence of local (primary pollutant) and regional (secondary pollutant)
acidity effects in ecosystems.
Table 7. Primary and secondary forms
of acidifying air pollutants.
Primary
Secondary
Sulfur oxides (S0X)
Sulfuric acid (H2SO4)
Nitrogen oxides (N0X)
Nitric acid (HNO3)
Hydrochloric acid (HC1)
Sulfur forms a number of oxides (SO2, S2O3, SO3, S2O7)
but only sulfur dioxide (SO2) and sulfur trioxide (SO3) are important
as gaseous air pollutants (Urone 1976). Usually only a small amount of
SO3 accompanies SO2 (1 or 2% of the SO?) and collectively the two
are designated SO- (Stoker and Seager 1976). The nitrogen oxides
(N0X) are composed of nitric oxide (NO) andnitrogen dioxide (NO2).
These primary pollutants are transformed by atmospheric reactions to
their corresponding acids and subsequently deposited, often hundreds of
kilometers from their source, through wet or dry removal processes. Wet
deposition is generally referred to as acid precipitation and accompanies
episodes of rain, snow, hail, sleet, and fog. The dry deposition of at-
mospheric acids is also substantial, particularly during dry spells, al-
though difficulties in the measurement of this component preclude a full
understanding of its relative contribution to total acid deposition. Pol-
lutant transport, transformation, and deposition processes are discussed
further in Chapter 3.0.
Acid precipitation is defined as rain and related events of pH less
than 5.6, a reference value for atmospheric water vapor in equilibrium
with naturally occurring concentrations of carbon dioxide (CO?) and its
derivative carbonic acid (H2CO3) (Cogbill and Likens 1974; Galloway
1979; Hales 1980). The relation of acid rain to the logarithmic pH scale
is depicted in Figure 5. Acid precipitation events are characterized by
elevated concentrations of dissociated hydrogen ions (H+) which results
in lowered measured pH values.I
18
-------
The increased acidity of precipitation was first observed on a re-
gional scale in Scandinavia, and attributed to the long-range transport of
air pollutants from the industrial centers of Europe and the United King-
dom (Barrett and Brodin 1955; Bolin et aU 1971; Brosset 1973). Acid pre-
cipitation was soon documented in the northeastern United States (Likens
et al. 1972; Cogbill and Likens 1974; Likens and Bormann 1974a; Galloway
et "aT. 1976; Likens 1976), and subsequently in Florida (Brezonik et al.
T98DJ and in many regions of the west (Powers and Rambo 1981), ihcTucITng
the states of Washington (Harrison et _al_. 1977; Dethier 1979), California
(McColl and Bush 1978; Liljestrand and Morgan 1978), and Colorado (Lewis
and Grant 1980a).
limes pure baking
vinegar rain soda ammonia
¦5S
m
2 3
V
acid rain
increasing
acidity
« yi6
f5.6
6 7 8 9 10 II 12 13 14
•neutra/'
increasing
alkalinity
pH SCALE
Figure 5. The pH scale with comparisons of acid rain
to common acid and alkaline substances. Below pH 7.0,
a unit decrease in pH is equivalent to a ten-fold in-
crease in hydrogen ion concentration. (From Glass 1979a).
19
-------
The primary constituents of acid precipitation are sulfuric acid
(H SO4) and nitric acid (HNO3), although hydrochloric acid (HC1) and
a variety of organic acids are minor contributors to precipitation acidity
(Likens 1976; Likens et al. 1979; Galloway and Likens 1981). The average
composition of strong aci^s in precipitation of the northeastern United
States is shown in Figure 6.
Figure 6. Typical northeastern U.S. acid
rain components averaged annually. (From
USEPA 1980c).
20
-------
Significant geographic variations are found to occur in the relative
composition of strong acids in precipitation. Sulfuric acid is observed
to be the dominant anion in precipitation of the northeastern and north-
western United States (Likens 1976; Dethier 1979). In California and
Colorado, the sulfuric acid fraction is less significant and the nitric
acid component can assume values of 50 to 80 percent of total acids in
precipitation (Lewis and Grant 1980a; McColl 1980; Morgan and Liljestrand
1980). The nitric acid component is increasing steadily in the north-
eastern United States, particularly in urban areas (Barnes 1979; Likens
et aj_. 1979; Brezonik et aj_. 1980). From 1964 to 1979, in a calibrated
watershed of New Hampshire, the importance of H2SO4 declined 30
percent relative to HNO3 while the contribution of HNO3 to acidity
increased 50 percent relative to H2SO4 (Galloway and Likens 1981).
There are also seasonal variations in precipitation chemistry. For
example, in New Hampshire the maximum contribution of H2SO4 to acidity
observed was 73 percent in summer and 59 percent in winter, whereas HNO3
contributed at most 31 percent in summer and 61 percent in winter (Gallo-
way and Likens 1981). In the northeastern United States, the nitric acid
fraction is more abundant in precipitation during winter than summer
(Galloway 1979). The pH of precipitation is also consistently lower in
summer than in other seasons of the year (Lioy 1979; Wolff et aJL 1979).
The average annual pH of precipitation in industrialized regions of
Europe and North America lies within the range of pH 4.0 to 4.5, and is
reported to average pH 4.1 in the northeastern United States (Likens 1976;
Likens et al. 1979; Hendrey and Lipfert 1980). Individual storms may be
considerably more acidic than this average and some storms have been re-
ported in the range of pH 2.1 to 3.5 (National Atmospheric Deposition Pro-
gram 1978). Figure 7 gives the weighted mean pH of precipitation over the
continental United States, for the years 1976 to 1979, from a compilation
of data obtained through eleven precipitation chemistry monitoring net-
works. In some cases precipitation may be alkaline; pH averages above 6.0
have been reported in prairie regions where large amounts of alkaline dust
are generated by winds (Cooper et aK 1976).
Recent efforts to improve ambient air quality by constructing tall
statks to discharge gaseous and particulate emissions have resulted in an
increase in the residence time of sulfur and nitrogen oxides in the atmos-
phere. With this increased residence time comes a greater potential for
acidifying transformations and a much larger area over which the pollu-
tants can be deposited. For this reason it is difficult, if not impos-
sible, to associate elevated levels of regional precipitation acidity with
specific pollutant emission sources.
In the United States, sulfur oxides are released primarily from fos-
sil fuel combustion (USEPA 1978b). Nitrogen oxide emissions, discussed
earlier as precursors to photochemical oxidants, are produced during the
21
-------
ro
ro
Chart Plotting Lt^end
• National Atmospheric Deposition Program (NADP)
~ Dtpwiwuil of Energy (DOE)
¦ Er«viroM»entai Protection AQency(EPA/NOAA/W«tO)
-~ University of CalHonila
-O CaMomia institute of Technology
A Molt^Sut* Atmospheric Power
Production Poftrtton Study (HAP3S)
A Electric Power Reeearch Institute (EPRI)
© Oak Huge National Laboratory
O Canadton Atmospheric Environment Service (CAN SAP)
^ University o< Arizona
Q University of Florida
Figure 7. Weighted mean pH of precipitation in the continental United States (1976-1979).
(From Wisniewski and Keitz 1982).
-------
high-temperature combustion of fossil fuels. Table 8 presents estimated
annual emissions of sulfur and nitrogen oxides, by source, for the nation
while Table 9 depicts the distribution of these emissions by EPA regions.
Estimated trends in sulfur dioxide emissions through the year 2000 are
given by source in Figure 8; similar trends in nitrogen oxide emissions
were shown in Figure 2.
Table 8. Estimated national anthropogenic
emissions of acid rain precursors in 1977.
Pollutant Source
Annual Production
(10^ metric tons/year)
i°x
Mx
Transportation
0.8
9.2
Fuel Combustion
(Stationary Sources)
22.4
13.0
Industrial Processes
4.2
0.7
Solid Waste Disposal
0
0.1
Miscellaneous
0
0.1
Totals 27.4
23.1
(From USEPA 1978b)
Primary sulfate and sulfuric acid emissions are also released during
fuel combustion, augmenting ambient levels of the secondary sulfates. For
a given sulfur content of fuel, the combustion of oil yields 5 to 10 times
more primary sulfate per unit of energy than coal burning, presumably due
to catalytic oxidation of SO2 by the excessive vanadium content of fuel
oil (Homolya and Fortune 1978). Moreover, flue-gas concentrations of
H2SO4 can be 3-8 times higher in an oil-fired boiler than a coal-fired
unit. Oil burning alone is responsible for 97 percent of all primary
sulfate emissions associated with combustion processes (Nader 1980).
23
-------
Table 9. Regional emissions of sulfur and nitrogen oxides
compared to population (percent of U.S. totals).
EPA
region
States
% of U.S.
population
% of U.S.
S0X total
% of U.S.
N0X total
1
CT, ME, MA, NH,
\l T
RI,
5.6
2.1
3.4
11
NJ, NY, PR, VI
12.9
5.3
7.0
III
DE, DC, MD, PA,
WV
VA,
71.T
15.0
10.7
IV
AL, FL, GA, KY,
NC, SC, TN
MS,
16.3
21.5
17.5
V
IL, IN, MN, MI,
WI
01,
20,6
29.0
23.0
VI
AR, LA, NM, OK,
TX
1Q.3
9.0
17.C
VII
LlJ
#»
O
t/O
5.3
6.6
6.5
VIII
CO, MT, ND, SD,
WY
VT,
2.9
2.9
3.7
IX
AZ, CA, HI, NV,
AS
GU,
11.7
7.3
8.1
X
AK, ID, OR, WA
3.3
1.3
3.1
(From GCA Corporation 1981)
24
-------
50 -
40 -
P
o
30 —
X
to
2 20
to
to
£
10 -
'Gross
~ Captured
¦
Net
1975 1985 2000
RESIDENTIAL/
COMMERCIAL
1975 1965 2000
COPPER
SMELTING
1975 1985 2000
INDUSTRIAL
COMBUSTION
1975 1985 2000
PETROLEUM
REFINING
1975 1985 2000
ELECTRIC
UTILITIES
1975 1985 2000
OTHER
—100
- 90
80
- 70
- 60
50
- 40
30
- 20
- 10
1975 1985 2000
TOTAL
EMISSIONS
Figure 8. Estimated trends in anthropogenic sulfur oxide emissions by source (high growth scenario).
(From USEPA 1980b).
-------
Natural sources of S0X include the oxidation of biogenic hydrogen
sulfide (H^S), dimethyl sulfide, and other reduced volatile compounds of
sulfur, as well as the bacterial reduction of sulfate (Galloway and Whelp-
dale 1980). Such emissions can be expected where anaerobic conditions
exist in swamps, marshes, bogs, and tidal flats. Gaseous flux of biogenic
sulfur have been quantified for a variety of these sources and for differ-
ent soil orders of the United States (Hitchcock 1976; Adams et aj_. 1980;
Aneja 1980). Volcanoes and forest fires sporadically release large
amounts of sulfur oxides. Near oceans, sulfates in sea spray are an addi-
tional source of atmospheric sulfur.
Globally, biogenic and other natural emissions exceed anthropogenic
emissions by a factor of two; man-made emissions are nevertheless project-
ed to equal or exceed naturally-derived atmospheric sulfur on a global
basis by the year 2000 (Kellogg et aK 1972; Cullis and Hirschler 1980).
On a regional scale, natural sources are reported to currently account for
about 4 percent of total sulfur emissions in eastern North America (Gal-
loway and Whelpdale 1980).
Sulfur dioxide and sulfate concentrations are greatest in urban and
industrial areas of high emission density. Sulfate levels can vary from
in excess of 80 ug/rrr in highly polluted localities to concentrations as
low as 0.04 ug/nr in remote areas (Whitby 1978).! Annual sulfate aver-
ages of 14-20 ug/nr are common in urban areas, while non-urban sites of
the northeastern United States average at least 5 ug/m^ (Altshuller
1976). Sulfate concentrations frequently exceed 15 ug/nF in the Ohio
River Valley and average 10-15 ug/m3 over large parts of the eastern
United States (Husar et al. 1978). As shown in Figure 9, ambient sulfate
levels are generally TvTgFest in summer (MacCracken 1979; Altshuller 1980).
Sulfur dioxide concentrations in urban air decreased 67 percent in
the period from 1964 to 1970; a 7 percent decline in SO2 emissions be-
tween 1970 and 1979 accompanied a 44 percent drop in national average
SO2 levels (USEPA 1980a). Improved ambient air quality has resulted
from restrictions of the sulfur content of fuels, better controls on ex-
isting sources, dispersal of sources from urban areas, and the building of
taller stacks. Nevertheless, the proliferation of point sources in rural
and remote areas has prevented the widespread attainment of federal SO2
standards (USEPA 1980a). Emissions are remaining at high levels (Glass et^
al. 1978; Glass 1979b) and are projected to increase both nationally (Alt-
shuller and McBean 1980; USEPA 1980b) and globally (Barnes 1979). Furth-
ermore, reported declines in S0p levels have been accompanied by only
modest decreases in concentrations of ambient sulfates (Altshuller 1980).
26
-------
70
ro
60
°2 50
X
CO
UJ
<10
r »
m
O*
m
30
20
10
O ITHACA
a PENN STATE
o VIRGINIA
_L
_L
SEP OCT NOV DEC
1977
JAN FEB MAR APR MAY J UN JUL
1978
Figure 9. Monthly mean concentrations of sulfate as a function of time
at Cornell University (Ithaca), Pennsylvania State University and the
University of Virginia. (From MacCracken 1979).
-------
3.0 ECOSYSTEM EXPOSURE TO AIR POLLUTION
Ambient concentrations of air pollution are generally considered to
be a function of rates of emission, dispersion, transport, transformation,
and deposition. In the past, ecosystem exposures via these processes were
only thought to occur in the immediate vicinity of point and area sources.
Air pollution, in the form or urban or industrial plumes, was regarded
strictly as a discrete, localized phenomenon causing frequent violations
of air quality standards at ground level. Efforts to improve local air
quality resulted in more efficient pollutant dispersion, longer transport
trajectories, and an increased residence time during which chemical trans-
formation of anthropogenic pollution can take place. Evidence of improved
ground-level air quality led to general assumptions that ecosystem expo-
sures were on the decline.
Ambient air concentrations continue to be determined by these same
fundamental processes, however, in the face of increased pollutant disper-
sion, the range and complexity of these processes has greatly expanded and
they now take place on a regional, and even intercontinental scale. Along
with growing recognition of the importance of LRTAP, and the complex in-
tegration of air pollutants within global biogeochemical cycling, has come
the understanding that injurious substances may be deposited over vast
regions, often entering remote ecosystems in quantities equal to or ex-
ceeding those deposited near sources. This understanding has led to wide-
spread concern that living resources and their supporting habitats are
subject to impact on a scale never before anticipated. This chapter de-
scribes in very general terms the atmospheric transport, transformation,
and deposition processes that lead to widespread ecosystem exposures to
air pollution. It also addresses the different pathways by which living
ecosystem components (i.e., fish, wildlife, and their habitats) are ex-
posed to deposited air pollutants.
3.1 ATMOSPHERIC TRANSPORT AND TRANSFORMATION
Atmospheric movements affecting pollutant mixing and transport range
from small gusts and eddies, a few meters wide and of minimal duration, to
wind systems that may extend over thousands of kilometers and last for 24
hours or more (Bolin et al. 1971). The long-range transport of air pollu-
tants reduces specific pFTlutant concentrations through mixing and deposi-
tion processes while allowing sufficient time for the transformation of
primary pollutants to secondary ones. The distance a pollutant travels
and potential transformations which can occur depend in large part on
site-specific factors such as microclimate and topography, as well as the
residence time of the pollutant in the atmosphere, the presence of cata-
lysts, and other factors affecting the efficiency of photochemical reac-
tions.
The formation of photochemical oxidants is a cyclical process, re-
curring daily in regions where there is inadequate air movement and dis-
persion of primary pollutants. The reactions producing the oxidants in-
volve the oxidation of hydrocarbons in the presence of nitrogen oxides,
28
-------
sunlight, and several oxidizing agents, of which the hydroxyl radical
(OH*) is the most important. Although extremely complex, the chemical
reactions that take place can be summarized as follows (Finlayson and
Pitts 1976):
N0X + HC—UV >03 + PAN + H202 + ...
where
HC = hydrocarbon
N0X = oxides of nitrogen
03 = ozone
PAN = peroxyacetylnitrate
UV = solar ultraviolet radiation providing the energy required
for the reaction
These substances are subject to short-range (urban-scale), inter-
mediate-range (mesoscale), and long-range (synoptic scale) transport, dur-
ing which ozone production per unit of precursor may actually be enhanced
(USEPA 1978a). More detailed information on the chemistry involved in
photochemical oxidant formation is provided by Altshuller and Bufalini
(1971), the National Research Council (1977), and the U.S. EPA (1978a).
The atmospheric reactivity of particulate matter is as diverse as its
chemical composition. Moreover, a significant portion of ambient concen-
trations, called secondary particulate matter, is formed by the conversion
of gaseous pollutants, already present in the atmosphere, to liquid aero-
sols or solid particles. For example, a variety of secondary trace organ-
ic compounds are formed by photochemical reactions involving the primary
hydrocarbons. Over twenty different organic micropollutants, some of them
known carcinogens, have been identified in polluted air masses from west-
ern Europe (Lunde and Bjorseth 1977; Lunde et al. 1977). Most of these
gas-to-particle conversion processes are depencfent on the presence and in-
tensity of solar radiation, the relative humidity and the concentration of
oxidizing substances in the air (National Atmospheric Deposition Program
1978).
The predominant reactants and physico-chemical processes involved in
particulate transformations are depicted in Figure 10. These processes
are related to particle size in order to indicate likely deposition pro-
cesses and the potential for long-range transport. The coarse particles,
along with materials adsorbed to them, are quickly deposited by gravita-
tional settling while the finer fractions may be carried considerable dis-
tances from their emission sources. Some of the fine particles, notably
29
-------
Figure 10. Particulate transformation and deposition processes in
relation to particle size. (Adapted from Whitby 1978).
30
-------
sulfates and nitrates, are hygroscopic in nature: they possess the abil-
ity to accelerate the condensation of water vapor. This property can
simultaneously affect their size, shape, reactivity, and pH. An excellent
review of particulate movement and reactivity in the atmosphere has been
prepared by Whitby (1978).
The primary acidifying gases, oxides of sulfur and nitrogen, can also
be transported great distances. While airborne they are transformed in
the atmosphere to secondary acidifying particulates. These aerosols may
be transported hundreds of kilometers further than the primary pollutants
since they deposit at much slower rates (Barnes 1979).
In clean air, SO2 oxidizes slowly to form sulfur trioxide (SO3),
which is generally present in only minor amounts in the atmosphere because
it reacts rapidly with moisture to form sulfuric acid (H2SO4). The
reactions producing sulfuric acid can be summarized as follows (Vermeulen
1978):
Dovland and Semb 1980; McMurray 1980):
• homogeneous oxidation processes, or the photochemical oxidation
of SO2 gas by thermally produced reactants or photochemically
generated free radicals (e.g., OH*, •HO2); and
• heterogeneous oxidation processes, or the catalytic oxidation of
S02 adsorbed to aqueous aerosols by metal ions (e.g., iron,
manganese) in solution.
Both of these conversion processes produce aerosols in the sub-micron size
range.
Atmospheric nitrates result from similar reactions which oxidize
nitrogen oxide (NO) and nitrogen dioxide (NOg) to nitric acid (HNO3)
and other inorganic and organic nitrates. The following general mechanism
for the conversion of N0X to nitric acid has been postulated (Vermeulen
1978):
31
-------
In addition, nitrate salts may be formed in the atmosphere through a var-
iety of reactions. The direct homogeneous capture of gaseous nitric acid
by gaseous ammonia:
NH3 + HN03 >NH4N03
may be a significant source of ammonium nitrate salt in the atmosphere if
ammonia levels in the polluted atmosphere are sufficiently high (National
Research Council 1977). Nitric acid also forms some mixed compounds or
complexes of considerable stability in sulfuric acid solutions, and may
become incorporated within sulfuric acid droplets at the lower tempera-
tures of the upper atmosphere. In general, much less is known of exact
mechanisms of nitrate formation as most research has focused specifically
on processes of atmospheric sulfate formation.
Considerable research has been undertaken to quantify rates of sulfur
dioxide conversion to sulfate in ambient air, with widely varying results
(MacCracken 1979). Ratios of sulfur dioxide to sulfate in urban and in-
dustrial plumes have been suggested as indicators of SO? transformation
rates; these ratios are greatest near emission sources and may serve as a
surrogate measure of the age of plumes (Dovland and Semb 1980). Homogen-
eous oxidation of SO? to SO4 ~ can occur at rates up to 1.0 percent
per hour in clean air during summer and may reach values of 5 percent per
hour in irradiated urban air (Eggleton and Cox 1978). Homogeneous oxida-
tion in plumes of coal-fired power plants typically occurs at rates less
than 1.0 percent per hour in clean air, however diurnal variations also
occur and maximum rates of 3 percent per hour have been recorded at mid-
day (Newman, L. 1980). Conversion rates for heterogeneous oxidation are
similar, averaging 1 to 2 percent in unpolluted air and 2 to 6 percent in
urban atmospheres (Dovland and Semb 1980). However, known reactions alone
are insufficient to explain SO? conversion to SO4 ~ in the patterns
and quantities observed, suggesting that one or several unknown mechanisms
may play a significant role in secondary pollutant formation (Budiansky
1980).
Comparisons of SO? emissions with ambient sulfate concentrations in
specific regions indicate that sulfate distribution is determined primar-
ily by transport and transformation processes and not local sources of
SO? (Altshuller 1976, 1980). Air parcel trajectory analysis has been
employed to determine the origins of elevated sulfate concentrations in
ambient air. The long-range transport of air pollutants has been found
to be greatest in air masses that have stagnated for several days over
heavily industrialized regions. In New York, elevated sulfate concentra-
tions and precipitation acidity have been correlated primarily with west-
erly and southwesterly transport from Canada and the Ohio River Valley
(Galvin _et al. 1978), while in Scandinavia they have been associated with
southerly arul southwesterly transport from industrial centers of western
Europe and the United Kingdom (Rodhe et al. 1972).
32
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3.2 ATMOSPHERIC DEPOSITION PROCESSES
The removal of primary and secondary pollutants from the atmosphere
occurs through complex physical and chemical processes referred to as at-
mospheric deposition, which may be defined as the transfer from air to
ground of any gas or particle via wet and dry removal processes; acid dep-
osition refers to transfers of strictly acidic substances by these same
removal processes (Interagency Task Force on Acid Precipitation 1981}.
Wet deposition is defined as the amount of material removed from the
atmosphere by rain, snow, or other precipitation forms; the term is also
used to refer to the process by which gases, liquids, or solids are trans-
ferred to the ground during a precipitation event (Interagency Task Force
on Acid Precipitation 1981}. Two separate removal processes are involved
(Hales 1972; Fowler 1980a, b):
• rainout (or snowout), in which pollutants serve as condensation
nuclei or are incorporated into hydrometeors, or water vapor con-
densates, before they begin to fall; and
• washout, the incorporation of pollutants into falling hydrometeors
below the cloud.
Collectively, these processes are referred to as precipitation scavenging
(SI inn 1977}.
The concentration of pollutants within droplets depends on the ini-
tial concentration in the collector particles, pollutant concentrations in
droplets added to the initial particle, find amounts lost prior to impact
(Hales 1972]. The time availaole for mixing in contaminated atmospneres
and the rate at which vapor droplets are scavenged from the cloud by
larger raindrops both affect the volume of pollutants depositee with pre-
cipitation .
As precipitation nears the earth's surface, it may be characterized
in several ways (Eaton et_ aj_. 1973; Galloway and Parker 1980):
t incident wet deposition is the rainfall transferred directly to
the ground by gravitational forces;
• throughfall is rain that has passed through a leaf canopy;
• net throughfall is the chemical composition of water that has
passed through a leaf canopy (i.e., the sum of materials in inci-
dent wet deposition plus substances leached or washed from leaves
minus materials retained within the leaf canopy); and
t stemflow is rainwater which flows down the branches and trunks of
vegetation.
33
-------
The special event of fog is another wet removal process which may be
significant in the deposition of air pollutants. Fogs are known to be
excellent scavengers of acid aerosols (Smith et _al_. 1980). Studies have
shown that the acidity of cloud water often exceeds that of precipitation,
suggesting that frequent fog, for example in mountainous areas, may be a
significant source of acidic inputs (MacCracken 1979).
Wet deposition processes also remove metal and trace elements from
the atmosphere. Exceptions are the vapor-phase metals and organics for
which gaseous adsorption is the dominant removal mechanism (Galloway et
al. 1981). Table 10 quantifies average concentrations of trace metals in
wet deposition from remote, rural, and urban areas. The ratios provided
in Table 10 indicate concentration factors by which urban and rural quan-
tities exceed those in remote areas.
Concentrations of sulfuric and nitric acid in precipitation have been
monitored during the past several years by a number of national and re-
gional precipitation chemistry networks. Patterns have emerged which per-
mit the characterization of trends in acid precipitation for all regions
of the continental United States. Weighted mean pH values of precipita-
tion across the United States for the period 1976-1979 were depicted in
Figure 7. Mean annual precipitation patterns, when combined with these pH
determinations, permit the calculation of mean annual hydrogen ion deposi-
tion in precipitation, as shown in Figure 11.
Dry deposition is defined as the aggregate of all materials trans-
ferred from the atmosphere to natural surfaces in the absence of precipi-
tation; its definition also includes the physical processes of transfer
(Interagency Task Force on Acid Precipitation 1981). Dry removal pro-
cesses include (Fowler 1980a, b; Galloway and Parker 1980):
• dry fallout (or sedimentation), the gravitational settling of
particles greater than 10 urn in diameter;
• aerosol impaction, the wind driven deposition of sub-micron
materials; and
• gaseous adsorption, the natural attraction between gases and
solid or liquid surfaces.
Because it is easy to collect precipitation, the wet deposition of
atmospheric pollutants has generally been easier to characterize on a
qualitative and quantitative basis than dry deposition, although many
questions remain about the exact nature of both processes (Galloway et al.
1979; Galloway and Parker 1980). Rates of dry deposition are virtually
impossible to predict with any degree of assurance since they are con-
trolled by numerous factors including some related to the pollutant, such
as size, shape, and chemical properties, as well as meteorological condi-
tions (McMahon et al. 1976; Sehmel 1980). Dry deposition rates cannot be
measured with certainty because the amount of material transferred will
also depend on the nature of the receiving surface.
34
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Table 10. Average concentrations of metals in wet deposition.
Ratios at remote, rural, and urban sites provided are factors
of concentration of urban and rural to remote levels.
Concentrations (ug/1) Ratios
Metal
Remote
Rural
Urban
Urban/Remote
Rural/Remote
Ag
0.008
0.023
3.2
400
3
As
0.019
0.04
-
-
2
Cd
0.008
0.6
0.7
87
75
Co
-
0.01
1.8
-
-
Cr
-
0.27
3.6
-
-
Cu
0.055
5.3
30
554
96
Hg
0.048
0.11
1.0
21
2
Mn
0.22
10.0
25
113
45
Mo
-
-
0.2
-
-
Ni
0.1
1.5
17
170
15
Pb
0.14
15
41
292
110
Sb
0.034
-
-
-
-
V
0.022
1.1
68
3,090
50
Zn
0.22
45
40
181
200
(From Galloway et aj_. 1981)
35
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CO
0*5
• National Atmospheric Deposition Program (NADP)
a Department of Energy (DOE)
¦ Environmental Protection Agency (EPA/NOAA/WMO)
University of California
o- California Institute of Technology
a Multi-State Atmospheric Power
Production Pollution Study (MAP3S)
A Electric Power Research Institute (EPRI)
© Oak Ridge National Laboratory
O Canadian Atmospheric Environment Service (CANSAP)
~ University of Arizona
0 University of Florida
Figure 11. Mean annual hydrogen ion (H+) deposition in precipitation over the continental United States,
1976-1979 (Kg/ha). (From Wisniewski and Keitz 1982)
-------
Despite the high degree of uncertainty in the understanding of dry
deposition processes, it is known that they are quite important. Adsorp-
tion is the most important removal process for gaseous pollutants, includ-
ing S0X, N0X, and HNCh. Particulate matter is deposited by impac-
tion and gravitational settling. In arid lands, as would be expected, dry
deposition processes are the dominant mechanism for pollutant deposition,
but in non-arid lands as well, the amount of matter deposited via dry re-
moval processes may exceed the amount deposited by wet removal processes
(Gravenhorst ert aj_. 1980).
Special events of dew and frost must be considered in connection with
dry depositon processes because of their potential for increasing the
acidity of deposited pollutants. Dews, for example, dissolve previous
gaseous and particulate deposits on plant surfaces and enhance their
transformation to strong acids, leading to a rapid decrease in the pH of
the wetted plant surfaces. The contribution of frost to acidity is only
significant when it melts and assumes the characteristics of dew because
ice does not usually incorporate foreign substancesinto its crystalline
structure (Smith et a]_. 1980).
3.3 PATHWAYS OF AIR POLLUTION EXPOSURE TO FISH, WILDLIFE, AND THEIR
HABITATS
Atmospheric contaminants enter the different compartments of aquatic
and terrestrial ecosystems through a variety of pathways. The degree of
exposure depends on pollutant emission rates, proximity to emission
sources, characteristics of pollutant transport, transformation and depo-
sition, as well as modes of pollutant uptake and distribution within
biotic or abiotic compartments. This section discusses pathways by which
fish, wildlife and living components of their habitats are physically
contacted by atmospheric pollutants.
The most direct exposures of terrestrial wildlife to air pollution
occur through the inhalation of gases and particulates. Fine particles
(<1.0 um in diameter), like gases, bypass respiratory filters and can be
deposited deep within animal lungs (Davidson et aK 1974; Natusch et al.
1974). Airborne particulates and reactive gases, such as ozone and sulfur
dioxide, may adsorb to the eyes of animals (Newman 1980). Direct expo-
sures of terrestrial vertebrates to acid precipitation are of little con-
sequence due to the dilute nature of dissolved acids and the protection
afforded body surfaces by scales, feathers, or fur.
Terrestrial wildlife may be indirectly exposed to air pollutants by
ingesting contaminated vegetation. Examples of such exposures include the
ingestion of plants with fluorides or arsenic deposited on their surfaces
(Lerman and Darley 1975). Animals may also ingest air pollutants with
drinking water. Similarly, indirect exposures can occur from the inges-
tion of food organisms that have accumulated various air pollutants
(Stickel 1975).
37
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Terrestrial plants comprising wildlife habitats are exposed to atmos-
pheric pollution through contact with above- and below-ground plant parts.
Exposure pathways depend on whether the pollutant is in gaseous or partic-
ulate form. Gaseous air pollutants contact internal leaf tissues after
entry through leaf stomata. This exposure mechanism has been verified for
sulfur dioxide (Mudd 1975a), nitrogen oxides {Taylor £t aK 1975), and
ozone (Heath 1975). Plant exposure to peroxyacetylnitrate has not been
thoroughly studied, however passage through leaf stomata appears important
(Mudd 1975b). Soils may act as a sink for gaseous pollutants which are
adsorbed by microorganisms or taken up by plants.
Aerial deposition is the primary mechanism of plant exposure to par-
ticulates (Lerman and Darley 1975). Fine particles, including trace ele-
ments, fluorides, and acid aerosols, are able to penetrate leaf tissues
via opened stomata (Chang 1975). Coarse particulates collect on leaf sur-
faces and may form hardened crusts through interactions with moisture on
leaf and stem surfaces. Acidic substances dry-deposited on leaves can be
dissolved by dew or frost (Smith et aK 1980), and these, along with
dilute acids in precipitation, coTTect in leaf depressions and enter via
epithelial cracks near specialized structures (e.g., stomata, veins, and
trichomes) of the leaf surface (Evans 1980).
Soil microorganisms may be exposed to particulate matter deposited on
soils; these substances may also be taken up by plant roots. Field stud-
ies have shown that approximately 90 percent of deposited metals are re-
tained in the top 15 centimeters of soils (Buchauer 1973). There they may
be readily absorbed by roots. Plants absorb and accumulate both essential
and non-essential trace elements from dilute soil solutions. Root expo-
sures may be further augmented by soil acidification and other pollutant-
induced alterations that increase the biological availability of trace
contaminants (Zimdahl 1975). Aerial portions of plants may ultimately be
affected by pollutants absorbed through the roots if the pollutants are
readily translocated within the plant.
Like plant roots and soil microbes, freshwater fauna may be exposed
to relatively unchanged forms of air pollutants, their transformation pro-
ducts and substances mobilized by pollution effects. For example, an in-
creased availability of toxic metals will result directly from wet and dry
deposition as well as indirectly through increased releases from sediments
and the watershed caused by acid precipitation (Beamish and Van Loon
1977). Trace elements found to increase in acidified waters include alu-
minum, cadmium, iron, magnesium, and manganese (Hall and Likens 1980; Hall
et al. 1980).
Aquatic organisms are exposed to these same types of pollutants pri-
marily through adsorption and uptake across gills and other permeable mem-
branes. This mechanism is particularly important for hydrogen ions, toxic
metals (e.g., aluminum and mercury), and the lipophilic organic substances
(Muniz and Lievestad 1980a). Other exposures to air pollutants occur
through the ingestion of contaminated food, water, detritus or sediment.
38
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Elevated exposures can be expected to coincide with spring snowmelt or in-
tense precipitation events; the acidity associated with such events in-
creases the abundance of biologically available forms of metals (Harvey
1979; Troutman and Peters 1980). Aerial exposures to ambient air pollu-
tion may be important for those organisms which inhabit or feed at the
water's surface.
Air pollutants contact emergent aquatic vegetation by direct deposi-
tion on plant surfaces or by entry through stomata. Submergent and float-
ing forms may adsorb contaminants in the water column to exterior surfaces
and transport them across membranes. For example, trace metals may be
adsorbed to cell walls and transported through them to internal plant tis-
sues (Dvorak et_ aj_. 1978). Rooted forms may also absorb and translocate
atmospheric contaminants trapped in sediments.
Exposure pathways are key determinants of the nature of biotic re-
sponses to air pollutants. Depending on its mode of entry and site of
deposition, a given pollutant may affect an organism in different ways.
Therefore, assessments of the biological impacts of air pollution and acid
rain should commence with a determination of the relative importance of
various exposure pathways among the organisms to be investigated.
39
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4.0 FACTORS AFFECTING ECOSYSTEM SENSITIVITY
Many interacting variables determine the vulnerability of aquatic and
terrestrial ecosystems to air pollution and acid rain phenomena. They are
of two basic types:
• factors related to the intensity and distribution of pollutant
loadings to ecosystems; and
• factors related to the inherent susceptibility of ecosystem struc-
ture and function to alterations induced by atmospheric deposi-
tion.
Ecosystem impact is usually, but not always, associated with the oc-
currence of both types of factors. Inherently susceptible ecosystems, for
example, may be defined as sensitive even in the absence of pollutant dep-
osition because they would be subject to alteration in the event that air-
borne loadings increase.
The primary factors regulating ecosystem sensitivity are (National
Atmospheric Deposition Program 1978; USEPA 1980c):
• meteorology;
• geology;
• pedology;
• hydrology;
• hydrochemistry;
• topography;
• biota; and
• human activity.
Each of these factors influences the extent of pollutant loadings or
ecosystem responses. Together, they may render certain ecosystems much
more vulnerable to adverse impacts than others. In general, the presence
or absence of these factors largely determines whether or not biological
effects will occur; the complexity of their interactions moderates biotic
effects in such a way that consistent patterns are difficult to establish
(Shriner 1979).
40
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4.1 METEOROLOGY
Meteorological factors contributing to ecosystem sensitivity are both
complex and varied. In all ecosystems, the dilution and dispersion of
primary pollutants are regulated by convection, turbulence and wind pat-
terns. Other influences include atmospheric stability and the frequency
of periodic thermal inversions. These factors determine the residence
time of pollutants in the atmosphere which in turn affects their distribu-
tion and long-range transport as well as the probability of their conver-
sion into secondary pollutants.
Humidity and solar radiation levels have a strong influence on the
rate of conversion of atmospheric gases to acids. At the time of deposi-
tion, wind and humidity together affect rates of gaseous adsorption and
particle impaction on terrestrial surfaces. Precipitation amounts and
frequencies, as well as pH and chemistry, determine the extent of ecosys-
tem pollution through rainout and washout. Seasonal variations in dry
deposition, rainfall, snowfall, and snowmelt can greatly influence the
magnitude of atmospheric pollutant inputs. Special events of dew, frost
and fog can also significantly augment pollutant loadings in areas of
frequent occurrence (Smith e^t a_h 1980). Moreover, snow, rain and special
events possess differing efficiencies in the removal and concentration of
airborne acids and other pollutants.
In all terrestrial ecosystems, the ratio of precipitation to evapo-
transpiration influences the extent of mineral release from soils through
leaching (National Atmospheric Deposition Program 1978). In the eastern
United States, where precipitation generally exceeds water losses to the
atmosphere, surface runoff and soil percolation may transport significant
amounts of leachates and unbuffered pollutant inputs from terrestrial to
aquatic ecosystems.
Several meteorological variables have been demonstrated or hypothe-
sized to mediate the extent of plant injury from ambient air pollution and
atmospheric deposition. These include (Evans 1979a; Shriner 1979):
• temperature;
• light intensity;
• relative humidity;
• duration and intensity of precipitation; and
• the time between precipitation events (i.e., the time available
for the accumulation of dry deposition).
Meteorological factors also influence diurnal and seasonal variations
in the sensitivity of plants to air pollutant loadings (Guderian 1977).
41
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In general, the direction of prevailing winds, the trajectories of
polluted air masses and the nature of precipitation patterns exert the
greatest influence on the magnitude of ecosystem loadings due to acid dep-
osition. These, along with other meteorological factors discussed above,
lead to disproportionately large amounts of atmospheric deposition in
mountainous regions, notably the Adirondacks of New York, the Sierras of
California, the Rockies of Colorado and the Appalachian range.
4.2 GEOLOGY
The geological characteristics of an ecosystem are generally consid-
ered to be the primary determinant of sensitivity to acid deposition
(Voigt 1980). Bedrock geology is especially significant where soils are
thin and runoff is substantial. In certain localities, however, soils may
largely override the influence of geological substrates (Hendrey et al.
1980a).
The extent of ecosystem impact from acid rain is determined by the
nature of dominant rock types as these substrates have widely varying cap-
acities to neutralize acid. Granite, quartz, basalt, and other alumino-
silicate materials, normally the most resistant to physical and chemical
weathering, are least capable of adequately buffering acidic inputs
(Abrahamsen et a}. 1979; Likens et al. 1979; Hendrey et cH. 1980a,b; Nor-
ton 1980). The result is an accelerated leaching of nutrient cations,
notably sodium, calcium, potassium and magnesium, as well as toxic metals
such as mercury and aluminum (Webb 1980). In fact, ecosystem losses of
calcium, magnesium and aluminum closely correlate with the amount of hy-
drogen ions passing through the watershed, and enrichment of drainage
waters with these substances may result (National Atmospheric Deposition
Program 1978).
A classification of rock types, based on their capacity to buffer
acid precipitation, is presented in Table 11. Figures 12 and 13 depict
two different methods of mapping geologically sensitive regions utilizing
this classification. Type 1 rocks are associated with extensive aquatic
ecosystem impact from acid precipitation while Type 2 substrates generally
facilitate the acidification of first- and second-order streams and head-
water lakes. Type 3 rocks are not thought to contribute to aquatic acid-
ification except where surface run-off drains over frozen ground. Type 4
rocks are considered to have "infinite" buffering capacity and for this
reason are in no way associated with aquatic ecosystem effects. Thus,
bedrock geology correlates directly with the acidification of waterways
and determines the extent of acid rain damage to aquatic ecosystems
(Hendrey elt aj[. 1980a; Norton 1980).
Earlier efforts to map geologically sensitive areas of the United
States, performed by Galloway and Cowling (1978) and Likens et j^. (1979),
were based on large scale geological maps (Figure 14). From these and
more recent studies, it has become apparent that significant portions of
the country are underlain by geologically sensitive bedrock. Those areas
of the U.S. underlain by granite bedrock include the Adirondack and
42
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Table 11. Classification of rock types used
to distinguish geological sensitivity.
Type 1 Low to no buffering capacity: Granite/syenite, granite
gneisses, quartz sandstones, or metamorphic equivalents.
Type 2 Medium to low buffering capacity: Sandstones, shales,
conglomerates, high-grade metamorphic felsite to intermediate
volcanic rocks, intermediate igneous rocks, calc-si1icate
gneisses with no free carbonates.
Type 3 Medium to high buffering capacity: Slightly calcareous, low
grade intermediate to mafic volcanic rocks, ultramafic and
glassy volcanic rocks.
Type 4 High buffering capacity: Highly fossi1iferous sediments or
metamorphic equivalents, limestones or dolostones.
(From Hendrey et al_. 1980a, b; Norton 1980)
Appalachian mountains and large parts of New England, Michigan, Wisconsin,
Minnesota, Washington, Idaho, Oregon, California and Colorado (Interagency
Task Force on Acid Precipitation 1981). Coastal regions with sand sub-
strates are also known to be vulnerable to aquatic and terrestrial eco-
system acidification (Glass et aj_. 1980). A large part of eastern New
Jersey is reported to have acidTfied streams and groundwaters where sand
substrates are deep (Johnson, A. 1979a, b).
4.3 PEDOLOGY
The ability of soils to neutralize acidification, retain other con-
taminants and prevent pollutant transfers to aquatic ecosystems depends
primarily on soil type, depth, age, parent bedrock, carbonate and organic
content, cation exchange capacity, glaciation history and the influences
of vegetation. For purposes of identifying vulnerable soils, however, the
following parameters are considered to be the most important (McFee
1980a, b; Klopatek et_ ah 1980):
• soil acidity (pH), which approximates the cation replacement
efficiency of additional hydrogen ion inputs;
• cation exchange capacity (CEC), or the total buffering potential
provided by the exchange sites of clay and organic matter;
43
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Figure 12. Areas of differing acid-neutralizing capabilities in
New York State, according to the geological buffering classifica-
tion of Table 11. (From Hendrey et a_L 1980a)
-------
Figure 13. Vulnerability map for the state
of Georgia, by county. The first number
indicates the percentage of rock type 1 (0,
0-9%; 1, 10-19%; 2, 20-29%; etc.) The second
number indicates the percentage of rock type
2 (refer to Table 11 for rock type classifica-
tions). (From Hendrey et al_. 1980a).
Figure 14. Geologically sensi-
tive regions of North America
with lakes susceptible to acid
precipitation. (From Galloway and
Cowling 1978).
45
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• base saturation of exchange capacity (BS), a measure of the degree
of soil development, and a function of pH;
• soil management systems, including cultivation, fertilization and
liming, as well as renewal through flooding and other processes;
and
t the presence or absence of carbonates within the soil profile.
Vulnerable soils are generally defined as those with potential for
decreased productivity in response to acid deposition. They are charac-
terized by low CEC; hence reduced clay and organic content, and fewer
sites for proton exchange (McFee 1979, 1980a, b). Sensitive soils possess
medium to high base saturation (30-50%) and pH values greater than 5
(Klopatek et_ aj_. 1980). This is because soils of low base saturation
(podzols) have low leaching rates while soils of high base saturation
(loams) have high leaching rates (Abrahamsen 1979). Like susceptible
bedrock, sensitive soils are prone to accelerated leaching of nutrient and
metal cations. The hydrogen ions in acid precipitation replace other
cations in the soil, notably calcium, magnesium, iron and aluminum, which
are dissolved in surface run-off and concentrated in the hydrographic
network.
A classification of soil sensitivity to acid precipitation is pre-
sented in Table 12. Its application to the mapping of sensitive soil re-
gions is exemplified in Figure 15. Noncalcareous, sandy soils of pH 5 are
generally agreed to be the most sensitive to acid precipitation (McFee
1980a, b; Wiklander 1980). Mature soils that are naturally acidic are
seldom subjected to further acidification from atmospheric deposition, yet
they are vulnerable to accelerated leaching, particularly of aluminum,
which can have long-term consequences for their sustained productivity
(McFee 1980b). These and other soil types are compared in Table 13 on the
basis of two major sensitivity parameters. Petersen (1980) provides a
detailed discussion of soil sensitivity based on recognized national and
international soil classifications.
Remaining soil sensitivity parameters include soil moisture content
and the extent of soil-water mixing as water percolates through the soil
profile (Voigt 1980). These factors, as well as soil texture, depth and
porosity, are particularly important, for if acidic precipitation cannot
achieve chemical equilibrium with soils, downstream water bodies will be
subject to considerable hydrochemical alteration (Bache 1980). Factors
mediating rates of soil acidification include (Wiklander 1980):
• calcium carbonate (CaC03) content;
• ferro-magnesium mineral content;
• soil texture;
46
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Table 12. Classification of soil sensitivity to
acid precipitation based on cation input and the
chemical characteristics of the top 25 cm.
Sensitivity Class
Cation Exchange
Capacity meq/lOOg
Other Relevant
Conditions
Non-sensitive (NS)
Any value or
>15.4
Free carbonates
present or subject
to frequent flooding
None.
Slightly sensitive (SS)
6.2 < CEC < 15.4
Free carbonates
absent; not subject
to frequent flooding
Sensitive (S)
> 6.2
Free carbonates
absent; not subject
to frequent flooding
Source: McFee (1980a, b)
Table 13. Soil sensitivity to acid precipitation based
on buffering capacity and hydrogen ion retention.
Calcereous
soils
Noncalcereous
clays
pH > 6
Noncalcareous
sandy soils
pH > 6
Cultivated
soils
dH > 5
Acid
soils
pH > 5
Buffering
capacity
Very high
High
Low
High
Moderate
Hydrogen
ion
retention
Maximal
Great
Great
Great
SIight
Overal1
sensitivity
None
Moderate
Considerable
None or
siight
SIight
(From Wiklander 1979)
47
-------
Key: NS - The area contains mostly non-sensitive areas
51 - Sensitive soils dominate the area
551 - Slightly sensitive soils dominate the area
52 - Sensitive soils are significant, but cover less than 50% of
the area
552 - Slightly sensitive soils are significant, but cover less
than 50% of the area
Figure 15. An example of soil sensitivity mapping based
on the soil sensitivity classification of Table 12.
(From McFee 1980b)
48
-------
• nature of soil litter; and
• water permeability.
Sulfate adsorption is a recently identified soil sensitivity parame-
ter of potential importance, yet more research is needed to clarify its
role in the prediction of soil vulnerability. In general, it has been
demonstrated that the leaching of sulfate and associated cations is depen-
dent on sulfate mobility, or, conversely, the sulfate absorption capacity
of soil (Johnson and Cole 1977; Johnson e_t jH. 1980; Singh 1980). Sul-
fates in precipitation are known to pass rapidly through the top soil hor-
izons and adsorb preferentially in the lower horizons (Singh 1980).
As a rule, excess positive charges in the soil will neutralize ad-
sorbed negative ions such as sulfate. As soils acidify, however, deposi-
ted hydrogen ions can replace these excess soil cations, freeing them to
bind with mobile sulfates. Thus, if sulfate loadings are greater than the
capacity of the ecosystem to utilize or store it, sulfate leaching will
occur, cations will be mobilized and the acidity of soil will increase
(Reuss 1976, 1980). Moreoever, as the sulfate adsorption capacity of soil
becomes saturated, further sulfate inputs will remain mobile and are sub-
ject to rapid transfer, along with associated cations, to the hydrographic
network.
Sulfate adsorption in soils is enhanced by such factors as increased
temperature, long equilibrium periods, decreased soil moisture, low soil
pH and moderated rates of sulfate loading (Singh 1980). The presence of
free iron and aluminum oxides also promotes sulfate adsorption while ele-
vated concentrations of organic matter increase sulfate mobility by block-
ing available adsorption sites (Johnson 1980; Johnson et a2- 1980). Sul-
fate mobility is difficult to predict, however, since any number of these
mediating factors may co-exist in the same soil.
Large portions of the eastern United States possess soils that may be
sensitive to acid precipitation (Glass et a_h 1980; Klopatek et al_. 1980;
McFee 1980a,b). These include the shallow and steep soils of-the Adiron-
dack and Appalachian mountains as well as the coarse, noncalcareous gla-
cial tills of New England. Moreover, considerable portions of the south-
eastern U.S. are located on highly weathered soils that are particularly
vulnerable to the leaching effects of acid precipitation. In agricultural
areas within sensitive soil regions, soil amendments and other management
techniques will largely negate any adverse effects of acid deposition. In
other areas of the country, the productivity of well-buffered soils may in
fact benefit from excess sulfur and nitrogen deposition (National Atmos-
pheric Deposition Program 1978).
49
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4.4 HYDROLOGY
The hydrological characteristics of watersheds are important determi-
nants of the sensitivity of ecosystems to atmospheric deposition. Dis-
charge volumes, velocities and pathways relate very closely with other
sensitivity factors, notably topography and soil properties. They deter-
mine rates of overland flow versus groundwater flow, the extent of soil
penetration, and the residence time of waters in different components of
the watershed (Norton 1980). The vulnerability of ecosystems may be en-
hanced by minimal soil contact, short residence times and rapid or volumi-
nous discharges (Hendrey et a"!. 1980a). Polluted waters may flow essen-
tially unaltered to groundwater and surface streams when topography is
steep, soils are shallow or particularly coarse, and root channels or
animal burrows are plentiful (National Atmospheric Deposition Program
1978).
Srtowmelts are an important hydrological factor in ecosystem vulnera-
bility as they may cause rapid release of the pollutants and acids that
have accumulated over the course of a season. The three stages involved
in the snowmelt process are all relevant to pollutant loadings in ecosys-
tems (Johannessen et aj_. 1980):
• the pressing out of soil water (piston flow);
• the preferential release of accumulated ions during the first
phases of snowmelt; and
• the dilation stage, where remaining water contains significantly
lower ionic concentrations.
The sensitivity of aquatic ecosystems to snowmelt loading is governed by
the time required for these stages to occur; rapid snowmelts are associ-
ated with more widespread and acute ecological effects. The extent of
mid-winter snowmelt, witli its preferentiat release of ions, and the degree
of snowpack stratification are other important factors mediating pollutant
loading of ecosystems (Wright and Dovland 1978; Overrein et ak 1980).
Furthermore, the ionic yields of snowmelt have been found to corre-
late inversely with amounts of snow cover (Lewis and Grant 1980b). When
snow cover is sparse, soil frost is more prevalent and ionic exports to
aquatic systems, especially of nitrates, are accelerated. Widespread soil
frost also preferentially increases terrestrial ecosystem losses of phos-
phates, potassium and other nutrients of biological demand that would
normally be retained by a variety of mechanisms before reaching the
aquatic environment (Lewis and Grant 1980b).
50
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4.5 HYDROCHEMISTRV
The physico-chemical properties of drainage waters largely determine
their susceptibility to acidification and other alterations in water
quality brought on by atmospheric deposition. As a general rule, water
chemistry is a function of the geological, pedological, and vegetative
characteristics of the watershed (Likens et £]_. 1979). The particular
sensitivity of these systems more often tFan not parallels the degree of
vulnerability observed in terrestrial components of the watershed (Likens
and Bormann 1974a).
Sensitive surface waters generally possess all the characteristics of
soft waters. They have low conductivities and pH values typically less
than neutral. Unlike hard, mineralized waters, they are low in alkalinity
and, consequently, poorly buffered.
The alkalinity of surface waters is generally agreed to be indicative
of the susceptibility of aquatic ecosystems to acidification (Henriksen
1979, 1980; Hendrey et aj_. 1980a). Alkalinity is defined as a measure of
the concentration of any available sink for hydrogen ions (Kramer 1976).
The principal buffering agent in natural waters is bicarbonate
(HCO3) or calcium carbonate (CaC03), although aluminum and other metal
complexes, and organic matter, may provide some acid-neutralizing capabil-
ity to softwaters (Hendrey et al_. 1980a; Norton 1980). Hendrey et al.
(1980a) suggest that alkalinTty values less than 500 ueq/1 are indicative
of potential freshwater sensitivity. The Ontario Ministry of the Environ-
ment (1979) provides the following hydrochemical parameter values below
which surface waters have a marked potential for acidification:
• alkalinity, 300 ueq/1 (15 mg/1) of CaC03;
• conductivity, 35 umho/cm at 25° C; and
• pH, less than 6.0.
Henriksen (1979, 1980) defines water acidification in terms of alka-
linity:
Acidification = pre-acidification alkalinity - present day alkalinity
Using historical water-chemistry data from 719 lakes in southern Norway,
he developed the "predictor" nomograph depicted in Figure 16. It is used
to predict the acid status of lakes where historical water-quality data
are lacking.
This model is based on the hypothesis that acidified waters result
from a large scale acid-base titration: cations released by weathering
processes in the watershed are titrated with acids from wet and dry depo-
sition (Henriksen 1979, 1980). The assumption is made that excess Ca'+,
51
-------
or Ca^+ and Mg^+, contained in waters reflects the weathering process
while the pH of precipitation, or lake concentration of SO4 , rep-
resent atmospheric acid inputs (all marine sources corrected for). It is
also assumed that Ca2+ and Mg2+ concentrations do not change signifi-
cantly in response to acidification, artd that these substances ^re accom-
panied by equivalent amounts of bicarbonate (Henriksen 1980). Regression
analyses led to the definition of three distinct lake classes:
• bicarbonate (HCO3) lakes, with pH 5.3;
7JD~ 5.0 A.j 4.5 4*4 4.3 4^2 71 4*0
pH of precipitation
Figure 16. Henriksen's nomograph for predicting the acid status of lakes.
Lake pH may be predicted from the sum of non-marine calcium and magnesium
concentration or calcium content alone and non-marine sulfate concentra-
tions in lake water or the weighted-average hydrogen ion concentration in
precipitation. (From Henriksen 1980)
52
-------
• transitional lakes, with pH between 4.7 and 5.3; and
• acid lakes, with pH <4.7.
The Schofield diagram, depicted in Figure 17 was developed to graphically
represent the distribution of lakes within a given region according to
these three lake classes.
To underscore the predictive capabilities cf this model, distinct
correlations have been made between judgments of fish population status
and the position of lakes on the nomograph. Results from the study of
684 Norwegian lakes are presented in Figure 18. Table 14 relates the fish
population status of 214 Adirondack lakes to the three lake classes, and
provides an indication of the applicability of the nomograph to this
region. In general, Henriksen's nomograph has been found to accurately
predict the acid status of lakes in various regions of the world; an ex-
ception is found in upland lakes of Scotland where sources of acidity
other than the atmosphere are thought to play a role (Wright et _al_. 1980).
Predictive models of this nature have yet to be developed for stream
acidification; at present, the measurement of water quality parameters
must be relied upon for the identification of acidifying streams.
Sensitive lakes are characterized by elevated concentrations of alu-
minum, manganese and other heavy metals; sulfate may be the predominant
anion (Wright et_ aj_. 1980). Moreover, aquatic sediments often reflect
increased concentrations of heavy metals indicative of elevated dissolu-
tion processes occurring in the watershed. Metal concentrations in the
water column way be augmented by decreased water pH that induces the
mobilization of toxic metals bound in sediments (Kramer 1976; Gahnstrom et
al. 1980). The relationship between aluminum concentrations and water pH
in lakes of the Adirondack mountains is presented in Figure 19 (note log
scale for Al concentrations). Reference metal levels in the water column
are (Ontario Ministry of the Environment, 1979):
• <50 ug/1 for Al; and
• <10 ug/1 for Mn.
Aluminum concentrations have been found to correlate well with judgments
of fish population status in acidified lakes of Norway (Overrein et al.
1980). ~~
Freshwater acidification has been found in streams and lakes ranging
from New England and New York to Florida, as well as in the Boundary
Waters Canoe Area of Minnesota (Hendrey £t aJL 1980a). It may be antici-
pated in mountainous terrains of the eastern United States as well as in
the Rockies and Sierras of the west (Lewis and Grant 1980a).
53
-------
Excess-S04 peq/
Figure 17. A Schofield diagram for Adirondack lakes, with sulfate levels
of 100-120 ueq/1, superimposed on Henriksen's nomograph. The thickness of
the diagram represents the number of lakes within each category of Ca^+
and concentration, while the shading indicates lake pH. (From
Wright et £L- 1980)
54
-------
1001 n=129
% 50
0-
1001
% 50
m
61
HCOg-lakes
n=332
0 - No fish
Transition lakes ^ " Sparse
2 - Good
3 • Overpopulated
100*1 n=223
Acid lakes
0 12 3
Fish status
Figure 18. Frequency histograms for fish status in 684 Norwegian lakes
separated according to their position on the nomograph of Figure 16.
(From Henriksen 1980)
55
-------
Table 14. Lake classification and fish population
status of 214 Adirondack mountain lakes based on the
nomograph of Figure 16.
Lake Class
Bicarbonate
Transitional
Acid
Expected pH
5.3
4.7-5.3
4.7
Calcium concentration (ueq/1)
110
73-110
73
Number of lakes
69
97
48
Number in pH range
61
65
27
Percent correctly classified
88.4
67.0
56.3
Percent of lakes without fish
2.9
54.6
91.7
(Adapted from Hendrey et aK 1980a)
1000-
Adirondtcks USA
134
Al
100-
10.
PH
Figure 19. Total dissolved aluminum vs. pH in lakes of the Adirondack
Mountains, New York. (From Wright etaj_. 1980)
56
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4.6 TOPOGRAPHY
Ecosystem topography is an important, though indirect, determinant of
general sensitivity to atmospheric deposition. Some relevant factors
determined by topography include the size and shape of watersheds, the
depth and surface area of lakes, the residence time of waters in lakes and
streams, and the distribution of overland versus groundwater flow (Hendrey
et^ _al_. 1980a). The ratio of watershed area to lake area and volume, for
example, influences the extent of interaction between atmospheric deposi-
tions and the geological components of the watershed (Likens and Bormann
1974b). Topography also influences rates and pathways of surface run-off
from precipitation events and snowmelt, especially over frozen soils
(Wiklander 1980). It moderates soil depth, stability and rates of soil
formation and weathering. Furthermore, it can significantly mediate
regional patterns of atmospheric contaminant deposition (Wright and
Dovland 1978; Schrimpff 1980).
In comparison with other ecosystems, mountainous terrains tend to
induce substantially greater amounts and frequencies of precipitation than
usually occur at lower elevations. As discussed above, these regions and
their headwaters tend to be more sensitive to the effects of acid rain due
to their smaller watersheds, thinner soils and unique meteorological con-
ditions. For example, aquatic ecosystem acidification in Norway was first
noticed in high-altitude lakes; only more recently has it been observed to
occur in sensitive downstream areas (Overrein et al_. 1980).
Topography also determines predominant types of vegetation, thus the
nature of biotic influences on ecosystem sensitivity. A north-facing
slope, for example, may support a greater number of conifers, with their
characteristically acidic organic layer, while the south-facing slope may
have a predominantly deciduous community, and more highly buffered soil
1itter.
Flatlands usually possess soils of sufficient depth and diversity to
counteract the accumulation and transfer of atmospheric pollutants.
Nevertheless, ecosystems situated in areas of past glaciation share many
of the same susceptibilities to damage as mountainous regions, except
where potential effects are modified by glacial till (Likens et al. 1979).
4.7 BIOTA
Plant cover is the major biotic factor governing terrestrial ecosys-
tem vulnerability to air pollutants. The height, type, and density of the
plant canopy directly influence the extent of gas adsorption and particle
impaction. The transpiration rate of plants modifies the ratio of pre-
cipitation to evaporation in the ecosystem; when evaporation rates are
elevated, acids and other foreign substances may concentrate on leaf sur-
faces (USEPA 1980c). Airborne pollutants may be absorbed and exchanged by
plant tissues, altering the chemical composition of precipitation and
57
-------
other forms of deposition reaching the soil. The nature of the soil, and
the biological processes which take place there, may then determine the
ability of the ecosystem to retain essential plant substances.
The degree to which plant roots absorb deposited sulfates and nit-
rates is another biotic property determining ecosystem sensitivity. The
stemflow component of incident rainfall is usually highly charged with
these ions in regions of acidic precipitation, and the solution is often
channeled directly to roots for uptake (Voigt 1980). These anions are
most readily absorbed from soils of low nitrogen and sulfur contents,
where a lack of these substances limits primary productivity. Plant up-
take helps to prevent them from accumulating in soils or being rapidly
transferred to aquatic ecosystems.
The composition and function of soil microflora is another biotic
factor regulating the sensitivity of terrestrial ecosystems to air pollu-
tion. Certain microorganisms involved in nitrogen and sulfur conversion
processes are sensitive to chemical alterations of their habitat and they
may prove unable to supply sufficient amounts of the correct forms of
these essential nutrients to plants (Alexander 1980a, b). Symbiotic re-
lationships which facilitate the uptake of nutrients by plants are signif-
icant cycling mechanisms in virtually all terrestrial ecosystems, yet they
are vulnerable to many different air pollution stresses (National Atmos-
pheric Deposition Program 1978). The fungal mycorrhizae which supply
nutrients and maintain the productivity of most forest trees cease to
function properly at low soil pHs. A similar effect is found with the
nitrogen-fixing bacteria associated with the root nodules of legumes and
other free-"living forms in the soil.
Microbial litter decomposition can compound the effects of atmospher-
ic loadings by adding to soil acidity. The nitrifying bacteria, for ex-
ample, routinely produce acidic nitrates through their mineralization
activities (Bache 1980; Rosenqvist et al_. 1980). Potential changes in
species composition and dominance from exposures to air pollution or acid
rain may also result in altered rates of acid or base production by soil
decomposers.
The decomposition of plant residues, primarily from conifers, mosses
and heather, is known to produce several organic acids which, in Norway,
are believed to play an important role in the formation of acidic podzols
of low base saturation (Rosenqvist et a_L 1980). The litter of Scots
pine, for example, has been shown to markedly augment localized soil
acidification directly beneath the foliar canopy (Farrell et ah 1980).
Biota can also influence ecosystem sensitivity to atmospheric deposi-
tion through physical modifications of soil structure. Vacated root chan-
nels and the activities of soil macro-fauna, particularly earthworms and
the tunneling vertebrates, all play a role in modifying the water perme-
ability of soils (Voigt 1980). This in turn determines the degree of in-
teraction between percolating run-off and the base exchange sites of
soils.
58
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The complexity of food webs and other expressions of species inter-
dependence is a biotic factor determining the potential for food chain
disruptions and resultant alterations in species composition. Both con-
sumer and detritus food chains may be affected. Drastic modifications in
food webs can occur when the dominant species, or those controlling energy
flow in the ecosystem, are most sensitive to altered conditions in the
habitat, or, as in northern climes, where food webs are simplified.
With respect to nutrient cycling in general, the degree of succession
attained by the ecosystem exerts pronounced influences on the stability of
individual cycles. Early successional stages, such as the flora which
colonizes abandoned farmlands, are characterized by low biomass and low
species diversity, and have "open" nutrient cycles. That is, system in-
puts and losses are elevated in comparison with storage in the biomass
due to a decreased biotic control of cycling in the microhabitat. These
"open" cycles are inherently unstable and perturbations, such as those
brought on by air pollution and acidification, may affect ecosystem func-
tions for a long time (Jordan et £]_• 1972). Late successional stages,
especially on well-buffered soTTs, are more resistant to the effects of
pollution because the essential nutrient cycles are "closed", and con-
sequently more stable. Low inputs and losses in relation to high biomass
are indicative of the ability of a productive and diverse flora to better
influence its microhabitat and recover more quickly from adverse pertur-
bations.
Ecosystems characterized by an abundance of nutrient pumps, or organ-
isms with significant nutrient cycling capabilities, may experience
increased vulnerability to the effects of air pollution if the reduced
pro- ductivity of the species disrupts nutrient cycles. Rooted shoreline
plants, for example, are often indispensible for the return of nutrients
from aquatic to terrestrial ecosystems. Their decline has been documented
in acidic waters of Scandinavia (Hendrey et al. 1976; Leivestad et al.
1976).
Ecotones, or transition communities occurring between two major eco-
systems, are often sensitive to air pollution loadings. Often separating
terrestrial and aquatic systems are extraordinarily productive wetlands
that provide habitat for a large diversity of flora, fauna, and sensitive
juvenile life forms as well as rare and endangered species. These habi-
tats may be dominated by flora of high or low tolerance to air pollution
effects, or may contain species that are subjected to additional environ-
mental stresses such as pesticide and metal accumulation.
4.8 HUMAN ACTIVITY
Many of the human activity factors which influence the nature and
extent of ecosystem loadings of airborne pollutants have been presented in
discussions of pollutant emission sources, dispersion patterns, and depo-
sition processes. Ecosystem sensitivity is in part a function of the lo-
cation and extent of urban and industrial development, their relationship
59
-------
to prevailing winds, and the seasonality of various emissions. Figure
20, for example, compares the relation between anthropogenic sulfate
loadings and decreasing pH in highly sensitive and less sensitive lakes
in Sweden.
Figure 20. Data from lakes in Sweden showing the relationship between
anthropogenic sulfate loadings and pH change for (1) very sensitive and
(2) somewhat less sensitive surroundings. (From Glass et al_. 1979)
60
-------
The variety of air pollutants generated by human activities may in-
teract synergistically to influence the degree of pollutant loadings in
ecosystems. Many of these, particularly ozone, may enhance rates of sul-
fur dioxide conversion to sulfates in the atmosphere (6CA Corp. 1981).
Moreover, the effects of acid deposition on plants, in the eastern United
States and in parts of the west downwind from urban areas, are believed to
differ with the intensity of exposure to ozone and other photochemical
oxidants (Jacobson 1979). Point, line, and area sources may emit a
variety of anthropogenic particulates that are deposited on soils; these
substances may be preferentially absorbed by vegetation over natural soil
constituents (Hutchinson 1979).
Land-use activities are often associated with watershed disturbances
that exert considerable influence on the suspectibi1ity of soils and sur-
face waters to acidification and other effects of air pollution, largely
through alterations in hydrology, soil structure and biota. Land use can
also significantly alter the chemical composition of rain and snow (Eisen-
reich et aj_. 1980). It should be noted, however, that comprehensive
studies in Norway have ruled out less disruptive former land-use patterns
(dairy cattle grazing) and recent changes in land use (reindeer grazing)
as contributing to lake acidification and declining fish populations
(Drablos and Sevaldrud 1980; Drablos et_ £l_. 1980; Overrein et £l_. 1980).
A.9 SUMMARY
The sensitivity of sr ecosystem tc atmospheric deposition depends on
its characteristic meteorology, geology, pedology, hydrology, tiydrochem-
istry, topography, biota and human activities. The relative importance of
these factors will differ from region to region. A synopsis of sensitiv-
ity factors which together render ecosystems most vulnerable to the ef~
fects of acid deposition is presented in Table 15. Areas of the United
States of potential ecosystem sensitivity to acid precipitation, on the
basis of soil substrata, are depicted in Figure 21.
In combination, these influences can reduce the inherent capacity of
ecosystems to control short-term functions of production and biogeochemi-
cal cycling, long-term processes of succession, and the continued suste-
nance of fish and wildlife populations. When they occur together,they
forebode serious ecological consequences for mountain ranges, pre-Cambrian
terrains, first- and second-order streams, and headwater lakes. There-
fore, consideration of these factors is essential to assessments of air
pollution impact on fish, wildlife and their habitats.
61
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Table 15. Factors indicative of potential ecosystem
sensitivity to acidifying air pollutants.
METEOROLOGY
§ Location downwind from emission sources
• Frequent precipitation, dew, frost or fog
• Precipitation exceeds evapotranspiration
GEOLOGY
• Igneous or metamorphic (alumino-silicate) bedrock
• High concentrations of aluminum, mercury and other heavy metals
• History of glaciation
PEDOLOGY
• Thin soils and organic layers (low buffering capacity)
• Mature, severely leached soils
• Low cation exchange capacity and base saturation
• Low sulfate adsorption capacity (high sulfate mobility)
• Absence of carbonates
HYDROLOGY
• Substantial snowpack accumulation
• Rapid discharges following precipitation or snowmelt
HYDROCHEMISTRY
• Oligotrophic surface waters
• Depressed water pH
• High metal concentrations
• Soft waters (low alkalinity, ionic content, conductivity and
buffering capacity)
TOPOGRAPHY
• High altitude
• Steep terrain
t First and second order streams
• Headwater lakes
BIOTA
« Prevalence of conifers
• Reduced litter generation
• Early successional stage (instable nutrient cycling)
HUMAN ACTIVITY
• Watershed disturbances (deforestation, urbanization,
fertilization)
62
-------
Figure 21. Areas of the United States with soils sensitive to atmospheric
deposition overlain with 1978-1979 pH isopleths from the National Atmospheric
Deposition Program. Black dots represent monitoring stations. (From Root et aK 1980)
-------
5.0 RESPONSES OF FISH, WILDLIFE, AND HABITAT
TO AIR POLLUTION AND ACID RAIN
The effects of air pollution and acid rain on fish, wildlife and
their habitats are varied and complex. This introductory dicussion of
biotic and ecosystem-level responses is designed to provide a general
overview of the current state-of-knowledge regarding potential and ob-
served air pollution impacts in terrestrial and aquatic ecosystems. For
each of the three pollutant categories, photochemical oxidants, particu-
lates, and acidifying pollutants, important research findings are high-
lighted and the pertinent literature is introduced. Detailed discussions
of biological and ecological response, along with references to more
specialized studies, are presented in the ecosystem-specific companion
reports of this series.
A summary of observed air pollution effects in plant and animal life
is presented in Table 16 for different levels of biological organization.
Included are ecosystem-1evel responses that play a role in biotic impact.
Although increased emphasis has recently been placed on the investigation
of biotic effects in natural settings, much of the available data comes
from laboratory studies performed under controlled conditions.
The responses of plants and animals to air pollution will be influ-
enced by the structural and functional characteristics of their ecosys-
tems. A given organism may exhibit certain physiological responses in the
laboratory but responses may be different when the organism is observed in
its native environment. The discrepancies between the responses of the
individual organism (autecology) and those of the organism as a part of
the biotic community in its environment (synecology) are due to the inter-
vention of a number of complex interactions and modifying variables as one
moves from the laboratory to the field. Perhaps the most notable of these
involves the ability of the organisms in aggregate to modify the environ-
ment according to their needs (Odum 1971). Therefore, a proper study of
the effects of a disturbance such as air pollution on plants, animals and
their habitats must necessarily include a cautious interpretation of the
results of both laboratory and field analyses.
The balance between the plant, animal, and the abiotic components of
an ecosystem can be disrupted by the loss of a few species which are sen-
sitive to air pollution leading possibly to major changes in the entire
ecosystem. Indirect impacts can be quite important. For example, animals
will be affected by changes in habitat and food availability resulting
from air pollution injury to plants and such changes would take place even
in the absence of any direct effects on the animals.
Many extensive reviews of air pollution and acid rain effects on
plants and animals have been prepared. Table 17 lists a number of these
with a brief description of the subjects covered.
64
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Table 16. A hierarchical classification of biotic and ecosystem-level responses to
air pollutant uptake or deposition.
Level of Organization
Cell
Tissue or Organ
Organism
Ecosystem
Alteration of the
cellular environment
Residue accumulation
Changes in growth and
appearance
Pollutant accumulation in plants
and animals as well as abiotic
ecosystem components (e.g., soils,
aquatic systems)
Changes in cellular
enzymes and metabolites
Altered growth, development
and energy requirements
Increased susceptibility
to biotic and abiotic in-
fluences
Residue transfer and biomagnifica-
tion through food chafns
Alteration of cell organ-
elles and inetabol ism
Altered photosynthesis, trans-
piration and respiration 1n
plants
Reduced plant productivity,
yield and quality
Shifts in inter- and intra-specific
competitiveness
Disruption of reaction
pathways
Altered blood chemistry and
physiology in animals
Abnormal animal behavior
Changes in population numbers,
species diversity and the spatial
distribution of biota
Teratogenic, mutagenic
and carcinogenic effects
Chlorosis, necrosis, death or
abscission of plant organs
Altered genetic resistance
Altered birth and mortality rates
in animal populations
Disruption and death of
cell
Degeneration and functional
disruption of lungs and other
animal organs
Death of organism
Disruption of decomposition,
nutrient cycling and energy
transfers
Reduction of ecosystem stability
and self- regulation
Degeneration of plant stands and
animal communities
(Adapted from Guderian 1977; Newman 1975, 1980)
-------
Table 17. Available reviews of the biotic and
ecosystem-level effects of air pollution and acid rain.
Effects of air pollution on
plants (illustrated)
Effects of air pollution on
terrestrial animals
Effects of acid deposition on
terrestrial and aquatic biota
(research reports and confer-
ence proceedings)
Effects of air pollution on
terrestrial ecosystems
Ecosystem structure and
function (background)
Jacobson and Hill (1970)
Lillie (1970), Newman (1975, 1980) and
Stick el (1975)
Impact Assessment Work Group (1981),
Michigan State University (1981),
Drablos and Tollan (1980), Hutchinson
and Havas (1980), Overrein et a]_.
(1980), Shriner et al. (1980J, Toribara
et al. (1980), U3EP7r( 1980c), A.S.A.P.
TT975"), Evans and Hendrey (1979), Wood
(1979), Hendrey (1978), National Atmos-
pheric Deposition Program (1978),
Dochinger and Seliqa (1976), Giddings
and Galloway (1976)
Woodwell (1970), Smith (1981)
Odum (1969), Likens and Bormann (1974b)
5.1 RESPONSE TO PHOTOCHEMICAL OXIDANTS
The photochemical oxidant complex constitutes a widely varying class
of pollutants. The composition will depend on the nature and availability
of specific precursors and the meteorological conditions prevailing during
formation. The ozone (O3) and peroxyacetylnitrate (PAN) components of
the photochemical oxidant complex are known to be very toxic to vegetation
in controlled laboratory and greenhouse environments (National Research
Council 1977). Oxidant injury to plants is widely reported in the field
as well.
Animals are affected by oxidants primarily in the eyes, lungs and
upper respiratory tract (Newman 1975, 1980). Ozone dissolved in water is
known to be toxic to fish. Although this phenomenon is associated with
water purification systems rather than airborne oxidant contamination
(Paller and Heidinger 1980), it is possible that impacts may occur to
aquatic organisms frequenting the water surface in areas of chronic oxi-
dant exposure (Taylor 1973). Since effects of airborne oxidants on
66
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aquatic organisms and environments have not been demonstrated, the follow-
ing discussion emphasizes anticipated effects on terrestrial organisms and
ecosystems.
5.1.1 Plant Response
Ozone is perhaps the most significant cause of air pollution damage
to plants in the U.S. Plant injury has been observed in many different
regions, affecting a wide range of vegetation including leafy vegetables,
grains, conifers and deciduous trees.
Ozone enters the leaves of plants through the stomata during normal
gas exchange. Once inside the leaf, ozone reacts with the moist cells
causing injury or death of cells and tissues; preferential attack occurs
on the palisade mesophyll cells. Recently matured leaves appear to be
most sensitive to ozone exposure (National Research Council 1977). The
response of plants can be discussed in terms of visible or subtle effects.
The visible effects of exposure to oxidants of broadleaf plants,
grasses and conifers typically include (Heck and Brandt 1977):
• stipple, fleck and chlorosis on upper leaf surface, premature
senescence or death ir broadleafs;
• scattered necrotic areas on both leaf surfaces and necrotic
streaking in grasses; and
• brown-tan necrotic needle tips and chlorotic mottling of needles
in conifers.
The "emergence tipburn" disease of eastern white pine is a well-
documented effect of photochemical oxidant pollution in the northeastern
United States (Berry and Ripperton 1963; Skelley jrt aH_. 1979). Concentra-
tions varying from 0.06-0.25 ppm O3 are sufficient to produce tipburn
symptoms, and primary roots are observed to die back in response (Berry
and Ripperton 1963, Costonis 1970). Symptoms on this species are enhanced
by excess leaf moisture common in this region (Sinclair and Costonis
1967). Other susceptible eastern species include larch, hemlock and pine
varieties, while red pine, firs and spruces demonstrate greater tolerance
(Davis and Wood 1973).
Chlorotic mottling of the needles is the major oxidant-related dis-
ease affecting ponderosa and Jeffrey pines in California (Taylor 1973;
Williams et aK 1977; Munn £t al_. 1977). Ozone levels of 0.5 ppm for nine
hours over 9-18 days produced these symptoms in controlled experiments
(Miller et al_. 1963). Subsequent research indicated that under natural
conditions exposures of 0.08 ppm O3 for 12-13 hours per day are suffi-
cient to injure ponderosa pines {Taylor 1973). Mid-elevation pines are
more severely damaged than those at higher and lower altitudes as oxidants
concentrate at the ceilings of thermal inversions (Williams et_ al_. 1977).
67
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The quality of growing sites also influences the extent of this disease,
although symptoms are found in the best growing sites. Other western spe-
cies affected by oxidant exposures include white fir, sugar pine, sequoia,
incense cedar, Lodgepole pine and black oak (Williams et_ ak 1977).
The subtle effects of ozone on plants are elicited by sublethal ex-
posures characteristic of chronic oxidant pollution and involve interfer-
ence in the normal physiological and biochemical processes (Pell 1974).
The results include reduced growth and yield, closure of stomates, genetic
abnormalities, reduced reproductive yield, and many more species-specific
responses (Heck and Brandt 1977).
Exposure of the unicellular plant Euglena to ozone reduces photosyn-
thetic rate and increases respiration (de Koning and Jegier 1968). Low-
level exposures to ozone are also known to reduce the chlorophyll, protein
and RNA content of duckweed species (Lemna), leading to destructive
changes in the enzymes and membranes of cell tissues (Craker 1971, 1972).
The effects of sub-lethal (0.1 ppm O3) exposures on duckweed development
include (Feder and Sullivan 1969):
• slower multiplication processes;
• lowered rates of frond doubling;
t prolonged flower production; and
• fewer flowers than controls.
Changes in photosynthesis, respiration, ATP and total adenate content
in the primary leaves of pinto beans have also been correlated with the
development of O3 toxicity symptoms (Pell and Brennan 1973). Exposures
of 0.3 ppm O3 for three hours inhibited internode elongation and the
initiation of new internodes after intially stimulating these parameters
at lower concentrations (Leone and Brennan 1975).
Carlson (1979) showed that maple, oak and ash trees suffered reduced
photosynthesis at ozone levels that do not cause visible damage. Eastern
deciduous trees are generally sensitive to ozone at 0.2 and 0.3 ppm for
two to four hours (National Research Council 1977). Histological and his-
tochemical changes in the needles of ponderosa pine are also known to re-
duce apparent photosynthetic rates (Miller et al. 1969; Evans and Miller
1972). Much remains to be learned, however, o7~~specific biochemical and
bioenergetic alterations induced by low-level oxidant exposures on all
groups of primary producers.
PAN is taken into the leaves of plants in the same manner as other
gaseous pollutants, i.e., through the stomata. As opposed to ozone, PAN
preferentially attacks the spongy mesophyll cells. Young, developing
leaves are most sensitive to exposure to PAN (National Research Council
1976, 1977). Visible symptoms include (Heck and Brandt 1977):
68
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• glazed, silvered, or bronzed appearance on leaf underside, bi-
facial collapse in a banded pattern and early senescence with
abscission in broadleafs;
• irregular collapsed banding of bleached yellow to tan color in
grasses; and
• needle blight with some chlorosis or bleaching in conifers.
The subtle effects on plant physiology induced by PAN exposure are
similar to those discussed for ozone. PAN is especially known to prefer-
entially inactivate enzymes with sulfhydryl groups in plants (Mudd 1963).
Field studies of PAN injury are rendered impractical by difficulties in
distinguishing its effects from those of ozone or total oxidant injury,
therefore little plant damage has been attributed directly to PAN.
A variety of abiotic factors have been found to modify the extent of
oxidant injury to plants (Heck 1968). They include:
• light intensity;
• temperature;
• relative humidity;
• soil characteristics; and
• diurnal and seasonal influences on plant physiology.
Synergism between oxidant and S0£ exposures has been shown to occur as
well; concentrations of SO2 and O3 that alone caused no injury brought
on damage when combined (Menser and Heggestad 1966). Ozone uptake in red
kidney beans was seen to increase several-fold as relative humidity rose
from 35 to 75 percent (McLaughlin and Taylor 1981); oxidant injury of to-
bacco has been correlated with increased humidity (Otto and Daines 1969).
The nutritional status of plants in part determines their susceptibility
to ozone toxicity (Craker 1971). Leone et_ aK (1966) found that tobacco
plants were most sensitive under optimum nitrogen conditions, but appeared
to acquire protection when too little or too much nitrogen was present.
Biotic factors regulating ozone tolerance include the age of the organism
and its tissues, and inherent genetic susceptibility relative to the
structure and function of stomata (Heck 1968). Biotic pathogens must be
considered as well in investigations of air pollutant effects on plants.
Ozone injury can enhance, inhibit or show no effect on bacterial, viral or
fungal infestations, while, paradoxically, some pathogens provide local-
ized protection to ozone injury (Heagle 1973).
Ethylene {C2H4) is at once a phytotoxic hydrocarbon pollutant of
occasional localized importance in urban areas and a natural plant hormone
69
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that plays an important role in regulating plant development. The symp-
toms of ethylene injury include (Craker 1971; National Research Council
1976):
• growth reduction;
• decreased apical dominance;
• flower deformities;
• leaf growth abnormalities;
• mitochondrial swelling;
• early senescence; and
• premature leaf abscission.
Stress ethylene refers to excessive amounts of this hormone produced in
response to various stresses such as air pollution (Tingey 1980). Stress
ethylene production rates are found to correlate with foliar injury from
ozone exposures; they accurately measure the extent of tissue response to
ozone stress in plants and predict their relative sensitivity (Tingey et
al. 1976). Indeed, some of the injury evidenced by oxidant exposure may
due to accelerated ethylene production by injured tissues (Craker
1971).
5.1.2 Animal Response
Information on the effects of photochemical oxidants on animals is
derived mainly from laboratory studies (Lillie 1970; Newman 1975, 1980;
National Research Council 1977). Ozone is the most commonly studied of
the compounds making up the photochemical oxidant complex because it is
known to be among the most toxic of gases. Other photochemical oxidants
may be biologically active but the studies of associated effects are few.
The only study of PAN reported led to the conclusion that it is less toxic
to animals than ozone (National Research Council 1977). Furthermore, in
laboratory studies of exposure to the complex of photochemical oxidants,
the observed effects are quite similar to those of exposure to ozone
(National Research Council 1977). This section therefore focuses on the
known effects of ozone.
Ozone enters the vertebrate system primarily during respiration. Due
to its relative insolubility in water, it can reach deep in the lungs and
therefore cause damage to central airways and terminal alveoli. Ozone is
a highly reactive oxidant and is likely to be absorbed by the lung tissues
(National Research Council 1977). The effects of ozone depend on the
length of exposure.
70
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Generally, acute exposures to ozone have been found to cause pulmon-
ary edema (excess fluid in the lungs), and vascular hemorrhaging in cer-
tain species. Changes in lung function have also been found to occur.
These include alterations in the elastic behavior of lung tissue, in-
creased resistance to air flow, and decreased carbon monoxide diffusion
capacity (National Research Council 7977j.
Several effects have been found to result from prolonged exposure to
ozone. These include (Stokinger 1962):
• effects on lung tissue function and morphology;
• acceleration of lung tumor formation; and
• acceleration of aging processes.
More specific potential effects of chronic low-level exposures include
(Kavet and Brain 1974):
• oxidation of structural proteins and membrane lipids;
• lung inflammation and edema;
• altered cellular enzyme activity; and
• inhibition of lung macrophage function.
Localized tolerance to ozone in the lungs is known to be acquired from
previous exposures, however altered enzyme activity can irreversibly sup-
press the cell growth and metabolism required for adequate pulmonary func-
tion (Kavet and Brain 1974).
Ozone is also known to impair normal lung defenses resulting in in-
creased susceptibility to infectious organisms. Enhanced susceptibility
in mice has been reported at doses as low at 0.08 ppm for three hours, a
level which could be achieved in rural areas downwind of major urban cen-
ters (National Research Council 1977). It is thought that ozone and other
oxidants inhibit or inactivate two normal respiratory functions: the
action of cilia in the nasal and upper respiratory passages that clears
particles and therefore prevents them from entering the lungs, and phago-
cytosis by macrophages found in the alveoli (Kavet and Brain 1974; Nation-
al Research Council 1977).
The differential tolerance of animal species to ozone is thought to
be primarily phylogenetically controlled; nevertheless, genetic inheri-
tance of ozone tolerance has been demonstrated as well within populations
of deer mice in California (Rlchkind and Hacker 1980). Deer mice from
areas of elevated oxidant pollution were found to be significantly more
tolerant to ozone than those inhabiting areas of less pollution. Their
laboratory-born progeny reflected the same trend, suggesting a genetic
71
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basis for resistance to ozone effects. Randomly bred populations are more
tolerant than inbred populations, indicating the importance of genetic
variability. The biochemical links have yet to be found, however, between
the degree of ozone toxicity and genetic expressions of tolerance in wild-
life exposed to photochemical oxidants.
A summary of photochemical oxidant effects on animals is presented in
Table 18. Further investigation is required to elucidate other potential
impacts of photochemical oxidants on wildlife. For example, it remains to
be learned whether oxidant air pollution alters the visual and olfactory
senses of vertebrate wildlife. Blindness and eye irritation are postulat-
ed effects of oxidants on bighorn sheep of the San Bernardino Mountains of
California (Taylor 1973). It is also likely that oxidant alterations of
specific habitats, or the ecosystem in general, may impact wildlife popu-
lations in ways that are difficult to detect or predict (Taylor 1973).
Table 18. The effects of photochemical oxidants on animals.
Ecological Physiological
Changes in population numbers
Changes in cellular enzymes
Altered birth and death rates
Altered blood chemistry or phys-
iology
Abnormal behavior
Lowered resistance to natural
environmental stresses
Altered genetic resistance
Birth defects, mutagenesis or
carcinogenesis
(Adapted from Newman 1975, 1980)
5.1.3 Ecosystem Response
Chronic photochemical oxidant pollution has been reported to produce
ecosystem changes analagous to but less severe than those observed by
Woodwell (1970) for radiation, heavy metal and sulfur dioxide pollution.
Westman (1979) observed that, in coastal sage scrub ecosystems, chronic
ozone exposures of 0.18 ppm elicited the following system-level responses:
• decreased total foliar cover;
• lower number of plant species;
72
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• increased number of dominant species; and
• more individuals of the dominant species.
Increases in mean annual oxidant concentrations were found to correlate
better than 43 other habitat variables in explaining the overall decline
of foliar cover, implicating oxidant pollution as the primary stress in
this ecosystem. Oxidant-induced shifts in species composition away from
dominant populations have been observed in the San Bernardino Mountains
of California (Miller et al. 1969), other areas of the west (Treshow and
Stewart 1973; Westman 1979]" and the deciduous forests of the east (Hayes
and Skelley 1977; Skelley et aj_. 1979).
Field studies performed by Treshow and Stewart (1973) determined that
oxidant levels of 0.15 ppm are sufficient to produce visible injury in the
dominant species of natural plant communities. A dominant grass species
was found to be the most sensitive of an oak-grassland community while the
dominant aspen was most sensitive in an aspen-conifer community. In re-
maining plants, ozone exposure results in a general reduction of plant
biomass and decreased root storage for assuring survival in following
years (Price and Treshow 1972).
These effects on species composition and biomass can lead to altered
nutrient cycling and energy relationships in terrestrial communities and
altered hydrology and water quality in the drainage basin (Taylor 1980).
Resultant patterns of succession are difficult to predict as each plant
reacts in a different way to changes in community dominance. Understory
species, for example, may lack the shade they require for growth in their
niche and may thus not proliferate if the dominant species is selectively
removed. Once changes in species composition are set in motion, the out-
come is uncertain and there is little basis for a prediction of future
species composition (Treshow and Stewart 1973; Harward and Treshow 1975).
Large-scale alterations in species composition and ecosystem succes-
sion may result if the oxidant-tolerant species replacing the more sensi-
tive dominants are themselves susceptible to another environmental stress.
For example, the oxidant-sensitive dominants of mountains in California
are well-adapted to frequent occurrences of fire; many of the conifers
that would replace them are unable to survive fires, and the net result
could be a rollback of successional patterns to a shrubland stage (Kickert
and Gemmill 1980).
Several specific ecosystem stresses may occur in conjunction with
chronic oxidant pollution. One of the best known is the enhanced suscep-
tibility of dominant pines in oxidant-contaminated California forests to
infestation by pine bark beetles (Miller and Elderman 1977; Dahlsten and
Rowney 1980). These trees are also predisposed to root pathogens (James
et al. 1980). A variety of other interactions between plants and their
patKogens, parasites and symbionts can occur in the presence of oxidants
(Heagle 1973). Ozone, for example, is believed to inhibit the
73
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fertilization of parasitic nematodes or destroy their favored sites for
plant entry, feeding or reproduction (Weber et a_L 1979). Soil and root
populations are protected from the direct action of oxidants yet may
flourish or decline in response to altered plant metabolism or metabolite
translocation.
As a general rule, oxidant damage to terrestrial habitats and changes
in specfes composition and interrelationships may be anticipated to lead
to altered population numbers and community diversity among wildlife,
largely through changes in the competitive status of each species. Wild-
life impacts of this nature are difficult to detect and quantify. For
example, a reduction in small mammal populations of oxidant-contamineted
forests in California has been reported although exact reasons are not
immediately clear (Miller and Elderman 1977). The subtlety and potential
extent of oxidant effects in terrestrial ecosystems remains poorly under-
stood (Skelley 1980).
5.2 RESPONSE TO PARTICULATES
Atmospheric particulate matter is a complex mixture of natural and
man-made substances that includes both primary and secondary air pollu-
tants. The most important components of particulate matter from an eco-
logical point of view are:
• anthropogenically generated dusts, ash, and soot;
• synthetic organic compounds;
• trace metals;
• radioactive particles; and
• non-metallic ions, including the acidic sulfates and nitrates to
be discussed in Section 5.3.
The particulate composition of the atmosphere also includes mineral dusts
and nutrient elements such as nitrogen, potassium, silica, sulfur, and
phosphorus, which may be beneficial or essential to the productivity of
terrestrial and aquatic ecosystems.
Biotic responses to particulates vary widely and depend on exposure
levels and duration, affected biota, and mediating environmental condi-
tions. As with most pollutants, acute effects are attributed to periods
of high exposure while chronic effects result from long-term, low level
exposures. The latter may occur as a result of long-range pollutant
transport but many particulates do not have as great a range as gaseous
pollutants. Trace metals and organic compounds, such as pesticides, are
the particulates most often associated with chronic effects. Even when
exposures remain below toxic thresholds, biomagnification in food chains
can lead to toxic effects in wildlife.
74
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5.2.1 Terrestrial Plant Response
Particulates can affect plants through external deposition on plant
surfaces, or they can enter plants through opened stomata. Particulate
matter deposited on the soil can also be absorbed by the root system.
Heavy particulate concentrations on leaf surfaces can reduce photosynthet-
ic activity (Heck and Brandt 1977) or block stomatal openings (Williams et
al. 1971). When sufficient moisture is present, hard crusts may form on
pTant surfaces (Darley 1969) and leaf cuticular materials may be dissolved
(USDA 1973). The exact effect on the plant will depend on the type of
pollutant. Caustic particles, such as fly ash, can burn plant tissues
(Dvorak et aj_. 1978), while the primary impact of alkaline dusts deposited
on leaves will be reduced plant growth (Brandt and Rhoades 1972, 1973;
Manning 1971; Lerman and Darley 1975). The result of external deposition
will be disruption of normal gas diffusion and a potentially greater vul-
nerability to gaseous pollutants and plant pathogens (Ricks and Williams
1974). Visible effects can include necrosis, chlorosis, and stunting or
deformation (Dvorak et ail_. 1978).
Some particulate matter, metals for example, are readily absorbed by
leaf and root systems. Depending on the plant and the metal, absorption
may take place preferentially through the aerial or underground portions
of the plant and metal translocation within the plant may or may not oc-
cur. For example cadmium is absorbed with greater efficiency by plant
roots than by leaves (Haghiri 1973). It is rapidly translocated to the
aerial portions of plants, although reproductive organs have been reported
to accumulate less Cd than vegetative tissues (Pietz et^ al. 1978). Al-
though foliar absorption of lead is the chief mechanism o7 accumulation in
the aerial portions of plants, lead absorbed through plant roots may be
translocated to above-ground plant parts (Motto et al. 1970; Lagerwerff et
al_. 1973).
Many particulates are accumulated and magnified in food chains.
Lichens and mosses are well-known accumulators of atmospheric metals and
have been intensively investigated for their potential in monitoring rates
of trace metal deposition. Table 19 lists some important references in
this area. Mondano and Smith (1974) showed that epiphytic mosses as well
as tree needles and twigs accumulate mercury to levels exceeding those
found in soils. The leaves and twigs of urban woody trees have also been
observed to accumulate excessive amounts of atmospheric metals (Smith
1972, 1973). Within two kilometers of a zinc smelter, tree parts have
been found with zinc levels of 4500 ppm and Cd levels of 70 ppm (Buchauer
1973). Metal accumulations usually render plants unsafe for animal con-
sumption.
75
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Tab^e 19. Selected references on the use of lichens and mosses
in monitoring the deposition of atmospheric trace metals.
Goodman and Roberts 1971
Grodzinska 1978
Groet 1976
Mondano and Smith 1974
Pilegaard et aj_. 1979
Gough and Erdman 1977
Little and Martin 1974
Pilegaard 1979
Rasmussen 1977
The degree of particulate accumulation will depend on a number of
varying factors. Plant leaf morphology is perhaps the most influential
biotic factor mediating particulate accumulation on plant surfaces (Little
and Martin 1972). Once on the surface, penetration will depend on the
environmental conditions which govern the opening of stomata. Soil chem-
istry will also influence root uptake. For example:
• cadmium uptake through plant roots is enhanced in acid soils with
low levels of organic matter to adsorb metals (Andersson and
Nilsson 1974);
• similar conditions in the presence of soluble lead facilitate up-
take of lead (Zimdahl 1976);
• the uptake of certain metals may be reduced by the presence of
other metals in the soil (McGrath et £l_. 1980);
• increased soil acidity can facilitate plant uptake of fluorides
accumulated in soils (Thompson et aj_. 1979).
Metal toxicity in plants usually results fron interference with res-
piration and photosynthesis caused by the binding of metals to chloro-
plasts and mitochondria (Zimdahl 1976). Cadmium, for example, accumulates
preferentially in chloroplasts, reducing growth and causing chlorosis,
leaf wrinkling, and red veins (Barber and Brennan 1974). Cadmium is
acutely toxic in small quantities; 2.5 ppm in soils is sufficient to
produce toxicity symptoms (Haghiri 1973).
Toxicity varies for different metals. Lead toxicity is generally
reduced in plants by the immobilization of the metal in structural tissues
(Rains 1971; Zimdahl 1976). While such mechanisms may reduce plant in-
jury, secondary effects can take place when lead-contaminated plants are
eaten by animals. Fluoride particulates are well-known phytotoxins whose
impacts have been observed primarily in forest ecosystems of the north-
western United States (Miller and McBride 1975; Weinstein 1977). The
toxicity of fluorides is related to their solubility in water; thus
76
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gaseous fluoride is generally more toxic than particulate forms (Jacobson
and Hill 1970). Moreover, the symptoms of fluoride injury may differ from
one ecosystem to another for the same plant species (Amundson and Wein-
stein 1980).
The threshold and extent of plant damage can be modified by a number
of factors. Nutritional deficiencies, meteorological conditions, and soil
variables affect the concentrations of metal required to produce toxicity
symptoms (Zimdahl and Arvik 1973; Zimdahl 1976; Gough et a2» 1979).
Stress ethylene production is stimulated by soil applications and may in-
fluence the extent of plant damage of Cd, Pb, and other metals (Rodecap
and Tingey 1981). Metal-induced foliar injury is enhanced in the presence
of SO2 although the interaction of SO2 and trace metals does not
appear to affect metal uptake rates (Krause and Kaiser 1977).
5.2.2 Terrestrial Animal Response
Animals may be exposed to particulate pollution directly through in-
halation or indirectly by ingestion. While some particulate matter is
biologically inert, other forms of particulates are directly toxic or may
adsorb or contain toxic components. Animal responses will vary widely
depending on the pollutant, the exposure pathway, and the animal's toler-
ance, life habits, and position in the food chain. The response of ani-
mals to particulates is discussed by Newman (1975, 1980) and Gough et al.
(1979). An indication of the diversity of the effects is given in TabTe
20, which summarizes some retrospective observations of wildlife damage in
the field.
Atmospheric particulates, notably the trace metals, fluorides, and
pesticides, bioaccumulate in both vertebrates and invertebrates. The
field studies listed in Table 21 document bioaccumulation of metals in
wild populations. These studies found different rates of trace metal ac-
cumulation among individuals, species, and groups of related species. The
most prevalent target organs for trace metal accumulation are the bones,
kidneys, liver and brain. Known target organs for constituents of atmos-
pheric particulates are given in Table 22.
The accumulation of non-essential metals, such as lead and cadmium,
increases with age in vertebrates. Biologically essential elements, on
the other hand, are maintained at stable levels through physiological reg-
ulation (Schlesinger and Potter 1974). In addition to age, the life hab-
its and nutritional requirements of each species will be important deter-
minants of particulate uptake and bioaccumulation. This has been demon-
strated for small mammals in a number of studies (Quarles et aj_. 1974;
Getz et^al. 1977b, ; Clark 1979; Wright et aK 1978). Inter-species var-
iability in accumulation has been clearly shown. Grasshoppers do not ac-
cumulate roadside lead while levels of lead in earthworms near roads cor-
relate with the lead content of the soils (Van Hook 1974; Scanlon 1979).
Earthworms are known to have a high affinity for cadmium (Ireland 1979).
In general, this metal is accumulated much more rapidly than lead or zinc
in soil invertebrates (Scanlon 1979).
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Table 20. Incidents involving the adverse effects of atmospheric particulates on vertebrate
wildlife.
Location
Species
Pollutant(s)
and source(s)
Reported atmospheric
levels or tissue
concentrations
Effects
References
Montana, USA
Mule and white-
tail deer
Fluoride, aluminum
plant
Up to 430 ppm in vegetation
Fluoros is
Kay et ah (1 y7b)
Washington, USA
Black-tailed deer
Fluoride, aluminum
3,900 ppm in metatarsals
Fluorosis
Newman & Yu (1976)
Ontario, Canada
White-tailed deer
Fluoride plant,
industrial complex
Up to 7,125 ppm in bone and
1,200 ppm in water
Fluorosis
Karstad (1967)
England
Sparrowhawks and
song thrushes
Cadmium, smelter
Up to 387 ppm in kidney
Food chain mag-
nification
Martin & Coughtrey (1976)
California, USA
Voles
Lead, urban sources
1.1 ppm in bone
Biological con-
trations
Hirao & Patterson (1974)
New Hampshire,
USA
Mice and shrews
Lead, cadmium, and
copper, industrial
regions
Body conc. for Pb, Cu, and Cd
of 2.7, 3.2, and 0.4 ug/g in
mice, 2.6, 2.9, and 0.4 ug/g
in shrews
Biological con-
centrations
Schlesinger & Potter
(1974)
England
Wood mice and
bank voles
Mercury, chloro-
alkali plant
37 to 124 ng/dm2/d deposition
Biological con-
centration
Bull et ah ( 1977)
Czechoslovak ia
Hares
SO? and fly ash,
power plants and
other industries
0.15 mg SO^/m^ in air and
300 t/km<7yr fly ash deposi-
tion
Hypocalcemia
and hypopro-
teinesis
Novakova & Roubal (1971)
Novakova et ah (1973)
(Adapted from Newman 1975, 1980)
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Table 21. Selected references on the bioaccumulation
of trace metals in wild animal populations.
Pb
Fluoride
Synthetic
organic
compounds
"a
Cu and Ni
Pb, Cd
and Cu
Pb, Cd
and Zn
Clark 1979
Getz et al_. 1977a, 1977b
Goldsmith and Scanlon 1977
Hirao and Patterson 1974
Jeffries and French 1972
Mierau and Favara 1975
Ohi et al. 1974
Tansy ancl Roth 1970
Ward and Brooks 1978
Welch and Dick 1975
Williamson and Evans 1972
Wright _et _al_. 1978
Carlson and Dewey 1971
Dewey 1973
Stlck el 1975
Bigler and Hoff 1977
Bull et al. 1977
Ranta et al.. 1978
Schlesinger and Potter 1974
Ireland 1979
Johnson jjt aik 1978
Martin and Coughtrey 1976
Scanlon 1979
Van Hook 1974
79
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Table 22. The major biological systems of animals affected by atmospheric particulates.
Major biologial systems of animals
Central
Circu-
Gastroin-
Pul-
Respir-
Skeletal
Atmospheric
particulates
nervous
system
latory
system
testinal
system
Hepatic
system
monary
system
atory
system
Renal
system
& dental
system
Asbestos
•
Arsenic
•
•
Barium
•
1
Beryl 1ium
•
•
Boron
•
•
1
Cadmium
•
•
•
9
Chromium
•
•
Fluoride
1
•
Iron
•
Lead
•
•
Manganese
•
•
Mercury
•
•
1
•
Molybdenum
•
Nickel
•
Selenium
1
•
Vanadium
•
•
•
t
Zinc •
(Adapted from Newman 1975, 1980)
-------
The tolerance of animals to the accumulation of particulate matter
varies widely. Many species of terrestrial animals accumulate lead and
other particulates to high levels with no apparent ill effects. Quarles
et aH• (1974) observed lead levels in small mammals that would produce
severe symptoms of lead poisoning in humans and other higher vertebrates
at similar body-tissue concentrations.
Particulate exposure and accumulation can cause effects in animals
ranging from irritation to debilitating disease, reproductive failure and
death (Lillie 1970; Newman 1975). When inhaled, many types of particu-
lates, in particular the fine particulates, are deposited on the respira-
tory surface of the lungs (alveoli) where they can cause symptoms ranging
in severity from slight difficulty in breathing, to emphesyma and lung
cancer (Lillie 1970). In sufficient concentrations, they are known to
depress the viability of protective lung-macrophage cells (Zarkower and
Ferguson 1978).
In addition to the effects on lungs, many types of particulate matter
are directly toxic. For example, the chronic effects of synthetic organic
compounds on the reproductive success of wild bird species are well docu-
mented (Stickel 1975). Several metallic compounds present as pollutants
have been observed to cause toxic, carcinogenic, and/or reproductive ef-
fects in experimental animals (Lillie 1970; Newman 1975).
Bacterial screening tests for mutagenicity have been applied to
several types of particulates and materials which adsorb to particulates.
Substances which demonstrated mutagenic potential include:
• fly ash (Fisher et al_. 1979);
• filtrates of fly ash (Chrisp et aU 1978);
• diesel soot (Li and Royer 1979);
• polycyclic aromatic hydrocarbons (Glass 1979b);
• uncharacterized substances in urban aerosols (Alfheim et al¦
1980); and
• organic micropollutants undergoing long-range transport (Alfheim
and Moller 1979).
The effects of atmospheric fluoride particles are well-documented.
Fluorosis, a disease causing deterioration of the bones and teeth, occurs
in livestock and wild vertebrates as a result of exposure to fluorides and
subsequent bioconcentration (Karstad 1967; Kay et^ £l_. 1975; Newman and Yu
1976; Krook and Maylin 1979). The concentration of fluoride in the femur
is believed to be the best indicator of uptake in vertebrates (Wright e^t
al. 1978). Reductions of arthropod populations have also been attributed
to fluoride bioaccumulation (Carlson and Dewey 1971; Dewey 1973).
81
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Many sub-lethal effects can result from exposures to atmosnhprir
ticulates. The ingestion of alkaline dusts, for example can chanap
ach and intestinal pH in midlife and livestock, potentially alter?™ th
composition of internal microflora, or caustic ash may irritate stnmarh
linings and predispose the animal to infection or produce symptoms of
gastro-ententis (Tendron 1964). Atmospheric lead can deprS level* r.f
specific enzymes in the bloocf of pigeons (Ohi et al. 1974) Combing n*
ticulate and sulfur dioxide pollution has beenThown to effect bln^n nE
terns, serum chemistry and urine pH, as well as reproductive succe^ ?n
hares (Novakova 1970; Novakova and Roubal 1971; Novakova et al 1973)
Table 23 summarizes the variety of responses anticipated TrorrTthe exoo
sure of wildlife populations to particulate air pollutants.
5*2.3 Terrestrial Ecosystem Response
The effects of intrinsically toxic particulates on biota can lead to
disruptions of ecosystem functioning both in the vicinity of point sour-
ces, where exposures may be acute, and in areas of chronic, low level deD
osft ion. Observed effects include changes in soil microbial activity
nutrient cycling, species growth and diversity, succession, and the 1nt*n
rity of food chains. Furthermore, the effects of widespread ecosystem
contamination from atmospheric particulates are probably not easily re-
versed (Purves 1972).
Jackson and Watson (1977) studied the effects of lead, cadmium cod-
per, and zinc emitted from a lead smelter and identified two distinct
stages of ecosystem disruption corresponding to the amount of pollutant
loading. In a far zone, 1.2-2.0 miles from the stack, soil and litter
microfauna were unaffected but forest litter was observed to accumulate
In this zone, the nutrient pool size was elevated by litter accumulation
Closer to the smelter, in a zone 0.4-0.8 miles from the stack re-
duced decomposer communities were observed. Here deposited metals'accumu
lated at a rate too great to be assimilated. High concentrations of met-
als in leaf litter inhibited decomposition by soil microflora and the si?
and quality of the nutrient pool was depressed (Jackson and Watson 1977)
Under the more severe conditions observed in the near zone, nutrient®
tend to remain bound in dead tissue. Soil acidity increases and the cat-
ion exchange capacity decreases as a result of reduced decomposition
leading to the mobilization and translocation of free soil nutrients' not-
ably calcium (Jackson and Watson 1977; Freedman and Hutchinson 1960a' c
d). Reduced availability of nutrients for plant uptake can result in a'
decreased nutrient content of living biomass and future forest litter j
sum, mineralization and decomposition rates are depressed at the same"tim2
that nutrient losses from the ecosystem are accelerated.
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Table 23. Documented responses of wildlife, domestic and laboratory animals to acute and chronic
exposures of atmospheric particulates.
Responses
Arse- Asbes- Bar- Beryl-
nic tos lum Hunt
Selected air pollutants
Cad- Chrom- Fluor- Manga- Her- Molyb- Selen- Vanad-
Boron nium 1um
Ide
Iron Lead nese cury denum Nickel
lura
Zinc
Changes in population
numbers
Changes in blood chem-
istry or physiology
Changes in cellular
enzymes
Changes in external
appearance
Change in population
distribution
Change in death rate (In
free living animals)
Change in birth rate
Change in growth rate
Change in genetic
resistance
Abnormal behavior
Physiological changes
observed in autopsy and
histological analysis
Lowered resistance to
natural environmental
stress
Residue accumulation in
body tissues
Teratogenic, mutagenic,
or carcinogenic effects
(Adapted rrom Newman 1975, 1980)
-------
As discussed above, depressed rates of decomposition are important as
an indicator of soil response to atmospheric pollutant loading. Reduc-
tions in microbial activity have beer documented in a number of ways:
• decreased weight loss in forest litter samples (Innan and Parker
1978);
• depressed CO2 evolution, soil respiration, substrate decomposi-
tion, and enzyme activity (Ruhling and Tyler 1973; Ebregt and
Boldwijn 1977; Freedman and Hutchinson 1980a, d);
• carbon nitrogen ratios in litter are inversely proportional to
microbial populations and their activity (Jackson and Watson
1977); and
• amino-sugar carbon concentrations correlate negatively with the
presence and function of both fungi and soil arthropods (Jackson
and Watson 1977).
The ecosystem responses to elevated metal deposition are reductions
in species diversity and altered plant succession. Gilbert (1975) report-
ed drastic changes in forest structure due to fluoride contamination while
Freedman and Hutchinson (1980b) documented reduced tree diversity and
biomass in forests near smelters. Desert grassland communities have been
observed to increase in species diversity with increasing distance from
point sources of copper and other metals (Wood and Nash 1976; Dawson and
Nash 1980). Dust accumulations in eastern deciduous forests have also
provoked shifts in species dominance and community composition (Brandt and
Rhoades 1972, 1973).
The observed shifts in plant dominance and diversity may in part be
due to pollution-induced changes in pathogen and symbiont communities.
Some particulates are known to inhibit plant pathogen activity, for ex-
ample, pine rust (Parmeter and Uhrenholdt 1975), while other infectious
organisms, notably fungi, may be stimulated (Manning 1971). Fluorides
have been observed to predispose natural plant communities to insect at-
tack (Carlson et aj_. 1974).
Biomagnification of substances contained in atmospheric particulates
is an important ecosystem response. Well-known examples of biomagnifica-
tion in terrestrial food chains include radioactive particles in reindeer
and pesticides in piscivorous and carnivorous birds (Stieke 1 1975). Food
chain biomagnification probably contributes to the elevated concentrations
of trace metals observed in small mammals and birds of intermediate tro-
phic levels (see references in Table 21).
The tendency for substances to biomagnify may be reduced by natural
purification mechanisms which can efficiently remove some accumulated
particulates from food chains. For example, in terrestrial ecosystems
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there are at least four obstacles to Pb biomagnification in food chains
(Oones and Clement 1972):
• the immobility of Pb in soils;
• restricted translocation from roots to shoots;
• restricted absorption of ingested lead by herbivores; and
• the preferential retention of Pb in inedible bone tissue.
Hirao and Patterson (1974) studied lead aerosol pollution in remote
areas of the California Sierras. They confirmed that while herbivores
tended to bioaccumulate Pb from leaf surfaces, it occurred in such a way
that Pb was not biomagnified at higher trophic levels. Natural purifica-
tion mechanisms involving snowmelt interactions with contaminated forest
humus prevented further Pb uptake by prey organisms at the base of food
chains.
Such processes may not be effective barriers for other types of par-
ticulate matter. Moreover, different ecosystems will have varying mecha-
nisms and capacities to protect food chain integrity. Should these be
non-existent or easily circumvented by certain metals and organics, food
chain biomagnification can readily occur.
The ecological effects of particulates appear to be most severe when
acidifying pollutants, especially sulfur dioxide, are present, although
few efforts have been made in field studies to distinguish the relative
influence of each. Biotic effects of this synergism are not well known,
however atmospheric metal deposition in combination with acid precipita-
tion can cause fundamental changes in soil structure and its organic con-
tent: cation exchange capacity and pH are both reduced, aluminum and
other cations are mobilized, and soil erosion is promoted in areas of sig-
nificant forest destruction (Hutchinson and Whitby 1974). All of these
conditions can substantially increase pollutant phytotoxicity, and impact
water quality and hydrology, exacerbating the effects of particulates on
aquatic biota.
5.2.4 Aquatic Plant Response
Impacts on aquatic ecosystems related to particulate pollution are,
for the most part, associated with trace elements, especially metals.
Water bodies accumulate metals from a variety of sources. Particulate
matter may be deposited directly in aquatic ecosystems through both wet
and dry removal processes, or matter deposited in the watershed may run
off into surface waters (Wright and Gjessing 1976; Galloway and Likens
1979; Overrein et ah 1980; Schindler et al. 1980a; Eisenreich et al.
1981). Trace elements may also be deriveT~from natural weatherTng pro-
cesses in the catchment or mobilized from sediments (Likens and Bormann
1974b; Norton 1980; Van Hassel et al. 1980).
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Releases of terrestrial and sediment-bound metals can be greatly ac-
celerated by acidic conditions in the watershed (Maimer 1976; Driscoll
1980; Gorham and McFee 1980; Norton 1980; Schindler _et a2« 1980a). In
fact, it is hardly possible to discuss the effects of atmospheric particu-
lates, and trace metals in particular, without considering the effects of
acid deposition (Section 5.3 of this report). Usually both types of pol-
lution are present together in field exposure situations. In addition the
acidification of soils and water has a strong effect on the mobility of
trace metals and hence the biological availability of these potential
toxins (Wright and Gjessing 1976; Aimer et aK 1978; Henriksen and Wright
1978; Driscoll et a_]_. 1980; Schindler et £]_• 1980a).
Studies of trace metal effects on aquatic flora must necessarily in-
corporate the synergistic effects of habitat acidification (Van Loon and
Beamish 1977). For example, Stokes et al. (1973) suggested that trace
metals had adversely affected the phytopTankton populations of two lakes
located near smelters. It was later concluded, however, that the observed
effects were probably due to the combination of excessive metal concentra-
tions with low water pH (Stokes and Hutchinson 1976).
Yan (1979) compared the phytoplankton communities of uncontaminated
lakes to those of partially acidified, metal-contaminated lakes. The
changes in community composition which were found strongly resembled those
observed in acidified waters with low metal concentrations (Section
5.3.4). These included the replacement of dominant chrysophyte genera
with a single dinoflagellate species, and a significant reduction in the
number of taxa with no concurrent decrease in biomass (Yan 1979). The
overall alteration may be summarized as one of:
• species replacements and shifts in dominance;
• decreased number of species; and
• expanded populations of remaining species.
Aquatic macrophytes are also subject to impact from atmospheric
metals, the synergistic influence of acidification and antagonistic or
synergistic effects of combinations of trace elements (Dvorak et al.
1978). Gorham and Gordon (1963) first reported that trace metaT levels of
toxicity to aquatic vegetation are achieved in lakes near the Sudbury,
Ontario, smelters as a result of emissions. Both submerged and emergent
forms are known to effectively bioaccumulate trace metals from surrounding
water and sediments (Tiffen 1977). Cladophora species of algae and the
cat-tail Typha accumulate the highest levels, although internal concentra-
tions are found to vary among the different organs of the same plant
(Wells et al. 1980). Algae may also play a significant role in the mobil-
ization ofTediment-bound metals as they possess the ability to sorb and
accumulate them in living tissues (Laube et al_. 1979).
86
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Trace element contamination in the absence of freshwater acidifica-
tion has been documented near roads. Van Hassel et (1980) were able
to correlate stream sediment concentrations of le"a
-------
significant component of atmospheric deposition, but rather is leached
from the watershed in sensitive areas receiving acid deposition (Abraham-
sen et _al_. 1979; Johnson, N. 1979; Driscoll et jjK 1980). This terres-
trial leachate will be discussed in connection with acid precipitation in
Section 5.3.
Mercury is a toxic trace metal found as an atmospheric particulate.
Concentrations of mercury can be elevated in areas receiving acid precipi-
tation. Local atmospheric emissions have been correlated with the pH of
lakes in Sweden as well as with elevated mercury concentrations in pike
(Jernelov 1980). Rain and snow in Quebec, Canada, are known to scavenge
20-100 mg/1 total mercury and deposit it in surface waters with background
concentrations of only 5-10 mg/1 (Tomlinson et al_. 1980). These studies
have shown atmospheric deposition in both Canada and Scandinavia to be the
major source of mercury in unbuffered waters not already subject to point
discharges of this metal. Most of the research on mercury in aquatic eco-
systems is focused on exposure pathways and bioconcentration, with rela-
tively little defining the effects on aquatic organisms.
Freshwater acidification is an important controlling mechanism of
mercury toxicity as it determines the availability of the toxic monomethyl
form (CH3Hg). Elevated concentrations of this form are found in fresh-
waters, primarily in spring following snowmelt, however other aspects of
acidification may account for increasing levels (Jernelov 1980; Wood
1980):
• waters of low pH and conductivity retain more mercury from pre-
cipitation than well-buffered waters;
• acidified waters revolatize less mercury than waters of higher
pH; and
• greater amounts of mercury are transformed to the monomethyl form
at pH levels below 7.0, either through increased bacterial produc-
tion or through interactions between inorganic mercury and the
humic and fulvic acids of freshwaters.
As a result, higher rates of mercury uptake are found in fish of low
pH waters than in fish of well-buffered waters (Scheider et £l_. 1979).
Several factors have been postulated to enhance mercury uptake in fish
under conditions of acidification (Jernelov 1980; Tomlinson et 1980;
Wood 1980):
• monomethyl mercury, like the organic micropollutants is lipophil-
ic; due to this property and its high affinity for sulfur-groups
present in blood, it is rapidly transferred through gill mem-
branes;
• decreased food availability in an acidified environment results in
more water being drawn through the gills, hence a greater mercury
uptake per unit increase in weight;
88
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• decreased biomass per unit of water volume, also an effect of
acidification, results in greater mercury availability per unit
weight of remaining biomass, and individual uptake is more rapid;
t acidification may result in the eradication of entire year classes
of fish, meaning that predators are likely to receive greater
amounts by consuming older, more contaminated prey.
Mercury tends to accumulate in fish so that body levels are greatest
in the older individuals of a population; levels also correlate with fish
length. Mercury is also subject to biomagmfication in fish at higher
trophic levels. For example, smallmouth bass in the Adirondacks were
found to accumulate mercury at greater rates than white suckers because of
their higher metabolic rate and differing food preferences (Bloomfield et_
al. 1980). The tendency of mercury to biomagnify in food chains was con-
TTrmed in a study by the Michigan Water Resources Commission (1972) which
reported that predatory fish of the Great Lakes have mercury concentra-
tions double those of bottom feeders. Of all trace metals assayed in this
study, only mercury was found to bioconcentrate in fish to levels poten-
tially harmful for secondary and tertiary consumers. Pike (Esox sp.) are
reported to be sensitive indicators of environmental contamination by mer-
cury as they concentrate the element in body tissues to levels 3000 times
higher than surrounding water (Johnels et ajN 1967).
Other trace metals affect aquatic fauna but wide differences have
been observed between metal levels causing fish mortality and those pro-
ducing sublethal stress effects or reduced growth (Sauter et al. 1976).
Continuous low-level exposures of copper (Cu) found to physiologically
impair fathead minnows were only 3-7 percent of the 96-hour median toler-
ance limit for adults (Mount 1968). Reproduction in the fathead minnow
was shown to be inhibited at zinc (Zn) levels that had no effect on the
survival, growth, and maturation of eggs and fry (Brungs 1969).
Few adverse effects on fish have been observed at levels below 10
ug/1 Cu and 90 ug/1 Zn in metal-contaminated lakes, although bottom-
feeders can be exposed to greater amounts and be affected at these water
levels {Van Loon and Beamish 1977). White suckers in lakes contaminated
with zinc and cadmium (Cd) were observed to grow faster at maturity and
die prematurely when compared to control fish. In addition, egg sizes
diminished and spawning success declined, resulting in a reduction of new
year-class recruitment and depressed population numbers (Franzin and
McFarlane 1980). Fry are generally the most sensitive life stage of fish
whereas eggs are often the most resistant to metals (Sauter et al. 1976).
Copper and cadmium are of comparable toxicity to fish and an orcler of mag-
nitude more toxic than lead or chromium (Sauter et 1976). Lead and
zinc levels in benthic insects and three fish species from roadside
streams were found to correlate with traffic density, yet no adverse ef-
fects were detected among fauna (Van Hassell et al_. 1980).
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Principle water chemistry variables influencing the availablity and
toxicity of inorganic metal ions include hardness, acidity, alkalinity and
rates of organic or inorganic complexation (Chakoumakos et al_. 1979). Re-
duced water calcium levels can enhance metal uptake in aquatic fauna
(Franzin and McFarlane 1980; Muniz and Lievestad 1980a). Zinc is found to
be more toxic to bluegill at depressed pH levels (Cairns et a^. 1972). At
a given pH, less Cu is required to reduce fish growth in soft waters than
hard waters; for given hardness and Cu levels, greater growth reductions
are observed in waters of low than neutral pH (Waiwood and Beamish 1978).
The temperature of surface waters is also found to influence metal toxi-
city. Salmon acclimated to cold water temperatures were observed to be
significantly more resistant to Zn toxicity than warm-acc1imated individu-
als (Hodson and Sprague 1975).
Aquatic insects are generally less sensitive than fish to metal con-
centrations in surface waters, however, the mayfly species and other
groups are acutely sensitive to many metals (Warnick and Bell 1969;
Nehring 1976). The crustaceans on which fish feed are believed to be ad-
versely affected at metal levels that also impair fish, however daphnia
are found to be more sensitive than fish (Biesinger and Christensen 1972).
Effects of particulates on most zooplankton and other invertebrates large-
ly have yet to be clarified, particularly for the case of aluminum en-
trapment and air pollutants such as mercury.
5.2.6 Aquatic Ecosystem Response
The potential effects of atmospheric particulates on aquatic ecosys-
tems range widely from beneficial fertilization with nitrogen, phosphorus
and potassium compounds to a variety of toxic effects from the trace met-
als, organic pollutants and, associated acids (Galloway and Cowling 1978).
However, very few ecosystem-level responses, beyond reductions in biotic
density and diversity, have been documented.
Despite the lack of detailed ecological assessments, metal-impacted
aquatic systems have been studied near point sources (Hutchinson and Whit-
by 1974; Franzin and McFarlane 1980; Keller et al_. 1980), as well as in
areas subject to regional deposition of pollutants transported long dis-
tances (Aimer et al_. 1974; Wright and Gjessing 1976; Ficke 1978; Galloway
1978; Schindler et . 1980a; Tomlinson et aj. 1980). Further impact sit-
uations have been characterized for the acid-induced mobilization of met-
als from sediments (Schindler et aK 1980b; Schindler 1980), for snowmelt
pulses of elevated metal concentrations (Schofield 1977; Franzin and
McFarlane 1980) and for chronic leaching of toxic metals from the water-
shed (Troutman and Peters 1980). These studies lead to the general con-
clusion that unusually elevated concentrations of metals are found primar-
ily in those lakes, streams and sediments in North America and Scandinavia
that receive chronic acid precipitation.
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The full scope of ecosystem impact associated with freshwater acidi-
fication will be described in the following sections dealing with acidify-
ing air pollutants. At present, the degree to which ecosystem alterations
can be attributed to metals is highly uncertain. In many cases, the ef-
fects of atmospheric particulates may be entirely overshadowed either by
impacts of acidification or by non-atmospheric point and non-point sources
of metal and organic pollution. Trace metal depositions seldom occur in
the absence of the acidifying air pollutants; thus, as discussed above, it
is likely that interactions between water pH and metal toxicity determine
the extent to which metals can alter ecosystem-level processes. Further
research is needed to demonstrate the direct and synergistic effects of
atmospheric metals on aquatic ecosystem functions.
In general, metal-induced changes of water chemistry and subsequent
biotic impoverishment can be expected to cause ecosystem alterations
paralleling those observed in terrestrial systems. These may include:
• increased organic litter accumulation;
• decreased decomposition and nutrient cycling;
• altered primary and secondary productivity;
• altered energy relationships between trophic levels; and
• changes in species interactions and competitive relationships.
Few, if any, of these potential impacts have been confirmed in a quantita-
tive fashion nor linked directly with effects from atmospheric metal depo-
sition. Although such effects have been observed in situations involving
extremely high levels of contamination, such as acid mine drainage (Par-
sons 1968), freshwater metal levels and acidities resulting from air pol-
lution are far less than those of acid mine drainage and the ecosystem-
level impacts taking place can be expected to be more subtle and difficult
to detect.
Aluminum inputs to aquatic ecosystems must continue to be studied as
an important special case since this common element of the earth's crust
is introduced as an undesirable by-product of acid deposition in sensitive
terrestrial and aquatic ecosystems. Again, the degree of interplay be-
tween acidity and Al toxicity may determine the importance of this sub-
stance in ecosystem impacts. Further research is needed to reveal whether
or not primary and secondary productivity can be reduced by Al complexa-
tion with nutrient calcium, phosphorus and magnesium, and subsequent pre-
cipitation to sediments. Also of importance will be knowledge of the
long-term effects of Al pulses accompanying snowmelt on microbial activ-
ity, nutrient cycling and sustained ecosystem productivity.
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5.3 RESPONSE TO ACIDIFYING AIR POLLUTANTS
The primary acidifying pollutants SO;?, N02> and C12* in their
unreacted states, generally produce acute effects on terrestrial biota
only in the immediate vicinity of their release from point sources. Ter-
restrial plant damages from acute and chronic exposures to these air pol-
lutants are well-documented while pollutant avoidance and other behavioral
alterations have been demonstrated in terrestrial animal populations.
As described in Section 3.1 of this report, when the primary acidify-
ing pollutants remain in the atmosphere following emission, they may be
transported over long distances and undergo chemical transformations to
secondary sulfates, nitrates and chlorides which contribute to both wet
and dry acid deposition. In the northeastern United States, mean annual
precipitation pH values are found in the range 4.0-4.5, while individual
storms of pH 3.0-3.6 have been recorded (National Atmospheric Deposition
Program 1978). Known effects include measurable damage to aquatic ecosys-
tems and severe impoverishment of sensitive soils; potential long-term in-
juries to forests are suspected. In general, the effects of acid deposi-
tion on terrestrial plants, animals, and ecosystem-level processes are
subtle and, in contrast to aquatic impacts, remain poorly understood.
Despite the scientific uncertainties, acid deposition has come to be per-
ceived as the primary environmental threat to terrestrial and aquatic
systems in large portions of Scandinavia, eastern Canada and the north-
eastern United States (National Atmospheric Deposition Program 1978;
Overrein et aj_. 1980; Impact Assessment Work Group 1981).
Both direct deposition, including wet and dry deposition, and surface
run-off of the secondary acidifying pollutants contribute to the acidifi-
cation of freshwater ecosystems. As a general rule, surface water acidi-
fication may be expected to occur in sensitive regions receiving chronic
acid precipitation of pH 4.7 or less (Schindler 1979). Changes in species
abundance and community composition appear at all trophic levels of the
aquatic ecosystem as a result of habitat acidification and subsequent al-
terations of water chemistry. As pH levels drop below 5.5, many common
plant, fish, and fish-food species are eliminated or reduced in numbers
while a small variety of acidophilic plants and animals become more abun-
dant (Hendrey £t jH. 1976). Aquatic biota may be subjected to direct
physiological and reproductive stresses as well as indirect effects aris-
ing from alterations of their abiotic environment. Ecosystem functions of
decomposition, mineralization and energy transfer can be disrupted by
acid-induced changes in water quality and the spatial and temporal distri-
bution of biota. The end results are often a simplification of biotic
community structure, reductions of ecosystem self-regulation and stabil-
ity, and arrested ecosystem succession and evolution.
With respect to aquatic and terrestrial ecosystem alike, little is
known of how alterations in precipitation chemistry and atmospheric depo-
sition can accelerate patterns of change taking place over long time per-
iods. As with particulate accumulation in biota and other ecosystem
92
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compartments, the short-term consequences of exposures to acidifying air
pollutants are often not readily apparent and the long-term potential for
harm remains largely unknown. Should such long-term alterations be taking
place, virtually nothing is known of their reversibility or the potential
for mitigation.
Thus, continued research into the effects of acid deposition must ad-
dress many fundamental questions beyond those pertaining to direct biotic
effects. Entire biogeochemical cycles, in particular, should be the sub-
ject of future research; linkages between terrestrial and aquatic ecosys-
tems are also an important concern since biotic and abiotic effects in
different ecosystems may be highly interrelated (National Atmospheric Dep-
osition Program 1978). Such studies should increase understanding of how
alterations induced in ecosystem functioning and the abiotic environment
can affect short-term and long-term plant and animal adaptivity, produc-
tivity, and survival, especially in relation to the top-order carnivorous
birds and fish. The complex nature of questions raised by the acid depo-
sition phenomenon tends to underline the intimate relationship between
air, water, and soil contamination in the assessment of impacts to fish,
wildlife, and their habitats.
5.3.1 Terrestrial Plant Response
The unreacted gaseous pollutants SO;?, NOj, and C12 enter plants
via stomatal openings and exert their primary damage on the cells within
leaves, creating a characteristic type of lesion for each pollutant
(Jacobson and Hill 1970). Within the plant the primary pollutants are
transformed to their corresponding acids by moisture in the leaf interior.
Plants possess a limited capacity to reduce the contaminants to less toxic
compounds, but plant injury occurs once this capacity is exceeded (Knabe
1976). Reductions in photosynthetic production, plant growth, and cellu-
lar enzyme function are subtle effects which may occur even in the absence
of visible tissue damage, premature senescence, or defoliation (Ziegler
1975; Malhotra ancf Hocking 1976).
In general, the effects of plant exposure to acidifying air pollu-
tants will depend upon the inherent sensitivity, stage of development, and
nutritional status of the plant, as well as environmental factors govern-
ing the opening and closing of stomata and the condition of the leaf cuti-
cle (Guderian 1977). For example, high light intensity and adequate soil
moisture will increase the susceptibility of exposed plants to damage by
stimulating stomatal opening and gas exchange (Glass 1979b). In many
plants, particularly in locations where elemental sulfur and biologically
available nitrogen are limiting, low-level exposures to primary gaseous
emissions may actually enhance plant growth and productivity (Bennett et
al_. 1974; Knabe 1976).
Lichen species are known to be particularly sensitive to chronic low-
level SOj) pollution and a wide variety of physiological responses have
been documented from laboratory and field exposures (Eversman 1978, 1980).
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Table 24 lists several studies in which epiphytic lichens or their tree-
bark substrates have been employed to monitor ambient levels of acidifying
air pollution.
Tat^e 2U. Studies of trie use of lichens and tree bark
in the monitoring of acidifying air pollution.
Gilbert 1970
Le Blanc et al. 1972
Creed et aT.~T973
Johnsen and Sochting 1973
Sundstrom and Hallgren 1973
Grodzinska 1976
Henderson and Seaward 1977
Skye 1979
Deposition of the secondary acidifying pollutants, or strong acids,
has yet to be linked with adverse plant effects in the field. However,
experimentation with simulated acid mists has suggested that terrestrial
plants may be adversely affected by acid precipitation in two distinct
ways {Jacobson 1980):
• by direct contact with foliage and reproductive organs, which can
injure living tissues and reduce productivity, growth, yield, and
forage qua!ity; and
• by reduced soil fertility that results from gradual alterations
of the physical, chemical, and biological properties of soils.
However, as noted above for low-level SO2 pollution, chronic sulfate and
nitrate deposition on plant and soil surfaces can significantly augment
plant growth and production in sulfur- and nitrogen-deficient soils (Abra-
hamsen 1980a; Overrein _ak 1980; Tamm and Wiklander 1980; Tveite and
Abrahamsen 1980).
The observed effects of foliar applications of simulated acid rain
(usually of pH less than 4.0) include:
• abnormal cell elongation and prol iferation (Evans et ^1_. 1977;
Evans 1980);
t erosion of leaf cuticle and epithelial waxes (Shriner 1975; Evans
1979a; Fowler jit aj.. 1980);
• leaching of plant nutrients from foliage (Fairfax and Lepp 1975;
Wood and Bormann 1975; Tukey 1980);
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• predisposition to attack by pathogens, parasites, and insects, and
inhibition of beneficial symbionts (Shriner 1976, 1978, 1980;
Weber et aj. 1979; Shriner and Cowling 1980);
• interference with sexual reproduction (Evans 1979b; Evans and
Conway 1980; Evans and Lewins 1980); and
• necrotic lesions on foliage and reproductive structures (Wood and
Bormann 1974; Irving 1978; Shriner 1978; Evans and Curry 1979;
Evans 1980; Haines et al. 1980; Horntvedt et al. 1980; Lee et al.
1980).
Other direct effects postulated to occur include interference with the
function of guard cells and alterations of root- and leaf-exudation pro-
cesses (Tamm and Cowling 1976). The former may result in water stress and
increased foliar uptake of gaseous pollutants, such as SO2 and ozone,
while the latter effect may reduce populations and activities of benefi-
cial leaf and root symbionts such as nitrogen-fixing bacteria and fungal
mycorrhizae.
Dissolved acids in rain, dew and fog are believed to collect in small
pools in leaf surface depressions. Depending on the plant species, a num-
ber of protective mechanisms, summarized in Table 25, may reduce pollutant
injury. Generally, however, acids are able to penetrate cracks in protec-
tive leaf cuticles near specialized structures such as veins, trichomes,
and stomata (Evans et al. 1977, 1978). Necrotic lesions, pitting,
chlorosis, spotting, aruf galls are observed most frequently near leaf
veins and trichomes; injuries are found to occur at both the cellular and
tissue levels, and to enlarge in a localized fashion (Shriner 1979; Evans
1980). A surficial pH of 3.4 is through to be the threshold for visible
injury (Evans 1979a).
Table 25. Processes and structural characteristics
that reduce plant sensitivity to acid deposition.
Exclusion
• leaf orientation and morphology
t chemical composition of cuticle
• flower orientation
• protection of sexual organs
t pol1ination mechanism
Neutralization
• salts on leaf surfaces
• buffering capacity of leaves
Metabolic Feedback
Reactions
• enzymatic reactions that consume hydro-
gen ions or yield alkaline products
(From Jacobson 1980)
95
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Expanding and newly expanded leaves and needles are more sensitive to
acid penetration and injury than immature and older leaves (Wood and Bor-
mann 1974; Evans et £2* 1978) while matured leaves are the most suscept-
ible to acid-induced foliar leaching (Tukey 1980). Monocotyledons are
generally thought to be less sensitive to leaf injury than dicotyledons,
conifers are comparatively less sensitive than deciduous trees, and the
foliage of woody, broad-leaved plants may be less sensitive than that of
broad-leaved, herbaceous plants (Evans 1980).
Plant effects caused by reduced soil fertility and increased acidity
are not as well understood as direct foliar impacts. While reductions in
available phosphorous have not been documented, accelerated soil leaching
of nutrient calcium, magnesium, potassium, iron and manganese is known to
occur (Berigari and Xerikos 1975; Reuss 1978, 1980; McFee 1979).
In soils of low pH, seedling establishment has been shown to be more
sensitive to acidity than the germination process (Lee and Weber 1979).
Seedling growth has been found to be enhanced in some tree species and
reduced in others (Wood and Bormann 1974, 1977). Field investigations of
forest tree growth have yet to provide evidence of adverse effects in
areas receiving acid precipitation (Cogbill 1976; Overrein e_t ah 1980).
Long-term research and improved research methods will be required to ade-
quately assess potential acid-induced impacts to native plant species in
their natural environments.
5.3.2 Terrestrial Animal Response
The effects of acidifying air pollutants on terrestrial vertebrates
are expected to be primarily indirect, involving alterations of soils and
vegetation which, in turn, may lead to modifications of food resources and
the habitats of animal communities (Dvorak et a/L 1978). For example,
arctic caribou are dependent on lichens as their sole source of nourish-
ment, yet these are known to be among the most sensitive of terrestrial
plant species to low-level SO2 exposures (Schofield et _aL 1970). Much
more research is needed to link the nutritional requirements and feeding
habits of wildlife species with the effects of the acidifying pollutants
on plants.
Information on the direct effects of sulfur and nitrogen oxides has
been derived almost exclusively from toxicological experiments conducted
in the laboratory (Lillie 1970; Kavet and Brain 1974). These studies gen-
erally do not provide well-defined exposure thresholds for physiological
injury or death, nor are they conducted at the low exposure levels most
commonly encountered in the field (Dvorak et a]_. 1978). On the other
hand, terrestrial wildlife populations are seldom found in areas with high
ambient concentrations of sulfur dioxide or sulfates. Acute effects on
animals are thus restricted to areas very near point sources of the acid-
ifying air pollutants. The most prevalent effects of these pollutants are
irritations of the eyes and respiratory tract (Newman 1975, 1980) which
can lead to emigration, a variety of behavioral alterations or reductions
in inter- and intra-specific competitiveness (Chilgren 1978).
96
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Laboratory studies undertaken using mixtures of air pollutants have
shown that there is often a synergistic or antagonistic effect between
sulfurous pollutants and other components of the atmosphere. Amdur (1975)
has shown that the bronchial irritant effect of SO2 is enhanced by the
presence of sodium chloride aerosols. Substances such as potassium chlor-
ide, ammonium sulfate, ferrous oxide, magnesium and vanadium were all
shown to act synergistically with SO2 in increasing lung airway resis-
tance. Materials with buffering or reducing capacities appeared to have a
protective effect in opposition to SOj and thereby prevented or mini-
mized observed reductions in pulmonary function.
Field work conducted in Czechoslovakia indicates that a variety of
other direct effects may occur in vertebrate populations near point sour-
ces of SO2 and particulate pollution. Changes in reproductive coeffi-
cients as well as physiological alterations of blood calcium and protein
levels, hematocrit and hemoglobin, and urine chemistry have been document-
ed in rabbit populations exposed to mixed industrial emissions (Novakova
1969, 1970; Novakova and Roubal 1971; Novakova et _al. 1973). The poten-
tial for these and other subtle impacts on animaTs remains to be investi-
gated in North America.
No direct impacts on terrestrial wildlife populations have been docu-
mented from exposures to acid precipitation. Although research on the in-
direct effects of diminished food supply and habitat degradation due to
acidification is just beginning, these effects are thought to be of con-
siderable importance. Table 26 provides a list of some predominantly ter-
restrial species which may be subject to indirect effects from aquatic
ecosystem acidification. Other species which may be impacted in a like
fashion include eagles, ospreys and other avian piscivores, as well as
bears, gulls, and other species that regularly exploit the annual salmon
runs in North American rivers and streams.
In the Adirondack Mountains of New York and the LaCloche Mountains of
Ontario, common loons continue to nest and fish at lakes that no longer
sustain viable fish populations (Harvey 1979; Trivelpiece et al. 1979).
Fully seventy percent of common loon and hooded merganser populations in-
habit areas receiving acid deposition while approximately half of the
range of the bald eagle is similarly threatened (Peakall 1979). Aimer et
al. (1978) have suggested that many piscivorous birds may be forced to
migrate from large areas impacted by acid deposition. The potential mag-
nitude and permanence of these wildlife effects will depend on the sever-
ity of aquatic animal responses to freshwater acidification, as discussed
below in Section 5.3.5.
97
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Table 26. Birds and mammals susceptible to
indirect effects of acid deposition.
Common loon
(Gavia immer)
Common merganser
(Mergus merganser)
Hooded merganser
(Mergus cucullatus)
Red-breasted merganser
(Mergus serrator)
Great blue heron
(Ardea herodias)
Belted kingfisher
(Megaceryle alcyon)
Common goldeneye
(Bucephalus clangula)
Ring-necked duck
(Aythya collar is)
Mai lard
(Anas platyrhynchos)
Black duck
(Anas rubripes)
Green-winged teal
(Anas carolinensis)
Northern pintail
(Anas acuta)
American widgeon
(Anas americana)
American mink
(Mustela vison)
River otter
(Lontra canadensis)
Muskrat
(Ondatra zibethicus)
(Adapted from Impact Assessment Work Group 1981)
5.3.3 Terrestrial Ecosystem Response
The dominant ecosystem-level effects of the acidifying air pollutants
and acid precipitation involve alterations of soil chemistry, fertility
and structure, which can diminish the productivity of habitat. As dis-
cussed above in relation to terrestrial plant responses, cations essential
to plant growth such as calcium, magnesium and potassium may be subject to
accelerated leaching from terrestrial ecosystems as a result of soil acid-
ification (Abrahamsen 1980a, b; Galloway et al. 1980; Wright and Johannes-
sen 1980). As soils become more acidic, TFTeTTiological availability of
phytotoxic metals may increase (Overrein et aj_. 1980). Abrahamsen (1979)
studied the relation of acid deposition to productivity in terrestrial
ecosystems and identified three important factors related to soil proces-
ses which require study:
98
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• the function of bicarbonate and aluminum systems in soils;
t the efficiency of cation exchanges with hydrogen ions; and
• comparisons of anion inputs in precipitation with leaching out-
puts to the catchment.
In discussing the fertilization effect of nitrate and sulfate depo-
sition in terrestrial ecosystems, Abrahamson (1980a) has warned that the
long-term effects of nitrate and sulfate enrichment may be comparable to
those of incomplete fertilization: plant growth may be stimulated by
nitrogen and sulfur inputs, while deficiencies of potassium, magnesium,
phosphorus and other micronutrients are simultaneously reinforced. Stud-
ies of anion budgets in well-buffered watersheds of Tennessee suggest that
the majority of deposited sulfur is retained within soils, while approxi-
mately 20 to 30 percent is drained into aquatic systems (Kelley 1980).
However, in young, poorly buffered soils of low cation exchange capacity,
sulfate adsorption in soils may be considerably reduced and atmospheric
sulfate inputs may equal drainage losses (Kerekes 1980; Skartveit 1980;
Wright and Johannessen 19P.Q). Nitrates, on the other hand, are largely
retained in both well and poorly buffered watersheds due to rapid uptake
and utilization by flora. Thus, sulfate mobility will largely influence
rates of soil cation leaching and can produce the equivalent of an oligo-
trophication effect, while nitrates will have a negligible effect.
Ecosystem productivity may also be adversely affected by acid-induced
changes in the structure and activity of soil microfloral communities.
Such alterations, due to increased soil acidity, have been documented for
individual populations of bacteria, fungi, and algae (Bryant et aj_. 1979;
Alexander 1980a,b; Baath jit al. 1980; Lohm 1980). Microbial processes
consequently affected may inFTude (Alexander 1980a,b; Francis et al.
1980):
• contributions to soil structure, such as soil aeration and the
biosynthesis of humus;
• the detoxification of aromatic substances and other phytotoxic
metabolites derived from incomplete decomposition;
• the detoxification of accumulated anthropogenic pollutants; and
• reactions fundamental to sustained plant growth, including the
decomposition of organic matter, the mineralization of organic
phosphorus and other nutrients, ammonification, nitrification,
and nitrogen fixation by blue-green algae and plant symbionts.
Macrofaunal responses are also important to ecosystem productivity as
these organisms aid in decomposition processes and contribute to humus
conditioning and to the maintenance of soil structure (Voigt 1980). Re-
ductions in species numbers and avoidance behavior have been shown to re-
sult from SOp fumigation in a variety of insects ranging from decomposer
99
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beetles to grasshoppers (Bromenshenk 1979, 1980; Leetham et a2« 1980a,b;
McNary et al. 1980). Population fluctuations have been eTTcited by exper-
imental soTT acidification with artificial acid rain; the abundance of
springtails (Collembola) in particular diminishes with increasing soil
acidity (Hagvar 1980; Overrein et al. 1980). These studies raised the
possibility that reproductive success in soil macrofauna is primarily a
function of soil pH.
Food chain effects, similar to those observed with particulate pollu-
tion, have not been specifically associated with exposures to acidifying
air pollutants. Nevertheless, reductions in primary productivity may have
repercussions on secondary productivity through a diminished carrying
capacity of habitat. Moreover, nutrient exports from aquatic to terres-
trial ecosystems may decline significantly in response to effects on
waterfowl and other piscivorous species noted in the previous section.
Such exports are important insofar as they comprise one of the primary
links by which nutrient matter drained into surface waters is returned for
recycling in terrestrial ecosystems.
5.3.4 Aquatic Plant Response
Freshwater acidification has been shown to cause drastic alterations
of all major aquatic floral groups. Altered water chemistry can directly
impair some species, notably phytoplankton, although toxicity mechanisms
and the individual effects of excess hydrogen, sulfate and metal ions re-
main to be clarified. Acidophilic species, including some plankton,
mosses, and filamentous algae, increase in abundance under acidified con-
ditions. In general, the nature of biotic impacts produced through modi-
fications of the abiotic environment determines the extent of indirect
effects on aquatic flora. Alterations of aquatic macrophyte populations,
for example, often occur in response to changes in species composition and
other aspects of ecosystem structure and function induced by direct im-
pacts on biota.
The phytoplankton typically respond to acidic conditions by increas-
ing or decreasing in numbers, species diversity, and proportion of commun-
ity biomass. Populations of many species decline or vanish while those of
a few proliferate, resulting in an overall decrease in the diversity of
taxa present. The exact nature of shifts in species numbers and community
structure remain obscured by conflicting observations in acidifying en-
vironments of differing locations (Kelso et al_. 1981; Schindler 1980).
Moss (1973) investigated the pH tolerances of freshwater phytoplank-
ton and found the lower growth limits for most species to lie between pH
4.5 and 5.1; however, as pH drops, the acidophi1ic Euglena sp. (a flagel-
late form) and Eunotia sp. (a diatom) survive to pH levels of 3.65 and
3.9, respectively^ Eunotia exigua, in particular, has been found to be
one of the most acid-tolerant species of diatoms (Van Dam et al. 1-980).
Figure 22 presents average minimum pH tolerances of major pfiylToplankton
groups inhabiting freshwaters of North America. These averages must be
used with caution, however, since responses of individual species may vary
100
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7.0
6.0 — diatoms (6.0)
5.0
4.0
3.0
- desmids (5.25)
— green and yellow algae (4.6)
— blue-green algae (4.5)
flagellates (3.1)
2.0
Figure 22. Lower limits of pH tolerance among the phytoplankton.
(Adapted from Eilers and Berg 1981)
101
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greatly, as noted above for diatoms. Virtually all of these groups con-
tain some acidophilic species (Eilers and Berg 1981).
In Scandinavian lakes, the most rapid reductions in species numbers
of phytoplankton were found to occur at pH 5.5 (Leivestad et al_. 1976) and
most population shifts took place within the 5.0 to 6.0 pH interval (Aimer
et al. 1974). This finding was supported by Kwiatkowski and Roff (1976)
wTToTeported that diversity indices of phytopl ankton communities in the La
Cloche Mountain lakes of Ontario declined significantly as water pH fell
below 5.6; the following reductions in species diversity were observed:
• green algae (Chiorophyta) were reduced from 26 to 5 species;
• yellow algae and diatoms (Chrysophyta) decreased from 22 to 5
species; and
• blue-green algae (Cyanophyta) declined from 22 to 10 species.
Blue-green algae were observed to predominate over green algae in these
acidified lakes, contradicting observations in Scandinavia (Aimer et al.
1974) and Florida (Crisman eit aj. 1980) where blue-green algae were found
to decline in acid lakes. Also in southern Ontario, populations of the
flagellate Dinophyceae were found to increase in response to declining pH,
largely replacing the diatoms and Chrysophyceae; when present as a signif-
icant fraction of producer biomass, dinoflagel1ates have been suggested as
a reliable indicator of freshwater acidification (Yan 1978; Yan and Stokes
1978; Yan 1979). Despite the conflicting reports on responses of many
phytoplankton species, there is general agreement that species numbers of
green algae and diatoms decrease in acid lakes (Conroy et _al_. 1976; Aimer
et al. 1978).
Reductions in numbers of phytoplankton taxa also have been recorded
in Adirondack Mountain lakes of decreasing pH (Hendrey 1980). Figure 23
shows the number of taxa observed in three lakes of varying pH. Crisman
et al. (1980) reported an average of 10.8 species in acid lakes of central
FTorida while non-acid lakes averaged 16.5 species; phytoplankton exhibit-
ed greater population declines than members of any other trophic level in
these lakes. It is generally difficult to attribute effects on these pop-
ulations to excess hydrogen ions alone, as metal toxicity or nutrient de-
ficiencies may also play an important role (Hendrey 1979; Hendrey et al.
1980c). ~~
Observations of the effect of acidification on overall biomass and
productivity are contradictory. In some acid lakes, phytoplankton biomass
and productivity have shown an overall decline (Hendrey et 1976;
Hendrey 1978). In Florida, Crisman et al. (1980) recorded mean annual
chlorophyll concentrations of 1.88 mg/m^~in acid lakes, whereas non-acid
lakes possessed 7.53 mg/m^; phytoplankton numbers were 5,700/ml and
14,000/ml respectively. In other acid lakes, these parameters are found
to remain comparable with surrounding non-acid, oligotrophic lakes (Yan
102
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Wood Lake Sagamore Panther
(pH 4.7) Lake Lake
(pH 5.5) (pH 7.0)
Adirondack Mountain Lakes
Figure 23. Number of phytoplankton species observed in
Adirondack Mountain lakes of different pH.
(From Hendrey 1979)
103
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1979), while in still others, phytoplankton biomass is observed to in-
crease (Malley et 1981). In acid lakes of New York, yearly increases
in phytoplankton biomass and productivity are found to be atypical of
those in nearby non-acidic lakes as they occur much later in the season
following spring thaw (Hendrey 1980).
The decline of phytoplankton productivity and biomass per unit of
water volume may be accompanied by an increase in biomass and productivity
per unit area. This happens because eliminations and reductions of some
species leads to increased water transparency, deepening the euphotic
zone. The remaining populations expand to deeper water strata (Aimer et
al. 1978). Kwiatkowski and Roff (1976) observed that high levels of pro-
cfuctivity on a volumetric basis were maintained above pH 4.8, yet subse-
quent reductions in standing crop were offset by the increased transpar-
ency and elevated areal production was maintained to pH 4.5. Thus, in
many cases, no net loss or gain will accompany increased hydrogen ion
loadings (Yan 1979; Hendrey et aj_. 1980c). Phytoplankton biomass and
productivity are therefore less reliable as an indicator of freshwater
acidification than changes in species abundance and composition (Yan and
Stokes 1978).
The recent pH history of lakes for which water quality data are not
available can be inferred from stratigraphic studies of the diatom compo-
sition of surface sediments (Aimer ejt ail_. 1978; Singer 1981). The per-
centage of acidophilic diatom species present in sediments of Norwegian
lakes was found to be inversely correlated with increasing depth of the
sample, thus indicating the time frame over which lake acidification had
progressed (Davis and Berge 1980).
Responses of mosses and benthic filamentous algae to freshwater acid-
ification are generally very similar. Both of these plant groups typical-
ly increase in numbers and proportion of community biomass. In the ab-
sence of aquatic acidification, benthic algae are usually found in small
numbers while the mosses, which are primarily terrestrial or semi-aquatic
species, may be absent or confined to shallow littoral areas. The in-
creased water transparency accompanying freshwater acidification permits
them to inhabit depths where sufficient light would not ordinarily be
available. However, the mechanisms which enhance the competitiveness of
these species are not yet clearly understood.
Both the mosses, primarily Sphagnum and Fontinalis species, and fil-
amentous Mougeotia, Batrachospermum and Zygogonium species form heavy mats
in association with fungi and organic matter. These mats effectively seal
profunda! sediments (Grahn et _al_. 1974; Hultberg and Grahn 1976; Hultberg
1978). The invasion of Sphagnum species has been observed to begin at pH
6.0 and intensify greatly below pH 5.0 (Hendrey and Vertucci 1980); it is
found to be greatest in shaded portions of the littoral zone (Grahn 1977).
The spread of filamentous algae first occurred at pH 5.5 in experimentally
acidified shield lakes of Canada (Schindler 1979). Hall and Likens (1980)
reported periphyton biomass to increase to 1.7 ug chlorophyll a/cm2 in
104
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acidified streams (pH 4.0) from a value of 0.3 ug chlorophyll a/cm^ in a
reference stream (pH 6.4). Dense growths of filamentous algae have also
been reported in lakes of low pH near Sudbury, Ontario (Keller et al.
1980). Both mosses and periphytic algae are abundant in acid streams of
Scotland (Harriman and Morrison 1980), in the lakes of the Adirondacks
(Hendrey and Vertucci 1980), and in Scandinavian lakes (Hultberg and Grahn
1976; Aimer et £l_. 1978; Overrein ej. 1980).
There is little evidence that aquatic vascular plants are directly
affected by freshwater acidification, at least in the absence of heavy
metals (Gorham and Gordon 1963), although reduced growth and flowering
activity in Lobelia dortmanna is a documented response to lowered water pH
(Leivestad et al. 1976). Tndirect effects, however, can be marked as sub-
mergent macrophytes are displaced by expanding populations of aquatic
mosses and algae that shade necessary light, deplete nutrients and render
stream and lake bottoms unfavorable for the growth of these species (Grahn
1977). In acid Lake Colden of New York, Lobelia, Littorella, and Isoetes
species were driven out by invasions of mosses, primarily Sphagnum pyla~
esii; other decreasers included Nuphar leuteum and Eleocharis acicularis
whi le numbers of Utriculari a and Eriocaulon species typically increased.
Some tolerant species may colonize suitable substrates at greater
depths. Juncus species are acidophilic and also increase in response to
elevated accumulations of organic matter (Hultberg and Grahn 1976). Fur-
thermore, the benthic production of macrophytes may be comparatively
greater in acid than non-acid lakes as it is for invading mosses and fil-
amentous algae (Aimer et al_. 1978). Nonetheless, the end result will be
an overall simplification of the macrophyte community (Singer 1981).
Responses of aquatic flora to acidification may be summarized as
follows (Burton et aj_. 1981):
• species shifts to acid-tolerant forms;
• lowered species diversity; and
• increased standing crops of some species concurrent with decreased
standing crops of others.
Changes in primary productivity cannot be predicted.
5.3.5 Aquatic Animal Response
Both vertebrate and invertebrate forms of aquatic fauna are suscep-
tible to adverse impact from freshwater acidification. Direct and in-
direct effects on fish and fish-food organisms occur at all trophic lev-
els. Common responses involve alterations of the osmoregulatory processes
freshwater organisms use to maintain the ionic strength of internal tissue
fluids, leading to subsequent changes in cell metabolism and energy use.
Indirect effects are also common in impacted aquatic ecosystems and occur
in response to modified plant and animal populations or other acid-induced
alterations of habitat.
105
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Communities of zooplankton and zoobenthos characteristically respond
to freshwater acidification by diminishing in species diversity and abun-
dance. Fewer taxa and lower mean numbers of zooplankton and benthic in-
vertebrates, compared with control lakes, have been documented in acid
lakes of Sweden (Aimer et al_. 1974), Norway (Hendrey and Wright 1976;
Hendrey et 1976; LeTvestad et al. 1976; Raddum 1980), Canada (Sprules
1975; Conroy et al. 1976; Roff and Kwiatkowski 1977; Yan and Strus 1980;
Keller et al .H^BO"), and Florida (Crisman et aj. 1980). The number of
dominant species in the zooplankton community is found to be less in acid-
ified waters (Keller et ak 1980); Sprules (1975) noted only one or two
dominants of 1-7 species present in acid lakes compared with 3 or 4 domi-
nants out of 9-16 species in non-acid lakes. Factors affecting inverte-
brate responses to acidification are thought to be very complex, however,
altered dominance patterns in acid lakes are caused primarily by (Overrein
et aT_. 1980):
• differences in the physiological tolerance of individual species
(direct effects); and
• the absence of predatory fish populations (indirect effects).
Some of the most sensitive fish-food organisms, including the tadpole
shrimp (Lepidurus arcticus) and common gammarus (Gammarus lacustris), are
eliminated below pH 6.0 in Scandinavian lakes (Borgstrom and Hendrey 1976;
Hendrey 1979). The opossum shrimp, Mysis relicta, and other crustacean
fish-food organisms of North American lakes are reported to be equally
sensitive (Schindler 1979; Malley et al_. 1981). Reductions of these pop-
ulations may serve as an early warning of ecosystem acidification.
Acidity tolerances of species representative of the major groups of
fish-food organisms are presented in Figure 24. An extensive review of
invertebrate tolerances to pH has been prepared by Eilers and Berg (1981).
Like the gammarus, molluscs are consistently shown to be absent from
waters of pH below 6.0 (Okland 1969; Okland, J. 1980; Okland, K. 1980;
Okland and Okland 1980). The freshwater louse Asellus aquaticus is more
tolerant of intermediate pH levels; its disappearance from a lake signi-
fies that fish populations are likely to be impacted. Populations of the
mayfly, Baetis rhodanis, an important food-chain link in oligotrophic
lakes of Norway, begin to disappear at pH levels below 5.0 (Overrein et
al. 1980). Oligochaete worms may be inhibited by metal accumulations in
TJTe sediments of acid lakes (Raddum 1980).
Calcium availability is thought to play a major role in the responses
of invertebrates, especially shell organisms and moulting crustaceans, to
acidification (Singer 1981). In a stream experimentally acidified to pH
4.0, Burton et al. (1981) found populations of the isopod Asellus inter-
medius reducecf to ten percent of their original numbers while the snail
Physa disappeared entirely. Decreased emergence of aquatic insects,
particularly midges (Diptera) and mayflies (Ephemeroptera), have also been
observed in laboratory and field experiments (Bell 1970, 1971; Fiance
106
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< 4 0
Freshwater shrimp
(Garrmarus lacustris)
Snails
Small mussels
Freshwater louse
(Asellus aquaticus)
pH range for important fresh water fish
Figure 24. Comparative pH tolerances of four groups of fish-food organisms.
(From Okland and Okland 1980)
-------
1978). Lowered densities and increased drift have been observed in ben-
thic insects of acidified streams (Hall et aj_. 1980; Hall and Likens
1980).
Some invertebrates, notably the shredder organisms (Trichoptera), are
acidophilic and are found to assume a greater proportion of total benthic
biomass in impacted ecosystems (Friberg et aj_. 1980; Burton et a]_. 1981).
The waterboatrnan (Corixidae) and other surface-dwelling insects of stream
pools and lakes are particularly tolerant of low pH and may proliferate in
the absence of fish predators (Rosseland et £]_. 1980). Nevertheless, the
caloric value of these fish-food organisms has been found to decline in
acidified waters as invertebrates in general must devote more energy to
the maintenance of osmoregulatory balances (Hendrey 1979; Hall and Likens
1980). A detailed review of benthic invertebrate responses has been pre-
pared by Singer (1981).
Amphibians are an essential food-chain link as predators of inverte-
brates and as prey for fish and waterfowl. Amphibian populations may be
particularly sensitive to acid deposition if they breed in temporary ver-
nal pools following snowmelt. Furthermore, the range of many North Amer-
ican amphibian species is found to overlap areas receiving acid precipita-
tion (Peakall 1979).
Some amphibious species which are susceptible to reduced reproductive
success resulting from habitat acidification are listed in Table 27.
Pough (1976, 1978) has demonstrated reductions of up to 90 percent in the
hatching success of the spotted salamander (Ambystoma maculatum) and Jef-
ferson salamander (Ambystoma jeffersonianum) at pH levels below 6.0.
Larvae and young of the shovel-nosed salamander (Leurognathus marmoratus)
have been shown to be more sensitive to acidic conditions than adults of
the species (Mathews and Larson 1980). Reproduction in frogs has also
been shown to be adversely impacted by depressed water pH (Gosner and
Black 1957; Overrein et _al. 1980).
The effects of acidifying air pollutants on fish populations are
generally well-documented and have been reviewed by Howells and Holden
(1979), Fritz (1980) and Fromm (1980). Observed reductions in fish popu-
lations have been linked to lake acidity in Norway (Leivestad et jH. 1976;
Muniz and Leivestad 1980a), Sweden (Aimer et al. 1974; Dickson~T975), the
Adirondacks (Schofield 1976) and Ontario (beamish and Harvey 1972; Harvey
1975, 1980). Sudden fishkills have been observed following rapid drops in
pH resulting from heavy autumn rains and spring snowmelt in rivers and
streams of New York (Schofield 1977) and Scandinavia (Jensen and Snekvik
1972; Leivestad and Muniz 1976; Hultberg 1977; Wright and Snekvik 1978).
The wide range of effects observed in individual fish, whole populations,
and communities are summarized in Table 28. However, the mechanisms of
acid-induced, physiological alteration taking place in the field are poor-
ly understood. Physiological stress, periodic mass mortality, and repro-
ductive impairment, as discussed below, all contribute to recruitment
failure and the extinction of fish species in areas sensitive to acid
deposition (Schofield 1980).
108
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Table 27. Amphibians susceptible to reduced reproductive
success from breeding-habitat acidification.
Species Breeding Habitat
Rana catesbiana Lakes
(bul Ifrog)
Eurycea bislineata Streams
(northern two-lined salamander)
Ambystoma maculatum Temporary meltwater ponds
(yellow-spotted salamander)
Ambystoma laterale
(blue-spotted salamander)
Bufo americanus
(American toad)
Rana sylvatica
["wood frog)
Pseudacris triseriata
(chorus frog)
Hyia crucifer
(northern spring peeper)
Notophthalmus viridescens Permanent ponds
(red-spotted newt)
Necturus maculosus
(mud puppy)
Rana clamitans
(green frog)
Rana pipiens
(TTortnern leopard frog)
Rana septentrionalis
[mink frog)
Hyja versicolor
(gray tree frog)
(Adapted from Impact Assessment Work Group 1981)
109
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Table 28. A summary of effects of freshwater
acidification on fish.
ORGANISMS
Acute mortality
Ionic balance stress
Reduced blood pH
Suffocation or anoxia
Gill damage
Spinal deformation
Histological changes
Disrupted calcium metabolism
Trace element bioconcentration
Reduced growth rate
Altered feeding behavior
Decreased resistance to
environmental stress
Disrupted reproductive behavior
Failure to reach spawning
condition
Delayed or reduced spawning
Delayed egg hatching
Reduced egg viability
Reduced survival of life stages
POPULATIONS
Altered density
Altered birth rates and death rates
Altered age pyramid
Extincti on
COMMUNITIES
Reduced population abundance
Reduced species diversity
Diminished autotrophic and
heterotrophic production
Reduced biomass
Species replacements
Species eliminations
Altered inter- and intra-specific
relations
ThP failure of fish to successfully regulate body salt content is
, I V pause of fish mortality in acidified waters.
MlieVH cndium flux is caused by changes in the ion permeability of fish
ins active salt uptake mechanisms are inhibited by low pH resulting
q- ' Sin Jit content of blood and body tissues (Leivestad et aU
}no£ rM unii^ and P0?ts 1978; Muniz and Leivestad 1980a, b) -A~aPid,
1976; of hvdroqen ions can cause acidosis in fish blood
preferential -mflu interfere with a variety of physiological mechanisms
whKh is pres.to interf,Qften inc1ude gill c1ogging with
(Fromm 1980). increased respiratory rates, and lowered blood
mucous, hyperventilation, inclfea^^e^' ]m)*
oxygen tension (Overrem et a].. 1980, Rosseiana iyuu;.
Research into the effects of acid precipitation on fish populations
foH In aluminum toxicity as a major cause of injury and death to
inScted rodents (Cronan and Shofield 1979). This
hTo£ esis was proposed wheS initial studies of freshwater acidification
revealed certain responses that »ere difficult to explain (Schofield
1976):
110
-------
• fish mortalities in natural waters were observed to occur at pH
levels shown not to be acutely toxic in the laboratory; and
• the impact on fish populations was found to vary considerably in
lakes of comparable acidity.
It became apparent that one or several additional toxic agents operate
synergistically with freshwater acidity to impair aquatic fauna, especial-
ly during the spring snowmelt. Elevated concentrations of aluminum have
since been discovered and documented in acid lakes of Scandinavia and
North America (Wright and Gjessing 1976; Beamish and Van Loon 1977; Hen-
riksen and Wright 1978; Galloway and Likens 1979), in groundwater (Grahn
1980), in the snowpack (Schofield 1977; Seip 1980) and in soils sensitive
to acid precipitation (Hall et al. 1980; Hermann and Baron 1980). Water
concentrations of aluminum have-Been found to correlate with freshwater pH
(Johnson, N. 1979; Cronan and Shofield 1979; Wright et_ £]_• 1980). Thus
aluminum toxicity is under extensive investigation as a major cause of
fish mortality in waters of low pH (Baker and Schofield 1980; Muniz and
Leivestad 1980a, b; Schofield and Trojnar 1980).
Gill destruction has been shown to be the primary biotic effect of
aluminum on fish (Muniz and Leivestad 1980a, b; Schofield and Trojnar
1980). At the time of snowmelt, aluminum toxicity may be sufficiently
severe to cause anoxia and mortality in fish populations (Schofield 1980).
All of these effects are found to occur at water pH values (4.4-5.2) that
of themselves would not produce gill damage or any other major physiologi-
cal stresses in fish; in fact, elevated aluminum concentrations may exert
an antagonistic effect on acid stresses to fish at pH values below 4.4
(Leivestad et £l_. 1980; Schofield 1980; Schofield and Trojnar 1980).
Experiments with brook trout showed that while sensitivity to low pH
decreases with increasing age of the individual, sensitivity to elevated
aluminum levels increases with increasing age; fry are the most sensitive
life stage to combined low pH and elevated aluminum while the effects of
low pH on eggs are largely mitigated by increasing aluminum concentrations
(Baker and Schofield 1980). In sum, the distinction of various forms of
aluminum present in water, as well as the timing of aluminum and hydrogen
ion pulses in relation to the presence of sensitive fish life-history
stages, are important considerations in the evaluation of potential alumi-
num effects on indigenous fish populations.
Acid-induced physiological stress can lead to diminished fish growth
and a variety of other sublethal impacts (Beamish et al. 1975; Ryan and
Harvey 1980). Physical deformities observed in faTReacf minnows (Mount
1973) and white suckers (Beamish 1972; Beamish et &]_. 1975) may be due to
calcium resorption from bone tissue (Harvey 1973J. Both epithelial cell
degradation and altered blood protein function can predispose fish to in-
fection or disease (Daye and Garside 1976; Falk and Dunson 1977; Fromm
1980). A generally weakened condition may diminish feeding efficiency and
other competitive advantages while increasing the likelihood of predation.
Ill
-------
Fish tolerance to depressed water pH is found to increase in waters
of high conductivity and plentiful calcium concentration (Howells and
Holden 1979; Muniz and Leivestad 1980a; Leivestad et. aK 1980). At the
same time, naturally acidified waters are more toxic to adult fish and
salmonid eggs, alevins, and fingerlings than water experimentally acidi-
fied with sulfuric acid, possibly due to elevated trace element levels.
Other factors modifying fish responses to acidification include (Howells
and Holden 1979):
• age, size, sex, condition;
• sexual maturity;
• acclimation;
• genetic tolerance;
• ambient temperature; and
• the duration of acid events.
Seasonal changes are also an important factor as overwintering fish
may be in their weakest condition during spring snowmelt, or fish may be
subjected to additive thermal stresses in summer (Fritz 1980). The exis-
tence of a genetic basis for pH tolerance and survival has been estab-
lished for brown trout, and the role of genetic variation in salmonid pop-
ulation responses is a subject of intensive research (Gjedrem 1976, 1980;
Edwards and Gjedrem 1979).
Reproductive failure in fish populations often results from an in-
ability of the female to reach spawning condition, either through altered
breeding behavior or disruptions of endocrine functions. Abnormal serum
calcium levels, observed in white suckers that failed to spawn in acidi-
fied waters, have been linked to the latter mechanism {Beamish et ^1_.
1975). Other reproductive effects involve the tolerance of fis"R~eggs and
developing embryos. The genetic material of developing fish ova is sus-
ceptible to acid damage and the maturation of ova in several species is
prevented by acid conditions (Menendez 1976; Ruby et aj_. 1977; Kennedy
1980). Schindler (1979) reported an elevated incidence of embryonic mor-
tality and deformity in fathead minnows at pH 5.8, despite the fact that
adults survived pH values above 5.5. The failure of population recruit-
ment, in spite of spawning, results from the increased sensitivity of fish
eggs, fry, and fingerlings, and may explain a major portion of observed
population reductions (Rosseland et a_L 1980). Perch, trout, and white
suckers are all found to vary in pH sensitivity at different stages of
their life history (Beamish et aj_. 1975; Harvey 1979).
The spawning of mature fish may also be disrupted by acid-induced
invasions of benthic flora that alter the suitability of breeding sites
112
-------
(Fritz 1980). Stream acidification is known to provoke avoidance reac-
tions in brown trout populations, possibly forcing tnem from preferred
nesting areas (Hall et _al_. 1980). As shown in Figure 25, this species
tends to gather in areas of most favorable water quality during pronounced
acidification events. Fish are particularly vulnerable to physiological
stress during spring when ice cover and depleted oxygen in deep waters
trap populations in surface layers while the early snownielt is increasing
the acidity of the water (Schofield 1980). The variability observed in
fish population responses to lake acidity may in large part be accounted
for by differences in the availability of areas protected from acidifica-
tion (Muniz and Leivestad 1980a).
In Norway, fish population losses first observed in the 1940s were
found to accelerate after 1960. As shown in Figure 26, an increasing
percentage of lakes in four size classes have lost their brown trout
population over time. Rates of population decline are greatest in south-
ern Norway; trends in fish population status indicate that 80 percent of
all trout will be lost by 1990 as a result of habitat acidification (Muniz
and Leivestad 1980a). Sixty percent of brown trout populations above 800
meters in Norway are now extinct and trout populations are absent from
over 200 high-altitude lakes of New York and Ontario where they once were
common (Harvey 1979). Species of lake trout, brown bullhead, white
sucker and several cyprinids have been eliminated from Adirondack lakes
(Schofield 1976). In streams and rivers, brook trout populations are
thought to be the species most impacted by acid deposition in the
Adirondacks (Pfeiffer and Festa 1980) and Great Smokey Mountain National
Park (Mathews and Larson 1980).
Failure in the recruitment of new age classes, due to physiological
and reproductive stress, is known to be the chief mechanism of fish popu-
lation extinction (Jensen and Snekvik 1972; Scnofield 1976; Ryan and Har-
vey 1977). Accompanying losses of older fish are also common, resulting
in large-scale alterations of formerly stable age distributions (Beamish
et _a_l. 1975; Harvey 1980; Rosseland et a_[. 1980). Severe snowmelt effects
Tn any given year may contribute to "tTTe selective absence of different
year classes in acidified surface waters (Hultberg 1977). Recruitment
failure may also lead to gene pool reductions which limit potential selec-
tion for genetic tolerance in surviving populations (Beamish and Harvey
1972). Indirect effects can result from altered inter- ana iritra-specific
competition (Henriksen jit jH. 1980). For example, the relatively acid-
tolerant rock bass has been observed to increase in size due to reduced
competition for available food (Ryan and Harvey 1977), while piscivorous
fish may face a loss of juvenile prey and resort to cannibalism of their
own young (Fritz 1980).
Many questions remain about the exact nature and extent of acid depo-
sition effects on aquatic fauna. Establishing dose-response relationships
for organisms is complicated by the need to control for other factors,
such as trace element concentrations and the duration of acidic episodes
(Fritz 1980). Additional research is needed to correlate fish population
113
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SUMMER
pH 5.2
NORMAL HABITAT SELECTION
SPRING MELT
pH L.S
GROUND-
WA
habitat selection during acio INClOENTS
Figure 25. Fish habitat selection in response to freshwater acidification.
(From Muniz and Lievestad 1980a)
114
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Sample
Lake Size Size
5-10 ha 361
10-40 ha 103
40-100 ha 167
> 100 ha 120
Figure 26. Extinction of brown trout populations in lakes of southern
Norway during the period 1940-1970. Populations lost in each ten-year
period are indicated for four different categories of lake size.
(Adapted from Muniz and Leivestad 1980a)
115
-------
status with pH, alkalinity, and aluminum concentrations in potentially
sensitive regions; monitoring for fishkills, blood chemistry stress and
changes in growth, age-class composition and reproductive success will
permit the determination of trends in aquatic impact over time (Harvey
1980). Kelso et al. (1981) suggest that subtle qualities of surface water
chemistry, bioTTcH'ifestyle and habitat modification must be considered
in a complete assessment of biological effects.
5.3.6 Aquatic Ecosystem Response
The effects of acid deposition in aquatic ecosystems are numerous,
complex, varied, and highly interrelated. In regions sensitive to acid
inputs, virtually every aspect of ecosystem structure and function is sub-
ject to alteration. The nature of ecosystem damage in many ways reflects
the variety and severity of responses observed in plant and animal popula-
tions. At the same time, the extent or rate of ecosystem degradation is
intimately linked to the physico-chemical qualities characterizing lakes,
rivers, and streams of different regions and continents. Effects may thus
be anticipated to result from a combination of acid-induced changes in
water chemistry and biotic composition. Much of the current state of
knowledge has been gained through long-term interdisciplinary research
initiated in Norway. Research currently underway in eastern North America
has confirmed many findings of the Scandinavian works, but regional dif-
ferences demand a great deal of care in applying the Norwegian experience
to North America. Also, many complex issues remain to be resolved among
the wide variety of scientific disciplines contributing to studies of
aquatic ecosystem-level impacts.
Freshwater acidification has been demonstrated to alter nutrient
cycling in streams and lakes by suppressing organic litter decomposition.
As water pH levels decline, the activity of aerobic bacteria is diminished
and functions of decomposition and mineralization are gradually taken over
by less efficient fungi (Grahn et al_. 1974; Hultberg and Grahn 1976;
Likens 1976). Reductions in microbial biomass have been measured by de-
clines in weight loss of leaves in litterbags placed in acid lakes (Hend-
rey et jH. 1976); depressed community respiration has also been shown to
occur-in response to increased acidity (Traaen 1980). As a result of de-
pressed microbial activity, organic litter accumulates at an accelerated
rate and fewer nutrients are made available to primary producers (Friberg
et al. 1980; Gahnstrom et al. 1980; Hall and Likens 1980; Hendrey and
VerTucci 1980). These effects may be particularly pronounced in streams,
where allochthonous material is the most important nutrient source
(Traaen 1980).
Nutrients in the water column can be further reduced by interactions
with metals in precipitation and surface run-off. Phosphorus and possibly
silica are precipitated to the sediments by aluminum and iron ions present
in waters of low pH (Dickson 1980). Organic carbon may be lost in a simi-
lar fashion (Muniz and Leivestad 1980a). Work in terrestrial ecosystem
116
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response also suggests that watershed acidification may result in reduced
exports of phosphorus and other nutrients to aquatic systems (Hendrey et
al. 1980a).
The effects of acidification on nutrient cycling are stimulated by
the spread of periphyton and aquatic moss to deeper substrates. These
plants form thick mats which effectively seal off the sediments from chem-
ical exchange with the water column, inhibiting the release of sediment-
bound nutrients and detritus (Grahn et al_. 1974; Hendrey and Vertucci
1980). Anaerobic conditions created in upper sediment layers suppress
bacterial and invertebrate populations alike while promoting the estab-
lishment of odor-producing, anaerobic bacteria (Hendrey 1979). Sediment
aeration is further inhibited by reductions of rooted macrophyte communi-
ties (Leivestad et a^L 1976).
The aquatic mosses, especially Sphagnum, may also contribute to
freshwater acidity and oligotrophication (Grahn £t aK 1974; Grahn 1977;
Hendrey and Vertucci 1980). Once established, these acidophilic plants
are believed to accelerate the acidification process by giving off hydro-
gen ions in exchange for available cations in the water column. This
process is enhanced by the abundance of trace metal ions in acid lakes,
however essential elements such as calcium and potassium are also readily
sorbed by Sphagnum. The result is a combination of increased acidity,
reduced buffering capacity, and diminished fertility which together con-
tribute to biotic impact and accelerate the oligotrophication of acid
lakes. A simplified flow diagram encompassing major portions of the over-
all process is presented in Figure 27. The Sphagnum invasion and its
chain of effects are thought to be initiated by the acid pulse of spring
snowmelt, when competing benthic flora and fauna are weakened or only
partially established (Hultberg 1977).
Energy flows between trophic levels may be disrupted or reduced when
primary productivity declines, or when variations occur in the relative
productivity of competitive species. As discussed above, sensitive fauna
may expend additional energy counteracting physiological stress, to the
detriment of their growth and successful reproduction. Food chains are
further altered and simplified through the elimination of some species and
decreased numbers of others. Benthic detritivores, for example, are known
to prefer to feed on detritus conditioned by bacteria rather than the rel-
atively cruder by-products of fungal decomposition; many also derive a
large portion of their energy from the bacteria contained in detritus
(Hendrey et jfL 1976). The reduction of bacterial populations accompany-
ing oligotrophication would adversely affect these detritivores.
Populations of invertebrates are further reduced by the loss of habi-
tat accompanying Sphagnum invasion (Grahn et al. 1974), potentially af-
fecting the variety and amount of food avaTTaFTe to fish, waterfowl and
117
-------
Figure 27. A schematic representation of the hypothesis of auto-oligotrophication
in acid lakes.
-------
other consumers at higher trophic levels (Hendrey 1978). The elimination
of fish populations, largely through alterations of prey-predator rela-
tionships, can also induce radical changes in energy transfer along the
food chain (Friberg et al. 1980; Henriksen et al. 1980; Muniz and Leive-
stad 1980a).
In sum, the effects of acid deposition on aquatic systems are cumula-
tive and they interact in a number of ways to compromise ecosystem struc-
ture and function. Successional processes that promote the gradual eutro-
phication of surface waters may be interrupted, causing the systems to
regress through less mature evolutionary stages to an increasingly oligo-
trophy state. The most pronounced effects will involve:
• reductions of nutrient cycling;
• disruptions of trophic relationships;
• alterations of the spatial and temporal distribution of species;
and
• arrested ecosystem evolution.
The long-term consequences of large-scale alterations are difficult to
predict with any certainty and the potential reversibility of these ef-
fects in aquatic ecosystems remains unknown.
119
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6.0 AIR QUALITV LEGISLATION
The most important piece of Federal Legislation dealing with air pol-
lution is the Clean Air Act, as amended in 1977 (42 USC 7401 et seq.).
In this chapter, the main features of the law are described. It should
be recognized that this discussion is by no means a complete presentation
of all aspects of this very complicated law. For additional details the
reader should refer to the summaries in Arbuckle et aJL (1979), Avery and
Schreiber (1979), and Quarles (1979). This chapter also describes the
Acid Precipitation Act of 1980 as well as other activities undertaken to
deal with the international problems associated with the long-range trans-
port of air pollutants. In specific cases local environmental regula-
tions, which are not discussed in this report, must also be reviewed,
since they may be more stringent than Federal and State regulations.
6.1 THE CLEAN AIR ACT
The principal features of the Clean Air Act are the requirements for
the Federal government to set standards for pollution levels and for the
states to develop State Implementation Plans (SIPs) to meet the standards.
Acronyms representing the major structural elements of this legislation
are provided in Table 29. A graphic overview of the Clean Air Act is
presented in Figure 28.
6.1.1 Federal Air Quality Standards
The Clean Air Act, as amended, embodies Federal efforts to establish
maximum permissible concentrations of major air pollutants throughout the
entire United States. National Ambient Air Qua!ity Standards (NAAQS) are
based on pollutant concentration per volume of air (ug/m^ or mg/m^)
and have been set for six criteria pollutants: sulfur dioxide (SO2),
nitrogen dioxide (NO2), carbon monoxide (CO), ozone (O3), hydrocarbons
(HC) and particulates. A comparable NAAQS has also been formulated for
lead (Pb).
The primary ambient air quality standards specify atmospheric concen-
trations of the criteria pollutants which, if surpassed, could adversely
affect human health. Secondary ambient air quality standards specify
those ambient concentrations which, if exceeded, could have deleterious
effects on public welfare. The concept of public welfare includes living
resources such as fish and wildlife, and their habitats.
Both long-term and short-term NAAQS have been set. The former spec-
ify pollutant concentrations which cannot be surpassed on an annual aver-
age, and the latter dictate levels not to be exceeded over a period rang-
ing from three to twenty-four hours. These standards are based on criter-
ia documents which collect the scientific evidence regarding the effects
of air pollution on health and welfare. Table 30 presents the primary and
secondary NAAQS. While these standards are relevant to ground-level air
pollution control, they do not directly address problems of atmospheric
deposition in sensitive ecosystems.
120
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Table 29. Acronyms for principal components of the Clean Air Act.
AQCR Air Quality Control Regions
BACT Best Available Control Technology
LAER Lowest Achievable Emission Rate
NAAQS National Ambient Air Quality Standards
NESHAPS National Emission Standards for Hazardous Air Pollutants
NSPS New Source Performance Standards
PSD Prevention of Significant Deterioration
SIP State Implementation Plan
121
-------
Emissions
Legislation Jurisdiction
Figure 28.
Regulatory
Standards
Coverage
Requirements
NESHAPS
All industries
emitting hazardous
pol1utants
Strict
Abatement
NSPS
For specific
industries:
all new sources
and modifications
to existing sources
BACT
LAER
PSD
NAAQS
NAAQS achieved:
all new sources
and modifications
to existing sources
Reconstruc-
tion ap-
proval
BACT
NAAQS not achieved:
all new sources
and modifications
to existing sources
Construc-
tion per-
mit
Offsets
LAER
An overview of the Clean Air Act.
-------
Table 30. National primary and secondary ambient air quality standards.
Air Quality Standards (micrograms/m3)
Pollutant
Primary
Annual Maximum Concentration
Mean (Allowed Once Yearly)
Secondary
Annual
Mean
Maximum Concentration
(Allowed Once Yearly)
Sulfur Oxides (SOy)
(measured as S02)
Particulates
Carbon Monoxide (CO)
Ozone (03)
Hydrocarbons (HC)
Nitrogen Dioxide (N02)
Lead
80
75
100
365
(during 24 hours)
260
(during 24 hours)
10 milligrams/m3
(during 8 hours)
40 milligrams/m3
(during 1 hour)
240
(during 1 hour)
160
{during 3 hours
6:00-9:00 a.m.)
1.5
(Averaged over
calendar quarter)
60
1300
(during 3 hours)
150
(during 24 hours)
Same as Primary Standard
Same as Primary Standard
Same as Primary Standard
Same as Primary Standard
Same as Primary Stantard
(From Council on Environmental Quality 1980)
-------
6.1.2 Federal Emission Standards
The Clean Air Act directs EPA to establish New Source Performance
Standards (NSPS) for emissions of criteria pollutants which apply to
selected industries. The owner of a proposed major source must notify the
state within which the source will be located before beginning construc-
tion or operation. The levels set by EPA are based on the best demon-
strated technology for controlling emissions, taking into account economic
feasibility.
No construction permit which the act requires can be issued for a
source in any of the specified categories unless the emissions will meet
the NSPS standards. However, the 1977 Amendments to the law contain addi-
tional regulations which must be incorporated into the SIPs to control
emissions from new sources and modifications to existing sources. These
approaches impose emissions limitations which are more stringent than the
NSPS. The NSPS is therefore the upper limit for emissions from new
sources in the specified industries.
Other emission standards under the law deal with hazardous air pollu-
tants. The National Emission Standards for Hazardous Air Pollutants
(NESHAPS) have been adopted to control asbestos, beryllium (Be), mercury
(Hg), and vinyl chloride. These emission standards are based strictly on
human health considerations. Unlike the NSPS, the NESHAPS apply to both
new and existing sources of the hazardous pollutants.
Title II of the Clean Air Act, also known as the National Emissions
Standards Act, placed emission limitations on mobile sources of ambient
air pollution. Mobile emissions are measured in grams of pollutant per
vehicle mile over the lifetime of the source. For legal purposes, emis-
sion limits are expressed by emission factors, also measured in grams per
mile, as shown in Table 31. These factors are often lower than actual
emission limits due to potential emission increases associated with the
deterioration of the vehicle.
6.1.3 State Responsibilities Under the Clean Air Act
The goal of Federal air pollution control legislation is to make en-
forcement activities the responsibility of the States. Under the Clean
Air Act, the States are directed to establish State Implementation Plans
(SIPs) which embody the necessary abatement procedures required to attain
the NAAQS. However, if the State and local governments are unable to
develop an acceptable SIP, the EPA is directed to impose the necessary
regulations.
The 1970 legislation required that the country be divided into Air
Quality Control Regions (AQCRs). These 247 regions are highly variable
in size and independent of jurisdictional boundaries. Some are contained
within state boundaries while others include portions of several states.
The AQCRs were selected on the basis of regional homogenity in existing
124
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Table 31. Mobile source exhaust emission factors for 1979, 1980
and 1985-1990.
Emission factors (grams/mile)
Nitrogen Oxides Hydrocarbons Carbon Monoxide
1985- 1985- 1985-
1979 1980 1990 1979 1980 1990 1979 1980 1990
Automobiles 1.5 1.5 0.29 1.13 0.13 0.13 18.6 3.0 1.4
Light duty 1.73 1.73 0.41 0.94 0.94 0.31 14.5 14.5 3.87
gasoline trucks
Heavy duty 9.1 9.1 4.0 5.22 5.22 1.46 191.9 191.9 15.4
gasoline trucks
Heavy duty 19.9 19.9 5.35 4.5 4.5 2.85 27.0 27.0 27.0
diesel trucks
(Adapted from MITRE 1978)
air quality data, urban concentrations, types of industry and nature of
emissions, terrain, and meteorological conditions (Arbuckle et £l_.
Quarles 1979). When an AQCR is found to be in violation of tFTe NAAQS, the
SIP must delineate a combination of control strategies intended to bring
the AQCR into compliance.
Primary standards are to be attained "as expeditiously as practic-
able" and secondary standards within a "reasonable time". Each SIP con-
tains:
• a description of the air quality in each AQCR;
• an inventory of emission sources;
125
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• emissions limitations and compliance schedules for each source;
• a permit program for new source construction; and
• procedures for monitoring, reporting, and enforcement.
The 1977 Amendments required the states to designate areas which
meet the NAAQS (attainment areas) and those which do not (nonattainment
areas). Because the AQCRs are defined without regard to political bound-
aries and cross state lines, the States have not used AQCRs in defining
attainment and nonattainment areas. As a result, the focus of activity
under the Clean Air Act has shifted to the regulation of sources in at-
tainment and nonattainment areas, reducing the significance of the AQCRs
(Arbuckle et jiJL 1979).
a. Nonattainment areas. The SIPs must incorporate the NSPS discussed
above, ensuring that all new sources meet these standards. In addition
the SIPS must describe the mechanism for review of new sources in areas
not meeting the NAAQS (nonattainment areas). Any new source or modifica-
tion of an existing source with potential emissions of 100 tons per year
or more of particulates, SO2, N0X, volatile organic compounds, or
carbon monoxide is subject to a construction permit requirement. In non-
attainment areas, building and operating a major source requires the con-
trol of pollutants to the Lowest Achievable Emission Rate (LAER) for a
given industrial category. The LAER will always be at least as stringent
as the NSPS. For construction and operation permits to be issued the
following conditions must be met:
• all other installations in the state of the same ownership must be
in compliance, or on an approved schedule to achieve compliance,
with the abatement portions of the SIP;
• any projected new emissions must be more than equally offset by
reductions from existing sources of air pollution in such a way as
to produce "a positive net air quality benefit" in the area; and
• new sources in nonattainment areas must meet the LAER defined for
each case.
b. Attainment areas. The Clean Air Act Amendments of 1977 provide the
legislative basis for actions to maintain air quality in areas where
ambient standards are already being met (attainment areas). The amend-
ments for the most part adopted regulations which had been promulgated
earlier by EPA, incorporating them into Part C (Sections 160-169) of the
Act. These Sections establish the program to Prevent Significant
Deterioration of Air Quality (PSD). Under Section 161 tne SIPs must be
revised to include any measures needed to prevent the degradation of air
quality in attainment areas, and Section 168 specifies that until the
SIPs are modified to meet this requirement, the regulations issued
earlier by EPA will remain in force.
126
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As was the case for nonattainment areas, a preconstruction review of
major sources and modifications to existing sources in attainment areas
must also be made under the PSD program (Quarles 1979). The PSD program
applies to stationary sources in a specified set of industrial categories
that will emit at least 100 tons per year of any regulated air pollutant.
Sources not included in the specified industrial categories are also co-
vered if their emissions will exceed 250 tons per year. Sources located
in nonattainment areas also fall within the scope of the PSD program if
their emissions will effect attainment areas.
Within the PSD program, areas of the country which meet the NAAQS are
divided into three classes: Class I for areas of restricted growth, Class
II for areas of moderate growth, and Class III for industrialized areas.
Permissible increases in ambient pollution levels depend on the designated
area classification, but in all cases those levels must remain below the
NAAQS. New plant construction approvals depend on the classification of
the affected area, the resulting increase in air pollutant concentration,
and a general requirement to employ the Best Available Control Technology
(BACT), as determined by the permitting authority for each case (Quarles
1979). BACT may not be less stringent than NSPS. Maximum increases in
concentrations of SO2 and particulates have been promulgated for each
class of area and are shown in Table 32.
In addition to the requirements for BACT and the specification of
maximum allowable increments, the applicant for a preconstruction permit
must provide the results of an analysis of the impact on soils, vegeta-
tion, and visibility of the proposed source and any associated develop-
ment. There is also to be a public hearing. The applicant must provide
air quality monitoring data for a period of one year prior to construc-
tion, and agree to monitor air quality after operation begins.
c. Review by federal officials. Section 165 of the Act provides that,
in Class I areas, Federal Land Managers (FLM) are responsible for
assuring that potential emissions in and around areas under their juris-
diction will not contribute to the deterioration of "air quality related
values (including visibility)" of these lands, where Federal Land Manager
means the Secretary of the department with authority over the land. If
the Federal official charged with direct responsibility for management of
any lands within a Class I area or the Federal Land Manager of such lands
finds that the new facility will have an adverse effect on air quality
related values, then the permit may be denied, even if the increase in
pollutant concentrations is less than the maximum allowable increase for
a Class I area. The owner of the proposed facility can appeal to the
Governor who may recommend a variance, and the permit may be issued if
the FLM agrees. However, in cases where the Governor recommends a
variance in which the FLM does not concur, their recommendations go
to the President for a final decision. If the increase in pollutant
concentration exceeds the maximum allowable for Class I areas, a permit
may still be granted if the FLM certifies that no adverse impact on air
quality related values will occur.
127
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Table 32. Prevention of significant deterioration regulations.
Maximum allowable increase over baseline concentrations.
Land Classification
Pollutant Class I Class II Class III
Sulfur dioxide (micrograms/m3
Annual geometric mean
24 hour maximum
3 hour maximum
Particulates (micrograms/m3)
Annual geometric mean
24 hour maximum
(From Avery and Schreiber 1979)
128
2
5
25
20
91
512
40
182
700
5
10
19
37
37
75
-------
It is possible for an area to be classified attainment for certain
pollutants and nonattainment for others. In addition, pollutants from a
proposed plant in an attainment area may cross boundaries into a nonat-
tainment area. Thus the PSD and nonattainment regulations may overlap
(Quarles 1979). The provisions of a SIP constitute a highly complex re-
gulatory process which varies from state to state. Moreover, area desig-
nations are subject to periodic revision and careful attention is there-
fore required to assure that the most up-to-date versions are employed in
the regulatory activities of Federal agencies.
6.1.4 Discussion
The rationale for Federal control of air pollutant emissions is
based, at least partially, on the transboundary nature of the long-range
transport of air pollution and on the inability of any state to control
pollution which is generated outside its borders. Pollution from mid-
western states, for example, is transported to states in the northeast,
possibly using up their PSD increments or rendering areas nonattainment.
This in effect would restrict the potential for economic development in
the receptor states. The impacts would also reduce the economic value of
wildlife resources used for recreational purposes.
Activities under a SIP are designed to achieve and maintain
satisfactory air quality within that state. The transport of air
pollution across state borders will contribute to the deterioration of
air quality in receptor states, and stricter emission controls will then
be required on sources in those states. The rigid pollution controls
required for northeastern development have prompted lawsuits by some of
these states to enforce similarly stringent controls of emissions in
source states. The strengthening of midwestern SIPs would, in turn,
impact the economies of those states by increasing pollution control
costs.
Effective regulatory control of acid precipitation in the future may
directly involve both the NAAQS and performance standards. Emissions
standards for industries emitting acid precursors should result in a re-
duction of both total emissions and atmospheric loadings of sulfur and
nitrogen compounds through the use of control devices or fuels which have
been cleansed of their sulfur. The recent NSPS for coal-fired power
plants are a positive step towards decreased total SO^ emissions in the
eastern United States (Wetstone 1980). However, a revised set of primary
NAAQS based on more than strictly health-related considerations would
require additional legislation. Such legislation would require interdis-
ciplinary research, quantified scientific evidence, and consistent public
testimony relative to the subtle ecological effects of atmospheric depo-
sition (Berry and Bachmann 1977).
6.2 THE ACID PRECIPITATION ACT
A recent piece of Federal legislation directed at limiting air pollu-
tion effects on wildlife and habitat resources is the Acid Precipitation
Act of 1980 (42 USC 8901 et £§£.). This act mandates the formation of
129
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Table 33. Federal departments and agencies participating
in the Interagency Task Force on Acid Precipitation.
DOA
Department of Agriculture
co-chair
EPA
Environmental Protection Agency
co-chai r
NOAA
National Oceanic & Atmospheric Administration
co-chai r
CEQ
Council on Environmental Quality
DOC
Department of Commerce
DOE
Department of Energy
DOI
Department of Interior
DOS
Department of State
DHHS
Department of Health & Human Services
NASA
National Aeronautics and Space Administration
NSF
National Science Foundation
TVA
Tennessee Valley Authority
(From
Interagency Task Force on Acid Precipitation 1981)
130
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the Interagency Task Force on Acid Precipitation. Composed of DOI, EPA,
and ten other Federal departments and agencies (Table 33), this group is
charged with the development of a ten-year research program, the National
Acid Precipitation Assessment Plan (Interagency Task Force on Acid Precip-
itation 1981).
6.3 INTERNATIONAL COOPERATION
Since the 1972 United Nations Conference on the Human Environment in
Stockholm, the European Economic Community (EEC), the United Nations (UN),
and the Organization for Economic Cooperation and Development (OECD) have
all undertaken efforts to create guideines for international programs for
environmental protection. The long-range transport of air pollution has
been the subject of numerous ethical statements from international organi-
zations as well as proposed treaties in northern Europe. The first inter-
national agreement on the long-range transport of air pollutants entered
into by the United States was the "Convention on Transboundary Air Pollu-
tion" of the Economic Commission for Europe, a UN organization. While it
provides no abatement schedule or other enforcement provisions, it devel-
ops routes of international cooperation in research relevant to trans-
boundary air pollution and its socioeconomic and ecological repercussions.
The United States and Canada have signed a "Memorandum of Intent on Trans-
boundary Air Pollution" for an intensive research program leading to the
preparation of bilateral agreement (External Affairs Canada 1980). The
U.S.-Canada border crosses large, isolated areas which are well suited
for the observation of transboundary pollution and its effects. This
agreement therefore could serve as a model for future international
agreements (Wetstone 1980).
131
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GLOSSARY
acid - A substance which can donate hydrogen ions.
acid rain - A popular term for acid precipitation, or rain principally
containing the hydrolyzed end-products of oxidized sulfur and
nitrogen substances (dilute strong acids) and of pH less than 5.6,
the minimum pH expected from equilibrium with atmospheric carbon
dioxide.
advection - The horizontal movement of an air mass due to atmospheric
pressure gradients.
adsorption - Adhesion of a thin layer of molecules to a liquid or solid
surface.
aerodynamic diameter - An expression of the aerodynamic behavior of an
irregularly shaped particle in terms of the diameter of a sphere of
unit density having identical aerodynamic behavior to the particle
in question.
aerosol - Solid particles or liquid droplets which are dispersed or sus-
pended in air.
air quality standards - Concentrations of air pollutants which cannot leg-
ally be exceeded during fixed time intervals within specified geo-
graphic areas.
alkalinity (total) - A measure of the concentration of all acid-neutral-
izing substances in solution.
ambient air - Air surrounding a given point.
anion - A negatively charged atom or molecule.
antagonism - A relationship in which the combined action or effect of two
or more pollutants is less than the sum of their individual effects.
area source - A geographic location from which pollutants are emitted and
transported by advection.
bioindicator - A plant or animal species sufficiently sensitive to a given
pollutant to be useful as an indicator of the presence of the same
pollutant.
buffer - A substance in solution capable of neutralizing both acids and
bases, thereby maintaining the original pH of the solution.
buffering capacity - The ability of an entity (e.g., body of water or its
watershed) to neutralize introduced acids.
176
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catalyst - A substance capable of increasing the speed of chemical reac-
tions without itself undergoing physical or chemical change.
cation - A positively charged atom or molecule.
cellular permeability - The ability of cells to absorb or evacuate sub-
stances across cell membranes; a sensitive indicator of injury to
deep-lung cells.
dose - A measured concentration of a toxicant for a known time period
during which a subject is exposed.
dry deposition - Matter transferred from the atmosphere to ground in the
absence of precipitation; also the process of such transfer, includ-
ing surface adsorption of gases, sedimentation, Brownian diffusion,
and particle impaction.
dust - Solid particles generated by physical alterations of organic or
inorganic substances.
ecosystems - The interacting system of a biological community and its
environment.
effluent - Any solid, liquid, or gaseous waste emitted by a process.
emission standards - Standards based on the concentration of pollutants
from stacks that cannot legally be exceeded during fixed time inter-
vals within specified geographic areas.
fluorides - gases or particles containing fluorine compounds.
fly ash - Suspended incombustible or partially incinerated matter carried
in the gaseous products of combustion.
fossil fuel - Fuel derived from decayed organic matter from past geologic
ages.
fumigation - The natural or controlled exposure of biota to toxic gases or
volatile substances.
haze - Fine dust, smoke or vapor which reduces the transparency of air.
heterogeneous process - A chemical reaction involving reactants of more
than one phase or state, such as one in which gases are adsorbed to
aerosol droplets where the reaction takes place.
hydrocarbons (HC) - Organic compounds of carbon and hydrogen derived main-
ly from fossil deposits and vegetative sources; along with N0X,
they are implicated in the formation of ozone and other
photochemical oxi- dants in smog.
177
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hydrometeor - A product of the condensation of atmospheric water vapor.
impaction - An impinging or striking of one object against another; also,
the force transmitted by this act.
inversion - A thermally stable atmospheric condition wherein cooler air is
prevented from rising by a layer of warmer air above.
isopleth - A line on a map connecting points at which a given variable has
a specified constant value.
leach - To dissolve out through the action of a percolating liquid.
LRTAP - The long-range transport of air pollution, often analagous to
transboundary air pollution.
mass median diameter (MMD) - The geometric median size of a distribution
of particles based on weight.
mobile source - A moving source of air emissions.
monitoring - The use of gas sensing instruments or other devices to meas-
ure the concentrations of pollutants.
mutagen - Any agent that induces inheritable genetic change in living or-
ganisms .
nitrogen oxides (N0X) - A class of nitrogen and oxygen compounds which
includes nitric oxide (NO), nitrogen dioxide (NOj), and nitrogen
trioxide (NO3); derived from many natural sources as well as fossil
fuel combustion, they participate with hydrocarbons in the formation
of photochemical smog, and with sulfur oxides in causing acid precip-
itation.
oxidant - A chemical compound which has the ability to remove electrons
from another chemical species, thereby oxidizing it; also, a sub-
stance containing oxygen which reacts in air to produce a new sub-
stance, or one formed by the action of sunlight on oxides of nitrogen
and hydrocarbons.
ozone (O3) - A colorless to faintly bluish, unstable, pungent gas pro-
duced by electrical discharge in air, by solar ultra-violet radia-
tion, or by other photochemical reactions of mixtures of certain
hydrocarbons and nitrogen oxides; a strong oxidizing agent that is
phytotoxic at low concentrations.
PAN - The acronym for peroxyacetyl nitrate (CH3C-O-NO2); the principal
constituent in a homologous series of compounds, referred to as per-
oxyacyl nitrates or PANs, and formed as a product of photochemical
reactions involving nitrogen dioxide and hydrocarbons.
178
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particulates - Fine liquid or solid particles, such as dust, smoke, mist
fumes or smog, found in the air or in atmospheric emissions.
pathogen - Any biotic or abiotic agent capable of causing disease.
photochemical oxidants - Primarily ozone, nitrogen dioxide and PAN, along
with lesser amounts of other compounds, formed as products of atmos-
pheric reactions involving organic pollutants, nitrogen oxides, oxy-
gen and sunlight.
plume - The path taken by pollutants emitted continuously from a point or
area source.
podzol - Any of a group of zonal soils that develop in a moist climate,
especially under coniferous or mixed forest; they are characteristi-
cally acidic and low in essential plant nutrients.
point source - A stationary emitting point of air pollution.
Q
pollutant (air) - Any gas, liquid, or solid contaminant present in the
atmosphere in such quantity as to cause undesirable effects on living
organisms or materials.
precipitation scavenging - The capture of air pollutants within (rainout)
and beneath (washout) clouds by a hydrometeor.
primary pollutants - Pollutants which are emitted directly from an identi-
fiable source.
rainout - Removal of particles and/or gases from the atmosphere by their
involvement in cloud formation (particles act as condensation nuclei,
gases are adsorbed by cloud droplets), with subsequent precipitation.
secondary pollutants - Pollutants produced in the air by reactions involv-
ing primary pollutants and/or other atmospheric constituents.
sink - A reactant with or absorber of another substance.
smog (general) - a mixture of smoke and fog.
(London type) - A mixture of coal smoke and fog, with sufficient sul-
fur dioxide to induce chemical reducing properties.
(Los Angeles type) - A mixture of photochemical oxidants, with suffi-
cient ozone to induce chemical oxidizing properties.
sulfur oxides (S0X) - A class of sulfur and oxygen compounds composed
primarily of sulfur dioxide (SO2) and sulfur trioxide (SO3); they
are derived mainly from anthropogenic sources and participate with
nitrogen oxides in the formation of acid precipitation.
179
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synergism - A relationship in which the combined action or effect of two
or more pollutants is greater than the sum of the effects of the in-
dividual pollutants.
total suspended particulates (TSP) - An aggregate measure of solid and
liquid particles present in the atmosphere; it may contain toxic
trace substances and interact synergistically with sulfur oxides.
toxicant - A substance that kills or injures living organisms by its chem-
ical or physical action, or by altering the environment of the
organism.
trace elements - Atomic species suspended in air at natural concentrations
of less than 1.0 ppm; toxic and possibly carcinogenic elements at low
concentrations include beryllium, cadmium, fluorine, lead, manganese,
mercury and selenium.
washout - The capture of gases and particles by hydrometers falling be-
neath clouds.
wet deposition - Matter transferred from the atmosphere to ground in pre-
cipitation; also the process of such transfer.
180
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30272-101
REPORT DOCUMENTATION
PAGE
lu REPORT NO.
FWS/0BS-40.3
4. Title »nd subtitle j\-jr PolTut 1 ori and Acid Rain, Report 3
The Effects of Air Pollution and Acid Rain on Fish,
Wildlife, and Their Habitats - Introduction
7. Author(s)
Peterson, M. A.
9. Performing Organization Name and Address
Dynamac Corporation
Dynamac Building
11140 Rockville Pike
Rockville, MD 20852
12. Sponsoring Organization Name and Address US Department Of the
Interior, Fish and Wildlife Service/Office of Bio-
logical Services; Eastern Energy and Land Use Team,
Route 3 Box 44, Kearneysville, WV 25430
3. Recipient's Accession No.
5. Report Date
June 1982
6.
8. Performing Organization Rept. No.
10. Project/Task/Work Unit No.
11. Contract(C) or Grant(G) No.
(C) 14-16-0009-80-085
(G)
13. Type of Report & Period Covered
Final
IS. Supplementary Notes
16. Abstract (Limit: 200 words)
This introductory volume synthesizes results of scientific research rela-
ted to air pollution effects on fish and wildlife resources. It is intended
for use as a general reference to provide background information for the 8
ecosystem specific reports in this series: Deserts and Steppes, Forests,
Grasslands, Lakes, Rivers and Streams, Tundra and Alpine Meadows, Urban Eco-
systems, and Critical Habitats of Threatened and Endangered Species.
Related air pollutants are classified in three categories. A general
summary of pollutant origins, atmospheric transport, transformation and dep-
osition is presented. The report describes plant, animal and ecosystem resp-
onses to air pollution as well as factors affecting the sensitivity of recep-
tive ecosystems. This volume also briefly describes relevant features of
air quality legislation.
The three categoreis of air pollutants classified are: photochemical oxi-
dants, particulates, and acidifying air pollutants.
i7. Document Analysis a. a tmospher i c pollution, pollutants, exhaust emissions,
acidification, precipitation, terrestrial habitats, aquatic habitats
b. identifi.™/openEnded Terms f~\ue dust, f 1 ue gases, fumes, faze, oxidizers, smog,
smoke, soot, air content, pH, ecosystems, ecology, environmental effects
c. cosATt Field/Group 4qb, G; 57C, H, U, Y
18. Availability Statement
Release unlimited
19. Security Class (This Report)
unclassified
20. Security Class (This Page)
unclassified
21. No. of Pages
198
22. Price
(See ANSI—Z39.18)
181
OPTIONAL FORM 272 (4-77)
(Formerly NTIS-35)
Department of Commerce
-------
As the Nation's principal conservation agency, the Department of
the Interior has responsibility for most of our nationally owned
public lands and natural resources This includes fostering the
wisest use of our land and water resources, protecting our fish and
wildlife, preserving the environmental and cultural values of our
national parks and historical places, and providing for the enjoy-
ment of life through outdoor recreation. The Department assesses
our energy and mineral resources and works to assure that their
development is in the best interests of all our people. The Depart-
ment also has a major responsibility for American Indian reservation
communities and for people who live in island territories under U.S.
administration.
U.S.
FISH & WILDLIFE
SERVICE
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