Biological Services Program
FWS/OBS-80/40.10
JUNE 1982
Air Pollution and Acid Rain,
Report No. 10
THE EFFECTS OF AIR POLLUTION AND ACID RAIN
ON RSH, WILDLIFE, AND THEIR HABITATS
URBAN ECOSYSTEMS
Office of Research and Development
U.S. Environmental Protection Agency
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
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For sale by the Superintendent of Documents, U.S. Government Printing Office
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FWS/OBS-80/40.10
June 1982
AIR POLLUTION AND ACID RAIN REPORT 10
THE EFFECTS OF AIR POLLUTION AND ACID RAIN
ON FISH, WILDLIFE, AND THEIR HABITATS
URBAN ECOSYSTEMS
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, DC
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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 - urban ecosystems. U.S. Fish and Wildlife
Service, Biological Services Program, Eastern Energy and Land Use Team,
FWS/OBS-80/40.10. 89 pp.
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ABSTRACT
Air pollution and 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 volume on urban ecosystems is part of a series synthe-
sizing the results of scientific research related to the effects of air
pollution and acid deposition on fish and wildlife resources. The other
accompanying reports in this series are: Introduction, Deserts, Forests,
Grasslands, Lakes, Rivers and Streams, Tundra and Alpine Meadows, and
Critical Habitats of Threatened and Endangered Species.
General aspects of urban ecosystems relevant to a discussion of air
pollution effects are presented along with an outline of various other
types of ecosystem stresses. The bulk of this report describes plant,
animal and ecosystem responses to air pollution within the following
pollutant categories: photochemical oxidants, atmospheric metals, acid-
ifying air pollutants and miscellaneous urban air pollutants. The poten-
tial use of biological indicators in monitoring ambient urban air pollu-
tion is introduced and the report closes with a discussion of relevant
topics for further research.
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CONTENTS
ABSTRACT iii
FIGURES vi
TABLES vii
1.0 INTRODUCTION l
1.1 Report Organization 1
1.2 Summary of Effects \ 2
2.0 WILDLIFE AND HABITAT IN THE URBAN ECOSYSTEM 4
2.1 Urban Wildlife 4
2.2 Urban Wildlife Habitat 6
2.3 Urban Wildlife Communities 8
2.4 Environmental Stresses of Urbanization 10
3.0 EFFECTS OF AIR POLLUTION AND ACID RAIN ON
WILDLIFE AND HABITAT OF THE URBAN ECOSYSTEM 12
3.1 Photochemical Oxidants 13
3.1.1 Effects on Plants 15
3.1.2 Effects on Animals 17
3.1.3 Ecosystem Effects 20
3.2 Atmospheric Metals 23
3.2.1 Effects on Plants 27
3.2.2 Effects on Animals 33
3.2.3 Ecosystem Effects 44
3.3 Acidifying Air Pollutants 45
3.3.1 Effects on Plants 50
3.3.2 Effects on Animals 52
3.3.3 Ecosystem Effects 53
3.4 Miscellaneous Urban Air Pollutants 54
3.4.1 Carbon Monoxide 54
3.4.2 Fluorides 54
3.4.3 Pesticides 55
4.0 BIOLOGICAL INDICATORS OF AIR POLLUTION
EFFECTS IN THE URBAN ECOSYSTEM 57
4.1 Plant Bioindicators 57
4.2 Animal Bioindicators 59
IV
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CONTENTS (continued)
5.0 TOPICS FOR FURTHER RESEARCH 63
5.1 Baseline Study 63
5.2 Plant Effects 64
5.3 Wildlife Effects 65
5.4 Soil Effects 66
5.5 Ecosystem Effects 67
5.6 Conclusion 68
REFERENCES 70
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FIGURES
Number Page
1 An example of contrasting food chains among
urban fauna 9
2 Hourly ozone frequency distribution for Azusa,
California, 1965-1972 14
3 Hypothetical relationship of habitat diversification,
oxidant dose, and the economic value of vegetation
along a transect in the southern coastal air basin
of California 22
4 Roadside lead distributions in vegetation and the
soil profile of two highways of different age 29
5 Lead levels in roadside grass at varying traffic
volumes 30
6 Lead concentrations in earthworms at varying
distances from roads of different traffic volumes 35
7 Lead levels in whole bodies of mammals at varying
distances from a roadside 40
8 Lead concentrations in whole bodies of small
mammals from roads of varying traffic volumes 41
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TABLES
Number Page
1 Two classifications of urban wildlife 5
2 Urban wildlife habitats and related land use 6
3 Examples of wildlife refuges in American cities 8
4 A summary of environmental stresses common to
urban wildlife and habitat 11
5 Approximate ozone sensitivity of important
western conifers 16
6 The site of action of pollutant gases in the
respiratory tract of animals 18
7 Median concentrations (nanograms/m3) of metals
in urban and remote atmospheres with ratios of
urban-to-remote concentrations 24
8 Median concentrations (i_ig/£) of metals in wet
deposition from urban and remote sites, with
ratios of urban-to-remote concentrations 25
9 Metal content of wet deposition collected in
samples along transects from a nickel smelter
(mg/m3/28 days) 26
10 Average trace metal deposition in dustfall at
three sites in New York City 26
11 Mean metal concentrations (ppm) in soils of
Grand Rapids, Michigan, related to land-use
patterns 32
12 Accumulation sites of atmospheric trace metals
in vertebrates 34
13 Ratios of lead in tissues of urban songbirds to
concentrations in rural songbirds 37
14 Lead concentrations (ppm dry weight) in tissues
of deer mice (Peromyscus maniculatus) from
roadside sites of different traffic volumes 38
15 Lead concentration (ppm wet weight) in bats,
rodents and shrews 42
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TABLES (continued)
Number Page
16 Effects of ingestion and inhalation in sheep
exposed to automobile exhaust 43
17 Characteristics of some acidic atmospheric
sulfates in the 0.1 to 1.0 ym particle range 46
18 Comparative urban/rural precipitation chemistry 48
19 Potential responses of animal indicators to
air pollution 60
vi n
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1.0 INTRODUCTION
This report on urban ecosystems is one of a series presenting current
knowledge about the effects of air pollution and acid rain on fish, wild-
life, and their habitats. The purpose of the series is to assist U.S.
Fish and Wildlife Service biologists in the early detection and identifi-
cation of pollution damage and to suggest fruitful lines of research.
The report is based on recorded damage by urban pollution and on the
effects that can be predicted from the characteristics of urban air pollu-
tants. Interpretation of research findings and prediction are complicated
by the many other environmental factors that affect the productivity of
urban ecosystems and by the paucity of specific research that has been
done on the subject. Complete understanding of the ecological effects of
air pollution and acid rain on urban biota requires firm knowledge of the
other impacts to wildlife and habitat arising from the many stresses of
urbanization. The scope of this report, however, is limited to air pollu-
tion and acid rain. Other effects will be discussed only briefly.
With respect to acid precipitation, only recently have monitoring and
research been undertaken in the urban ecosystem. One reason for this is
the pressing need to understand acid deposition effects on sensitive sur-
face waters and remote ecosystems of the country. Another arises from
analytical problems brought on by the many different localized air pollu-
tants in urban atmospheres. These factors make it difficult to extrapo-
late research findings in remote ecosystems to the situations in cities.
Another subject of interest concerns the effects of air pollution and
acid precipitation on urban materials and structures, paints and finishes,
artwork, property values, human health, and drinking water supplies. Al-
though these are closely associated with the socioeconomic impacts of air
pollution and acid deposition in the urban ecosystem, a presentation of
these aspects is beyond the scope of this document.
1.1 REPORT ORGANIZATION
This dicussion of ecological effects on urban wildlife and habitat
will cover three general categories of air pollutants: the photochemical
oxidants, metals and other particulates, and the acidifying air pollu-
tants. A detailed description of the individual pollutants within these
categories is found in the introductory volume of this series, along with
a presentation of pollutant sources, transport, transformation, deposi-
tion, and fate in aquatic and terrestrial ecosystems. Other reports in
this series present the ecological effects of air pollution and acid pre-
cipitation within a variety of specific ecosystems.
The following chapter provides a brief introduction to urban wild-
life, habitat, community structure, and environmental stresses inherent to
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the urban ecosystem. The purpose of this discussion is to highlight im-
portant distinctions between man-modified and natural ecosystems relevant
to an ecological assessment of air pollution and acid rain.
Chapter 3 discusses the wide range of observed and potential effects
of air pollution and acid rain on urban plants, animals, and ecosystem
function. Material presented in the chapter extends in many instances
beyond the confines of the urban ecosystem for two reasons. First, direct
information on air pollution and acid rain effects in the urban ecosystem
is incomplete. Field observations in nonurban ecosystems, as well as ex-
periments involving laboratory animals and economic crops, can be used to
fill some of the information gaps.
The second reason is that urban-generated air pollution may in large
part be responsible for considerable disruption in downwind rural and re-
mote ecosystems. Despite the uncertainties of extrapolation, research
findings in these ecosystems may be indicative of effects to be antici-
pated in cities from similar mixtures of air pollutants.
The fourth chapter discusses selected plants and animals as bioindi-
cators of urban air pollution effects. This discussion is more practical-
ly oriented to the needs of field biologists as it is restricted to ob-
served, identifiable, and, in many instances, quantifiable biological ef-
fects. The purpose of this section is to synthesize information of use in
future efforts to design and implement biological monitoring systems sen-
sitive to the quality of the urban environment.
The report concludes with a discussion of needed research into air
pollution and acid rain effects in the urban ecosystem. Five major re-
search areas are proposed: baseline study of wildlife and habitat in the
urban ecosystem; effects on plants; effects on animals; effects on soils;
and effects on the structure and function of entire ecosystems. Each of
these subjects requires further elucidation before an integrated assess-
ment of air pollution and acid rain effects can be developed and applied
to the protection of living resources in the urban ecosystem.
1.2 SUMMARY OF EFFECTS
As a general rule, the biotic effects of air pollution and acid rain
are primarily influenced by the genetic constitution of both individuals
and populations. These effects in the urban ecosystem are summarized in
the following paragraphs.
• Photochemical oxidants may be responsible for acute visible injury
to plants during peak episodes as well as both visible and meta-
bolic injury from chronic low-level exposures. Animals may suffer
eye irritation from peak episodes; however, the effects of chronic
exposures, if any, are unknown. Small mammal populations from
urban areas have been demonstrated to acquire a degree of genetic
resistance to ozone toxicity.
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• Atmospheric metals are absorbed, translocated, and accumulated by
plants, and most metals stimulate plant growth at low concentra-
tions. Visible and metabolic injuries rarely occur at ambient
urban concentrations, although there are many recorded instances
of selective or total plant elimination due to gross metal pollu-
tion near emission sources, particularly base metal smelters.
Animals also accumulate atmospheric metals and other toxic partic-
ulates via inhalation and ingestion. These substances may accumu-
late at variable rates in different body tissues, yet pathological
effects similar to classic lead and mercury poisoning from inges-
tion have not been observed.
• Acidifying air pollutants are associated with direct visible in-
jury to plants during peak episodes as well as metabolic injury
from continuous low-level exposures. They are well known for
causing the elimination of lichens and mosses from city centers.
The deleterious effects of acidifying substances on the physiology
and reproduction of amphibians and aquatic organisms are well doc-
umented; however, evidence of direct effects on terrestrial inver-
tebrates and vertebrates is scarce.
t Miscellaneous urban air pollutants include carbon monoxide, and,
in many cases, fluorides and pesticides. While no demonstrable
effects on the well-being of urban plants and animals from ambient
carbon monoxide exposures are known, both fluorides and pesticides
may contaminate plants used by wildlife for nutrition. Resulting
animal effects include a wide variety of severe pathological dis-
orders.
Air pollutants may be responsible for widespread disruption of the
structure and function of urban ecosystems. Major ecological effects to
be anticipated include alterations in soil chemistry and texture; micro-
bial activity; biogeochemical cycling and nutrient exchange; the energet-
ics of food webs; species abundance, diversity, distribution, and inter-
action; and general patterns of ecosystem succession and evolution. These
complex, interactive effects result from direct air pollution injury to
biotic ecosystem components as well as impacts to abiotic components
brought on by the chronic accumulation of deposited metals and acids.
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2.0 WILDLIFE AND HABITAT IN THE URBAN ECOSYSTEM
For purposes of ecological assessment, the urban ecosystem may be
perceived as a spatial continuum radiating outwards from the focal point
of the downtown commercial district. It encompasses suburban residential
and industrial developments, parks and greenbelts, various rights-of-way,
and other forms of land use extending to the rural or agricultural out-
skirts of cities. Along this continuum of ecosystem components, one finds
a more or less limited range of wildlife as well as pet, farm, and zoo
animals. Species diversity tends to increase along this continuum, al-
though suburban areas may harbor a greater diversity than outlying agri-
cultural areas, and other exceptions, such as mid-city parks and wildlife
refuges, are frequently encountered. In many cases, rapid urbanization
and industrial development at the city fringe may have the effect of re-
versing this spatial continuum.
Terrestrial habitats range from natural and disturbed woodlots and
open spaces to the highly ornamental plantings and landscapes character-
istic of many cities and their suburbs. Aquatic habitats range from chan-
nelized streams, river fronts and man-made impoundments to relatively un-
disturbed lakes and drainage systems. The physical location of wildlife
and supporting habitats within this continuum significantly determines the
nature and degree of most impacts to biota arising from the many environ-
mental stresses of urbanization.
2.1 URBAN WILDLIFE
The fauna inhabiting American cities ranges from the truly wild spe-
cies of surrounding ecosystems to highly adaptive urban species that are
no longer abundant in the wild. As a group, the larger carnivores have
been virtually eliminated from metropolitan areas, while a significant
number of smaller carnivores (foxes, coyotes), omnivores (opossums, rac-
coons, skunks), and herbivores (rabbits, deer) are occasional residents of
the urban ecosystem. Migratory birds and waterfowl are a seasonally im-
portant fauna, and over two hundred different bird species are native to
American cities (Stearns and Montag 1974). The more permanent inhabitants
include the rodents, the burrowing and tree-dwelling small mammals, vari-
ous members of the lower vertebrates (lizards, salamanders) and inverte-
brates (insects, earthworms), pets, livestock, and zoo animals. Urban
aquatic fauna ranges from the usual constituents of naturally occurring
stream, river, and lake communities to those associated with elevated or-
ganic or toxic pollutant loads, and in some cases includes introduced
exotic species.
Several different schemes have been devised to classify the many
kinds of wildlife found in the urban environment. Some are based on simi-
larities in the life habits of groups of species, while others are founded
on human perceptions of their desirability or function in city life. Such
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classifications are integral to an ecological assessment of air pollution
effects since they facilitate an understanding of the different biological
responses typical of various wildlife groups. Two classifications of ur-
ban wildlife and the animals which they comprise are presented in Table 1.
A third, offered by Stearns (1972), is based on relationships of wildlife
to man, and includes:
• species adapted to life with man and partially dependent upon him
for specific habitat factors (food, water, cover, breeding sites);
• species tolerant of man which sometimes take advantage of special
land uses; and
• species which shun contact with man and whose habitat requirements
are not fully satisfied within urban areas.
Table 1. Two classifications of urban wildlife.
Category
Major constituents
Classification I
Domestic Species
Nuisance Species
Wild Species
Classification II
Desirable Species
Unavoidable Species
Adaptive Species
Pet birds and mammals
Livestock and other farm animals
Urban arthropods
Pigeons, starlings, crows, and house
sparrows
Rats and house mice
Feral mammals
Terrestrial and aquatic invertebrates
Amphibians, reptiles, and fish
Migratory birds and songbirds
Small and large mammals
Domestic pet and farm animals
Rats, mice, and a variety of household
insects which persist in human settle-
ments and thrive on wastes
Pigeons, sparrows, crows, bats, and
other species which are well adjusted
to man-modified habitats
(Adapted from Leedy et al_. 1978; Sudia 1978)
5
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None of these classifications is exhaustive, and the distinctions be-
tween groupings are not always clear. For example, feral and stray domes-
tic animals are considered to be a serious nuisance in metropolitan areas
and may often outnumber pets. On the other hand, some adaptive birds tra-
ditionally considered to be a public nuisance are more and more frequently
perceived by the public as desirable cohabitants. These classifications
are nonetheless useful in determining the locations and basic habitat re-
quirements of the different groups of urban fauna. Moreover, they facili-
tate planning efforts aimed at the acquisition of data relevant to an as-
sessment of air pollution effects, since indirect effects of air pollution
on wildlife through habitat changes are more likely than direct effects.
2.2 URBAN WILDLIFE HABITAT
Habitat for wildlife is most simply defined as the cover, food, and
water required by a species for its nutrition, protection, and reproduc-
tion. The term also encompasses a variety of other conditions required
for the maintenance of wild species (Stearns 1972). Depicted in Table 2
are the major categories of natural habitat available to urban wildlife,
along with a few of the many land uses associated with them. These habi-
tats range from natural forests, grasslands and marshes at the city fringe
to the variety of disturbed or landscaped areas, such as gardens, cemeter-
ies, vacant lots, and construction sites, which accompany urbanization.
Together, they represent a wide variety of successional communities.
Table 2. Urban wildlife habitats and related land use.
Natural Habitat Type Exemplary Urban Land Use
Open - grassland Railroad and canal banks
Open - arable Gardens
Parkland Institutional grounds
Woodland edge Golf courses
Cliff and ravine Quarries
Wetland Sewage farms
Aquatic Reservoirs
(From Gill and Bonnett 1973)
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For most wild urban species, vegetation provides the important habi-
tat requirements of food and shelter. Its primary productivity forms the
trophic base that supports wildlife through herbivorous, carnivorous, and
detritivorous food chains. The diversity of this vegetation is one of the
factors which encourage or limit the distribution of urban fauna. In gen-
eral, as one moves along the continuum from city center to greenbelts and
other outlying areas, vegetation increases in diversity, and the essential
habitat requirements of a greater number of wildlife species are satis-
fied. This greater diversity of vegetative cover is important in protect-
ing wildlife populations from catastrophic occurrences, such as the loss
of a prime food crop, and is essential for providing nesting sites, pro-
tection for young, and spaces for travel or rest (Stearns 1967, 1972).
Diversity may also serve to reduce the risk of potential impacts from air
pollution in comparison with areas of decreased vegetative diversity. On
the other hand, the distribution of domestic animals and adaptive or
nuisance species, which rely on urban structures as a significant
component of their habitat, is much less influenced by the diversifica-
tion of the urban flora. The effects of introduced exotic plants on
species distribution in urban areas is virtually unknown.
While this concept of increasing habitat diversity with increasing
distance from the metropolitan center is descriptive of the situation en-
countered in most American cities, there are exceptions. In the Los An-
geles basin, for example, steep inaccessible slopes of dense chaparral
pervade the metropolitan area. This unique topography greatly facilitates
wildlife movement, while providing virtually no habitat transition between
the wild and urbanized ecosystem (Gill and Bonnett 1973).
Habitat continuity, like diversity, is an important feature regulat-
ing the density and distribution of urban wildlife populations. Unplanned
metropolitan growth tends to reduce suitable wildlife habitats and lead to
spatially isolated pockets. Although urban ecosystems include rights-of-
way, such as roads, railways, utility lines, and watercourses which link
vital habitats, discontinuity severely lowers the value of habitat for
sustaining vigorous urban wildlife populations (Stearns 1967).
The importance of maintaining continuous habitats has come to be in-
creasingly recognized in urban planning. Thillmann and Monasch (1976)
refer to them as "environmental quality corridors" and suggest that hydro-
graphic networks serve as the main structural element connecting public
parks, woodlots, marshes, and other wildlife refuges. In a number of
American cities, as exemplified in Table 3, significant amounts of the
natural ecosystem have been left intact, expressly to accommodate the hab-
itat requirements of migratory birds and other esthetically valued wild-
life. Environmental quality corridors may also improve the urban environ-
ment for human inhabitants. The effectiveness of habitat preservation in
supporting wildlife communities nevertheless remains a function of the
intensity of air emissions and water pollution from adjacent development
(e.g., roads, sewer outfalls, storm drains) and other stresses inherent
to the urban ecosystem.
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Table 3. Examples of wildlife refuges in American cities.
City Wildlife Area Size (acres)
Chicago, IL Cook County Forest Preserve (woodland) 40,000
New York, NY Jamaica Bay Refuge (tidal marsh) 11,840
Philadelphia, PA Tinicum National Environmental Center 898
(tidal marsh and estuary)
Washington, DC Rock Creek Park (wooded river valley) 1,754
(From Gill and Bonnett 1973)
2.3 URBAN WILDLIFE COMMUNITIES
Considerable variations exist in the nature and organization of urban
wildlife communities. For purposes of impact assessment, the adaptive or
nuisance species are best perceived in terms of individual populations
because their interactions with other animal groups are often minimal.
Moreover, they are largely capable of short-circuiting natural food chains
by utilizing the gratuitous wastes of man. As a result, they are less
sensitive to disturbances of natural habitat caused by air pollution.
Their flexible behavior patterns permit them to adjust rapidly to the
rigors of urbanization over the course of their lifetime, without the
necessity for genetic change (Gill and Bonnett 1973).
Wild species, on the other hand, remain significantly dependent on
natural or slightly modified food chains and are best studied in terms of
their niche in the overall wildlife community. In comparison with the
simple, relatively stable food chains of nuisance and domestic species,
the complex trophic webs upon which wild urban species depend may contain
components that are highly vulnerable to environmental stresses. Air pol-
lution or acid deposition may indirectly impact urban wildlife by elimin-
ating or reducing populations of food organisms and otherwise altering the
energetics of established food chains. This contrast is illustrated in
Figure 1 for the case of food chains.
Substantial modifications in the community structure and trophic re-
lationships of urban wildlife have been observed by comparing them with
rural counterparts. Gill and Bonnett (1973) have reported that birds
feeding primarily on rodents in rural areas may convert to a diet of small
birds, with an occasional rat or mouse, while living in the city. They
also state that predators may shift from their usual prey to alternative
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Pigeon
Cockroach
D
Mouse
Food Scraps
Stored Foods
Adaptive
or
Nuisance
Species
Wild Species
Figure 1. An example of contrasting food chains among urban fauna.
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food supplies afforded by human activities, thereby adopting an essential-
ly omnivorous habit. Some species, notably the insects, flourish in the
absence of many of their natural predators. Urban development often af-
fords unique opportunities for organisms to advantageously modify rela-
tions with their environment, sometimes leading to genetic adaptation. A
thorough understanding of urban-induced community alteration will include
the effects of air pollution as one of the many environmental stresses.
2.4 ENVIRONMENTAL STRESSES OF URBANIZATION
A straightforward assessment of air pollution effects in the urban
environment is complicated by the many other environmental stresses im-
pacting wildlife and habitat. Table 4 provides a concise summary of other
major influences. Any combination of these circumstances can compound or
override the potential effects of air pollutants on urban wildlife and
habitat; hence, a thorough understanding of them is important.
Habitat destruction is clearly the source of greatest stress to the
viability of urban wildlife populations. The peculiarities of urban cli-
mate are significant in reducing the productivity of vegetative habitat
and are important factors in the predisposition of several plant species
to air pollution damage (Gill and Bonnett 1973). Water pollution and
other urban residuals are an ever-present stress and along with soil con-
dition affect the availability of desirable food organisms and generally
regulate the extent to which wildlife can make full use of various habi-
tats. The remaining stresses are attributable to urban culture and the
technologies employed by man in the daily functions of city life.
The many environmental repercussions of urbanization serve to empha-
size that air pollution is but one of several causes of animal mortality
or decline in metropolitan areas. Stearns (1967) has suggested that these
stresses can account for subtle changes in the territoriality, adaptabil-
ity and competitive relations of wildlife species, the carrying capacity
of habitats and the many interdependences of wildlife on the physical en-
vironment. Clearly, the extent of these baseline impacts can be exacer-
bated to varying degrees by chronic or acute episodes of air pollution and
acid precipitation.
10
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Table 4. A summary of environmental stresses common
to urban wildlife and habitat (air pollution and acid
rain excluded).
HABITAT DESTRUCTION
removal of vegetative cover
earthmoving, extractive processes, and agricultural development
land drainage or inundation
road and lot paving
solid waste disposal
stream channelization (increased runoff and lowered water table)
dredging or filling of coastal areas
replacement of older buildings with modern architecture
URBAN CLIMATE (relative to rural environments)
decreased total and ultraviolet radiation
increased amounts and frequency of cloudiness, fog, and precipitation
decreased relative humidity
increased annual mean temperatures and winter minimum temperature
altered wind speed and gustiness
WATER POLLUTION
domestic waste and other organic pollution
liquid industrial wastes
refractory substances in toxic runoff (metals, biocides)
heat and radioactive wastes
erosion and si Hat ion
SOIL CONDITION
• nutrient and mineral leaching
• drying and compaction
• metal and biocide accumulation
t acidification and metal mobilization
STATIONARY STRUCTURES (towers, guy wires)
REFLECTIVE SURFACES (plate glass)
INTRODUCTION OF EXOTIC SPECIES
MOTOR VEHICLE STRIKES
(Adapted from Gill and Bonnett 1973; Leedy ert aj_. 1978; Stearns and Montag
1974)
11
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3.0 EFFECTS OF AIR POLLUTION AND ACID RAIN ON
WILDLIFE AND HABITAT OF THE URBAN ECOSYSTEM
Air pollution and acid rain have become ubiquitous environmental
stresses to wildlife and habitat in many urban areas of the United States.
Their biological effects may be classed as (Smith 1974):
• acute injury - direct impacts evidenced by obvious symptoms; or
• chronic injury - direct or indirect impacts with no noticeable
external symptoms.
Urban air pollutants, however, are not found in isolation from one ano-
ther, and a variety of synergistic impacts from ambient mixtures of air
pollutants and other urban stresses can be anticipated. Indirect effects
of depressed primary production, habitat degradation, and the contamina-
tion of food chains are also common wildlife stresses of air pollution in
the city (Newman 1980).
Inherited genetic susceptibility has been recognized as the most im-
portant factor regulating biotic tolerance to air pollution effects, as
evidenced by differential responses among plant varieties (Heggestad 1968)
and phylogenetic groups of mammals (Richkind and Hacker 1980). Inherent
genetic tolerance to air pollutants, however, may be of limited value to
the survival of species which are especially susceptible to injury from
other environmental stresses. The most tolerant trees of the San Bernar-
dino National Forest, for example, are also the most susceptible to de-
struction by fire, whereas the fire-resistant pines, responsible for re-
newed succession in this fire-adapted community, are rapidly succumbing
to the effects of urban air pollution (Kickert and Gemmill 1980).
The airborne contaminants of complex urban atmospheres as a whole
exert genetic selection pressures on the plant and animal life of cities.
Together with urban land-use patterns, they constitute a major factor con-
trolling ecosystem succession. Under conditions of chronic air pollution
or other urban disturbances, sensitive plant species decline and associ-
ated community alterations occur until plants which are more adapted to
stress become dominant (Glass 1979). Community diversity and biomass de-
crease while the ratio of respiration to photosynthesis rises in both dom-
inant species and the community as a whole. Accompanying changes in the
soil microflora from chronic low-level air pollution and acidification may
reinforce patterns of successional setback, although the reversibility of
such processes and the implications for ecosystem recovery are not as yet
known.
The following discussion is devoted to the observed and potential
effects of the photochemical oxidant complex, atmospheric metals, and the
acidifying air pollutants on urban plants, animals, and related habitat
factors of ecosystems. For more information, Jacobson and Hill (1970)
12
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provide a concise and comprehensive pictorial review of plant symptoms
generated by these air pollutants, and physiological effects in animals
have been reviewed by Lillie (1970), Newman (1975, 1980), and Gough et a/L
(1979). Discussions of air pollution and acid rain effects on forest,
grassland, lakes, and streams, which are often components of urban areas,
are provided in companion reports of this U.S. Fish and Wildlife Service
series.
3.1 PHOTOCHEMICAL OXIDANTS
The photochemical oxidant complex, or smog, has for many years been
recognized as a major component of harmful air pollution in most urban
areas of the United States. Fueled by emissions of the primary reac-
tants - nitrogen oxides (NOX) and hydrocarbons (HC) - photochemical re-
actions generate ozone and a variety of secondary products. Important
among these are peroxyacetylnitrate (PAN) and like compounds, formalde-
hyde, other aldehydes, ethylene, acrolein, hydrogen peroxide, and a wide
range of sulfate-, nitrate-, and carbon-based aerosols.
Generation of this mixture is enhanced by several factors. Thermal
inversions coupled with intense solar radiation have long characterized
severe episodes of oxidant pollution in the cities of southern California.
The high-altitude city of Denver receives sunlight of sufficient intensity
to rapidly drive the various photochemical reactions. The components of
smog are often generated as well in suburban and rural areas downwind of
urban emissions, where ambient concentrations frequently exceed those of
the city. While many of these chemical are short-lived and others are
quickly removed from the atomsphere, the long-range transport of urban
oxidants and associated pollutants to agricultural and remote areas has
been observed and studied in different regions of the country (Fankhouser
1976; Cleveland and Graedel 1979).
Many cities and suburbs, indeed the greater part of the northeastern
United States, are in repeated violation of federal ozone standards (0.12
ppm). This highly reactive gas, of documented toxicity to plants and ani-
mals, is normally found at background concentrations of 0.01-0.06 ppm in
the lower atmosphere (National Research Council 1977). Elevated 1-hour
average concentrations ranging from 0.13 to 0.58 ppm have been recorded in
metropolitan areas of the United States. Figure 2 presents a frequency
distribution of hourly ozone levels in a California city for a 7-year
period. Concentrations as high as 0.3 ppm are seen to occur approximately
1 percent of the time (National Research Council 1977). Such values are
often sufficient to bring about eye irritation and respiratory distress in
people and other vertebrates. Moreover, they may be contrasted with
threshold exposures for large-scale visible plant damage unofficially set
in California at approximately 0.15 ppm (Hay 1970) and in the eastern
United States at approximately 0.13 ppm (Wester and Sullivan 1970).
13
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0.60s-
0.01
0.01
10
20 30 40 50 60 70
FREQUENCY, %
Figure 2. Hourly ozone frequency distribution
for Azusa, California, 1965-1972.
(From National Research Council 1977)
14
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3.1.1 Effects on Plants
The primary phytotoxins among the photochemical oxidants are ozone,
PAN, nitrogen dioxide, and aldehydes, although indirect evidence supports
speculation that other plant toxins are present in smog at lesser concen-
trations (Heggestad 1968; National Research Council 1977). Oxidant expo-
sure to plant tissues occurs through the opened stomata of leaves. Plant
sensitivity to oxidant injury is thus in part influenced by genetic and
environmental factors regulating stomatal function. Once inside the
leaves, oxidants attack membranes and other lipid components of cells,
damaging chloroplasts when they are physiologically active (National Re-
search Council 1977). The result is chronic or subacute injury often, but
not always, leading to visible damage, and a reduction in photosynthesis
and yield, especially among cultivars of the more sensitive species.
Ozone toxicity has been investigated in a wide range of herbaceous
and woody plants. It is found to accelerate senescence and to cause ex-
tensive injury to foliage, flowers, and fruit, evidenced as flecks,
stipple, chlorotic patterns, and necrotic lesions (National Research
Council 1977). Sensitive tree species used by urban wildlife for food and
cover include the white ash (Fraxinus americana), quaking aspen (Populus
tremuloides), and white oak (Quercus alba) (Davis and Wilhour 1976Ti
Evidence of a synergistic effect between ozone and sulfur dioxide on
leaves of sensitive tobacco strains was established by Menser and Hegge-
stad (1966). The threshold concentration of 03 necessary for leaf in-
jury was drastically reduced when SO? gas was present. Heck (1968) ob-
served that a combination as low as 0.03 ppm 03 and 0.1 ppm $62 could
injure sensitive tobacco. Otto and Daines (1969) reported that the sensi-
tivity of tobacco to 03 rose under conditions of increasing humidity and
postulated this factor as an explanation of the greater ozone sensitivity
of plants of the eastern United States relative to those inhabiting the
southwest.
In New York City, injuries to the foliage of lilac bushes have been
attributed to ambient ozone (Heggestad 1968). Symptoms included the usual
chlorosis, necrosis, premature leaf abscission and leaf bronzing, as well
as a distinctive rolling of the leaves. Extensive damage was observed in
several trees, shrubs, and ornamentals of Washington, D.C., during a 4-day
thermal inversion (Wester and Sullivan 1970). In the presence of hourly
peak oxidant concentrations as high as 0.22 ppm, a 40 to 70 percent loss
of foliage, followed by significant dieback, occurred in individual cot-
tonwoods, weeping willows, and white pines. Many individuals of the sen-
sitive species showed no effects, indicating the possible inheritance of
resistance. Wester and Sullivan (1970) provide a comprehensive listing of
the sensitivity and foliar symptoms of native trees, shrubs, and herbace-
ous plants to ozone toxicity from this episode.
Conifers are known to be especially susceptible to chronic oxidant
stress since they retain their photosynthetic tissue much longer than the
15
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deciduous trees. Costonis (1970) demonstrated the extreme sensitivity of
eastern white pine (Pinus strobus) to ozone at 0.2-0.3 ppm for 2 hours.
The chief visible symptom was tip necrosis, or needle dieback. Berry and
Ripperton (1963) produced symptoms of "emergence tipburn" in eastern white
pine by exposing the trees to 0.06 ppm 03 for 4 hours. Comparative
studies with S02 indicated that ozone had the lower phytotoxicity of the
two in this species (Costonis 1970).
The chlorotic decline of western pine stands in the San Bernardino
National Forest and Sierra Nevada, due to the long-range transport of oxi-
dants from the Los Angeles basin, has been studied intensively (Hay 1970,
1971; Taylor 1973; Munn j2t _al_. 1977). Major foliar symptoms include de-
creased terminal and diameter growth, yellow mottling of the needles, and
loss of all but the youngest needles (Taylor 1973). Chlorotic mottling of
Ponderosa pine (Pinus ponderosa) has been shown to occur from doses of 0.5
ppm 03 for 9 hours over a 9 to 18 day period (Miller _e_t al. 1963) while
exposures as low as 0.08 ppm 03 for 12 or more hours a dly are suffi-
cient to visibly injure this species (Taylor 1973). Over time, root de-
terioration and eventual death of susceptible trees occurs, a presumed re-
sult of tissue drying, altered histological and histochemical relation-
ships, and physiological changes in the ploem. Sensitive individuals, for
example, had relatively more stomata on each needle, their chloroplasts
aggregated to one side of the cell, and they exhibited depressed levels of
reserve sugar in the phloem (Taylor 1973). Table 5 lists species of
western conifers by their relative susceptibility to oxidant damage.
Table 5. Approximate ozone sensitivity of important western conifers.
Woodland chaparral
Conifer zone zone
San Bernardino San Bernardino
National Forest Sierra Nevada National Forest
Sensitive Ponderosa pine Western white pine Big-cone Douglas fir
Jeffrey pine Monterey x knobcone
White fir pine cross
Moderately Coulter pine Red fir Knobcone pine
Sensitive Incense cedar
Rocky Mountain
ponderosa pine
Tolerant Sugar pine Giant sequoia
(From Taylor 1973)
16
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Experimental exposure of conifers to ozone for 18 days demonstrated
the Deduced chlorophyll content of impacted needles relative to those of
ambient air controls (Taylor 1973). Damaged trees that were moved to
filtered-air environments recovered well, suggesting that chronic oxidant
injury can be reversible. Less reversible are the infestations of pine
bark beetle to which chronically injured trees are predisposed. These
pests have been observed to preferentially attack oxidant-stressed trees
(Miller and Elderman 1977; Dahlsten and Rowney 1980).
Plant injuries induced by PAN are similar to those of ozone and are
usually indicated by the glazing, silvering, or bronzing of leaf surfaces
(Heggestad 1968; National Research Council 1977). PAN causes leaf chloro-
sis at concentrations of 0.05 to 0.1 ppm, as well as early senescence and
leaf abscission (National Research Council 1976, 1977). Broad-leaved
plants are the most sensitive, and younger plants appear to be more sus-
ceptible to damage than the mature individuals. Concentrations of 0.02 to
0.05 ppm PAN for a few hours have been shown to be sufficient to induce
injury in highly sensitive plants (Heggestad 1968).
Ethylene has been proven to damage orchid flowers at 0.1 ppm for 6
hours, producing symptoms identical to general oxidant injury (Heggestad
1968). It was further reported that flower buds failed to open on these
plants. Ethylene is a common growth hormone in plants, and internal con-
centrations may be increased by exposures to air pollutants or toxic sub-
stances (Tingey ejt aj_. 1978; Tingey 1980). Referred to as wound or stress
ethylene, these increased levels can injure plants by altering growth and
aging processes. Most of the many other hydrocarbons of urban atmospheres
are not believed to occur at ambient concentrations sufficient to bring on
plant damage (Smith 1974).
Oxidant effects on lichens and mosses remain largely unknown; how-
ever, ozone possesses a proven toxicity to fungal spores under moist con-
ditions (Heggestad 1968). Lists of the relative susceptibility of plants
to oxidants are given by the National Research Council (1977) for orna-
mental varieties and cultivars; Davis and Gerhold (1976) reported the rel-
ative sensitivity of urban coniferous and broad-leaved trees to ozone
while Benedict ejt aj_. (1971) provided a comprehensive listing of the de-
gree of resistance of crops, ornamentals, shrubs, and trees to ozone, PAN,
and general oxidants. More detailed discussion and reference are present-
ed in the report of this series on forest ecosystems.
3.1.2 Effects on Animals
The effects of photochemical oxidants on animals are manifested pri-
marily by irritation of or injury to the eyes and respiratory system (New-
man 1975). Ozone, formaldehyde, acrolein, and PAN are common smog compo-
nents known to induce eye irritation in experimental animals (National Re-
search Council 1976). The effects of gaseous air pollutants on animal
17
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lungs are numerous and varied, and depend on the degree to which pollu-
tants penetrate the respiratory system. Table 6 provides a comparison of
the sites of pulmonary exposure to ozone and other pollutant gases charac-
teristic of urban atmospheres.
Table 6. The site of action of pollutant
gases in the respiratory tract of animals.
Ambient
Gas concentration Target site3
(ppm)
S02
N02
NO
03
CO
Formaldehyde
0.01
0.05
0.05
0.05
1.0
0
- 0.5
- 0.5
- 2.0
- 0.5
- 100
.3
URT and large bronchus
CA and A
CA and A
CA and A
A
URT
aURT = upper respiratory tract; CA = ciliated airways; A = alveoli and
alveolar airways.
(Adapted from National Research Council 1977)
Since ozone is not readily soluble in water, it is able to penetrate
far into the nonciliated, terminal airways and alveoli of the lungs (Kavet
and Brain 1974). Once inhaled, little is likely to be exhaled due to its
reactivity with lung tissues. Lesions are known to appear in the bronchi-
oles and alveolar ducts while squamous cells are stripped from the bifur-
cations of pulmonary airways. Structural proteins and membrane lipids are
oxidized, creating enzyme imbalances in cells. Enzymes necessary for sus-
tained growth and metabolism can be depressed, thereby preventing the nor-
mal function of pulmonary tissues. Laboratory animals exposed to less
than 1.0 ppm 03 for short time periods have been shown to undergo these
cellular and tissue alterations (National Research Council 1977). A value
of 1.0 ppm 03 was reported as well to represent the threshold for pul-
monary irritation in cats and dogs (Lillie 1970). Exposures of at least
0.2 to 0.5 ppm have diminished lung elasticity in laboratory animals, in-
dicating the onset of chronic pulmonary disease (National Research Council
1977). In general, however, the pulmonary effects of low-level chronic
exposures are reversible. Moreover, evidence points to the development of
18
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localized tolerance to ozone in the lung, as well as reduced susceptibil-
ity to the toxicological effects of similar irritants (Lillie 1970; Kavet
and Brain 1974; National Research Council 1977).
Exposures to ozone are also known to increase the susceptibility of
animals to respiratory infection and disease by reducing the overall via-
bility of lung defense mechanisms and impairing the biocidal capabilities
of alveolar macrophages (Kavet and Brain 1974). Mucociliary streaming
that normally clears foreign materials from the lung is also inhibited.
Exposures to above-ambient concentrations have been demonstrated to reduce
normal resistance to infectious organisms introduced directly to the lung
of experimental animals (National Research Council 1977). In contrast to
the acquired tolerance of respiratory effects in animals, previous expo-
sures to oxidants have not been shown to reduce pulmonary sensitivity to
infection.
Peroxyacetylnitrate and hydroxy free radicals generated in smog con-
tribute to the oxidizing character of polluted air masses; however, their
toxicity to animals is reported to be less than that of ozone (National
Research Council 1977). Experiments employing complex automobile exhaust
mixtures clearly demonstrate that aldehyde and total oxidant concentra-
tions are augmented by irradiating the mixture. Irradiation also intensi-
fied observed effects of reduced voluntary activity and elevated carboxy-
hemoglobin formation in laboratory animals (National Research Council
1977). These experiments revealed differential biological effects from
varying ratios of oxidant and aldehyde levels as well. A low oxidant-to-
aldehyde ratio produced effects characteristic of upper respiratory tract
irritation including reduced breathing frequency, while a high ratio
brought on increased breathing frequency and other effects indicative of
deep lung penetration (National Research Council 1977). Long-term expo-
sures to oxidant concentrations simulating diurnal highs and lows in urban
atmospheres showed no effects on mouse mortality, growth rate, or histol-
ogy? yet were observed to produce reversible conditions of reduced fer-
tility and infant survival (National Research Council 1977).
The photochemical oxidant complex has been associated with mutagen-
icity in bacteria or tissue culture as well as carcinogenesis and birth
defects in the higher animals (National Research Council 1976, 1977).
Ozone has been shown to increase chromosome aberration and breakage in
lymphocytes of experimental animals at exposures of 0.2 ppm for 5 hours
(National Research Council 1977). Accelerated lung tumorigenesis was ob-
served in strains of sensitive mice, and increased incidence of jaw de-
formities occurred in offspring. Neonatal mortality was shown to in-
crease, as newborn individuals were found to be more susceptible to ad-
verse effects of 03 exposure than their parents. Motor activity was
reduced by half, and an increased potential for protein cross-linking,
indicative of an accelerated aging process, was observed.
A variety of complex organic aerosols are often found at elevated
concentrations in the presence of the photochemical oxidant complex.
19
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Benzene and its derivatives are reported to have mutagenic and carcinogen-
ic potential in domestic and experimental animals (Lillie 1970). Primary
polycyclic aromatic hydrocarbons (PAHs), which condense on fine particu-
lates in urban atmospheres, and diesel exhaust soots, are of known poten-
tial carcinogenicity (Glass 1979; Li and Royer 1979). Moreover, it has
been shown that many other uncharacterized substances contribute to the
mutagenicity of urban aerosols (Alfheim et_ aJL 1980). Carcinogens asso-
ciated with the photochemical oxidant complex have been suggested as con-
tributors to greater instances of lung cancer in a wide variety of mammals
and waterfowl of the Philadelphia zoo (Snyder and Ratcliffe 1966; Newman
1975).
Evidence of the direct effects of photochemical oxidants on urban
animal populations is lacking; however, eye irritation, blindness, and
changes in the cornea have been reported for bighorn sheep in highly pol-
luted areas of the San Bernardino National Forest downwind of Los Angeles
(Taylor 1973). This study further suggested respiratory difficulties in
birds and mammals of high metabolic rates as potential effects of elevated
oxidant levels. Evidence also indicates that small mammal populations may
be comparatively reduced in areas of chronic oxidant exposure (Miller and
Elderman 1977). Impaired visual and olfactory senses, which can modify
food-gathering efficiency and competitive interactions among animals, were
reported as possible effects in wild species under influences of urban air
pollution (Taylor 1973; National Research Council 1977).
Responses of urban wild animal populations to oxidant air pollution
have been investigated by Richkind and Hacker (1980). Deer mice from
high-pollution areas of the Los Angeles basin were shown to be signifi-
cantly more resistant to ozone toxicity than those from areas of low oxi-
dant pollution. Laboratory-born progeny demonstrated similar response
patterns, indicating a genetic basis for 03 tolerance. Inbred deer mice
were found to be more susceptible to 03 exposures than randomly-bred
mice, suggesting a relationship between reduced genetic variability and
degree of 03 toxicity. Age differences were not observed to influence
responses of deer mice from high-pollution areas; however, an increased
susceptibility of younger individuals to 03 toxicity was shown among the
populations from low-pollution areas. Genetic factors protecting older
mice from 03 toxicity were therefore presumed to become active at an
earlier age in populations exposed to high, chronic oxidant concentra-
tions. The fact that wild deer mice were significantly more tolerant of
0^ than were laboratory mice and rats yet experienced toxicity levels
similar to those of hamsters was taken as evidence that susceptibility to
03 injury in animals is fundamentally influenced by phylogenetic (evo-
lutionary) factors. The acquisition of tolerance to 03 in small mammals
is nevertheless though to be a graduated effect and should not be taken to
infer that sub-lethal exposures may be totally harmless.
3.1.3 Ecosystem Effects
A wide diversity of observed and potential ecosystem effects are as-
sociated with chronic oxidant pollution in metropolitan centers and their
20
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outlying regions. Nevertheless, the specific ecosystem effects of oxi-
dants on the man-modified habitats of cities are often obscured by the
multitude of other environmental stresses to which they are subject. In
natural habitats, virtually all parameters determining the stability of
ecosystem structure and function can be affected. Energy storage and
flux, biogeochemical cycling, and established patterns of ecosystem suc-
cession are altered by depressed primary productivity, lowered species
diversity, reduced availability of forage for herbivores, and changes in
prey-predator, plant-pathogen, or other competitive and interactive rela-
tionships (Munn et_ aj_. 1977). Many such effects have been observed or
postulated to occur in the San Bernardino National Forest of California.
This ecosystem has acquired perhaps the greatest potential as a natural
laboratory for the study of urban oxidant effects, although similar situ-
ations may be found to occur in the future in American cities contiguous
to mountain ranges.
The long-range oxidant contamination of the San Bernardino forest is
believed to have begun in the 1940s and has since extended to the southern
Sierra Nevada Mountains. During this time, it also became increasingly
difficult for sensitive trees, ornamental plants, and leafy vegetation to
be grown in the ambient air of Los Angeles (Hay 1970). Figure 3 illus-
trates the progressive habitat diversification of the west coast ecosystem
along a transect from the coast to the inland desert. A characteristic
pattern of ambient oxidant concentration, the result of simultaneous
transport and photochemical reaction, is related to the potential economic
value of dominant plant communities in the natural and man-modified eco-
systems. The coniferous forest is seen to receive more total oxidants
than cities and suburbs, explaining why sensitive tree species in this
forest experience significant visible damage (Taylor 1973). Direct tree
injury is in large part attributable to the function of mature forest can-
opies as efficient sinks for reactive oxidants, unlike the shorter domi-
nants of agroecosystems. This process is reflected by the diminished oxi-
dant concentrations prevailing in the downwind inland desert.
Many changes have occurred within this coniferous ecosystem as a re-
sult of direct plant injury. Depressed primary productivity has resulted
in a reduction of food and habitat for wildlife (Taylor 1973). Forest
litter generation may increase at the same time that nutrient deficiencies
in oxidant-damaged foliage reduce the inventory of essential elements
stocked within the system (Taylor 1980). While the effects of ozone on
the mineralization activity of the soil microflora are not known, oxidants
exhibit rapid flux to soils and this may be more influential on microbial
exposures than ambient ozone concentrations. Oxidants are known to read-
ily react with soil and water surfaces, however, affording a certain de-
gree of protection to biota (Glass 1979). Many potential effects of re-
duced tree foliage have been hypothesized to occur in streams of this
forest (Taylor 1973). Diminished shade and leaf-litter production may re-
sult in possible alterations in temperature regime, dissolved oxygen
levels, biological oxygen demand (BOD), and the availability of fish-food
organisms.
21
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Ecosystems
CLIMAX WOODY PERENNIAL IRRIGATED LOW HIGH COASTAL
VEGETATION SHRUBS GRASSES PASTURE VALUE VALUE LANDS
AND WEEDS CROPS CROPS
VEGETATIVE COMPOSITION
SIMPLE AGROECOSYSTEMS-
-COMPLEX NATURAL ECOSYSTEMS
Figure 3. Hypothetical relationship of habitat diversification,
oxidant dose, and the economic value of vegetation along a tran-
sect in the southern coastal air basin of California. The only
real data are oxidant doses (dotted line), which define the pol-
lutant concentration gradient from the coast to the inland
desert. (From National Research Council 1977)
22
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Additional observed changes include the predisposition of damaged
trees to infestation by insect pests, as discussed earlier, or to in-
creased activity of fungal root pathogens (James et aj_. 1980). Moreover,
squirrels were found to preferentially collect the cones of tolerant trees
as those of damaged trees become scarcer, another response potentially re-
ducing the regeneration of oxidant-tolerant forest growth (Taylor 1973).
The possibility has also been raised that the oxidant-tolerant species may
be less adapted to other natural ecosystem stresses and thus experience
difficulty in replacing oxidant-sensitive dominants. Similar alterations
in metropolitan areas, and their relation to oxidant contamination, are
subjects requiring further investigation. Even in natural systems, the
cumulative effects of such changes and their potential for irreversibility
are unknown. Ozone is now thought to threaten remote forests on a region-
al basis in several portions of the conterminous United States (Armentano
et_ a]_. 1980), and has already been shown to result in shifts of species
composition away from eastern white pine dominance in the Blue Ridge
Mountains of the eastern United States (Hayes and Skelley 1977).
3.2 ATMOSPHERIC METALS
Metal particulates generally achieve their highest atmospheric con-
centrations in and around metropolitan centers. They are directly gener-
ated by point, line, and area sources, of which stationary industries,
roadways, and city centers are examples. The primary metal constituents
of urban atmospheres are presented in Table 7 along with their median con-
centrations in urban areas, background concentrations in remote areas, and
ratios of urban-to-remote levels. These ratios, or concentration factors,
indicate the potential magnitude by which urban values exceed ambient con-
centrations in remote areas, primarily as the result of anthropogenic
emissions.
Lead is seen to be the most abundant trace metal in urban air, and
for this reason it has been the most thoroughly studied. The highest am-
bient levels of atmospheric Pb in the United States have been found to
occur in the city of Los Angeles, where values as high as 5 ug/m3 are
recorded (Hall 1972). Large cities of about two million inhabitants aver-
age approximately 2.5 ug/m3 Pb in ambient air, whereas smaller communi-
ties of under 100,000 rarely experience Pb levels over 1.7 ug/m3 (Hall
1972). Ambient concentrations in urban air have been demonstrated to cor-
relate closely with amounts of gasoline consumed locally as well as with
Pb levels measured in soils and surface waters. On a seasonal basis, am-
bient levels of cadmium, copper, zinc, and lead all increase during summer
when wind speed is comparatively low; concentrations of nickel and vanadi-
um are found to peak in winter, despite the influence of elevated wind
velocity (Kneip et_ _al_. 1970).
The mode of deposition of atmospheric metals is largely a function of
their particle size and the degree to which they combine with other atmos-
pheric aerosols. Metal particulates are found in urban air in a wide
23
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Table 7. Median concentrations (nanograms/m3) of metals in urban
and remote atmospheres with ratios of urban-to-remote concentrations
Metal
Ag -
As -
Be -
Cd -
Co -
Cr -
Cu -
Hg -
Mn -
Mo -
Ni -
Pb -
Sb -
Se -
V -
Zn -
Silver
Arsenic
Beryllium
Cadmi urn
Cobalt
Chromium
Copper
Mercury
Manganese
Molybdenum
Nickel
Lead
Antimony
Selenium
Vanadium
Zinc
Urban
1.1
25
0.14
2.0
10.0
40.0
100
20
150
2
30
2000
30
4.7
50
1000
Remote
0.01
0.2
- -
0.1
0.05
0.3
0.2
0.5
0.4
0.3
0.36
1.0
0.2
0.1
1.0
0.5
Concentration Factor
Urban : Remote
110
125
—
20
200
133
500
40
375
7
83
2000
150
47
50
2000
(Adapted from Galloway e^ aj_. 1981)
24
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range of particle sizes. The larger particles (> 10-20 urn) settle and ac-
cumulate in the vicinity of point and line emission sources, invariably
along a gradient of decreasing concentration as distance from the source
increases. The finer particles (<-10-20 urn) may remain suspended in stag-
nant air or undergo long-range transport, often binding to other aerosols
or catalyzing photochemical reactions. Evidence for the long-range trans-
port of urban lead aerosols to remote locations has been established
through analyses of Greenland ice; Pb concentrations in the icepack have
been shown to increase arithmetically from the late 17th century and expo-
nentially since the 1940s (Hall 1972).
Metal particulates of all sizes contribute to the wet and dry compo-
nents of trace metal deposition in cities and downwind regions. Table 8
presents median concentrations of selected metals in wet deposition (rain,
snow, and ice) for urban and remote areas, as well as the concentration
factor by which urban values exceed those in remote areas. Lazrus et a1.
(1970) have reported metal ion concentrations in precipitation specTFicTto
32 cities of varying size across the continental United States. Metals in
precipitation have beem measured by Hutchinson and Whitby (1974) along
transects from nickel smelters in the Sudbury region of Ontario. Their
results are summarized in Table 9. The extent to which atmospheric metals
are removed by dry deposition is not well known; however, Table 10 pro-
vides a quantification of the metal content of dustfall in New York City.
Table 8. Median concentrations (ug/1) of metals in wet deposition from
urban and remote sites, with ratios of urban-to-remote concentrations.
Ag
Cd
Cu
Hg
Mn
Ni
Pb
V
Zn
Metal
- Silver
- Cadmium
- Copper
- Mercury
- Manganese
- Nickel
- Lead
- Vanadium
- Zinc
Urban
3.2
0.7
30
1.0
25
17
41
68
40
Concentration Factor
Remote Urban : Remote
0.008
0.008
0.055
0.048
0.22
KO.l
0.14
0.022
0.22
400
87
554
21
113
>170
292
3090
181
(Adapted from Galloway et_ _al_. 1981)
25
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Table 9. Metal content of wet deposition collected in samples
along transects from a nickel smelter (mg/m3/28 days).
Distance from Smelter (km)
Metal
Cu
Ni
Co
Al
Zn
Mn
Fe
1.6
122
271
8.5
91
5.7
4.8
144
1.9
85
95
4.2
216
7.4
0.8
34
7.4
22
16
0.5
60
4.9
0.7
47
13.5
3.4
2.2
0.2
2
2.3
0.3
5
19.3
2.1
8.1
0.1
3
5.4
0.4
7
(Adapted from Hutchinson and Whitby 1974)
Table 10. Average trace metal deposition
in dustfall at three sites in New York City.
Metal
Pb
Fe
Cd
Cu
Cr
Mn
Ni
Zn
V
Site 1
25
160
0.16
8
1
3.0
4
25
3
Average Deposition
Site 2
31
227
0.30
15
3
6.7
3
25
7
(mq/m^/mo)
Site 3
18
120
0.24
17
4
3.7
4
26
4
(Adapted from Kleinman et_ _al_. 1977)
26
-------
Apart from the inherent toxicity of certain metals, the particle size
of atmospheric metals is an important factor regulating biotic exposures
via inhalation. Up to 70 percent of Pb aerosols in urban and suburban air
may measure less than one micron and are thus in the respirable range (Lee
et _al_. 1968). An average particle size of 0.25 microns was observed in
the Pb aerosols of 59 selected cities (Robinson and Ludwig 1967). Never-
theless, some of these fine metal aerosols may adsorb to hygroscopic par-
ticles, which greatly enlarge during periods of high relative humidity,
and in this way are removed from the respirable range of particle size
distributions. In contrast, the accumulation of metals in soils and
plants, as well as biotic exposures through ingestion, are processes
largely independent of particle size.
3.2.1 Effects on Plants
Plants accumulate lead and other metals, sometimes to toxic levels,
through two types of exposure (National Research Council 1972):
• root absorption and uptake from soil solution; and
• foliar adsorption and uptake of atmospheric deposits.
The uptake of heavy metals is a physical process resulting in deposi-
tion on the cell walls and protoplasmic membranes of roots, shoots, and
leaves (Rains 1971; Zimdahl 1976). They may also bind to the membranes of
chloroplasts and mitochondria, thereby interfering with the electron
transfer reactions of respiration and photosynthesis (Zimdahl 1976). Low
levels are known to stimulate plant growth while higher levels exhibit
widely varying toxicities depending on factors of soil type and acidity,
precipitation frequencies, light availability, temperature and nutrient
status (Zimdahl 1976). Typical symptoms of lead toxicity (reductions in
photosynthesis, mitosis, and water absorption), for example, occur at
lower Pb concentrations in plants experiencing phosphate deficiencies,
with injury thresholds varying among the different species. Soil levels
of 1,000 ppm Pb are generally required to produce symptoms of plant toxi-
city under optimal growing conditions; however, corn has demonstrated re-
duced germination and root elongation with soil amendments of 100 ppm
(Zimdahl 1976). Background concentrations of lead in plant tissues are
reported to be approximately 0.5 ppm (Motto jet _aj_. 1970).
Lead and other nonessential metal ions are readily absorbed and re-
tained within the underground portions of plants. Cadmium and zinc are
accumulated in preference to lead via root absorption (Lagerwerff and
Specht 1970). Cadmium has been demonstrated to translocate rapidly
through plant tissues and to be preferentially retained as follows: stems
leaves reproductive organs (Haghiri 1973). Uptake via soil solution
was observed to be 16 times greater than with foliar applications, and
translocation was greatly enhanced. The uptake of cadmium by corn from
soil was also shown by Pietz ^t al. (1978) to result in minimal trans-
27
-------
location to the grain relative to other plant parts. Lead uptake through
roots is known to occur independently of soil concentration. It is
mediated by the concentration of soluble lead in the soil (Motto et al.
1970). Both uptake and plant toxicity are enhanced by depressed soiT~pH
(Zimdahl 1976). Some translocation to aboveground plant portions occurs
but does not lead to Pb accumulation in edible parts (National Research
Council 1972).
Lead levels in roadside soil and vegetation have been demonstrated to
correlate with prevailing winds, soil depth, distance from the roadway,
traffic volume, and the age of the road. Corresponding soil Pb inputs
from point and area sources are a function of wind, soil depth, distance,
intensity of emissions, and the age of the installation. Some of these
relationships are depicted in Figure 4 through a comparison of two major
highways. Prevailing winds tend to concentrate metal deposition in dis-
tinct patterns about their sources, as shown in both soil and vegetation
near the roadways in Figure 4. There is also a marked tendency for metal
concentrations to decline exponentially as the distance from emission
sources is increased (Daines .et _al_. 1970). This property has been demon-
strated for at least eight metallic emissions from primary smelters
(Little and Martin 1972; Buchauer 1973; Hutchinson and Whitby 1974). Cad-
mium, nickel, lead, and zinc levels are all found to decline with increas-
ing distance from roadways and with increasing depth in soil (Lagerwerff
and Specht 1970; Scanlon 1979).
A significant relationship between lead concentrations and traffic
density has been observed in roadside soils (Singer and Hanson 1969;
Lagerwerff and Specht 1970; Page and Ganje 1970; Williamson and Evans
1972). Cadmium and zinc concentrations in roadside soil are also found
to correlate with traffic volume (Scanlon 1979). The relationship of
plant lead accumulation to traffic volume is shown in Figure 5. Between
roads of high traffic density, comparative age is an important factor de-
termining overall accumulation rates in the roadside environment (Chow
1970). Figure 4 illustrates the extent to which cumulative traffic on a
50-year-old road has produced much greater plant and soil accumulations
than the traffic of a 30-year-old road.
While lead may accumulate to phytotoxic concentrations in the upper
soil horizon, long-term penetration to deeper soil layers is not thought
to produce toxic levels (Chow 1970; Williamson and Evans 1972). The
mobility of lead in soil depends primarily on (Zimdahl and Arvik 1973):
• soil pH;
• soil type and composition;
• percentage of organic matter;
• cation concentrations; and
0 soil drainage characteristics.
28
-------
West
Prevail ing Wind"
•-300
Lead
concentration
(ppm)
400
••200
East
--100
30
15 7.6 7.6 15
Distance (m)
30
Key:
Measurement Site
Soi 1 :
-5 cm
New Road
Baltimore-Washington
Parkway, Bladensburg,
Maryland
(Opened in 1954)
Old Road
Route #1, Beltsville,
Maryland
(Opened in 1920)
Figure 4. Roadside lead distributions in vegetation and the soil profile of
two highways of different age. (Adapted from Chow 1970)
29
-------
250-
200-
150-
100-
50-
Lead
concentration
(pom)
25
75 125
Distance (ft)
175
225
Key:
Traffic Volume
HIGH
54.7 x 103
vehicles/day
LOW
19.7 x 103
vehicles/clay
Vegetation
Mashed
Unwashed
Figure 5. Lead levels in roadside grass at varying traffic volumes
(Adapted from Motto et al_. 1979)
30
-------
Linzon j2t aj_. (1976) reported the following average Pb concentrations
in soil and vegetation from 65 sites in a metropolitan area:
Soil (ppm) Vegetation (ppm)
Depth: 0-2.5 cm 10-15 cm Condition: Unwashed Washed
292 148 63.8 31.5
These levels were found to remain well beneath tentative phytotoxicity
values of 600, 150, and 75 ppm Pb in soil, unwashed foliage, and washed
foliage, respectively. Soils and willow foliage in the immediate vicinity
of urban industries were shown in one instance to have the following Pb
concentrations:
Soil (ppm) Vegetation (ppm)
Depth: 0-5 cm Condition: Unwashed Washed
21,000 3,500 2,700
after which levels declines exponentially with increasing distance from
the source (Linzon ert jil_. 1976).
The following average concentrations of available trace metals have
been reported in urban and rural soils (Purves 1972):
Metal Concentration (ppm)
Metal Urban Soil Rural Soil
Boron 1.81 0.7
Copper 15.8 2.8
Lead 11.2 0.65
Zinc 52.4 2.9
A study comparing urban land-use patterns with metal concentrations in
soils of Grand Rapids, Michigan, showed consistently higher levels of most
metals to occur in industrial areas and the vicinity of an airport than in
residential or agricultural areas of the city (Klein 1972). Levels as
high as 1.44 ppm Hg and 314 ppm Zn were recorded in the study area. Table
11 gives mean metal concentrations by land-use type for the urban soils
sampled. A related study demonstrated that soils surrounding coal-fired
industrial facilities may be similarly enriched with these metals, and
that plants may accumulate Cd, Fe, Ni, and Zn from these emissions (Klein
and Russell 1973). Both of these studies suggest that mercury from com-
busted coal disperses over a much greater area than most metals; deposi-
tion patterns tend to be less well-defined for mercury than other metals
released from coal combustion.
Foliar adsorption of atmospheric lead depositions is the primary
cause of Pb accumulation in the aerial portion of roadside plants (Motto
et al. 1970). The entry of lead particles into leaf stomata is thought to
Fe Tpassive or chance process, as up to half of all surface deposits are
easily washed from leaves (National Research Council 1972). No evidence
31
-------
Table 11. Mean metal concentrations (ppm) in soils of
Grand Rapids, Michigan, related to land-use patterns.
Metal
Industrial
Airport
Land use
Residential
Agricultural
Ag
Cd
Co
Cr
Cu
Fe
Hg
Ni
Pb
Zn
0.37
0.66
2.8
8.5
16.3
3,100
0.14
8.3
47.7
56.6
0.29
0.77
7.9
17.6
10.4
6,200
0.33
12.3
17.9
36.6
0.13
0.41
2.3
3.2
8.0
2,200
0.10
5.4
17.9
21.1
0.19
0.57
2.7
4.6
8.8
2,600
0.11
5.6
15.4
22.1
(From Klein 1972)
has been reported for the translocation of airborne lead to the roots.
In fact, atmospheric lead exposures have been shown to decrease the extent
to which soil-derived Pb is translocated upwards in the plant (Lagerwerff
e^a]_. 1973).
The accumulation of Pb in the aerial portions of wild oats due to
primary smelter emissions from the San Francisco area has been reported by
Rains (1971) to have caused death in livestock. Foliar uptake was ob-
served to be least during early growth stages but increased significantly
over the summer season until levels of 500 ppm were recorded in winter.
As the plants aged, washing the leaves became an increasingly ineffective
means of removing airborne lead, since the older leaves tightly bind in-
corporated lead in their tissues. Graham and Kalman (1974) observed con-
centrations of 950 ppm in roadside forage grasses in Stanford, California.
These concentrations are of known toxicity to animals, and represent a
200-fold accumulation over background levels.
Urban and woody trees have been shown by Smith (1$72) to act as a
long-term repository for atmospheric lead. Greater than normal levels of
aluminum, chromium, nickel, iron, and zinc were also observed in urban
trees and shurbs (Smith 1973). Leaves and current twigs from a sugar
maple removed 60, 140, 5,800, and 820 mg of Cd, Cr, Pb, and Ni, respec-
tively, from ambient air during one growing season (Smith 1974). Twigs
and shoots generally demonstrated higher levels thatn leaves. Mondano and
Smith (1974) reported that mercury accumulation in mosses and conifer
parts in an urban area were below toxic levels and evidently of atmospher-
ic origin since Hg was more concentrated in plants than in the soil. Sim-
ilar resonses of trace metal accumulation have been documented in vegeta-
tion surrounding primary smelting facilities and other heavy industries
32
-------
(Little and Martin 1972, 1974; Buchauer 1973; Hutchinson and Whitby 1974;
Freedman and Hutchinson 1980).
Evidence for the long-range transport of lead and other metal aero-
sols has been documented in white pine stands of central Massachusetts.
Over the period from 1962 to 1978, Pb and Zn levels in soil litter were
shown to have increased 81 percent and 8 percent, respectively, while cop-
per ions essential to plants showed a 37 percent decrease (Siccama et al.
1980). Copper ions are presumed to play a role in the detoxification of
other metals that accumulate in plants (National Research Council 1972).
Lead levels in forest litter of the Great Smoky Mountains National Park
(GSMNP) have been shown to range from 250 ppm at low elevation sites to
over 450 ppm at higher elevations (Wiersma et^ jj_; 1980). Particle charac-
terization indicated high-temperature combustion as the primary lead
source while observed particle size distributions supported the hypothesis
that these contaminants had been transported long distances. A comprehen-
sive review of lead in plants and soils is presented by Zimdahl and Arvik
(1973), while Smith (1976) provides a thorough discussion of lead in the
roadside ecosystem.
3.2.2 Effects on Animals
The effects of ambient atmospheric metal concentrations on animals
are again mostly associated with progressive tissue accumulation from
chronic exposures rather than acute toxicity, although many instances of
lead poisoning in livestock of nonurban areas have been reported (Lillie
1970; National Research Council 1972; Newman 1975, 1980). Metal accumula-
tion occurs via the inhalation of ambient aerosols or the ingestion of
contaminated food supplies and, according to the mode of exposure, is de-
posited in differential amounts throughout the various tissues of higher
animals. Essential elements such as copper are physiologically regulated
at stable body concentrations, but the nonessential metals tend to accumu-
late (Schlesinger and Potter 1974; Gough jet _al_. 1979). Metals known to
accumulate in different organs of vertebrates are presented in Table 12.
The potential mutagens and carcinogens among the atmospheric metals in-
clude berryllium, cadmium, nickel and selenium (Newman 1975, 1980). Some
nonmetallic particulates such as asbestos are both accumulating and car-
cinogenic. Both passive and active effects may be manifested in biota
ranging from bacteria to the higher vertebrates from exposure to atmos-
pheric metals in air, water, and soil.
Insects collected along roadsides demonstrate highly variable rates
of metal accumulation. Grasshoppers were not observed to concentrate lead
above 4 ppm, presumably due to their short life spans, whereas earthworms
were shown to accumulate significant amounts of lead in soils of high Pb
concentration (Van Hook 1974; Scanlon 1979). Lead levels in earthworms
near roadways closely reflect soil concentrations and their relationship
to traffic density is depicted in Figure 6. As for plants and soil,
levels of cadmium, nickel, lead, and zinc in earthworms also decrease with
33
-------
Table 12. Accumulation sites of atmospheric trace metals in vertebrates.
Target
Organ
or
Tissue
Metals
As Bea B Cda Fe Pb Nia Mn Hg Mo Sea V Zn
Liver
Kidney
Bones
Blood
Lung
Brain
Muscle
Skin
Spleen
Nails
Fat
1
I
1
1
I
I
1
1
I
1
I
I
I
I
I
I
1
1
1
1
1
I
1
1
I
I
•
I
I
I
t
Potential mutagens or carcinogens
(From Newman 1975)
34
-------
JO
CL.
50
40
30
20
10
0
-£
32.10
1,085
8.51
12
_. . , ,
Distance (m)
18
Figure 6. Lead concentrations in earthworms at varying
distances from roads of different traffic volumes.
(Adapted from Goldsmith and Scan!on 1977)
35
-------
increasing distance from the roadway (Gish and Christensen 1973). Earth-
worms have been shown, however, to exhibit the greatest affinity for cad-
mium (Ireland 1979). The following ratios (concentration factors) of
metal levels in earthworms to those in the top 10 centimeters of soil were
reported by Van Hook (1974):
Lead Zinc Cadmium
0.11-0.30 3.0-13.0 11.6-22.5
Martin and Coughtrey (1976) found that earthworms concentrated cadmium
from even lightly contaminated soil by a factor of 8, while snails ab-
sorbed 41 times the Cd levels of vegetation. Wood!ice are shown to con-
centrate Pb and Zn by only a small factor over vegetation, while cadmium
is accumulated at levels 21 to 48 times those of plants (Martin and
Coughtrey (1976). Maximum Pb concentrations of 50 ppm in earthworms, 80
ppm in millipedes, and 700 ppm in woodlice have been observed (Williamson
and Evans 1972; Scanlon 1979). They are in large part explained by ele-
vated Pb concentrations and the preferential concentration of Pb in in-
sects with hard calcareous exoskeletons. The effects of ambient exposures
of atmospheric metals on other invertebrates, amphibians, and reptiles are
not well known.
Both wild and adaptive bird species of urban environments are report-
ed to concentrate lead and other metals in internal tissues. Historical
documentations of this effect have been reviewed by Newman (1980). Getz
et al. (1977a) reported Pb concentrations to be much greater in urban
FTrcTs" than in their rural counterparts, although levels remained below the
300 ppm threshold of toxicity. Tissue lead concentration factors are pre-
sented for four bird species in Table 13 along with average values record-
ed among the urban birds sampled. For example, feathers of urban star-
lings had an average content of 225.1 ppm Pb, or 35.2 times the average
feather content of rural starlings. These values reflect Pb exposures
from both the inhalation of ambient air and the ingestion of contaminated
food. Variations are substantial between the four species and their dif-
ferent tissues. Adults consistently showed greater lead levels in their
various organs than juveniles.
Similar results were obtained in a comparative study of urban and
suburban pigeons by Ohi et_ _al_. (1974) in Tokyo. While no behavioral symp-
toms were observed, birds were shown to have marked reduction in the ac-
tivity of a specific enzyme, to accumulate lead in the blood and to con-
centrate it in bones. Pigeons in downtown Philadelphia demonstrated blood
levels comparable to rural pigeons, but accumulated lead to much greater
concentrations in body tissues: levels of 800 ppm in kidneys, almost 600
ppm in feathers, 500 ppm in bones, as well as high levels in the beak,
nails, and liver were reported (Tansy and Roth 1970). No evidence, how-
ever, would lead to the conclusion that there are any direct effects of
ambient metal levels on bird mortality, reproduction, or behavior in urban
areas (Getz et_ jfl_. 1977a). This may in part result from the considerable
difficulty of quantifying parameters such as bird mortality and reproduc-
tion.
36
-------
Table 13. Ratios of lead in tissues of urban
songbirds to concentrations in rural songbirds.
Tissue
Liver
Gut
Kidney
Femur
Lung
Feathers
Pectoral
muscle
Sparrow
20.0
(12.0)a
11.4
(26.2)
9.7
(33.9)
7.7
(130.4)
7.7
(6.9)
5.9
(158.3)
2.3
(2.1)
Starling
4.0
(16.1)
4.6
(6,0)
27.4
(98.5)
16.6
(213.0)
1.9b
(5.2)
35.2
(225.1)
3.0b
(2.4)
Grackle
4.8
(12.1)
7.3
(10.2)
3.9
(13.5)
2.9
(62.8)
1.2b
(2.7)
2.3
(81.4)
1.8
(1.4)
Robin
4.4
(10.5)
7.7
(24.5)
3.4
(25.0)
3.2
(133.7)
4.7
(10.3)
3.2
(79.7)
1.2b
(1.2)
a Tissue concentrations of lead in urban songbirds are given in
parentheses in ppm dry weight
Not statistically significant
(Adapted from Getz et al_. 1977)
37
-------
A variety of small mammals have been studied in urban environments
and are shown to accumulate lead and other metals in excessive quantities.
For example, urban squirrels have been demonstrated to contain signifi-
cantly greater kidney concentrations of Pb, Cd, and Zn than rural squir-
rels (McKinnon _e_t _al_. 1976; Bigler and Hoff 1977). Cadmium accumulation
in the kidney was found to correlate with increasing age of individuals to
a greater degree than did lead or zinc accumulation. No relationships
could be established between lead levels and urban land-use patterns; how-
ever, squirrels trapped in relatively lower socio-economic areas had sig-
nificantly greater lead concentrations, suggesting that animal exposures
to lead and other metals are not uniform throughout the urban ecosystem.
Bull et_ _al_. (1977) reported elevated mercury concentrations in the brain,
kidney, liver, and hair of woodmice and bank voles near a chlor-alkali in-
dustry. Welch and Dick (1975) found that deer mice inhabiting roadsides
of high traffic density have the highest concentrations in bones, kidney,
an,d liver, with the lowest in brain, lung, and muscle. Tissue lead levels
in deer mice, and their ratios in high-to-low volume traffic, are given in
Table 14. Mierau and Favara (1975) concluded from a similar study of
roadside deer mice that traffic volumes of 100,000 vehicles per day would
be required to produce Pb intoxication, and over 200,000 to cause severe
Pb poisoning.
Table 14. Lead concentrations (ppm dry weight) in tissues of
deer mice (Peromyscus maniculatus) from roadside sites of dif-
ferent traffic volumes.
Vehicles per Day
Ratio or Concentration Factor
Body Tissue
38,000
4,200
(high-to-low traffic volume)
Bone
Stomach
Kidney
Lung
Brain
L iver
Muscle
106
29.5
23.0
3.3
2.7
4.6
2.7
5.05
2.9
2.55
0.42
0.4
0.7
0.42
21.0
10.2
9.0
7.9
6.8
6.6
6.4
(Adapted from Welch and Dick 1975)
The accumulation of lead and other atmospheric trace metals by small
mammals is largely a function of their position in the food web. Insecti-
vores generally accumulate the greatest quantities of lead, omnivores in-
termediate amounts, and herbivores the least (Getz ert _al_. 1977b). Other
factors determining lead accumulation rates in small mammals may be sum-
marized as follows (Quarles et al. 1974):
38
-------
• metabolic rate (average cc G£ consumption/g/day);
• food consumption (g/g/day);
• home range (hectares);
• movements;
• diet; and
• life span.
For example, the short-tailed shrew (Blarina brevicauda) is essentially
insectivorous and has a higher metabolic rate, and consequently greater
lead levels than the meadow vole (Microtus pennsylvanicus) or field mice
(Peromyscus spp.) (Jeffries and French 1972; Quarles et al. 1974;
Goldsmith and Scanlon 1977). This relationship is shown in Figure 7 for
small mammals at different distances from a roadway, and in Figure 8 for
roads of different traffic density. Table 15 permits a comparison of
body and gut concentrations of Pb in small mammals from roadside and
control locations,
as well as lead level parameters for two species of bats of varying life
habits. While both bat species inhabit upland locations away from major
roads, differences in tissue lead concentrations were attributed to the
habit of little brown bats to seek food near the roadway; big brown bats
feed primarily in upland areas (Clark 1979).
Other studies with small mammals have shown similar responses of cad-
mium accumulation in the kidney and liver (Martin and Coughtrey 1976;
Johnson et al. 1978). Studies of lead consumption in small mammals lead
to generaT agreement that comparable lead loadings in the diets of humans
and large vertebrates would be acutely toxic (Quarles et_ _al_. 1974). Clark
(1979) hypothesized that altered mortality and reproduction rates, or pos-
sible renal abberations, should appear in wildlife populations at these
levels, yet they were not detected. Nevertheless, subtle physiological
alterations in individuals, or effects on population distribution and
abundance, are extremely difficult to detect and quantify.
Studies of vertebrate exposures have been made in relation to ob-
served and potential manifestations of livestock mortality, and a variety
of symptoms are associated with chronic and acute exposures (Lillie 1970).
More relevant to ambient pollution loadings was an investigation of lead
accumulation in the organs of sheep foraging near roadsides (Ward et al.
1978). Ratios of tissue lead levels in sheep exposed to lead exhausts via
inhalation and ingestion to those of control sheep, column A in Table 16,
show substantial lead residues in bones, kidney, liver, and the gastro-
intestinal tract. Ratios of lead levels in exposed sheep to those removed
for six months from the roadside (column B) indicate the extent to which
tissue lead accumulations, notably in the kidney and gastrointestinal
tract, are reversible by excretion mechanisms. The difference in tissue
lead accumulation associated with inhalation versus ingestion exposures
39
-------
4->
.C
O)
3
>>
S-
s.
Q.
CQ
Q.
26
24
22
20
18
16
14
12
10
8
6
4
2
0
J
22.7
Blari,na
breviaauda
(short-tailed
shrew)
Key:
0-10 ft. from road
>25 ft. from road
16.3
Microtus
pennsyIvanious
(meadow vole)
6.8
3.9
Peromyscus
leucopus
(White-footed
mouse)
Figure 7. Lead levels in whole bodies of mammals at varying distances from a
roadside. (From Quarles et al. 1974)
40
-------
30 -
-C
01
o>
20 H
"O
CD
JlO
-Q
CL.
o-
11.6
6.5
,085
34.8
Short-tailed shrew
(Blarina brevieauda)
Meadow vole
(Miavotus pennsylvanicus]
12.1
8,120
Vehicles/day
21,040
Figure 8. Lead concentrations in whole bodies of small mammals from roads of
varying traffic volumes. (Adapted from Goldsmith and Scanlon 1977)
41
-------
Table 15. Lead concentration (ppm wet weight)
in bats, rodents and shrews.
A. Big brown bat vs. little brown bat in control location
Control location
Parameter
Big Brown Bat Little Brown Bat a
Body 46.55-male 16.97
31.49-female
Embryo
Stomach
contents
Guano
0.16
3.8
61
2.38
26
65
B. Voles, mice and shrews in roadside location vs. control location
Roadside location Control location
ra i ainc LCI
Body
Stomach
contents
Voles
9.4
1.45
Mice
9.3
4.91
Shrews
240
26.2
Voles Mice
1.5 1.0
0.84 1.16
Shrews
6.8
1.85
a This bat coramanly feeds in the roadside environment
(Adapted from Clark 1979)
42
-------
Table 16. Effects of ingestion and inhalation in sheep
exposed to automobile exhaust. Organ lead concentration
factors are given for various conditions of exposure.
Ratio (Concentration Factor)
Organ
Bone-vertebrate
Bone-shoulder
Kidney-cortex
Liver
Rumen
Intestine-small
Intestine-large
Muscle-abdomen
Brain-cerebrum
Lung
Muscle-shoulder
Muscle-heart
A
280
203
193
28.6
11.9
10.3
7.8
5.5
3.4
3.3
3.2
2.9
B
1.1
0.86
12.8
4.0
2.2
2.7
3.6
1.2
1.1
1.2
1
1.3
C
1
0.84
0.09
0.48
0.92
0.48
-
1
-
4.0
1.2
1.4
Key:
A - Sheep grazing beside roadside : Control sheep,
B - Sheep grazing beside roadside : Sheep removed
from roadside for six months.
C - Sheep exposed to auto emissions but fed
uncontaminated forage : Sheep not exposed to
emissions but fed forage from roadside.
(Adapted from Ward ejt aj_. 1978)
43
-------
are shown in column C. These ratios show, for example, that lung tissue
levels were rendered four times greater by aerial exposures whereas
kidney levels increased by a factor of 10 when contaminated forage was
ingested.
The deaths of seals and sea lions in a pool near a highway in a De-
troit zoo were attributed to Pb poisoning from atmospheric sources, and
lung anthracosis has been reported in several urban zoos (Newman 1975).
Lead aerosols were suspected of causing Pb toxicity in the large cats of a
city zoo, presumably through their grooming habits, yet most studies cor-
relate zoo animal poisoning with the ingestion of lead-based paints (Zook
et. _al_. 1972; Zook and Paasch 1980). As a result, little evidence is yet
available to implicate atmospheric lead as a major cause of acute mammali-
an Pb toxicity in urban areas.
Metal particulates have been shown to elicit pathological responses
in lungs of experimental animals by altering alveolar functions and the
immune response mechanism (Zarkower and Ferguson 1978). Laboratory stud-
ies have indicated the following preferential retention of metals in the
lungs of vertebrates: Pb > Cd > Ni > Cr, although little is known of
their fate once in lung solution (Henderson and Henderson 1979). Cadmium
is bound by proteins produced in the lung, and low-level exposures appear
to lead to extensive cadmium retention through the gradual saturation of
these sites (Benson and Henderson 1979). Pathological lung damage from
the inhalation of other metal particulates at ambient concentrations is
not presently known to occur. Nevertheless, Natusch and Wallace (1974)
found toxic trace metals and condensed organic substances to concentrate
perferentially in particles of respirable size, and more than half of
emitted hydrocarbons in urban areas are associated with particles in the
respirable size range (DeMaio and Corn 1966; Ciaccio _et _aJL 1974).
3.2.3 Ecosystem Effects
Habitat degradation is known to occur when soils are overloaded with
atmospheric metals; reductions in species diversity and altered ecosystem
successional patterns can result (Woodwell 1970; Jackson and Watson 1977).
The contamination of urban soils with lead, copper, and zinc may be large-
ly irreversible (Purves 1972). Natural forest ecosystems, such as those
of greenbelts and parklands, function to remove Pb from air and water via
soil and foliage; lead levels in forest litter may exceed Pb retained in
the biomass by a factor of 14 (Siccama and Smith 1978). Metal particu-
lates have been shown to depress litter decomposition in coniferous for-
ests, to inhibit spore germination and enzyme activity in fungi, and to
depress soil respiration (Ruhling and Tyler 1973). These effects on bio-
geochemical cycling are accelerated at low soil pH, a common condition in
urban soils, and accompanied by substantial leaching of calcium, magnesi-
um, potassium, and manganese.
A major impact on urban-influenced ecosystems of excessive atmospher-
ic metal deposition is the accumulation of these substances in terrestrial
44
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food chains. Small mammals and soil invertebrates have been shown to rap-
idly accumulate atmospheric lead and other metals in urban areas. Most of
these metals can be biomagnified in the body tissues of omnivorous and
carnivorous vertebrates. Hirao and Patterson (1974) have documented such
an occurrence among small mammals of remote Sierra Mountain forests down-
wind of Los Angeles. These areas received transported lead aerosols in
sufficient concentrations to overload natural ecosystem mechanisms, chief-
ly soil fixation, which exclude lead from food chains. Metal accumula-
tions were characteristic of those in small mammals living around line,
point, and area sources. Elevated metal loading can also be expected to
impact synergistically with other pollutants, notably oxidants, as well
as other environmental stresses. Such effects may be particularly acute
in the suboptimal portions of the habitat range of species (Newman 1980).
3.3 ACIDIFYING AIR POLLUTANTS
The major acidifying air pollutants in urban atmospheres are oxides
of sulfur and nitrogen, their derivative acids and, to a lesser extent,
hydrogen chloride. The primary reactants, sulfur oxides (SOX) and
nitrogen oxides (NOX), originate in large part from the combustion of
fossil fuels in stationary and mobile emission sources. While these gas-
eous substances number among the dominant components of urban air pollu-
tion, they are especially recognized for their propensity to undergo long-
range transport with concurrent photochemical transformation into strongly
acidic sulfate and nitrate aerosols (Altshuller and McBean 1980). These
gaseous and particulate secondary products are known to be scrubbed from
the air by precipitation or deposited as fallout over remote regions which
may be several hundreds of kilometers removed from the original emission
sources (Galloway and Cowling 1978)- In urban areas of the United States,
both local and regional sources can be responsible for elevated rates of
hydrogen ion deposition and resulting ecosystem acidification. Both types
of sources augment ambient levels of particulate sulfates in the 0.1 to
1.0 urn range which reduce visibility in urban atmospheres (Altshuller and
McBean 1980). Table 17 gives the atmospheric sources and chemical prop-
erties of four sulfates common in urban air.
During the past decade, a general trend towards lower S02 emissions
and ambient concentrations in city air has been attributed to the reloca-
tion of major air pollution sources away from urban areas, as well as to
the increased use of tall stacks (National Research Council 1978). This
situation has resulted in the achievement of greater compliance with Fed-
eral ground-level air quality standards (0.03 ppm for S02, 0.05 ppm for
N0£) but is responsible as well for widespread regional air pollution
and an increasing trend in ambient sulfate concentrations throughout the
entire northeastern United States (Altshuller 1973; Cogbill and Likens
1974). Air parcel trajectory analyses, for example, have shown that both
New York City and rural areas of upstate New York receive the greatest
portion of their atmospheric sulfates from a westerly to southerly direc-
tion and that the most elevated concentrations accompany slow-moving
45
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Table 17. Characteristics of some acidic atmospheric
sulfates in the 0.1 to 1.0 urn particle range.
Name
Sources
Chemical
Properties
Sulfuric acid
Atmospheric oxidation
of S02
Strong acid;
very hygroscopic
Ammonium bisulfate
Oxidation of SOa with
ammonia addition
Strong acid;
hygroscopic
Triammonium acid
disulfate
(NHlt)3H(S04)2
Oxidation of S02 plus
ammonia
Acidic
Ammonium sulfate
Oxidation of S02 plus
ammonia
Weak acid;
water soluble
(Adapted from National Research Council 1978)
46
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high-pressure systems originating in midwestern states (Calvin et a 1.
1978; Wolff, G. et^ aj_. 1979). Depressed precipitation pH was shown to
correlate well with winds prevailing from these same directions (Wolff, G.
et_ a]_. 1979). Sulfate concentrations of 5 ug/m3, observed when air par-
cels had passed over relatively unpolluted areas, increased to 20 ug/m3
or more in air that had stagnated over the Ohio Valley and other industri-
al areas (Galvin et_ al. 1978). Ambient sulfate levels above 5 ug/m3 are
considered to be indicative of the long-range transport phenomenon in both
urban and rural localities (Altshuller 1973).
Large regional differences exist in the behavior of acidifying pollu-
tants in the metropolitan United states. For example, Huff (1976) ob-
served that the rate of dry sulfate deposition over St. Louis, Missouri,
was less than ten percent of the wet deposition rate. In San Francisco,
su'ifate deposition via wet removal processes was found to be slightly
greater than dry sulfate deposition only during the short rainy season,
after which dry deposition became the dominant removal process (McColl
1980). Within precipitation, the relative contributions of sulfuric and
nitric acid to total acidity have been shown to differ markedly between
the northeastern and western portions of the United States (USEPA 1979).
Moreover, the atmosphere of the northeastern United States, apart from
major urban centers, is virtually devoid of windblown dust and other alka-
line particulates of natural origin which are dominant features of the at-
mosphere throughout the west and midwest (Winkler 1976). These substances
exert a neutralizing effect on precipitation acidity, thereby minimizing
potential impacts, and may even result in consistently alkaline rainfall.
Contrasted in Table 18 are annual average precipitation chemistries
for a northeastern city, Yonkers, New York, and a rural reference loca-
tion, Hubbard Brook, New Hampshire, as well as two western cities, Los
Angeles and Pasadena, California, and a less urbanized site at Riverside,
California. A brief review of precipitation chemistry values in the
northeastern United States reveals that while rainwater at the rural ref-
erence station contains significantly less sulfate, nitrate and chloride
anions, and associated hydrogen ions, mean pH values and hence acidity are
similar. The relationship reflects the tendency for rainwater acidity to
be uniformly elevated throughout the entire northeastern United States,
regardless of rural or urban location, and lends further credence to the
hypothesis that long-range transport of acidifying air pollutants is the
major source of precipitation contaminants, particularly sulfates. Never-
theless, the wide discrepancy in strong acid concentrations (SQq 2~,
N03 ~, Cl~) between these two sites immediately raises the question
as to how the two rainwater pH values can be so similar. Localized an-
thropogenic emissions of alkaline particulates in the urban area probably
exercise a significant neutralizing effect on the elevated concentrations
of strong acids and prevent depression of average pH values. Finally,
differences in the relative proportion of nitrate-to-sulfate concentra-
tions in these two locations indicate the important of nitrate ions in ur-
ban precipitation. Presumably derived from the photochemical oxidant com-
plex, elevated nitric acid proportions suggest that local transportation
47
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Table 18. Comparative urban/rural precipitation chemistry.
Monitoring Locations
Parameters
PH
H+
SO,
N03
Cl~
F"
Ca2
Mg2
K+
Na+
NH.
P(K
HCO
yeq/1
2~ mg/1
mg/1
mg/1
mg/1
mg/1
mg/1
mg/1
mg/1
mg/1
3 mg/1
3 mg/1
Urban Rural
Yonkers Hubbard Brook
New Yorka New Hampshire
4.1 4.15
148 72.4
4.8 2.9
4.4 1.47
1.2 0.47
<0.1 -e
0.16
0.09
0.07
0.12
0.22
0.008
0.006
Los Angeles
California0
4.49
32.1
2.67d
2.11
1.42
NG
0.292
0.128
0.191
0.785
0.647
<0.005
Urban
Pasadena
California
4.41
38.5
1.85d
1.94
.998
NG
0.133
0.0869
0.0657
0.553
0.379
<0.005
""
Urban
Riverside
California0
4.97
10.74
1.60d
2.02
1.06
NG
0.345
0.102
0.0943
0.578
0.601
<0.005
—
Average values - 1974 (Jacobson e^ a_l_. 1976)
b Average values - 1963-1974 (Likens et_ a]_. 1976)
c Average values - Fall 1978-Spring 1979 (Morgan and Liljestrand 1980)
Corrected for sea salts
e NG = negligible; = no data
48
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emissions exert a significant influence on levels of this component of
rainwater in urban areas of the northeastern United States. At the rural
reference station, as in most other portions of the northeast, sulfuric
acid is the primary proton donor to precipitation.
Precipitation chemistry measurements in Table 18 describe a signifi-
cantly different situation in the western United States. The nitric acid
component of precipitation is dominant and depressed rainwater pH values
more closely reflect localized (urban) air pollution, i.e., the transpor-
tation-related photochemical oxidant complex. Average precipitation chem-
istries in Pasadena reflect this trend, as do data from the comparatively
rural reference station in Riverside. Sulfates predominate in the acidic
composition of Los Angeles rainwater because of localized fossil fuel
combustion.
Morgan and Liljestrand (1980) have established a firm spatial trend
of increasing precipitation pH from the east Los Angeles basin (Los Ange-
les and Pasadena) to the west Los Angeles basin (Riverside) and on up to
more remote areas, such as Big Bear, where average annual rainfall pH is
5.42. This situation contrasts sharply with the uniformly depressed rain-
fall pH observed in the northeast and indicates the overriding influence
of localized emission sources, as well as neutralizing NH3+ and particu-
lates, in episodes of acid precipitation throughout the western United
States. Morgan and Liljestrand (1980) also reported that acidity was
greatest at the beginning of precipitation events, after which it tended
to fluctuate in lower acidity ranges. They further noted that drought
conditions result in lower average rainfall pH values because scavenged
contaminants are not diluted to the same extent as under conditions of
abundant precipitation.
The work of McColl (1980) in San Francisco further documented the
strong correlation between the hydrogen ion content of precipitation and
levels of nitrate, copper, and manganese ions. The ratio of nitrogen-to-
sulfur in rainwater was reported to decline from 2.38 in San Jose to 0.53
in a downwind rural reference point. These observations indicate that
nitric acid is the principal contributor to rainfall acidity in the west-
ern United States and suggest a significant role for metal ions in the
catalysis of photochemical reactions involving nitrogen and sulfur oxides.
The increasing prevalence of the long-range transport of acidifying
air pollutants in the western United States has been documented by Lewis
and Grant (1980). They report that the average pH of precipitation in a
remote area of the Continental Divide in Colorado has declined 0.80 unit
(from 5.43 to 4.63) over a period of three years. The average rate of
hydrogen ion loading in this watershed rose from 0.037 to 0.179 mg/m^/
week over the same period, representing almost a 10-fold raise in hydrogen
ion input to this remote ecosystem. In summary, a great deal of the evi-
dence collected to date substantiates the contention that acid precipita-
tion is ubiquitous to urbanized environments as well as to entire regions
of the continental United States.
49
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3.3.1 Effects on Plants
The toxicity of sulfur dioxide ($02) and nitrogen dioxide (N02)
to plants is a function of both the dosage and the duration of exposures
and is determined directly by genetic and environmental factors (Heggestad
1968). Both mature and young leaves are known to be susceptible to in-
jury, although impacts may be reduced during dormancy periods of low meta-
bolic activity (Glass 1979). Chronic effects may appear as chlorosis at
low concentrations, while sufates accumulate in leaves under conditions of
rising ambient levels. Dicotyledonous plants are generally more sensitive
than monocots (Lee _et _a_l_. 1980).
Ambient levels of 0.5 ppm N02 over a range of exposure times ap-
proximate the threshold of injury to sensitive herbaceous plants; compar-
able levels of S02 are 0.48 ppm for 4 hours or 0.28 ppm over 24 hours
(Heggestad 1968). Sensitive eastern white pines show visual damage at
0.05 ppm or more, while 0.25 ppm S02 is throught to represent the upper
threshold value of tolerant individuals (Costonis 1970). Low ambient
levels of S02 have also been shown to decrease the regeneration of mos-
ses as well as depress C02 fixation in conifers (National Research Coun-
cil 1978). Thresholds for plant injury are considerably reduced in the
presence of pollutant combinations such as S02, N02, and 03 (Hegge-
stad 1968). Foliar symptoms occasioned by acidifying air pollutants are
presented in detail by Jacobson and Hill (1970).
The deleterious effects of sulfur dioxide on lichens of urban and in-
dustrial areas are well documented. They consist primarily of reduced
photosynthesis and abnormalities in morphology and physiology (Pearson and
Skye 1965). Lichen populations have been observed to decline even where
adjacent vascular plants exhibit minimal effects (Eversman 1978). Factors
determining the densitivity of lichens and epiphytic mosses to S02 in-
clude (Robitaille et_ aj_. 1977).
• ambient sulfur dioxide concentrations;
• epiphyte pH;
• stemflow and bark pH;
• buffering capacity of bark and epiphyte; and
• relative percentage of S02 derivatives (bisulfite ions and sul-
furous acid) in and around the epiphyte.
Sulfur dioxide injury to lichens increases as the pH of plant tissues
declines. Toxic products generated by reaction of S02 and water are
concentrated in foliage as the acidity of lichen tissue rises. Chloro-
phyll levels are depressed by S02 exposures and photosynthetic processes
are disrupted completely at pH 3 (Turk and Wirth 1975; Skye 1979). A
positive correlation has been established between lichen growth and the pH
50
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of bark substrates (Skye 1965, 1979). Along a transect of elevated S02
concentrations downwind from a smelter, no lichens were observed to occur
on bark below pH 4 (Robitaille et jil_. 1977). The buffering capacity of
bark against acid was shown to decrease while capacity to buffer bases in-
creased. Bark pH averaged 3.66 near the smelter and 4.03 at the end of
the transect. Bark substrates are known to lose buffering capacity
through increased leaching of bound calcium and magnesium ions (Robitaille
e^aj_. 1977).
The potential effects of acid precipitation on urban plants may be
summarized as follows (Heagle 1973; Tamm 1976; Shriner 1981):
• visible foliar injury (e.g., necrosis) and altered foliar growth;
• accelerated erosion of the protective waxy cuticle;
• accelerated leaching of foliar nutrients;
• altered relations with pathogens, symbionts, and saprophytes;
• reduced seed germination and seedling growth in conifers; and
• indirect injury from pollutant-induced alteration of habitat.
Urban plants may be particularly subject to the deleterious effects of
soil acidification resulting from chronic acid deposition. Nutrient pools
(e.g., Ca, K, Mg) may be diminished in upper soil horizons, depriving
seedlings and shallow-rooted plants of needed minerals, while toxic metals
(e.g., Al, Mn) may accumulate in deep soil layers where they are made
available to deeply-rooted plants (Norton e_t £l_. 1980). Nutrient-depleted
soils may thus in part be responsible for graduated declines of habitat
productivity in urban areas.
Few if any observations have been made of foliar damage in urban
trees directly attributable to acid precipitation. Reduced growth and
foliar lesions have, however, been reported on deciduous forest trees ex-
posed to simulated acid rain (Wood and Bormann 1974, 1977; Evans et al.
1978). Under similar conditions, foliar symptoms and related effects have
been observed in various garden and agricultural crop species (Ferenbaugh
1976; Evans £ta]_. 1977; Irving 1978; Evans and Curry 1979).
Information is lacking on which to base a ranking of differential
plant susceptibilities to acid rain injury, although Haines et^ a]_. (1980)
have determined thresholds of leaf damage in some woody species. Davis
and Gerhold (1976) provide a listing of the relative susceptibility of
urban trees to SO? injury, while Benedict et^ aj_. (1971) report the sen-
sitivities of ornamentals, shrubs, and trees to both 563 and N0£. The
National Reasearch Council (1978) has published a listing of S02 concen-
trations reported to be below the injury thresholds of most plants.
51
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3.3.2 Effects on Animals
In general, animals are able to tolerate much greater doses of acidi-
fying air pollutants than can plants. Although sulfuric acid (I^SC^)
aerosols have been found to be more toxic to animals than SO;? gas (Lil-
lie 1970), neither these substances nor N02 and its acid derivatives are
thought to frequently occur at ambient levels sufficient to cause physio-
logical alterations or pathological responses in urban animals (Glass
1979). Moreover, naturally occurring elevated concentrations of ammonia
in the upper respiratory tract of animals have been postulated to provide
a degree of protection against inhaled acid particles (Wolff, R. et al.
1979).
Sulfur dioxide is a mild respiratory irritant which is absorbed
throughout the upper and lower respiratory tracts. Deep-lung penetration
occurs with greater frequency in small animals than in large animals be-
cause of their shorter lung airways (Dahl 1979). At toxic levels, SO^
provokes the reaction of broncho-constriction in lungs (Kavet and Brain
1974) followed by more severe systemic alterations (Lillie 1970). Isolat-
ed cases of severe debilitating symptoms in livestock, due to sulfuric
acid emissions of nearby industries, have been reported (Lillie 1970).
The potential effects of S02 are therefore more likely related to the
nature of the sulfites and sulfates into which they transform. Thus, ef-
fects may be substantially influenced by the particulate composition of
urban atmospheres.
Inhaled sulfate particles which are water-soluble dissolve and become
diluted in airways and tissues of the alveolar ducts. Insoluble particles
are cleared to an extent by mucous secretions, collected permanently in
the pulmonary lymph nodes, or they remain embedded in the recesses of the
lung. Little is known of the effects of chronic exposures; however, acute
irritancy potential has been observed to increase as particle size de-
creases, presumably due to the greater surface area of smaller particles
and variations in the site of pulmonary deposition (Kavet and Brain 1974).
In general, the fine particulates pass directly to the alveoli, while
coarser particulates deposit in the upper respiratory tract where lung
defense mechanisms act to remove them.
The synergistic action of S02 and atmospheric metals has been dem-
onstrated in pulmonary responses of laboratory animals (Amdur 1975). Zinc
ammonium sulfate ZnNh^SO,^), for example, has been shown to produce lo-
cal neurological reactions in portions of animal lungs (Kavet and Brain
1974). Novakova and Roubal (1971) and Novakova et_ jH_. (1973) reported
physiological alterations such as reduced blood calcium and protein levels
in hares exposed to airborne S02 and fly ash. Studies on the synergism
of S02 and particulate matter have led to the conclusion that particle
acidification in the atmosphere, rather than simple S02 adsorption, best
explains observed synergistic effects in animal lungs (Kavet and Brain
1974).
52
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The direct effects of strong acids in wet or dry deposition on most
urban fauna are unknown. Declines in urban earthworm populations have
been attributed to soil acidification brought on by chronic acid precipi-
tation (Gill and Bonnett 1973). In general, however, little is known of
the potential effects of acid precipitation on habitat factors such as
cover, food and water, nor of subsequent indirect effects on faunal dis-
tribution, abundance, or competitiveness.
The varied and substantial impacts of habitat acidification on fish
and amphibian reproduction in the United States have been established by
Schofield (1976) and Rough (1976). Assuming the absence of significant
organic pollution, toxic runoff, and other urban stresses to aquatic habi-
tat, such effects may be anticipated in cities located within regions of
geological and hydrochemical sensitivity to acidification (Hendrey et al.
1980). In most cities, however, many forms of water pollution are pres-
ent, and their interaction with freshwater acidification remains to be
studied.
3.3.3 Ecosystem Effects
Widespread sulfur dioxide injury to vegetation can be responsible for
reductions in the primary productivity of terrestrial ecosystems, decreas-
ed amounts of energy available to heterotrophic organisms, and general
patterns of biotic impoverishment and successional setback (Woodwell
1970). Responses of this magnitude to acid deposition have never been ob-
served in urban areas or other terrestrial ecosystems, yet a variety of
subtle impacts are becoming the focus of intensive research (National At-
mospheric Deposition Program 1978).
Of particular relevance to urban areas are observed changes in the
microclimate of soils (Glass 1979). The derivative acids of sulfur and
nitrogen oxides accelerate natural processes of soil acidification, de-
crease soil respiration (microbial activity), and facilitate the leaching
of nutrients, minerals, and toxic metals from soils (Malmer 1976). They
have been shown to inhibit nitrification by soil microflora, thus reducing
the availability of inorganic nitrogen to plants (Labeda and Alexander
1978). Moreover, acid precipitation is suspected to alter, in many dif-
ferent ways, relations between organisms and their pathogens, parasites,
saprotrophs, or symbionts (Shriner 1978; Shriner and Cowling 1980). It has
also been suggested that repeated exposure to acid precipitation decreases
not only the acid-buffering capacity of soils but of vegetation as well,
and that this capacity is indeed diminishing in urban and industrialized
areas (Bland 1977). Each of these effects has the potential for reducing
habitat productivity or altering trophic relationships in the urban eco-
system, although their cumulative impact and degree of reversibility are
not known.
53
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3.4 MISCELLANEOUS URBAN AIR POLLUTANTS
Some urban air pollutants do not fit neatly within the three major
categories above; yet a few of these substances demonstrate a potential
for biological impact on urban wildlife or habitat. The particulate mat-
ter of urban atmospheres includes many chemical substances which are sus-
pected of possessing a degree of biological reactivity. For example,
coarse and fine fractions, as well as serum-soluble fractions, of fly ash
from a coal-fired power plant are known to produce mutagenic alterations
in bacteria (Chrisp et. a]_. 1978; Fisher et a]_. 1979). A thorough investi-
gation of the potential biological effects of uncharacterized urban par-
ticulates is beyond the scope of this report. However, because of their
importance to urban air pollution, carbon monoxide (CO), fluoride (F~),
and pesticides are discussed below.
3.4.1 Carbon Monoxide
Atmospheric emissions of carbon monoxide (CO) from fossil fuel com-
bustion or industrial processes are primary absorbed by soils. Plants
have been shown not to absorb CO (Ingersoll 1971). A few cultivars of
herbaceous plants are known to be sensitive to ambient CO concentrations;
however, the threshold of toxicity to tree species is never achieved in
urban atmospheres (Smith 1974). Pulmonary effects of ambient CO inhala-
tion have never been observed in animals, although it is possible that
greatly elevated CO levels may reduce the availability of oxygen to body
tissues by increasing carboxyhemoglobin formation in blood (Kavet and
Brain 1974).
The extensive role of microorganisms in facilitating CO flux from the
atmosphere to soil has been studied in detail (Ingersoll 1971). The CO
uptake capacity of soils is observed to decline over prolonged exposures,
raising the possibility that chronic levels may reduce selection pressures
in microbial populations that would enhance mechanisms of CO removal from
air (Levy 1970).
3.4.2 Fluorides
Fluoride toxicity in wild and domestic animals has been well studied
and comprehensively documented (Karstad 1967; Kay et aj_. 1975; Krook and
Maylin 1979; Newman 1975, 1980). It is not a ubiquitous pollutant of ur-
ban atmospheres and is believed to affect animals and habitat only within
the immediate vicinity of point sources (Smith 1974).
Fluoride concentrations in plants correlate with those in soils, as
well as with dietary uptake levels and fecal excretion concentrations in
small mammals (Wright et aj_. 1978). Fluorides are cumulative poisons in
some plants, and levels of 0.5 ppm can be injurious to sensitive species;
however, other plants accumulate fluorides with no sign of injury to
54
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levels above 4,000 ppm (Heggestad 1968). Detailed information on the ef-
fects of fluorides in plants, including a listing of sensitive species, is
provided in the companion report of this series on forest ecosystems.
Exposure pathways to animals are primarily via the ingestion of con-
taminated vegetation. Fluorides may accumulate in the femur of mammals up
to values of 5,000 to 7,000 ppm or more; the preferential retention of
fluoride in tissues is reported to be: femur > kidney > liver ? muscle
(Wright et^ _al_. 1978). Severe teeth and bone disorders are symptomatic of
fluorosis, or fluoride poisoning, in vertebrates.
3.4.3 Pesticides
The extensive range of formerly abundant pesticides, through their
propensity for bioaccumulation and lengthy environmental persistence, may
have exerted a substantial influence on urban wildlife populations. Ap-
plications of pesticides in urban and suburban landscapes are thought to
equal agricultural uses on a per-acre basis (Leedy £t a]_. 1978). The dis-
appearance of carnivorous birds from the downtown areas of cities is a
well-observed phenomenon which may in part have been due to the devastat-
ing effects of persistent pesticides on the reproductive success of these
species (Hickey 1970).
Massive bird kills have resulted in urban areas of the United States
from the indiscriminate use of these substances for pest control. Hickey
and Hunt (1960) reported a fatality rate of 88 percent among robins, and
numerous fatalities in other species, just a few days after a spray was
applied to control the Dutch elm disease vector in southern Wisconsin.
Similar observations were made in Hanover, New Hampshire, by Wurster et^
al. (1965). Bird mortalities were attributed to rapid biomagnification
through the trophic chain of residues from pesticide-laden soil inverte-
brates. Earthworms rapidly accumulated pesticides to toxic concentra-
tions, resulting in a tendency for them to writhe at the soil surface be-
fore dying. This increased the likelihood that contaminated worms would
be ingested in large quantities by attracted birds (Stickel 1975).
The formerly common organochlorine pesticides had a long biological
half-life and were potent neurotoxins. They tended to accumulate in fatty
tissues where they remained until, under conditions of starvation or ex-
haustion, they were mobilized to levels causing rapid death. Chronic ex-
posures disrupted normal calcium metabolism in birds, leading to the well-
known effect of eggshell thinning and reduced reproductive success
(Stickel 1975). As a result, the use of several classes of these sub-
stances has since been banned or restricted, and formerly impacted bird
populations are thought to be adequately recovering.
Many of the new-generation pesticides (e.g., carbonates, organophos-
phates) do not persist for long periods in the environment, and therefore
are less likely to accumulate in individuals or food chains. They may
55
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nevertheless pose short-term threats of direct toxicity to exposed wild-
life species, particularly if their biological availability is enhanced by
habitat acidification. Moreover, the vast number of different chemicals
available raises many uncertainties as to their potential individual or
synergistic effects in biota and food chains.
56
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4.0 BIOLOGICAL INDICATORS OF AIR POLLUTION
EFFECTS IN THE URBAN ECOSYSTEM
The usefulness of plants as biological indicators of ambient air pol-
lution in urban centers has been recognized for many years. The potential
utility of animals as indicators of the biological activity of air pollu-
tants largely remains to be developed, although efforts have been made to
use livestock or various components of their habitat to detect releases of
toxic substances to the environment (Buck 1979).
The design of an adequately controlled bioindicator is difficult, and
most methods advanced to date require expertise in a wide variety of spe-
cialized fields. With the exception of some elaborate indicator systems
employing various groupings of plants, proposed biological indicators usu-
ally involve one or several observations of a single species. Such indi-
cator systems are generally directed toward highly specific objectives
which may or may not be relevant to different types of pollutants or their
synergistic qualities. Few efforts have been devoted to the development
of integrated plant-animal indicator systems, based on the individual and
population responses of several species in a community, despite their
potential to render a more detailed recognition of the scope of possible
impact. The following discussion offers a review of potential bioindica-
tors of urban air pollution which offer some promise as components of an
integrated biological assessment of the quality of urban environments for
wildlife.
4.1 PLANT BIOINDICATORS
The majority of efforts to develop plant indicator systems have in-
volved the use of lichens and epiphytic mosses. Patterns of lichen elimi-
nation from cities and potential bioindicator applications in Denmark are
discussed by Johnsen (1980). Sulfur dioxide is known to elicit a wide
variety of physiological responses in lichens, and measured reductions of
net photosynthesis, nitrogen fixation, and respiration can be correlated
with SOp accumulation (Sundstrom and Hallgren 1973). Sulfur dioxide
fumigations also lead to the physical degradation of chlorophyll in epi-
phytic mosses (National Research Council 1978). Measurements of changes
in these physiological parameters can be used to characterize long-term
S02 and particulate pollution in urban areas (Gilbert 1970; Creed et al.
1973). Indices of atmospheric purity, based on the number, coverage fre-
quency, and resistance factors of the epiphytes of single tree species,
have been developed and applied for the purpose of mapping long-term pol-
lution trends in Canadian cities (LeBlanc and De Sloover 1970; LeBlanc ert
al. 1972). Lichens have also been demonstrated to be sensitive indicators
cTF photochemical oxidant pollution in the San Bernardino Mountains of
California (Nash and Sigal 1980; Sigal and Nash 1980).
57
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Mosses in Polish national parks have been observed to concentrate
metals from emissions of nearby industrial developments (Grodzinska 1978).
Metal levels were found to correlate significantly with precipitation vol-
umes, confirming the atmospheric transport of deposited metals. Pilegaard
et aj_. (1979) found that a system using two lichens and one moss accurate-
ly reflects the atmospheric deposition of Cd, Cr, Cu, Ni, Pb, V, and In,
but not of Fe or Mn. Epiphytic lichens and bryophytes have been shown to
provide a good indication of airborne Hg levels in the United States
(Mondano and Smith 1974). Little and Martin (1974) were able to correlate
Cd, Pb, and Zn loadings in moss bags with emission rates from a base metal
smelter and prevailing wind patterns in the study area. They suggested
that the use of biological materials to monitor the dispersal of industri-
al metal emissions can take local topography and microclimate into account
in a more cost-effective fashion than complex modeling studies or expen-
sive monitoring networks.
Tree bark analyses may also be useful in the detection and quantifi-
cation of ecosystem acidification and atmospheric metal deposition. Mea-
sured parameters include bark pH and buffering capacity. Grodzinska
(1976) found the acidity of bark to correlate with monthly ambient concen-
trations of SO? in both air and precipitaiton. Over twice as much NaOH
was required to neutralize tree bark from city centers than from control
trees. Acidity was noticed to vary with bark depth, suggesting that only
external layers can be employed for bioindication purposes. Moreover, ob-
served seasonal variations in bark acidity indicate that the time samples
are made is an important control factor in potential applications of this
bioindicator.
An abundance of literature is available on the relative sensitivities
of herbaceous plant species to specific air pollutants. The tobacco
plant, specifically the Bel W3 cultivar, has been found to be highly sus-
ceptible to visual ozone damage (Heggestad 1968; Feder and Manning 1978).
Groupings of variably resistant tobacco plants have been employed to moni-
tor the severity of oxidant pollution episodes (National Research Council
1977). Pinto beans and white-flowered petunias are thought to have the
greatest potential for indicating elevating concentrations of PAN via
foliar damage (Heggestad 1968).
Two blue-flowering clones of Tradescantia have been employed as an
indicator of potential genetic injury to plants (Sparrow and Schairer
1977). The mutagenic response to air pollutants in this plant is a change
of stamen hair color from blue to pink. Mutagenicity was shown to be
weak, yet statistically significant for elevated exposures to SOp,
NO?, and 03.
The measurement of stress ethylene production in urban plants has
been recognized as a potential bioassay parameter for the detection of
many different pollutant stresses. Exposures to sulfur dioxide, ozone,
chlorine, metal salts, and complex organic pollutants have all been shown
to result in a short-lived production of ethylene above hormonal require-
ments (Tingey ejt _al_. 1978; Peiser and Yang 1979; Tingey 1980; Rodecap and
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Tingely 1981). Although excessive ethylene levels are known to cause
foliar injury, measurements of initial increases of this substance can in-
dicate subacute pollution stresses in the absence of physical damage.
Plants that are highly sensitive to metal and fluoride accumulation
readily exhibit visual foliar damage from exposures to low-level concen-
trations. Heggestad (1968) found the white-flowered gladiolus to show the
greatest foliar sensitivity to hydrogen fluoride (HF) fumigation. Plants
known to actively accumulate atmospheric metals and fluorides can provide
relative indices of gross pollution levels if appropriate controls are
applied. Regular monitoring of the heavy metal content of aerial portions
of wild oats has been recommended as a management practice for the protec-
tion of grazing species in portions of the western United States influ-
enced by urban or industrial air pollution (Rains 1971).
Taylor (1973) suggested that the extent of pine bark beetle infesta-
tion in a forest stand accurately indicates oxidant air pollution stress.
For example, relatively low populations of the western pine beetle (Den-
droctonus brevicomis) have been observed to cause greater injury in
oxidant-stressed stands than comparatively larger populations in healthy
stands (Dahlsten and Rowney 1980). Beetle population expansion and sub-
sequent rates of tree mortality were both found to be greater in oxidant-
stressed stands. Altered relationships between plants and their patho-
gens, parasites, and symbionts may also be indicative of an air pollution
stress (Heagle 1973; Shriner 1978; Treshow 1980). Research undertaken to
elucidate altered biotic interactions under conditions of oxidant air pol-
lution has provided considerable groundwork in the development of an even-
tual integrated forest monitoring system based on key ecosystem parameters
(Taylor 1973, 1980).
4.2 ANIMAL BIOINDICATORS
Newman (1975, 1980) has compiled an extensive literature survey of
recorded manifestations of air pollution injury in wildlife and has ex-
plored the potential suitability of wildlife species for use as bioindi-
cators. He found that wildlife species can generally be assigned to one
or more of five bioindicator categories (Newman 1975, 1980):
• sentinel species;
• bioassay organisms;
t general detectors;
• thrivers; and
• accumulators.
The wide variety of responses elicited by air pollution in animal bioin-
dicators are listed in Table 19.
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Table 19. Potential responses of animal
indicators to air pollution.
Sentinel Species
increased mortality or emigration
• reduced population numbers
Bioassay Organisms
• growth retardation
• physiological changes in cells
• changes in cell enzyme functions
• changes in blood chemistry
• altered energy requirements
• differential effects on life stages
• mutagenicity, cancer, or birth defects
General Detectors
• avoidance of pollutant
• abnormal behavior
• morphological alterations
• changes in abundance and distribution
• reduced tolerance to background stresses
Thrivers
increased population numbers
• increased natality or immigration
• genetic resistance
Accumulators
• residue accumulation in body tissues
(Adapted from Newman 1980)
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Sentinel species are highly sensitive to slight increases in air
pollution intensity and thus serve an early-warning function. Bees, for
example, have been suggested as sentinel indicators of rising ambient
S02 and fluoride concentrations as they succumb to the effect of these
pollutants well before plants and other animals are affected (Newman 1975,
1980; Bromenshenk 1980a,b). Bioassay organisms are those which exhibit
known responses to specific pollutants that can be quantified. These
organisms can be used to detect the presence of atmospheric contaminants
or to indicate ambient concentrations. Managed game species and common
urban inhabitants, whose ecology and physiology are well documented, best
serve this bioassay function. General detectors are animals whose life
habits are less understood, but which respond nevertheless in a predict-
able fashion to elevated air pollution levels. The population abundance
and distribution of urban birds and insect species have been suggested as
general detectors of air pollution stress in city centers and industrial
areas (Novakova 1969; Newman 1980). Thrivers are animals whose popula-
tions proliferate rapidly in response to elevated levels of a given pollu-
tant. Phytophagous insects and plant pathogens are prime examples. The
accumulators are the best known groups of animal bioindicators. They con-
centrate metals, fluorides, and lipophilic organic contaminants preferen-
tially in body tissues and are thus indicative of elevated pollutant
levels in plants, soil, waters or rain, depending on the mode of exposure.
Any of these groups could be used for the development of urban indi-
ces of environmental quality. The monitoring of sentinel species may be
most appropriate during periods of peak pollution levels or at the onset
of episodes. A large selection of potential bioassay organism are found
in metropolitan areas, and specific parameters indicative of excessive air
pollution are numerous (Table 19). Pigeons and sparrows are perhaps the
most suitable of the vertebrate organisms.
Chosen bioassay parameters should be developed for both general and
specific monitoring of the major pollutants of urban atmospheres (Newman
1980). The work of Ohi et^ _aj_. (1974) suggests that depressed levels of a
specific blood enzyme in pigeons are indicative of high lead exposures.
Histological aberrations in vertebrate lungs appear to be sensitive bio-
assay parameters as well. McArn et al. (1974) found increased levels of
granule-laden macrophages in the Tungs of sparrows exposed to the elevated
particulate concentrations of urban atmospheres. Monitoring undertaken
with avian populations in the vicinity of coal-fired power plants in Col-
strip, Montana, showed counts of particulate-laden lung macrophages to in-
crease over time and to correlate with distance from emission sources
(Kern eit al. 1980). Blood anemia and basophilic stippling are common
symptoms oT lead poisoning which may be of diagnostic significance in cer-
tain bioassay organisms (National Research Council 1972).
Invertebrate species whose responses to ambient air pollution are de-
termined by a single gene make excellent bioassay organisms. Creed (1974)
reported that melanism in ladybugs is a gene-specific response indicative
of soot levels in English cities. Ratios of black to red bugs permitted
comparisons of ambient pollution levels between cities.
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Extensive work with rabbits in Czechoslovakia has broadened the
potential use of this species for spatial determinations of the extent of
air pollution (Novakova 1969). The blood erythrogram, urine reaction, and
reproduction coefficients were all found to be sensitive bioassay parame-
ters. In addition, hemoglobin and hematocrit levels decreased under con-
ditions of heavy particulate pollution, whereas increases were noted for
exposures to gaseous and pesticide pollution. Alkaline particulate pollu-
tion from a cement factory produced alkaline urine in rabbits, while gase-
ous and pesticide pollution caused urine to be excessively acidic (Nova-
kova 1970). Reproduction coefficients were observed to decline by 30 per-
cent in areas of mixed SO? and particulate emissions, and by 55 percent
in areas contaminated with pesticides (Novakova 1969).
General detectors and thrivers can provide a red-alert function to
present conditions of atmospheric degradation. Abnormal behavioral char-
acteristics or changes in species abundance and distribution are the major
parameters monitored (Newman 1975). Both quantitative and qualitative im-
poverishment, as well as changes in species dominance, have been observed
among forest arthropods in industrial regions of Czechoslovakia (Novakova
1969). The arthropod communities of specific tree types exhibited 60 per-
cent impoverishment in zones of high ambient S02 pollution relative to
controls, and community structure was grossly altered. The number of spe-
cies remained constant; however some thrived while others became scarce.
Phytophagous species dominated the arthropod community where pollution
levels were elevated, whereas zoophagous species were dominant in control
areas (Novakova 1969). A variety of arthropod populations have been ob-
served to decline in response to controlled SO? exposures in experiment-
al plots of the Colstrip, Montana, study area (Leetham et_ _al_. 1980a,b).
Accumulator organisms are especially pertinent for the detection and
quantification of atmospheric metal deposition around line, point, and
area emission sources. As discussed above, earthworm concentrations are
highly indicative of increased Pb, Cd, and Zn contents of soils (Van Hook
1974). Lead is also known to concentrate greatly in insects with calcare-
ous exoskeletons. Pigeons in cities are reported to accumulate Pb to
highly significant levels in their kidneys, bones, and feathers (Tansy and
Roth 1970). Kidney concentrations are also the most sensitive indicator
of Pb and Cd accumulation in small mammals (Johnson et al. 1978). McKin-
non et_ _a]_. (1976) and Bigler and Hoff (1977) reportecTtFat in grey squir-
rels of urban areas kidney levels were the best indicator of lead, while
hair was extremely sensitive to Hg accumulation. Meadow voles are thought
to be the best wild indicators of lead pollution around cities because of
their limited range; however, insectivores can be monitored as well since
they display a tendency to accumulate metals in greater amounts (Quarles
et_ aj_. 1974). These types of organisms without doubt can contribute sig-
nificantly to an integrated bioenvironmental monitoring program of urban
air pollution effects.
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5.0 TOPICS FOR FURTHER RESEARCH
The effects of air pollution and acid precipitation on wildlife and
their habitats in metropolitan areas of the United States have never been
assessed in a comprehensive and systematic fashion. For this reason, re-
search priorities are difficult to assign. At least five broad subject
areas warranting further investigation can be identified as major compo-
nents of a research program. They are:
• a baseline study of the multiple environmental stresses of urban-
ization and the general condition of wildlife and habitat in
cities and suburbs;
• an investigation of the direct, indirect, and synergistic effects
of air pollution and acid rain on urban plants, and their resul-
tant viability as wildlife habitat;
• an examination of the effects of air pollution and acid rain on
urban wildlife populations, community structure, and the integrity
of food chains;
• a study of the direct and indirect effects of airborne contami-
nants on the structure, composition, and functional integrity of
natural and stressed urban soils;
• an evaluation of the impacts and repercussions of air pollution
and acid precipitation on the structure, function, and evolution
of whole ecosystems, including the many interrelated habitats of
urban landscapes as well as remote ecosystems subject to the in-
fluence of urban air pollution.
5.1 BASELINE STUDY
Any attempt to understand potential air pollution effects on urban
wildlife must be predicated on a firm knowledge of wildlife habits and in-
teractions as well as their responses to the many unique stresses of the
urban environment. Towards this end, much remains to be learned of the
productivity, breeding density, food and habitat requirements, and general
ecology of valued urban wildlife species, as well as specific modifica-
tions brought on by the impacts of urbanization. The urban setting offers
unique opportunities to study the extent to which wildlife species and
habitat associations are tolerant of combined stresses and their potential
for further stress-induced adaptation. Based on the complexity of the
problem, the development of bioindicators to monitor total urban stresses
will likely be a prerequisite to their use for the detection of air pollu-
tion effects.
Methods need to be devised for monitoring the state of the urban en-
vironment such that, in future, diachronic, or time-phased, analyses can
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be used to quantify changes in the status of wildlife habitat and other
essential attributes of the urban ecosystem. The bark of long-lived
trees, for example, can be monitored to provide baseline data on which
trends in atmospheric metal deposition or ecosystem acidification can be
based. Also for this purpose, unique wildlife species and habitat types,
as well as threatened or endangered species and their critical habitats,
should be identified and their distribution clarified.
Environmental monitoring efforts of the type employed in remote re-
gions are potentially adaptable to urban areas and, in addition, provide
meaningful data for comparison. The Great Smoky Mountains National Park
is an example of a remote area increasingly exposed to photochemical oxi-
dants, sulfates, heavy metals, and acidic precipitation (Herrmann et_ _aJL
1978). Integrated monitoring efforts undertaken in the Park include:
• data on pollutant inputs and cycling;
• precipitation analysis and weather data;
• regular collection of chemical and biological data;
• controlled monitoring of change in major plant communities;
• population monitoring of selected wildlife species; and
• population censusing, including endangered and exotic species.
These data provide the necessarily broad information base needed to sup-
port long-term and continuing evaluations of environmental quality.
A final prerequisite to an assessment of biological effects is a bet-
ter characterization of urban air pollution. Specifically, more research
is needed on rates and mechanisms of photochemical oxidant and strong acid
formation, variations in ambient concentrations of secondary pollutants,
deposition rates and patterns, and the overall dependence of urban air
pollution intensity on the meteorological characteristics unique to each
city. Undertaking such studies in the cities is advantageous because of
the greater data base on meteorological and air-quality parameters in
those areas. Modeling efforts may be particularly useful for organizing
data and analyses in cases of specific pollution problems, and should be
extended for use in general assessments of potential impact from various
urban pollutant mixtures. The acquisition of all such baseline data
should stress biological components and be carefully planned to meet the
requirements of the following major research areas.
5.2 PLANT EFFECTS
The assessment of plant effects, traditionally focused on economic
species (ornamentals and agriculture crops), should be extended to include
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the wide range of native urban vegetation. Of particular importance are
the dominant species of urban plant communities (e.g., deciduous trees,
forage grasses) that provide wildlife with habitat and otherwise exert a
measure of control over stable ecosystem function. Necessary investiga-
tions must be conducted at both meso- and micro-scale levels of detection.
In addition to their value in ecological assessment, such studies are more
likely to provide an accurate basis for evaluating the economic impact of
plant damage caused by urban air pollution.
With respect to research methodology in meso-scale studies:
• short-term exposure analyses must be complemented with studies of
potential plant impacts occurring over entire growing seasons; and
• protocols must be devised to permit controlled exposures of plants
and plant assemblages to mixtures of pollutants, particularly sul-
fur dioxide and ozone, that more closely reflect the ambient
composition of urban air.
Research is required to develop the utility of physiological parame-
ters indicative of the chronic injury and growth suppression caused by
various pollutants so that visible damage need not serve as the dominant
criteria of impact. Alterations in cell chemistry and metabolism, the
function of organelles, and plant hormone production afford many possibil-
ities for alternative impact monitoring. Such studies would lead to the
identification of bioindicator plants which can be used in the formulation
of an index of biological impact to urban ecosystems. They should provide
quick correlations with the status of valued wildlife and habitat species.
5.3 WILDLIFE EFFECTS
In general, very little research has been undertaken on the direct
responses of urban wildlife to air pollution, and even less on indirect
effects. A major prerequisite to these studies will be information de-
scribing animal exposures to air pollutants; for example, the way diet and
genetically influenced life habits affect the degree to which urban animal
populations are exposed to airborne contaminants. Also, sites of pollu-
tant deposition, biological activity, or passive retention are not as well
known in wildlife species as in humans and domestic animals. More needs
to be learned of the transfer and uptake in the lungs of reactive gases,
acid aerosols, and complex particulate matter, as well as their mechanisms
of excretion or rentention in body tissues.
In urban ecosystems, as elsewhere, work is needed to characterize
subacute manifestations of pollutant injury; for example, blood chemistry
changes, lung cell responses, or specific enzyme inhibitions. Bioindica-
tors, such as the ladybug and its single-gene response or the rabbit with
its several bioassay parameters, should be developed to monitor such al-
terations. Further effort is required in an assessment of the potential
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mutagenicity, cytotoxicity, or other lung responses of inhaled atmospheric
contaminants. Relationships between exposure levels and the carcinogeni-
city of bioaccumulating pollutants require further elucidation.
Research needs to be directed at determining the nature and extent of
potential effects on soil invertebrates and other key organisms in urban
food webs. Special efforts are required to detect and understand the more
subtle repercussions of air pollution and acid rain on trophic relation-
ships and competitive/interspecific interactions among urban wildlife, in-
cluding aquatic fauna. Also more must be known of the differential ef-
fects of air pollution on wildlife species over their entire habitat
range. Methods for appraising the extent to which pollution exerts selec-
tive pressures on biota, or alters their acquired adaptability to the ur-
ban environment, could become valuable long-term research tools. In gen-
eral, impacts on small mammals and predators must be well understood be-
fore progress can be made in the detection of community and ecosystem-
level effects.
Animals should be selected for study on the basis of their relative
importance in the maintenance of urban food webs. Research must therefore
be focused on the invertebrates, herbivores, and omnivores that provide
vital links in the food chain as well as significant contributions to
total biomass. Species of economic value, such as livestock, waterfowl,
and game fish, also deserve study in those urban areas where they are
prevalent. There is considerable justification for using the urban eco-
system as a natural laboratory for the study of air pollution effects on
wild animals species. Urban fauna are exposed to widely varying mixtures
of ambient air pollution, and much can be learned of potential acute and
chronic effects since exposures are generally elevated in comparison with
more pristine areas of the country. In many cases, studies of the effects
of single pollutants may be pertinent if pollutant interactions are mini-
mal .
5.4 SOIL EFFECTS
The potential effects of atmospheric metal deposition and acid pre-
cipitation on urban soils warrant extensive investigation because they
have been impacted by air pollution more severely and over a longer time
period than soils in remote areas. Moreover, much remains to be learned
of the fate and reactivity of atmospheric gases in soil. Effects on soil
texture and composition, nutrient leaching, and metal solubilities, espe-
cially under conditions of declining soil pH, all require further resolu-
tion. Study of urban soils could be a useful research tool for under-
standing the mechanisms of soil alterations.
Special efforts are required to monitor the effeciency of decomposi-
tion processes under various types of air pollution stress. Measurable
parameters include rates of soil respiration and organic substrate utili-
zation. This knowledge should complement an assessment of the dependence
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of wildlife and habitat assemblages on detritus food chains in cities.
Research is also needed to investigate the potential effects of pollutant
deposition on the functions of plant symbionts and other beneficial soil
organisms in cities. Especially important among these are the nitrogen-
fixing bacteria and fungal mycorrhizae which associate with plant roots
and aid their nutritional function.
In general, very little documentation exists on the vulnerability of
different urban soil types to air pollution exposures or acid and metal
deposition. The appraisal of soil sensitivity to acidification and the
establishment of relationships between projected emission levels and an-
ticipated future soil acidification all remain to be performed in urban
areas. Acid precipitation effects on urban plants and soils require im-
mediate emphasis since plant sensitivity is substantial and plant communi-
ty diversity is an important habitat attribute for urban fauna.
5.5 ECOSYSTEM EFFECTS
A variety of ecosystem-wide investigations are required in order to
determine the potential effects of atmospheric contaminants on the mainte-
nance of stability in ecosystems of the urban-rural continuum. Factors
determining ecosystem stability are known to differ greatly between rela-
tively simple urban landscapes and more complex natural habitats. For
example, in pioneer communities characteristic of disturbed habitats, food
chains are usually abbreviated while nutrient cycling and energy flow are
less conservative than in mature, climax communities. Ecosystem effects
should thus be more readily detected at these lower stages of ecosystem
organization.
These stability factors require greater elucidation in all stages of
ecosystem succession so that parameters indicative of ecosystem stability
can be developed and applied. Critical subsystems and functional mechan-
isms need to be identified to serve as specific indicators of ecosystem
stress. Such studies can form the basis for diachronic analyses permit-
ting a characterization of trends among the more sensitive ecosystem par-
ameters.
Research is needed to assess the impacts of localized urban air pol-
lution on mature forest stands and greenbelts in and around cities. More
should be known as well of the direct and indirect effects of the long-
range transport of urban air pollutants in remote ecosystems, for purposes
of comparison with the urban ecosystem. Photochemical oxidant alterations
of the forest ecosystem, well studied in the west, require special examin-
ation throughout the entire northeastern United States and the Appalachian
Range. Many seemingly minor items, such as the palatability of contamin-
ated vegetation to herbivores, should be incorporated in these studies.
The effects of acid deposition on urban ecosystems are also poten-
tially far-reaching as they result in the alteration of several abiotic
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ecosystem components and their functional relationships. Much information
is needed to understand mechanisms by which direct soil or plant effects
lead to habitat degradation and indirect effects on urban fauna. With
respect to urban aquatic systems, potential interactions between fresh-
water acidification and organic, metal, or toxic pollutant loadings, their
influence on water quality, especially acid-buffering capacity, and their
combined impact on aquatic flora and fauna all remain to be studied.
Dominant habitat types and configurations, as well as characteristic
flora and fauna, differ greatly among urban ecosystems of the United
States. As a result, generalized assessments of urban air pollution im-
pacts are of preliminary use only. On a regional basis, however, there
are many basic ecosystem similarities to be discovered and clarified.
Nevertheless, these cities may experience considerable variations in the
composition of ambient air pollution. In general, for virtually every
type of atmospheric contamination, the potential reversibility of effects
on normal or modified succession patterns in urban ecosystems, as well as
anticipated rates of recovery, could be much better understood.
5.6 CONCLUSION
Although analysis of the urban environment is complicated by the
multiplicity of environmental stresses, of which air pollution is just one
aspect, many aspects of cities favor their use as laboratories for inves-
tigation of air pollution impacts:
• higher levels of exposure increase likelihood of observable
effects;
• accessibility to research personnel and facilities;
0 a range in complexity of wildlife-habitat relations;
• history of monitoring data on air pollution;
• recent emphasis on acid precipitation monitoring in urban areas.
In view of the excellent research opportunities afforded by the natu-
ral laboratories of cities, as well as the extent of developmental impacts
rarely experienced elsewhere, urban areas would appear to be ideal loca-
tions for a preliminary, integrated assessment of air pollution and acid
rain effects on wildlife and their life-support systems. Many specific
plants and animals may be studied within the city; however, the pressing
need is to integrate these studies with knowledge of abiotic impacts
(soil, bedrock, waterways) in order to arrive at a multidisciplinary
understanding of total ecosystem impact, trends for future impact, and the
implications of such trends.
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In the past, the urban ecosystem has rarely been perceived in terms
of its living resources and the capacity of its varied habitats to sustain
wildlife. At present this trend may be reversing as the aesthetic and
recreational value of wildlife becomes increasingly recognized and greater
numbers of people consider wildlife vigor when judging the quality of the
urban environment for man. A highly visible program of air pollution and
acid rain research in cities could significantly raise the consciousness
of urban dwellers to issues of pollution in general, the long-range trans-
port of air pollution (LRTAP) in particular, and other facets of urban
environmental quality, thereby generating support for objectives of wild-
life and habitat conservation throughout the United States.
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50272 101
1. REPORT NO. I 2.
FWS/OBS-80/40.1(J
REPORT DOCUMENTATION
PAGE
4. Title and Subtitle
Air Pollution and Acid Rain, Report 10
The Effects of Air Pollution and Acid Rain on Fish,
Hi 1 d11fj_^Bd_jQTgjLrJjab_itats -Urban Ecosystems
7. Author(s)
i. A. Peterson
3. Recipient's Accession No.
5. Report Date
June 1982
8. Performing Organization Rept. No.
9t Performing Organization Name and Address
Dynamac Corporation
Dynamac Building
11140 Rockville Pike
Rockville, MD 20852
10. Project/Task/Work Unit No.
11. Contract(C) or Grant(G) No.
to 14-16-0009-80-085
(G)
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
13. Type of Report & Period Covered
Final
14.
15. Supplementary Notes
16. Abstract (Limit: 200 words)
Report 10 of the series synthesizing the results of scientific research related
to the effects of air pollution and acid deposition on fish and wildlife
resources deals with the urban ecosystem.-
General aspects of urban ecosystems relevant to a discussion of air pollution
effects are presented with an outline of various other types of ecosystem
stresses. Thds report describes plant, animal, and ecosystem responses to air
pollution within the following pollutant categories: photochemical oxidants,
atmospheric metals, acidifying air pollutants and miscellaneous urban air pol-
lutants. The potential use of biological indicators in monitering ambient urbe
air pollution is introduced and the report closes with a discussion of rele-
vant topics for further research.
17. Document Analysis a. Descriptors
atmospheric pollution, pollutants, exhaust emissions, acidification, precipi-
tation, terrestrial habitats, aquatic habitats
b. Identifiers/Open-Ended Terms
flue dust, flue gases, fumes, haze, oxidizers, smog, smoke, soot, air content,
pH, ecosystems, ecology, environmental effects
c. COSATI Field/Group 48B , G', 57C , H, U, Y
18. Availability Statement
Unlimited
19. Security Class (This Report)
Unclassified
20. Security Class (This Page)
Unclassified
21. No. of
94
22. Price
(See ANS|f-Z39.18>
87
OPTIONAL FORM 272 (-1-77)
(Formerly NTIS-3.S)
Department of Commerce
U. S. GOVERNMENT PRINTING OFFICE : 1982 379-346
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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.
us.
PISH & WILDLIFE
SERVICE
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