Biological Services Program
FWS/OBS-80/40.10 Air Pollution and Acid Rain,
JUNE 1982 Report No. 10
THE EFFECTS OF AIR POLLUTION AND ACID RAIN
ON FISH, WILDLIFE, AND THEIR HABITATS
URBAN ECOSYSTEMS
Office of Research and Development
U.S. Environmental Protection Agency _JBHP
Fish and Wildlife Service
U.S. Department of the Interior
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The Biological Services Program was established within the U.S. Fish and
Wildlife Service to supply scientific information and methodologies on key
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Projects have been initiated in the following areas: coal extraction and
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I*'or sale by the Superintendent of Documents, U.S. Government Printing Office
Washington, D.C. 20402
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FWS/0BS-80/40.10
June 1982
AIR POLLUTION AND ACID RAIN REPORT 10
THE EFFECTS OF AIR POLLUTION AND ACID RAIN
GN FISH, WIl&UFE, MH1 THEIR HABITATS
URBAN ECOSYSTEMS
by
M. A. Peterson
David Adler, Program Manager
Dynamac Corporation
Dynamac Building
11140 Rockville Pike
Rockvilie, MD 20852
FWS Contract Nimbsr 14-16-0009-80-035
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-80/4Q.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
accompany!ng 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.
iii
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CONTENTS
Paae
ABSTRACT 111
FIGURES V1"
TABLES vii
1.0 INTRODUCTION 1
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.~ EFFECTS OF AIR POLLUTION AND ACID RAIN OH
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 1?
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)
Pa^e
5.0 TOPICS FOR FURTHER RESEARCH 63
5.1 Basel ine 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
v
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FIGURES
Number Pa9e
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
vi
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TABLES
Number fiSฎ.
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 (yg/ฃ) 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
vii
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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 urn particle range 46
18 Comparative urban/rural precipitation chemistry 48
19 Potential responses of animal indicators to
air pollution 60
vi i i
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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.
2
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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.
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.
3
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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
4
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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
Pet birds and mammals
Livestock and other farm animals
Nuisance Species
Urban arthropods
Pigeons, starlings, crows, and house
sparrows
Rats and house mice
Feral mammals
Wild Species
Classification II
Terrestrial and aquatic invertebrates
Amphibians, reptiles, and fish
Migratory birds and songbirds
Small and large mammals
Desirable Species
Domestic pet and farm animals
Unavoidable Species
Rats, mice, and a variety of household
insects which persist in human settle-
ments and thrive on wastes
Adaptive Species
Pigeons, sparrows, crows, bats, and
other species which are well adjusted
to man-modified habitats
(Adapted from Leedy et aj_. 1978; Sudia 1978)
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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
(tidal marsh and estuary)
898
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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Adaptive
or
Nuisance Mild Species
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
t 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 siltation
SOIL CONDITION
nutrient and mineral leaching
drying and compaction
metal and biocide accumulation
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 et. a^. 1978; Stearns and Montag
1974)
II
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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
t 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 Li Hie (1970), Newman (1975, 1980), and Gough et al.
(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 (N0X) 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 aniT
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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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 1976]\
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 O3 necessary for leaf in-
jury was drastically reduced when SO2 gas was present. Heck (1968) ob-
served that a combination as low as 0.03 ppm O3 and 0.1 ppm SO2 could
injure sensitive tobacco. Otto and Daines (1969) reported that the sensi-
tivity of tobacco to O3 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 SO2 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 _et jH. 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 O3 for 9 hours over a 9 to 18 day period (Milleret al. 1963) while
exposures as low as 0.08 ppm O3 for 12 or more hours a day 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
National Forest
Sierra Nevada
San Bernardino
National Forest
Sensitive
Ponderosa pine
Jeffrey pine
White fir
Western white pine
Big-cone Douglas fir
Monterey x knobcone
pine cross
Moderately
Sensitive
Coulter pine
Incense cedar
Rocky Mountain
ponderosa pine
Red fir
Knobcone pine
Tolerant
Sugar pine
Giant sequoia
(From Taylor 1973)
16
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Experimental exposure of conifers to 02one for 18 days demonstrated
the reduced 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 et ak 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 et a]_. (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.
Gas
Ambient
concentration
(ppm)
Target sitea
S02
0.01 - 0.5
URT and large bronchus
no2
0.05 - 0.5
CA and A
NO
0.05 - 2.0
CA and A
03
0.05 - 0.5
CA and A
CO
O
O
1
O
A
Formaldehyde
0.3
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 O3 for short time periods have been shown to undergo these
cellular and tissue alterations (National Research Council 1977). A value
of 1.0 ppm O3 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 arid 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 O3 exposure than their parents. Motor activity was
reduced by half, and an increased potential for protein cross-1inking,
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 al_. 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 O3 tolerance. Inbred deer mice
were found to be more susceptible to O3 exposures than randomly-bred
mice, suggesting a relationship between reduced genetic variability and
degree of O3 toxicity. Age differences were not observed to influence
responses of deer mice from high-pollution areas; however, an increased
susceptibility of younger individuals to O3 toxicity was shown among the
populations from low-pollution areas. Genetic factors protecting older
mice from O3 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
O3 than were laboratory mice and rats yet experienced toxicity levels
similar to those of hamsters was taken as evidence that susceptibility to
O3 injury in animals is fundamentally influenced by phylogenetic (evo-
lutionary) factors. The acquisition of tolerance to O3 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^ al_. 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 WOOCY PERENNIAL IRRIGATED LOW HIGH COASTAL
VEGETATION SHRUBS GRASSES PASTURE VALUE VALUE LANDS
ANO WEEDS CROPS CROPS
VEGETATIVE COMPOSITION
simple agroecosystems ป
-ซ 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 irreversibi1ity
are unknown. Ozone is now thought to threaten remote forests on a region-
al basis in several portions of the conterminous United States (Armentano
e_t a_l_. 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/m^ Pb in ambient air, whereas smaller communi-
ties of under 100,000 rarely experience Pb levels over 1.7 ug/nw (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, aj_. 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 metal s in urban
and remote atmospheres with ratios of urban-to-remote concentrations.
Metal
Urban
Remote
Concentration Factor
Urban : Remote
Ag
- Silver
1.1
0.01
no
As
- Arsenic
25
0.2
125
Be
- Beryllium
0.14
- -
Cd
- Cadmi urn
2.0
0.1
20
Co
- Cobalt
10.0
0.05
200
Cr
- Chromium
40.0
0.3
133
Cu
- Copper
100
0.2
500
Hg
- Mercury
20
0.5
40
Mn
- Manganese
150
0.4
375
Mo
- Molybdenum
2
0.3
7
Ni
- Nickel
30
0.36
83
Pb
- Lead
2000
1.0
2000
Sb
- Antimony
30
0.2
150
Se
- Selenium
4.7
0.1
47
V
- Vanadium
50
1.0
50
Zn
- Zinc
1000
0.5
2000
(Adapted from Galloway et al_. 1981)
24
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range of particle sizes. The larger particles (> 10-20 um) 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 um) 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 _al_.
(1970) have reported metal ion concentrations in precipitation specTfic to
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.
Metal
Urban
Remote
Concentration Factor
Urban : Remote
Ag - Silver
3.2
0.008
400
Cd - Cadmium
0.7
0.008
87
Cu - Copper
30
0.055
554
Hg - Mercury
1.0
0.048
21
Mn - Manganese
25
0.22
113
Ni - Nickel
17
scO.l
>170
Pb - Lead
41
0.14
292
V - Vanadium
68
0.022
3090
Zn - Zinc
40
0.22
181
(Adapted from Galloway et aK 1981)
25
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Table 9. Metal content of wet deposition cpllected in samples
along transects from a nickel smelter (mg/m3/28 days).
Metal
Distance
from Smelter
(km)
1.6
1.9
7.4
13.5
19.3
Cu
122
85
22
3.4
2.1
Ni
271
95
16
2.2
8.1
Co
8.5
4.2
0.5
0.2
0.1
Al
91
216
60
2
3
Zn
5.7
7.4
4.9
2.3
5.4
Mn
4.8
0.8
0.7
0.3
0.4
Fe
144
34
47
5
7
(Adapted from Hutchinson and Whitby 1974)
Table 10. Average trace metal deposition
in dustfall at three sites in New York City.
Average Deposition (mg/m^/mo)
Metal
Site 1
Site 2
Site 3
Pb
25
31
18
Fe
160
227
120
Cd
0.16
0.30
0.24
Cu
8
15
17
Cr
1
3
4
Mn
3.0
6.7
3.7
Ni
4
3
4
Zn
25
25
26
V
3
7
4
(Adapted from Kleinman et al. 1977)
-------
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 jst aK 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 et 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 sofFpH
(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 metalTTc 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
soil drainage characteristics.
28
-------
West
Prevailing Wind"
- - 400
""300
Lead
concentration
(pprr)
"200
East
- -100
r
30
15 7.6 7.6
Oistance (m)
15
30
Key:
Measurement Site
Grass
Soil:
0-5 cm
5-10 cm
New Road
Ba1ti more-Was h 1 ngton
Parkway, Bladensburg,
Maryland
(Opened in 1954)
Old Road
Route #1, Beltsvil le,
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
-------
Key:
Traffic Volume
Vegetation
Washed
Unwashed
HIGH
54.7 x 103
vehicles/day
LOW
19.7 x 103
vehicles/day
Figure 5. Lead levels in roadside grass at varying traffic volumes.
{Adapted from Motto et al^. 1979)
30
-------
Linzon et 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 phototoxicity
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 et a]_. 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 aj_. 1970). The entry of lead particles into leaf stomata is thought to
"Be a passive 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.
Land use
Metal
Industrial
Airport
Residential
Agricultural
Ag
0.37
0.29
0.13
0.19
Cd
0.66
0.77
0.41
0.57
Co
2.8
7.9
2.3
2.7
Cr
8.5
17.6
3.2
4.6
Cu
16.3
10.4
8.0
8.8
Fe
3,100
6,200
2,200
2,600
Hg
0.14
0.33
0.10
0.11
Ni
8.3
12.3
5.4
5.6
Pb
47.7
17.9
17.9
15.4
Zn
56.6
36.6
21.1
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
et 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 tiqhtly 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 (1072) 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 al_; 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 et jjL 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 Be B Cd Fe Pb Nid Mn Hg Mo Sea V
Zn
Liver
Ki dney
Bones
Blood
Lung
Brain
Muscle
Skin
Spleen
Nails
Fat
1
1
1
1
1
,
a Potential mutagens or carcinogens
(From Newman 1975)
34
-------
50
40
K
30
CL
n
20
CL
10
0
*
50'H 2LQ40
veh U
:1es/day
32.10
1,085 vehicles/day 11.65
8.51
12
Distance (m)
18
Figure 5. Lead concentrations in earthwoms at varying
distances from roads of different traffic volumes.
(Adapted from Goldsmith and Scanlon 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 aK (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 al_. 1977a). This may in part result from the considerable
difficulty oT^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 Sparrow Starling Grackle Robin
Liver
20.0
4.0
4.8
4.4
(12.0)
(16.1)
(12.1)
(10.5)
Gut
11.4
4.6
7.3
7.7
(26.2)
(6.0)
(10.2)
(24.5)
Kidney
9.7
27.4
3.9
3.4
(33.9)
(98.5)
(13.5)
(25.0)
Femur
7.7
16.6
2.9
3.2
(130.4)
(213.0)
(62.8)
(133.7)
Lung
7.7
1.9b
1.2b
4.7
(6.9)
(5.2)
(2.7)
(10.3)
Feathers
5.9
35.2
2.3
3.2
(158.3)
(225.1)
(81.4)
(79.7)
Pectoral
2.3
3.0b
1.8
1.2b
muscle
(2.1)
(2.4)
(1.4)
(1.2)
a Tissue concentrations of lead in urban songbirds are given in
parentheses in ppm dry weight
b Not statistically significant
(Adapted from Getz et a_K 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 et 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, TTver, 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,
aryd 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
106
5.05
21.0
Stomach
29.5
2.9
10.2
Kidney
23.0
2.55
9.0
Lung
3.3
0.42
7.9
Brain
2.7
0.4
6.8
Liver
4.6
0.7
6.6
Muscle
2.7
0.42
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 et &]_. 1977b). Other
factors determining lead accumulation rates in smalTmammals may be sum-
marized as follows (Quarles et al. 1974):
38
-------
metabolic rate (average cc O2 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 &]_. 1978). Studies of lead consumption in small mammals lead
to general agreement that comparable lead loadings in the diets of humans
and large vertebrates would be acutely toxic (Quarles et aJL 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
-------
26
24
22
20
18
16
14
12
10
8
6
4
2
0
e 7.
Key:
. a
22.7
0-10 ft. from road
>25 ft. from road
16.3
5.8
Blarina
brevieauda
(short-tailed
shrew)
Miarotus
pennsyIvanious
(meadow vole)
6.8
Pevomyseus
leuoopus
(White-footed
mouse)
Lead levels in whole bodies of mammals at varying distances from a
roadside. (From Quarles et aK 1974)
40
-------
30 -
20 -
4->
-E
CD
r-
01
5
ฃ
-o
CD
5 io
Q_
0 -
11.6
*r
085
34.8
Short-tailed shrew
(Blapina bvevioauda)
Meadow vole
{.M-Lerotus pennsylvanious)
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 Scan!on 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
Parameter
Control location
Big Brown Bat
Little Brown
Bat a
Body
46.55-male
31.49-female
16.97
Embryo
0.16
2.33
Stomach
contents
3.8
26
Guano
61
65
B. Voles, mice
i and
shrews in roadside
location vs. control
location
Roadside location
Control location
Voles
Mice Shrews
Voles Mice
Shrews
Body
9.4
9.3 240
1.5 1.0
6.8
Stomach
contents
1.45
4.91 26.2
0.84 1.16
1.85
a This bat commonly 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 I
(Concentration
Factor)
Organ
A
B
C
Bone-vertebrate
280
1.1
1
Bone-shoulder
203
0.86
0.84
Kidney-cortex
193
12.8
0.09
Liver
28.6
4.0
0.48
Rumen
11.9
2.2
0.92
Intestine-small
10.3
2.7
0.48
Intestine-large
7.8
3.6
-
Muscle-abdomen
5.5
1.2
1
Brain-cerebrum
3.4
1.1
-
Lung
3.3
1.2
4.0
Muscle-shoulder
3.2
1
1.2
Muscle-heart
2.9
1.3
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 |