PB91-14846I
Ecological Exposure and Effects of Airborne Toxic
Chemicals: An Overview
(U.S.) Corvallis Environmental Research Lab., OR
Jan 91
U.S. DEPARTMENT OF COMMERCE
National Technical Information Service
mrs
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EPA/600/3-91/001
January 1991
ECOLOGICAL EXPOSURE AND EFFECTS OF
AIRBORNE TOXIC CHEMICALS: AN OVERVIEW
edited by
'Thomas J. Moser, 'Jerry R. Barker, and 'David T. Tingey
'ManTech Environmental Technology, Inc.
8 US Environmental Protection Agency
Environmental Research Laboratory
200 S.W. 35th Street
Corvallis, Oregon 97333
NOTICE
The information in this document has been funded wholly (or in part) by the U.S. Environmental
Protection Agency. It has been subjected to the Agency's peer and administrative review, and It has
been approved for publication as an EPA document. Mention of trade names or commercial products
does not constitute endorsement of recommendation for use.
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. TECHNICAL REPORT DATA
(fleae ntd /Movr/uw on At rtvtne before temple
1. REPORT NO.
EPA/600/3-91/001
2.
PB91-148460
[4. TITLE AND SUBTITLE
Ecological Exposure and Effects of
Airborne Toxic Chemicals: An Overview
B. REPORT DATE
January 1991
ft. PERFORMING ORGANIZATION CODE
7. AUTMOR(S)
Thomas J. Moser, Jerry R. Barker, and
David T. Tingey
B. PERFORMING ORGANIZATION REPORT NO.
». PERFORMING ORGANIZATION NAME AND ADDRESS
1st two, NSI, ERL-Corvallis; Tingey,
US EPA, ERL-Corvallis
10. PROGRAM ELEMENT NC
11. CONTRACT/BMANT NO.
12. SPONSORING AGENCY NAME AND ADDRESS
US Environmental Protection Agency
Environmental Research Laboratory
200 SW 35th Street
Corvallis, OR 97333
13. TVPE DP REPORT AND PERIOD COVERED
Pllhl -i ghoH Popr-i-rt-
14. SPONSORING AOENCY CODE
EPA/600/2
is. SUPPLEMENTARY NOTES
1990. U.S. Environmental
Laboratory, Corvallis, OR.
Protection Agency, Environmental Research
It. ABSTRACT
Since the release of the Environmental Protection Agency's (EPA's)
toxic release inventory (TRI) estimates for 1987, there has been a
heightened concern over the nation's air quality. Primarily, this
concern has been directed at human health effects in industrial-urban
areas. The fact that many airborne chemicals pose hazards to human
health is only one aspect of the problem. The continued deposition of
airborne toxic chemicals pose threats to both terrestrial and aquatic
ecosystems, this discussion is limited to terrestrial vegetation.
7.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.lDENTIPIERS/OPEN ENDED TERMS
c. COS AT i Field/Croup
S. DISTRIBUTION STATEMENT
Release to Public
t.SECJJ
UncTc
RITY.
assi
((.ASS (Tlui Htportj
fied
21. NO. O* PAGES
170
M. SECURITY CLASS
Unclassified
22. PRICE
•PA P»rm 1110-1 (S-7S)
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TABLE OF CONTENTS
Table of Contents ii
Preface iii
Atmospheric Transport of Toxic Chemicals and Their Potential Impacts on Terrestrial 1
Vegetation: An Overview (Thomas J. Moser, Jerry R. Barker, and David T. Tingey)
Organochlorines and Polycydic Aromatic Compounds in the High Arctic 8
(Greg W. Patton and Terry F. Bidleman)
Air-Sea Exchange of Hexachlorocydohexane in the Bering and Chukchi Seas 20
(Daniel A. Hinckley, Terry F. Bidleman, and Clifford P. Rice)
Trace Metal Deposition and Cycling in Forested Ecosystems (Andrew J. Friedland) 32
Atmospheric Deposition of Trace Metals to Lichens and Effects of Trace Metals 40
on Lichens (Thomas H. Nash III)
Fate and Effects of PAHs in the Terrestrial Environment: An Overview 48
(Nelson T. Edwards)
Effects of Organic Chemicals in the Atmosphere on Terrestrial Plants 60
(Jeffrey R. Foster)
Effects of Atmospheric Pollutants on Peatlands (Noel R. Urban) 90
Use of the PHYTOTOX Database to Estimate the Influence of Herbicide Drift 102
on Natural Habitats in Agroecosystems (James E. Nellessen and John S. Retcher)
Biological Markers in Environmental Sentinels to Establish Exposure to, and 107
Effects of, Atmospheric Toxicants (John F. McCarthy and Timothy J. Tschapiinski)
Air Toxics Ecological Effects Research Recommendations (Noel R. Urban and 128
Jerry R. Barker)
References 130
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PREFACE
It is a fact that numerous anthropogenic sources emit a large variety and quantity of
chemicals, many of them toxic, Into the atmosphere. These chemicals are commonly referred to as
air toxics with groupings In five broad categories: 1) synthetic Industrial organlcs, 2) agricultural
pesticides, 3) trace metals, 4) organometallte compounds, and 5) non-metallic Inorganics. It is also a
fact supported by scientific evidence that through atmospheric transport and deposition, toxic
chemicals find their way to rural locations, as well as remote areas such as the Arctic, high-elevation
forests, oceans, seas, and peatiands. Given the number and strength of air toxic emission sources
with their subsequent atmospheric transport and deposition, an increasing concern is that adverse
biological effects ranging from the biochemical to the ecosystem level of organization may be
occurring.
The slow biodegradability of many toxic compounds allows them to remain ecologically
harmful for long periods of time. The persistent nature of these compounds can result in adverse
biological effects by the incorporation and accumulation into food chains and disrupting ecological
processes. The effects from the chronic deposition of airborne toxic chemicals on various levels of
ecosystem organization and their potential interaction with natural stresses (e.g., insects, drought,
disease) to induce antagonistic to synerglstic effects are unknown.
The publication of the U.S. Environmental Protection Agency's (EPA) Toxic Release Inventory
for 1987 and 1988 has heightened the concern over the nation's air quality in regards to air toxics.
Although this concern is primarily human health related, more attention is currently being devoted to
potential environmental impacts from air toxics, as is evident by debates over amendments to the
Clean Air Act.
Little information is available concerning the regional distribution, exposure (concentration,
duration, frequency), bioavailabllity or potential effects of airborne toxics on sensitive biota and
ecosystems. Research is needed to focus on documenting chronic/cumulative impacts and mitigating
ecological problems which may result from the chronic, persistent environmental loading of a large
array of airborne toxic chemicals. In this regard, a credible, scientific data base is needed to support
the policy in the control of air toxics compounds and the listing of chemicals on the Toxic Release
Inventory.
The purpose of this document is to provide an overview of the current knowledge of
ecological exposure, fate and effects of airborne toxic chemicals, particularly at locations distant from
emission sources. The document consists of 11 chapters that were prepared by scientists who were
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invited to participate in air toxics-ecological effects sessions at the following conferences: 1) Air Waste
and Management Association, 82nd Annual Meeting (June 25-30, 1989, Anaheim, California, USA),
2) Society of Environmental Toxicology and Chemistry, 10th Annual Meeting (October 28 -
November 2, 1989, Toronto, Ontario, Canada); and 3) Environmental Protection Agency/Air and Waste
Management International Symposium on Measurement of Toxic and Related Pollutants (April 30 -
May 3, 1990. Raleigh, North Carolina, USA).
The emphasis of this document is on airborne chemical exposure, fate and effects in
terrestrial ecosystems, particularly vegetation. However, many of the concepts presented wDI apply
also to aquatic ecosystems. Chapter 1 provides an Introduction and overview of our knowledge of air
toxic emission sources, atmospheric transport and deposition, and potential impacts on terrestrial
vegetation. Chapters 2 and 3 are technical papers addressing the measurements of agricultural
pesticides and other toxic organics in arctic air and water, and demonstrates the ability of the
atmosphere to transport and deposit contaminants to environments far from emission sources.
Chapter 4 provides an overview of the sources of trace metals in forest ecosystems, evidence
implicating their atmospheric deposition, and their potential effects on soil processes. Chapter 5
discusses atmospheric deposition of trace metals and their potential effects on lichens which are
important components of many terrestrial ecosystems. Chapter 6 provides a thorough overview of the
sources, ambient concentrations and fate of polycyclic aromatic hydrocarbons in the terrestrial
environment. Chapter 7 provides a broad overview of the atmospheric deposition, fate and effects of
synthetic organics on terrestrial plants. Chapter 8 focuses on evidence of atmospheric transport and
deposition of airborne toxic chemicals into peatland ecosystems, and the effects of acid deposition
and airborne toxics on peatiands structure and function. Chapter 9 discusses the potential threat of
herbicide drift to natural plant communities associated with agroecosystems, and the use of the
PHYTOTOX data base to estimate the threat of specific herbicides on native norrtarget plant species.
Chapter 9 also demonstrates the paucity of scientific literature that addresses the ecological effects of
airborne toxic chemicals on natural vegetation. Chapter 10 provides a thorough description of the
potential applications, advantages and difficulties of using animal and plant biomarkers to evaluate and
monitor the exposure and effects of air toxics. Finally, a concluding chapter presents research needs
to document the effects of ecosystems exposure to airborne toxic pollutants.
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ATMOSPHERIC TRANSPORT OF TOXIC CHEMICALS AND THEIR POTENTIAL
IMPACTS ON TERRESTRIAL VEGETATION: AN OVERVIEW
Thomas J. Moser1, Jerry R. Barker1, and David T. Tingey2
ManTech Environmental Technology, Inc.1
U.S. Environmental Protection Agency2
U.S. EPA Environmental Research Laboratory
Corvallis, OR
INTRODUCTION
Numerous anthropogenic sources emit a large variety and quantity of toxic chemicals into the
atmosphere. Through the processes of atmospheric transport and deposition, toxic chemicals have
found their way into remote environments far from emission sources. Recent data strongly suggest
that the enriched concentrations of several contaminants detected in the air, water, soil, and biota of
rural and remote environments are the result of long-range atmospheric transport from sources
located in temperate and sub-tropical latitudes of North America and Eurasia. Many of these
chemicals are persistent, and they bioaccumulate and remain biologically harmful for long periods of
time. Although air toxics have been primarily considered an urban health problem, there is Increasing
concern among scientists that adverse ecological Impacts may result from their deposition into
terrestrial ecosystems and the subsequent exposure of plants. The chronic exposure of vegetation to
low concentrations of air toxics may result in sublethal effects such as decreased plant productivity
and vigor, that may culminate in changes in plant communities and ecosystem structure, composition,
and function.
Airborne pollutants can be broadly defined as any chemical occurring In the atmosphere that
may pose a threat to human health or the environment This broad definition includes an array of
chemicals ranging from the well-studied criteria pollutants to the less-studied radiatively important
trace gases. However, to provide a more focused definition for this paper, air toxics refers to the
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following groups of airborne substances: (1) synthetic Industrial organics, (2) agricultural pesticides,
(3) trace metals, (4) organometallic compounds, and (5) nonmetalltc inorganics.
Since the release of the Environmental Protection Agency's (EPA's) toxic release inventory
(TRI) estimates for 1987 (US EPA, 1989), there has been a heightened concern over the nation's air
quality. Primarily, this concern has been directed at human health effects in Industrial-urban areas.
The fact that many airborne chemicals pose hazards to human health is only one aspect of the
problem. The continued deposition of airborne toxic chemicals on a regional to global scale will
impact public welfare If It results in adverse impacts to the structure and function of sensitive
terrestrial and aquatic ecosystems. Although airborne toxic chemicals pose threats to both terrestrial
and aquatic ecosystems, this discussion Is limited to terrestrial vegetation.
EMISSION SOURCES
Airborne chemicals are emitted into the atmosphere from a large number and variety of point-
and area-sources. Anthropogenic emissions emanate from industrial, urban, and agricultural sources
such as chemical, metal, plastic, and paper/pulp industries; fossil fuel processing plants; motor
vehicles and aircraft; municipal waste incinerators; agricultural practices such as pesticide usage and
field burning; and small businesses such as dry cleaners. Emissions of toxic chemicals into the
atmosphere may occur directly by the deliberate or inadvertent releases from Industrial or urban
sources, or indirectly through volatilization following the deliberate or accidental discharge of
chemicals into water or soil resources. Also, considerable amounts of toxics enter the atmosphere
from wind drift and volatilization after agricultural chemical applications.
Industry is probably the major anthropogenic source of airborne toxic chemicals.
Approximately 65,000 chemicals are used worldwide for industrial purposes (Schroeder and Lane,
1988). Many of these chemicals eventually are emitted into the atmosphere. The 1987 TRI reported
that Industries within the United States emitted more than 1.2 billion kiograms of toxic chemicals into
the atmosphere (US EPA, 1989). The TRI underestimated the actual air emissions as It did not
include air emissions from numerous area sources (e.g., agriculture, households, motor vehicles),
Industrial categories such as petroleum tank farms, companies with less than 10 employees, or urban
businesses (e.g., dry cleaners).
A secondary source of airborne toxics is the application of pesticides. Atmospheric loads of
pesticide residues resulting from wind drift, as well as volatilization from sol and plant surfaces after
their application to agricultural, forest, range, industrial, and household lands are also significant.
More than 455 million kDograms of pesticide active ingredients are used annually on 16% of the total
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land area of the United States (Pimentei and Levttan, 1986). Agricultural lands account for
approximately 75% of the pesticide usage in the United States. The knowledge of magnitude of
pesticide active ingredients entering the atmosphere during and following application is limited, but is
estimated to range between 30 and 55% (Waddel and Bower, 1988).
ATMOSPHERIC PROCESSES
When chemicals are airborne, they are subjected to the prevailing atmospheric conditions.
Wind, precipitation, humidity, clouds, fog, solar irradiation, and temperature influence the
environmental fate of air toxics (Bidleman, 1988; Schroeder and Lane, 1988). The complex reactions
within the atmosphere that are driven by chemical processes such as hydroxyl scavenging or solar
Irradiation may result in the formation of products that can be as toxic, or more so, than the parent
compounds. On the other hand, transformation reactions may also render a toxic substance
harmless.
The atmosphere is a major pathway for the transport and deposition of the air toxics from
emission sources to the terrestrial ecosystem receptors - vegetation and soD (Bidleman, 1988). The
prevailing meteorological conditions and the physiochemical properties of the chemicals will dictate
atmospheric residence times and deposition velocities to the receptors (Schroeder and Lane, 1988).
Atmospheric residence times depend on such characteristics as mode and rate of emission,
atmospheric transformations, physical state (gas, solid, liquid), particle size, and chemical reactivity
(Schroeder and Lane, 1988). Thus, airborne pollutants may be deposited dose to their sources in
urban or agricultural ecosystems, or be carried great distances before being deposited into remote
ecosystems.
The movement of airborne chemicals downwind from point sources has received a great
amount of attention since the early 1900s when the damaging effects on vegetation that occurred
within plumes were recognized (Gordon and Gorham, 1963). During the last 10 to 20 years, however,
the phenomenon of long-range atmospheric transport has been documented in the wide distribution of
anthropogenic contaminants on regional and global scales. Industrial organic compounds, trace
metals, and pesticide residues have been detected in the vegetation of remote terrestrial ecosystems
such as the Arctic (Thomas, 1986), Antarctic (Bacci et a!., 1986), forests, (Eriksson, et al., 1989), and
peatiands (Rapaport and Eisenreich, 1988).
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ENVIRONMENTAL PARTITIONING AND VEGETATION EXPOSURE
Terrestrial plants are exposed to toxic chemicals through the environmental media of air,
water, and soO. The atmosphere, however, is of prime importance because of Its potential for
pollutant dispersal on a regional-to-giobal scale, Its abltty to move pollutants rapidly, and Its dynamic
nature. After pollutant deposition to terrestrial ecosystems, the fate of the toxic compounds depends
on their partitioning coefficients.
The environmental partitioning of pollutants within ecosystems wfll dictate their potential
ecological impact to vegetation and other biota (Weinsteln and Blrk, 1969). For example, trace metals
tend to accumulate on sol surfaces via their adsorption to organic matter. Trace metal accumulation
may reduce plant growth and vigor through the disruption of nutrient uptake by the roots and
decreased organic matter decomposition. Gaseous chemicals reside in the atmosphere with the
potential to disrupt plant-leaf biochemical processes after absorption through the stomata or cuticle.
Because of the lipophilic nature of many synthetic organics, the waxy cuticle of vegetation may
accumulate high levels of these substances. Transfer of toxic chemicals among ecosystem
compartments wfll occur. Trace metals may be absorbed by plant roots or deposited onto the leaves
and then transferred to the soD through deciduous tissue loss and decay. Contaminants may be
passed along food chains through herbivory with the potential for bfomagnlfication. The deposition of
airborne toxic chemicals Into agricultural ecosystems has the potential to contaminate human food
resources.
IMPACTS ON TERRESTRIAL VEGETATION
Scientists have recognized that airborne pollutants can adversely impact agricultural and
natural plant communities by reducing plant production and altering successional pathways
(MacKenzie and EI-Ashry, 1989; Weinstein and Birk, 1989). Emissions of sulfur dioxide, hydrogen
fluoride, trace metals, and other toxics from pulp and paper mills, ore smelters, and power plants
have severely reduced vegetation cover, biodiversity, and ecosystem integrity downwind from point
sources (Gordon and Gorham, 1963; Weinstein and Birk, 1969). In addition to local plume effects,
atmospheric pollutants also can cause regional damage to plant communities through exposure to
chemical oxidants such as ozone, peroxyacetyl nitrates, or add precipitation (Guderian, 1985;
MacKenzie and EI-Ashry, 1989).
The potential biological effects of air toxics on terrestrial vegetation are numerous and are
mediated through individual plants to the community and ecosystem (Weinstein and Birk, 1989). The
type and magnitude of these effects wil depend on the pattern of exposure (e.g., duration,
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concentration, frequency, season) individual plants receive, their sensitivity to the pollutant, and the
phytotoxictty of the chemical. When an airborne toxic chemical is Introduced into a plant community
some plants wfll be more affected than others depending on individual tolerances endowed by their
genotype, as well as their phenology and various modifying microclimatic variables. The sensitive
plants or species are no longer able to compete adequately with the tolerant plants or species and
wfll be partially or completely replaced. Those plants that do survive and persist in contaminated
habitats are the result of their tolerance or mterohabltat protection.
After pollutant absorption by the plants through the leaves or roots, biochemical processes
are the first site of action of the pollutant If enzymatic degradation detoxifies the pollutant, then no
injury wHI occur. However, If enzymatic action cannot render the pollutant or Its metabolites harmless,
then alterations in plant metabolism may result in foliar Injury, altered carbohydrate and nutrient
allocation, and reduced growth and reproductive capabPlty (Guderian, 1985). The degree of impact to
the plant wOl depend on the toxicity of the pollutant and Its exposure pattern. Acute exposures
usually cause observable morphological damage, such as leaf lesions, stunted growth, or even death.
Plant death resulting from acute exposure is usually localized when It does occur; resulting from an
inordinate amount of toxic chemical exposure through an accidental release or pesticide wind drift.
However, chronic, sublethal exposures may not induce observable morphological damage, but
rather alter biochemical pathways, which can result in decreased vigor and productivity, altered
phenology, loss of tissue, or reduced reproductive potential. Altered physiological processes will
cause a loss of vigor and render the plant more susceptible to insect damage or disease. Decreased
reproduction wfll impact the population through the loss of new recruitments to the plant community.
With continual exposure, even at sublethal concentrations, sensitive plant populations may decrease in
numbers allowing tolerant species to become dominant. Thus, shifts in plant community structure
and composition could result in decreased biological diversity and altered ecosystem functions.
Plant damage resulting from acute air toxic exposures are usually limited in time and space
as a result of control technology and legislation. However, sublethal, long-term plant exposure to
airborne pollutants may predispose vegetation to other natural stressors and induce damage or
mortality. Even though air toxic damage may not cause permanent functional loss, the diversion of
biochemical resources to repair the injury wfll inhibit normal plant functions. Thus, air-toxic-induced
physiological stress may predispose a plant to other stressors such as frost, drought, insects, or
disease. Some scientists propose that the widespread forest tree decline is not the result of a single
agent, but an interaction among chronic exposures of air pollutants and natural stresses.
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Air toxics may also indirectly affect vegetation by directly affecting other organisms that are
critically associated with the plants. Sol microorganisms and Invertebrates are critical In ecosystems
for litter decomposition and nutrient cycling. An accumulation of trace metals within the *O' horizon
of the soD may limit organic matter decomposition and nutrient avaBabVity to plants. Many plants rely
on insects for pollination. Airborne pollutants from the use of Insecticides can reduce nontarget insect
populations, resulting In inadequate flower pollination and subsequent seed set
The effects of air toxics on vegetation can be extended lo the animals within the ecosystem
(Schreiber, 1985). Reduced plant cover and habitat quality wll result in animals being more
susceptible to predatton and disease. Adequate birthing sites may be reduced because of changes In
vegetation structure and cover. Animal forage that is contaminated may result in decreased
population size or starvation. Toxic chemicals may be passed along food chains with the potential to
reduce the health of herbivores or may bioaccumulate within predators. Animal populations will
respond to such habitat changes through decreased reproduction, emigration, or mortality. Even
though the populations of certain species may increase because of favorable habitat changes, the
biological diversity of the overall ecosystem still may decrease.
RESEARCH NEEDS
Air toxic chemicals introduced into plant communities can produce effects ranging from the
biochemical level to changes in plant community structure and composition. The effects of acute
exposure to plants are well documented. However, a paucity of information exists on the effects of
long-term, chronic exposures of air toxics to vegetation. The following areas of research would
provide data to quantify vegetation responses to chronic air toxic exposure.
Identify and prioritize the most critical airborne contaminants and sensitive ecosystems. A
comprehensive computer-based system would be useful for conducting preliminary risk
assessments of the numerous airborne toxic chemicals and their effects on vegetation and
would provide research guidance.
Quantify and model the exposure, deposition velocity, and absorption of air toxics to plants.
Determine the biochemical and physiological responses of plants to chronic exposures to air
toxic chemicals and develop exposure-response functions. This research then could be
extended to quantify the response of plant populations and simulated plant communities.
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Initiate long-term studies to determine sensitive elements of plant community structure and
function that would lead to significant change and degradation from air toxic exposure. This
research would identify unacceptable change in plant communities and identify early warning
signals.
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ORGANOCHLORINES AND POLYCYCUC AROMATIC COMPOUNDS
IN THE HIGH ARCTIC
G.W. Patton1 and T.F. Bidleman8
Environmental Science Dept., Battelle - Pacific Northwest Laboratory, Rtehland, WA
8Dept. of Chemistry and Marine Science Program, University of South Carolina, Columbia. SC
INTRODUCTION
The Arctic is one of the most remote and isolated areas In the world. Despite the region's
seemingly pristine character, polar ecosystems are polluted by anthropogenic emissions from
temperate latitudes. Organochlorine (OC) pesticides and polychlorinated biphenyts (PCBs) have been
well .documented In the biota of the region (Bowes and Jonkel, 1975; Addison and Zinck, 1986;
Norstrom and Muir, 1988). The harsh polar environment requires that marine animals store large
amounts of fats, thus enabling them to accumulate high levels of lipophilic OC compounds. Native
people still rely heavily on marine animals for their food. A recent study reported the Insecticide
toxaphene (polychlorinated camphenes, PCCs) in fish from inland lakes of the Northwest Territories,
Canada (Muir et al., 1990). Long range atmospheric transport appears to be a major pathway of OC
pesticides and PCBs into the Arctic. Several studies have demonstrated or suggested that aerial
concentrations of these compounds can be linked to marine ecosystems (Norstrom and Muir, 1988;
Muir et al., 1990; Patton et al., 1989; Hargrave et al., 1988; Bidleman et al., 1989). A better
understanding of the concentrations, sources, and fate of organic contaminants In this region is
needed.
The polar atmosphere is a unique place to study air transport of organic compounds, due to
the large seasonal differences in temperature, photoperiod, and paniculate loadings. Most of the
pollution studies have centered around the arctic haze aerosol. Eastern Europe and Asia are the
8
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most likely sources of the haze, which is composed largely of suKate, elemental carbon, and other
combustion products. The haze pollution is seasonal with a maximum in the winter-early spring and
a minimum in the late summer-fall (Barrie, 1966).
There is a limited but growing body of Information on the types and levels of OC pesticides
and PCBs in arctic air. Tanabe and Tatsukawa (1980) measured DDT and hexachlorocydohexanes
(HCHs) over the Bering Sea. Workers from the Norwegian Institute for Air Research (Oehme and
Stray, 1982; Oehme and Ottar, 1984; Pacyna and Oehme, 1988) investigated concentrations and
temporal changes in PCBs, HCHs, chlordane, and heavy chlorobenzenes in air over Spttzbergen. Bear
Island, and Jan Mayan. Hoff and Chan (1986) measured chlordane isomers in the atmosphere over
Mould Bay, Northwest Territories, Canada. The research group (Patton et al., 1989; Bidleman et al.,
1989) used a floating ice island In the Beaufort Sea as a platform to collect HCHs, hexachlorobenzene
(HCB), chlordanes, the DDT group, PCBs, and PCCs in arctic summer air. Hargrave et al. (1988) also
measured OC levels in air, water, snow, sediments, and biota at the ice island.
Very little information exists on the types and levels of polycydic aromatic hydrocarbons
(PAHs) in the Arctic. Daisey et al. (1981) measured PAHs in spring and summer air in Barrow, AK,
and spring air at Narwahl Island. Fluoranthene (FLA) has been reported in the atmosphere of the
Norwegian Arctic (Pacyna and Oehme, 1988). PAHs and arctic haze emanate from combustion
sources and should be correlated. However this has not yet been established.
The Canadian Government maintains a joint weather station and military base at Alert, on the
northeastern shore of Ellesmere Island [(82.5 N, 62.3 W (Figure 1)]. We collected high volume air
samples at Alert during •February - Aprfl 1988 to determine the types and level of OC chemicals and
PAH in Arctic winter air.
METHODS
Air volumes of 2050 to 2490 m3 were pulled through two glass fiber filters (GFF) and two
polyurethane foam (PDF) plugs at flow rates of 0.5 to 0.6 m'/min using high volume pumps (Billings
and Bidleman, 1983). Collections were made on an elevated plateau about 150 m above the base,
approximately 5.5 km from the base. Electrical power for the pumps was supplied from the base's
generator. Contamination from the generator plume should have been minimal due to the prevailing
downward flow of air from the sampling hll to the station and a tight inversion layer over the camp
during periods of calm wind. Determining the extent of contamination from the base was not
9
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Arctic Ocean
Alert
\
FIGURE 1. Location of Alert, Northwest Territories, Canada, and the Ice Island during August 1986 to
June 1987 (Patton et al., 1989).
10
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practical during this experiment. Temperatures and daylight hours during the sampling period ranged
between -17 to -42 *C and 0 to 24 h, respectively.
PUP plugs were extracted in a Soxhlet apparatus with petroleum ether, and GFF were refluxed
with dichloromethane. Sample extracts were concentrated into isooctane and analyzed for
dibenzofuran PBF) and biphenyt by capBlary gas cnromatography • mass spectroscopy (GC-MS)
using electron impact ionization (El) and selected ion monitoring (SIM). DBF and biphenyl were
quantified against external standards. The extracts were then cleaned up and fractionated using
alumina-silicic acid column chromatography following the methods of Bllings and Bidleman (1983).
PAHs were analyzed by capillary GC-MS using El and SIM. The PAHs were quantified using
deuterated internal standards. The extracts were then treated with 18 M surfuric acid and analyzed for
OC and PCBs using capillary GC with electron capture detection (ECO). Pesticides and PCBs were
quantified by external standards. PCCs were analyzed by GC-negative ion-MS (GC-NIMS); ions
monitored were 309, 311, 343, 345, 379, 381, 413 and 415.
The low temperatures during the sampling period allowed the GFF-PUF system to retain
compounds as volatile as pentachlorobenzene (PeCB) and biphenyl with very low breakthrough to the
back PUF. The more volatile 1,2,4,5-tetrachlorobenzene was found in equal amounts in both PDF
traps and thus no quantitative results could be obtained. Recoveries of PeCB, HCHs, HCB, fluorene
(FLE), phenanthrene (PH), DBF, and biphenyl averaged 75% and results were corrected for this yield.
Recovery of all other compounds was 80 to 115% and no corrections were made.
ORGANOCHLORINE RESULTS
Concentrations of gaseous + paniculate chlorobenzenes and OC pesticides are given in
Table 1 and a comparison of our results to those of others Is presented in Table 2. PeCB, HCB, a-
HCH, and 7-HCH were found in all samples. Pentachlorobenzene and HCB levels are similar to other
values reported in the northern hemisphere (Bidleman et al., 1987). Compared to the ice-island study,
the level of cr-HCH was decreased by a factor of 2, whOe 7-HCH was fairly constant for both trips.
Beta-HCH was found in 4 of 5 samples at concentrations an order of magnitude lower than
THCH.
HCH insecticides are applied as either a technical mixture (55 to 80% cr-HCH, 5 to 14%
THCH, 8 to 15% /J-HCH) (Metcaff, 1955), or as pure rHCH (Undane). In North America and Europe
pure rHCH is used, whereas Asia, the Middle East, and Mexico use the technical mixture (FAOUN,
1978-1986). Tanabe et al. (1982) suggested that the heavy use of technical HCH by countries in the
eastern hemisphere is responsible for the predominance of cr-HCH. For this study, the a-HCH: 7-HCH
11
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TABLE 1. AIRBORNE ORGANOCHLORINES AT ALERT, CANADA (1988)*
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TABLE 2. COMPARISON OF MEAN REPORTED ORGANOCHLORINE LEVELS IN CANADIAN ARCTIC
AIR
(pg/ms)
Dates
HCB
a-HCH
rHCH
TC
CC
TN
IChlordane
ZDDT
PCBb
PCCe
Alert
This Work
Feb-Apr.
1988
120
160
23
.1.1
1.2
0.99
3.3
2.6
38
35
ICE IS.
USC(1)'
Sept.
1986
189
546
31
1.1
2.8
1.5
5.4
1.0
15.1
44
ICE IS.
USC(1)'
June
1987
>147
340
45
2.3
4.0
3.7
10
5.2
17.6
36
ICE IS.
BIO(2)'
May
1986
73
425
70
-
-
-
3.6
<1
-
~
ICE IS.
BIO(2)'
Sept.
1986
63
253
17
-
-
-
1.9
<1
™
Mould
Bay(3)'
June
1984
_d
-
-
0.96
1.5
1.3
3.8
-
-
•
• References: (1) Patton et al., 1989.
(2) Margrave et al., 1988.
(3) Hoff and Chan, 1986.
" PCB calculated as the sum of individual congeners.
c PCC calculated as toxaphene (from GC-MS data).
" No data available.
ratio. Also, the sampling period only extended five weeks into the light season (polar sunrise at Alert
occurred on March 5, 1988) and there may not have been sufficient time for the photochemical
conversion to occur.
Concentrations of trans-chlordane (TC), c/s-chlordane (CC) and trans-nonachlor (TN) are given
in Table 1. The overall concentrations of the chlordanes are lower than previously reported summer
air concentrations In the Canadian Arctic (Table 2). The mean TC:CC ratio was 0.95 and no trend in
this ratio with photoperiod was evident. Chlordanes are applied as a technical mixture. Trans-
chlordane would be thought to predominate in air because It is the most abundant and volatile isomer
in the technical mixture. Using Raoutt's law and the isomer composition of the mixture, the predicted
ratio of TC:CC In air is 1.7. Figure 2 compares TC:CC ratios for several locations and the ratio
appears to decrease away from application areas (e.g., Columbia, SC). The TC:CC ratio for this
winter study was twice as high as our previous value for arctic summer air.
13
-------
re
1
1
2.0
1.8-
1.6-
1.4-
1.2-
1.0-
08 -
0.6-
0.4-
0.2"
0.0
ice
Island
Columbia
Sweden
Sable
Island
Alert
Mould
Chlordane
Vapor
FIGURE 2. Comparison of trans-chlordane to c/s-chlordane (TC:CC) ratios in air. Sources: Ice Island
(Patton et at., 1989); Mould Bay (Hoff and Chan, 1966); Alert (this work); Sable Island (Bidleman and
Addision, unpublished data); Sweden (Bidleman et al.. 1967); Columbia, SC (Forman and Bidleman,
1987).
14
-------
BkJIeman et al. (1990) suggested that TC:CC ratios may be altered by selective degradation/removal
of TC. and these ratios may provide dues to the transit time of OC in the atmosphere. The higher
winter time TC:CC ratio (relative to summer) may indicate new or continuing Input of chlordane Into
the arctic atmosphere during this period, whle the lower summer ratio may indicate a stagnant (non-
input) period for the arctic summer where selective degradation/removal coukJ occur.
The levels of p,p'-DDT and p,p'-DDE in Alert air are given In Table 1. Concentrations of
p.p'-DDT were roughly double the p,p'-ODE levels In all of the samples. With GC-ECD results there is
a possibflity that DDT levels were inflated by coeluting PCC congeners. .The presence of p,p'-DDT in
the samples was confirmed by several GC-MS methods (GC-NIMS and GC-hlgh resolution-MS);
however, quantification of the GC-MS results was not attempted and the GC-ECD values for
p,p'-DDT and p,p'-DDE should be taken as upper limits. Other DDT-group compounds found by both
GC-ECD and GC-MS were o,p'-DDT and p,p'-DDD. Levels of p,p'-DDT and p,p'-DDE were similar to
those found during the last part of the June 1987 Ice Island study (Table 2).
Levels of PCBs in Alert air are presented In Table 1. PCBs were calculated as the sum of
individual congeners. The mean PCB concentration was 38 pg/m9 about twice as high as the Ice
Island result (Table 2). The more volatile PCBs dominated the congener distribution.
Concentrations of PCCs for three samples (calculated from GC-NIMS data) are given In
Table 1. The mean PCC level was 35 pg/m9, similar to the mean concentration (44 pg/m9) from the
Ice-Island study (Patton et al., 1989). The more volatile PCCs (6-7 chlorinated camphenes) were more
abundant in the samples than the heavier PCCs (8-9 chlorinated camphenes). Of the OC pesticides
identified in Alert air, PCCs were second in abundance (after HCHs) which may explain the high
concentrations of PCCs in fish from Northwest Terretories, Canada (Muir et al., 1990).
POLYCYCUC AROMATIC COMPOUND RESULTS
Concentrations of PAHs and biphenyl in Alert air are given In Table 3. DBF was found in all
samples PBF was not determined in PV*12) and was present almost exclusively in the vapor phase.
The mean DBF level was seven times larger than the value reported for North Pacific air (Atlas and
Giam, 1989). Biphenyl was also found in the vapor phase of all samples at an average level of 1250
pg/m9 (Tables 3 and 4). Full scan mass spectra of DBF and biphenyl from a sample extract are
shown in Figure 3.
A comparison of the PAH levels to those of other researchers is given in Table 4. The winter
concentrations were well below levels reported for winter air at Barrow, AK, and were more similar
15
-------
TABLE 3. AIRBORNE POLYCYCUC AROMATIC COMPOUNDS AT ALERT. NORTHWEST
TERRITORIES, CANADA (1988)
Sample #
DBF'
Biph
FLE
PH
AN
2-m-PH
FLA
PY
B(a)A
CR
B(b)F
B(k)F
B(e)P
B(a)P
Kcd)P
B(ghi)P
PV4
1600
1250
110
6.6
-
0.16
7.7
3.0
-
4.0
6.4
-
1.8
1.0
2.0
1.7
PV5
870
580
200
26
-
3.2
12
1.9
0.75
5.8
8.0
-
2.0
-
-
1.2
(pg/m3)
PV12
_b
670
440
130
-
-
74
50
8.3
24
23
13
8.0
3.1
5.5
8.6
PV14
2800
1400
340
120
0.78
2.9
74
49
3.0
21
19
-
5.6
3.4
5.6
3.6
PV 15
1700
700
100
34
0.47
0.76
30
14
3.6
11
17
-
4.0
3.2
4.1
4.5
* DBF = dibenzofuran, Biph = biphenyl, FLE = fluorene, PH = phenanthrene, AN = anthracene,
2-m-PH = 2-methyl-phenanthrene, FLA = fluoranthene, PY = pyrene, B/aJA = benzofajanthracene,
CR = chrysene, B[b]F = benzofcjfluoranthene, B[k]F = benzoftyfluoranthene.
B(e]P = benzofejpyrene, B[a]P = benzofa) pyrene. \(cd)P = lndeno(1,2,3-cd)pyrene,
B(ghi)P = benzo(ghi)perylene.
b No data available.
to levels reported for April at Narwahl Island (off the northern shore of Alaska) and summer air at
Barrow (Daisey et al., 1981). PAH levels at Alert were higher than those reported In the North Pacific
(Atlas and Glam, 1989). The low temperatures and higher particle loads at Alert dramatically raised
the partide:vapor ratios relative to the ratios reported In warmer North Pacific air (Table 4).
PAH In the volatility range between FLE to benzo/gW/ perylene (B/grty/ P) are presented in
Tables 3 and 4. FLE was found exclusively in the vapor phase at a mean level of 240 pQ/ms.
Phenanthrene (PH), FLA, pyrene (PY) and chrysene (CR) were found both in the gas and paniculate
phases. The amount of these compounds in the paniculate phase was operationally defined (front
GFF - back GFF) and the vapor phase was defined as ((2 x (back GFF) + PUF)).
16
-------
TABLE 4. COMPARISON OF MEAN SELECTED GAS-PHASE AND PART1CULATE POLYCYCLJC
AROMATIC COMPOUNDS IN REMOTE ATMOSPHERES
(pQ/m3)
Alert
(This Work) '
Gas Pan.
Compound
DBF 1700
Biph 920
FLE 240
PH 23
FLA 5
PY 3
CR 1
Benzofl uoranthenes
B[a] P
B/e; P
B[ghi] P
I/coy P
.'
-
40
35
21
12
28
2.7
4.3
3.9
4.3
North
Pacific (1)'
Gas Pan.
248
16
14
2.0
3.0
1.0
-
-
-
-
"
-
0.1
2.2
1.3
1.1
1.8
2.1
0.4
1.0
1.0
0.7
Alaska (2)*
Barrow Narwahl I.
Paniculate Only
(Aug.) (Mar.) (April)
-
-
17
28
29
5
-
10
-
10
'
-
-
120
290
310
80
-
30
210
70
"
-
-
20
60
150
-
-
-
-
-
"
" References: (1) Atlas and Giam, 1989.
(2) Daisey elal., 1981.
b No data available.
These corrections were made in an attempt to correct for gas adsorption to the GFF matrix (Ugocki
and Pankow, 1989). For the PAHs with vapor pressures below CR, the compounds were found
entirely on the front GFF.
CONCLUSION
The finding of OC and PAHs in arctic winter air provides additional evidence for the long-
range, atmospheric transport of anthropogenic pollutants to polar regions. In general, these
chemicals are persistent and toxic. The OC have bioaccumulation potential. Despite low atmospheric
concentrations, these pollutants impact the arctic environment and high levels can accumulate in polar
food webs. This study has provided the first measurement of both paniculate and vapor phase PAH
17
-------
1 —1
CO
I
I
168
DIBENZOFURAN
139
84
H H 76
I I I I I I I I I I I I I | I I
•6 M 106 116 126 196 146 165 1K
1—,
0>
c
76
BIPHENYL
MM I
U « 7B
1M
163
152
Ill Ml I I I
106 116 126 136 146
MASS/CHARGE
FIGURE 3. Full scan mass spectrum of the dlbenzofuran and biphenyl peaks from an Alert air
sample.
18
-------
in the Canadian Arctic. Despite the extreme cold, some PAHs have a sizable vapor phase
component.
ACKNOWLEDGMENTS
This work was supported in part by the National Science Foundation, Division of Atmospheric
Chemistry. Thanks to L Barrie of Atmospheric Environment Service, Downsvtew, Canada, for
extending the invitation for us to participate in the Polar Sunrise Experiment and for providing travel
and salary support. GC-MS analysis of toxaphene and DDT-group compounds were provided by
M.D. Walla and W.E. Gotham, Jr. A special thanks goes to the many people at the Alert weather
station and the 'Frozen Chosen' at C.F.S., Alert, who helped make this research possible.
19
-------
AIR-SEA EXCHANGE OF HEXACHLOROCYCLOHEXANE IN THE BERING
AND CHUKCHI SEAS
Daniel A. Hincktey', Terry F. Bidleman2 and Clifford P. Rice9
1 Ebasco Environmental, Bellevue, WA
2 Dept. of Chemistry and Marine Science Program, University of South Carolina, Columbia, SC
3 U.S. Fish and Wildlife Research Center, Laurel. MD
INTRODUCTION
Recent evidence documents long-range atmospheric transport and subsequent deposition of
organochlorine contaminants to terrestrial and aquatic ecosystems (Pacyna and Oehme, 1988;
Hargrave et al., 1988; Patton et al., 1989; Gregor and Gummer, 1989). As a result, these
contaminants have been found in arctic fish, marine mammals, birds, and plankton (Norstrom et al.,
1988; Muir et al., 1988). This raises concern because the food chain in these cold, remote locations
is relatively simple and may lead to high concentrations of pollutants in the diet of native Inuit (Muir et
al.. in press; DeWailly et al., 1989).
The third American-Soviet joint expedition to the Bering and Chukchi Seas in 1988 took place
on the R/V Akademik Korolev. This was the 47th oceanographic cruise for the Akademik Korolev
and the cruise was commonly termed the AK-47. The primary goal of the AK-47 was to characterize
the ecology of these waters. The Soviet Union and United Slates share a border between territorial
waters in this region and, as a result, most scientific studies have been on either the Soviet or
American side. This study was co-sponsored by the U.S. Department of Interior and the U.S.S.R.
Academy of Science. Therefore, the joint sponsorship allowed the first complete environmental survey
of the area. Several research groups consisting of Soviet and American scientists were established to
investigate the ecology of the Bering and Chukchi Seas among other things. This paper will focus on
the air-sea exchange of hexachlorocydohexane (HCH).
20
-------
The insecticide HCH is used throughout the world and is avalable in two formulations,
technical-HCH and lindane. Technical-HCH Is a mixture of five isomers In the following proportions
(Metcalf. 1955): (1) a, 55 to 80%, (2) A 5 to 14%, (3) 7, 8 to 15%, (4) 6, 2 to 16%. and (5) €, 3 to
5%. Although all isomers are toxic, only the gamma isomer is (nsecticidal and It is produced in pure
form as the insecticide lindane. Technical-HCH is now banned In the United States, the Soviet Union,
and western Europe, but continues to be heaviy used In Asia. Limited data are available about the
usage of HCH. Tanabe et al. (1982) reported the use of 77,000 metric tons of technical-HCH in India
from 1975 to 1978 and the production of 20,000 metric tons/year In one plant in China. The Food
and Agricultural Organization (FAO) Production Yearbook (1986-87), shows 23,400 metric tons/year
technical-HCH used in India between 1980 and 1985. Undane usage has been reported as 1,440
metric tons/year in Italy from 1980 to 1984 (FAO Production Yearbook, 1986-87), and 29 metric
tons/year in Scandinavian countries (Pacyna and Oehme, 1988). Together, the HCH Isomers are the
most abundant of the heavy organochlorines in the northern troposphere and surface waters.
Because of their high concentrations, low volumes of air (15 to 20 m3) and water (2 to 4 L) can be
sampled and analyzed for levels of a-HCH and rHCH.
The deposition of organic compounds from the atmosphere is controlled by their physical
properties (BkJIeman, 1988). HCH physical properties behave differently from most other
organochlorine pollutants. Vapor pressures and water solubilities of HCHs are sufficiently high that
they remain primarily as gases in the atmosphere or dissolved in the water column with relatively
small fractions sorbed onto particles. The exchange of HCH between the atmosphere and water
depends on the concentration gradient at their interface and its Henry's Law constant. HCH has a
low Henry's Law constant, facilitating transport from the atmosphere into water. A knowledge of air
and surface water HCH concentrations, along with the appropriate Henry's Law constant, allows the
determination of the equilibrium state for HCH in the water.
Consequently, surface (2 m) water concentrations of a-HCH and rHCH were determined at
19 stations of the AK-47 cruise, and air concentrations were measured at 16 of these same stations.
Henry's Law was applied to atmospheric HCH levels and the equilibrium between the atmosphere and
surface water was evaluated. In addition, HCH concentrations in the microlayer (top 200 pm), surface
water (2 m), and 1000 m seawater were compared for two of these stations.
21
-------
EXPERIMENTAL METHODS
Cruise Track
The cruise originated and terminated at the deep water port of Dutch Harbor, AK. and lasted
from July 26 to September 2, 1968. The cruise track, shown in Figure 1, was a convoluted route in
the Gulf of Anadyr, Southern Chukchi Sea, and Chirikov Basin. This is a shelf region with a maximum
depth of about 80 m. Two polygons were sampled in the southern Bering Sea where depths up to
4000 m were encountered (Figure 1). Sampling duration at each oceanographic station was relatively
short, lasting from 0.5 to 2 h. At these stations a specific conductMty/temperature/depth (CTD)
probe was sent to the bottom for water column analysis, and nutrient and primary productMty studies
were conducted. Sampling at each ecological station took place over 3 to 12 hours. In addition to
CTD, nutrient and primary productMty measurements and water samples were taken throughout the
water column for the determination of pollutant levels; and sediment, bacterial, phytoplankton,
zooplankton, and benthic Investigations were performed.
Sample Collection
Surface and deep water was collected using a variety of methods. Niskin and Go-Flo bottles
were sent to the required depth on hydrowire or Kevlar line and triggered by sending a messenger
down the wire. In addition, a water sampler, based on a design by Keizer et al. (1977), was
constructed using two solvent bottles mounted on a wooden frame. Teflon elbows were cemented
into the bottle caps, with a glass tube connecting the two bottles through the elbows. The
'messenger* sent down the wire broke the glass tube, which allowed water to fill the solvent bottles.
This sampler was used to collect surface water from 2 to 10 m and the bottles filled within 5 min.
Microlayer water samples were collected at three stations from a small boat at least 1 km from the
ship using a stainless steel screen technique (Garrett, 1965).
Air was pulled through two or three polyurethane foam (PUF) plugs (4.8 cm diameter, 3.2 cm
thickness) in a dean thick-walled glass tube (4.0 cm ID, 15 cm length) using a brushless pressure-
vacuum pump (MOIipore Corp., MBford, MA). Air was sampled continuously for 8 to 24 h at a flow
rate of 20 L/min, which yielded volumes of 10 to 30 m3. now rates were determined using an In-line
Top-Trak Model 820 flow monitor (Siena Instruments, Atlanta, GA). Breakthrough of analytes from
front to back PUF plugs was monitored by the separate analysis of each plug. Because of the low
air volumes sampled, only a-HCH, 7+ICH and hexachlorobenzene (HCB) were quantified.
22
-------
89-88.2
65C
60°
55e-
170° 180° 170s
I
—• _o_«— o—••
\. niA
»
Cruise Track
• Oceanographic
Station
O Short Ecological
Station
Full Ecological
Station
^f
Bering Sea
-69°
-67°
-65C
-63°
-61°
165°
FIGURE 1. Cruise track of the AK-47, July 26. 1988 to September 2, 1988. The three digit numbers
indicate stations.
23
-------
Preconcentration and Oeanuo
Analytes from 3.5 L water were preconcentrated by two methods: (1) liquid-liquid extraction
Into 300 ml methyiene chloride, and (2) adsorption onto C, bonded-phase cartridges (Hinckley and
Bidleman, 1989). The methyiene chloride extract was reduced in volume using a rotary evaporator
and the solvent was replaced with pesticide-quality hexane prior to gas chromatographic (GC)
analysis. The C, bonded-phase cartridges were eiuted with 3 ml 1:1 ethyl ether-hexane, and blown
down Into pure hexane with nitrogen. Poiyurethane foam plugs were extracted for 6 h in a Soxhlet
apparatus with petroleum ether, the petroleum ether volume was reduced using the rotary evaporator,
and the solvent was replaced with hexane or isooctane prior to GC analysis. All extracts were treated
with concentrated sutfuric acid for cleanup.
Gas Chromatographic Analysis
Gas chromatographic analyses were conducted using a Hewlett Packard (HP) 5840, Varian
3700, or Carlo Erba 4160 chromatograph with *Ni electron capture detectors (GC-ECD). The
instruments contained 25-m bonded-phase fused silica columns (polydimethyisiloxane, 5% phenyl,
Hewlett Packard or SGE Corp.). Carrier gases were hydrogen or helium at 30 to 40 cm/s, the
injector temperature was 240<€ and the detector was 320'C. Samples were injected using a splltless
Grob technique. Chromatographic data were collected using the HP-5840 integrator, an HP-3390A, or
a Shimadzu Chromatopac CR3A integrator. Alpha-HCH and rHCH were confirmed by GC-electron
impact mass spectrometry (GC-MS) with selected monitoring of the 181 and 217 ions using an HP
5890 GC with a mass selective detector and the same type of column used for GC-ECD work.
Henry's Law Constant Determination
Air-sea exchange is controlled by Henry's Law:
Cfl=KMxC, (1)
where Cg and C, are equBibrium concentrations in the air and surface water, respectively, and KH is
the distribution coefficient. KH can be further defined:
KM - H/RT (2)
where H is the Henry's Law constant (atm-m3/mol), R is the gas constant
24
-------
(8.2 x 10* m'-atm/mol-K), and T the surface water temperature (K). Lastly, the Henry's Law constant,
H, can be thermodynamically defined as:
H = p°/c° (3)
where p° is the vapor pressure and c° the aqueous solubility of the compound. Use of Equation 3
allows calculation of H; however, the aqueous solubitties of HCHs are not well characterized as a
function of temperature and we opted for direct determination of H in the laboratory.
Henry's Law constants have been determined in the laboratory by the gas stripping method
(Mackay et al., 1979). In this technique an aqueous solution of the HCHs in a thermostat-controlled
vessel is purged with air to strip the HCH vapors. The equation for the loss of HCH from solution
with the time is
In (C,/C0) = -(HG/VRDt (4)
where C, and C0 are the aqueous HCH concentrations at time = t and at time = 0, G Is the gas flow
rate (m'/h), and V is the solution volume (m3). The H value Is determined by the regression of
(C,/C0) on t (h) which has a Slope of -(HG/VRT) (Mackay et al., 1979).
Henry's Law constants were determined at 23 *C (n=3) using artificial seawater (Home, 1969).
The seawater was stripped with a cleaned air stream for 95 to 120 h. The gas-stream flow
(350 L/min) was monitored daily using a bubble flow-meter. Daily water samples were taken for
analysis of a-HCH and rHCH. Initial concentrations were approximately 6 fjg/L for both analytes.
After 120 h the concentration of a-HCH and rHCH reduced to 1 ug/L and 3ug/l_ respectively. The
state of equilibrium was monitored by varying the depth through which the bubbles rose. The
Henry's Law constant determined at 25 cm depth was the same (within experimental error) as that
found using 45 cm depth.
Quality Control
Spike recovery of 17 ng of a-HCH and rHCH, shown In Figure 2, ranged from 80% to 95%.
HCH concentrations reported in this paper have not been corrected for recovery. Air sample
breakthrough from front to back PUP averaged 32% and 18% for rHCH and a-HCH, respectively,
and 12% for HCB. Limits of detection, based on the area of lowest observable peaks near the area
of interest, were 0.15 ng/L in water and 0.03 ng/m' in air for both a-HCH and rHCH. Liquid-liquid
25
-------
100
95
90
Recovery
85
80
75
Water-BP
Water-LL
Air-PUF
alpha-HCH
gamma-HCH
FIGURE 2. Spike recovery of 17 ng a-HCH and THCH from bonded-phase cartridges (BP), liquid-
liquid extraction (LL), and poiyurethane foam plugs (PDF) obtained on the AK-47 cruise.
26
-------
and Ce cartridge methods of preconcentratkxi were compared at four stations (003, 007, 009, and
019; Figure 1). As comparable levels of HCH were found, regardless of the preconcentration method,
water at all remaining stations was analyzed using the liquid-liquid extraction method.
RESULTS
Water and Air Concentrations
Microlayer and surface water concentrations of cr-HCH and rHCH for station 003 are shown
in Table 1. No enhancement of HCH was found in the microlayer. There was a significant decrease
in levels of HCH with depth at station 110. The decrease from 2 m to 1000 m for a-HCH and rHCH
was 81% and 44%, respectively. These results are similar to those reported by Tanabe and
Tatsukawa (1980) In the North Pacific. Transport of HCH from surface to deep water is hindered by
low particle association of HCH (Tanabe and Tatsukawa, 1983). Consequently, HCH entering surface
water from the atmosphere will tend to stay at the surface and not be removed to bottom water on
particles, resulting in lower HCH concentrations in deep Bering Sea water.
TABLE 1. COMPARISON OF HCH' IN MICROLAYER, SURFACE, AND DEEP WATER
Concentration of HCH (ng/L)
Station
003
110
Microlayer Surface (2 m)
a 7 a
2.07 0.52 2.56
2.25
7
0.63
0.43
Deep (1000 m)
a 7
b
0.23 0.24
" No data available
Levels of a-HCH and rHCH in surface water (19 stations) and air (17 stations), and HCB in
air are given in Table 2, and are compared with literature values in Table 3. HCH concentrations
found during the AK-47 are simlar to levels found by Kawano et al. (1988) in the Bering Sea.
Beaufort Sea locations farther north appear to have higher HCH concentrations. Blanca Bay,
Argentina, has been included in this table as an example of local estuarine contamination.
27
-------
TABLE 2. SUMMARY OF AIR AND SURFACE WATER HCH CONCENTRATIONS
Air (ng/ms) Water (ng/L)
o-HCH
n 17 19
mean ± sd 0.21 ± 0.06 2.35 ± 0.36
range 0.12 - 0.34 1.78 - 2.92
THCH
n 17 19
mean ± sd 0.13 ± 0.05 0.52 ± 0.10
range 0.08 - 0.29 0.34 - 0.69
HCB
n 17
mean ± sd 0.17 ± 0.04
range 0.12 - 0.27
• No data available
TABLE 3. COMPARISON OF a-HCH AND THCH CONCENTRATIONS IN AIR AND WATER FROM
THIS STUDY WITH REPORTED VALUES
Location (Ref.)
Bering and Chukchi Seas
(this work)
N. Pacific and Bering Sea
(Kawano et ah, 1988)
Bering Sea (Tanabe and
Tatsukawa, 1980)
Beaufort Sea (Hargrave
et al.. 1988)
Beaufort Sea (Patton
et al.. 1989)
Hudson Bay (McCrea
and Fischer, 1986)
Blanca Bay, Argentina
(Sericano and Puccl, 1984)
Date
8/88
7/81
7/79
6/86
6/87
1980-81
1980-81
Water
o-HCH
2.35
2.75
(ng/L)
rHCH
0.52
0.65
Air (ng/m3)
a-HCH THCH
0.20 0.13
_b
3.9'
4.40
7.1
6.42
48.2
0.57
0.81
0.86
54.2
0.45 0.07
0.34 0.05
-
-
' Sum of a-HCH and
b No data available.
28
-------
An interesting aspect of the atmospheric HCH data is the 0/7 ratio. This has been suggested
as a marker for recent atmospheric transport of these pollutants (Pacyna and Oehme, 1988). During
this cruise the 0/7 ratio ranged from 1 to 3 and averaged 1.4. This is a much lower value than that
found in the Canadian Arctic, and is actually lower than in most regions (Pacyna and Oehme, 1988).
This may Indicate recent atmospheric transport of lindane from the Soviet Union and Asia over the
Bering and Chukchi Seas.
Henry's Law Constants and Air-Sea Exchange
The Henry's Law constants for a-HCH and rHCH in artificial seawater at 23*0, along with
literature values, are shown in Table 4. Comparison of experimental and literature values is favorable,
indicating the stripping system is at equilibrium. The Henry's Law constants at other temperatures
have not yet been determined, and temperature dependence of H for the HCHs was assumed to be
similar to that of PCBs (Patton et al., 1989; Burkhard et al., 1985). Station 061 in the Chukchi Sea,
where the surface water was 5*C, had air concentrations of 0.19 ng/m'and 0.13 ng/m3for
a-HCH and rHCH, respectively. Using H=1.24 x lO^atm-m'/mol for a-HCH and H=6.27 x 10*
atm-m'/mol for 7-HCH at 5*C and applying Equation 1, equilibrium surface water concentrations were
calculated. The results reveal a-HCH in apparent equilibrium between the air and water, while rHCH
exhibits a large disequilibrium, with lindane fluxing from atmosphere to surface water (Figure 3). This
may imply that fresh lindane, coming from the Soviet Union and Asia, Is a source of the pollutant into
Bering and Chukchi Sea surface water.
TABLE 4. HENRY'S LAW CONSTANTS (atm-m3/mol) FOR HCH IN SEAWATER
Isomer Temp, (*Ł) This Work" Other Work"
a-HCH
y-HCH
23
23
6.5 ± 1.0x 10*(n=4)
3.3 ± 0.9 x 10* (n=4)
1.1 x 10"5
2.5 x 10" e
" Mean ± standard deviation.
b Atlas et al., 1982; Fendlinger and Glotfelty, 1988.
c Calculated from reported fresh water value, assuming a 25% decrease of solubility In seawater
(Masterson and Lee, 1972).
29
-------
5.0
4.0
*
ID
1,0
5
1.0
0.0
67°20'N, 169°45'W
Equilibrium Actual
•Ipha-HCH
Equilibrium Actual
gamma-HCH
FIGURE 3. Comparison of surface water HCH concentrations in equilibrium with air (see text) with
actual measured concentrations for station 061.
30
-------
CONCLUSIONS
As a part of the 47th cruise of the R/V Akademlk Kordev, cosponsored by the U.S.
Department of Interior and the U.S.S.R. Academy of Science, air and surface water concentrations of
o-HCH and rHCH have been determined for 16 locations In the Bering and Chukchi Seas.
Concentrations of HCH were not enriched In microlayer water. Comparison of surface water (2 m) to
deep water (1000 m) in the Bering Sea showed an 80% decrease in o-HCH and a 40% decrease in
rHCH, Indicative of an atmospheric source. Surface water concentrations were 2.35 ± 0.36 ng/L for
o-HCH and 0.52 ± 0.10 ng/L for rHCH, slmBar to levels found by Kawano et al. (1988) in July 1981.
Air concentrations were 0.21 ± 0.06 ng/m'and 0.13 ± 0.05 ng/m* for cr-HCH and rHCH,
respectively, and 0.17 ± 0.04 ng/m3 for HCB. Levels of o-HCH and rHCH were slmBar to those
found In the Canadian Arctic (Hargrave et al., 1988; Patton et al., 1989); however. rHCH
concentrations in air were 2-4 times higher over the Bering and Chukchi Seas. Using Henry's Law
constants determined in the laboratory and assuming temperature dependence of H simDar to the
PCBs, rHCH appears to be fluxing from the atmosphere to the ocean, whereas o-HCH in surface
water appears to be in equilibrium with the atmosphere. This, along with higher 0/7 air ratios than
those usually found in the northern hemisphere, suggests recent atmospheric long-range transport of
lindane from the Soviet Union and Asia to these waters.
Because of their relatively high water solubilities, bioconcentration factors (BCF) of the HCHs
are lower than those of chlordanes and DDTs (Kawano et al., 1988). However the lower BCFs are
offset by the much higher concentration of HCHs in seawater. As a result, concentrations of HCHs in
fish and zooplankton from the Arctic are similar to or exceed those of PCBs and other organochlorine
pesticides (Muir et al.. 1988; Kawano et al., 1988).
ACKNOWLEDGMENTS
Special thanks to the U.S. Department of Interior, Rsh and Wildlife Service, and the U.S.S.R.
Academy of Sciences for sponsoring the 47th cruise of the R/V Akademlk Korolev. The U.S.
Environmental Protection Agency, Great Lakes Program Office, funded the laboratory Henry's Law
constant determinations.
31
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TRACE METAL DEPOSITION AND CYCUNG IN FORESTED ECOSYSTEMS1
Andrew J. Friedland
Environmental Studies Program
Dartmouth College
Hanover, NH
INTRODUCTION
Trace metals are.released into the atmosphere through a variety of human activities such as
fossil fuel combustion, smelting, and manufacturing, as well as through natural processes such as
volcanoes. Both short- and long-range atmospheric transport and deposition occur. Forests are one
ecosystem where trace metals accumulate. The accumulation of trace metals in some forests,
specifically in the organic-rich horizon of the soil, is well documented. The effects of trace metals on
biological processes are not well understood. This report will discuss the evidence for trace metal
deposition and accumulation in forests of the northeastern United States and the potential for adverse
effects on forest processes.
STATEMENT OF THE PROBLEM
Concentrations and amounts of lead, and possibly copper, zinc, nickel and cadmium, in many
forests are substantially higher today than they were earlier this century. The primary source of the
lead and partially for other elements is air pollution. The effect that the accumulation of these trace
metals have oh nutrient cycling, trace metal cycling, soi biota diversity and abundance,
decomposition processes, and plant root growth is not known.
1 Portions of this report have appeared in the following publications: Friedland et al., 1984a,b; Friedland
et al., 1986b; and Friedland, 1989.
32
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EVIDENCE FOR ATMOSPHERIC TRANSPORT AND DEPOSITION OF TRACE METALS
Sources of Trace Metals In Forest Ecosystems
There are three major sources of trace metals in most terrestrial ecosystems: (1) the
underlying parent material, (2) the atmosphere, and (3) the biosphere. Biotic sources of trace metals
are originally obtained from one of the other two sources. In different systems, the relative input from
each of these three sources varies.
Parent Material Inputs
In a natural, undisturbed ecosystem, the primary source of most trace metals is the underlying
bedrock (Adriano, 1986) or surface material transported via the atmosphere from another location
(Davidson et al., 1981). Except in areas of ore deposits and other unusually high concentrations of
metals, the trace metal content in most parent material is quite low. The average or range of
concentrations for llthospheric material and sedimentary rocks for six trace metals is shown in Table
1. Trace metals in soils usually are derived from the underlying parent material and are often similar
to parent material concentrations. Organic, soil-metal concentrations are generally higher than the
underlying mineral soil (Friedland et al., 1986a) (Table 1).
Bedrock is not necessarily always the parent material. In certain environments, glacial till or
fluvial and other sediments can be transported hundreds and even thousands of kilometers to new
locations. The subsequent soil that is formed will have a very different physicochemical
characterization than adjacent areas not covered by the transported soil (Birkeland 1984).
Weathering is the process by which rocks are slowly broken down into the mineral content.
The weathering processes that occur over time-both physical and chemical-lead to the breakup of
bedrock or surface rock, and release of elements, including trace metals, into the environment. These
metals then may enter rivers, streams, and other aquatic systems as well as terrestrial systems.
Atmospheric Inputs
Emissions of trace metals as participates and gases from volcanoes, forest fires, crustal
material, and continental dust have always been a natural input to soils and ecosystems (Davidson et
al., 1981). It is possible that before industrialization the combustion of dung and wood by humans
was a source of trace metal pollution pavWson et al., 1985). During the last 5000 years, human-
related emissions of trace metals have become increasingly important. In considering the relative
contribution of natural and human sources, many investigators have ignored vapor deposition of trace
metals because of a paucity of data. However, data are available on the deposition of atmospheric
mercury vapor, for example, in specific geographic areas, and to a lesser extent for the
33
-------
TABLE 1. AVERAGE OR RANGE OF TRACE METAL CONCENTRATIONS IN CRUSTAL MATERIAL
AND SOILS FROM A VARIETY OF LOCATIONS IN THE WORLD".
Element
Nickel
Copper
Cadmium
Zinc
Mercury
Iron
Uthosphere
75
24
0.5
70
0.243
31.5
Sedimentary
Rocks
48
30
1.0
20-97
0.040-0.055
6.7-49
Mineral
16
20
0.2
50
0.070
17
Snik
Organic
70
350
0.9
66
0.161
44
'All concentrations are in parts per million. Source: Adriano 1986; Nriagu 1979a,b, I980a,b.
Modified after Friedland (1989).
entire globe, but they are often excluded from global estimates of mercury cycling (LJndberg, 1987;
Galloway et al., 1982; Lantzy and Mackenzie, 1979).
Galloway et al. (1982) combined vapor emissions data with the dust paniculate emissions data
of Lantzy and Mackenzie (1979) and determined the ratio of atmospheric emission from natural and
human-related sources for a number of trace metals. As expected, aluminum has a very low
anthropogenic-to-natural emission ratio (approximately 0.15). Continental dust is the greatest source
of aluminum in the atmosphere; industrial processes and fossil fuel combustion contribute only slightly
to atmospheric aluminum. In some global estimates the anthropogenic-to-natural emission ratio for
mercury is almost as low as the value for aluminum. Lead has perhaps the highest anthropogenic-to-
natural emission ratio (approximately 346); automobile emissions and smelting are the two major
sources of technogenic lead In the atmosphere. In the past decade, as North America and some
European countries have used less leaded gasoline, the importance of automobile emissions as a
source of lead has decreased.
Ecosystems receive trace metals through wet (e.g.. rain) and dry (e.g., paniculate) deposition.
The relative contribution of precipitation, doud water, and dry deposition to a forest, and the effect of
the forest canopy on throughfall chemistry vary depending on site factors including elevation,
topography, and vegetation (Lovett et al.. 1982; Lovett et al., 1985). In high-elevationa! forests, cloud
water may be a significant source of trace metals (Lovett et al.. 1982). The relative contribution of
34
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wet versus dry deposition also may influence the relative solubilities of metals. In general, trace
metals are less soluble in dry than in wet deposition.
Biotic Inputs
Inputs to a system from existing vegetation occur In a number of ways: (1) inputs from
above-ground biomass, (2) inputs from roots and other below-ground biomass, and (3) leaching and
wash off from leaf surfaces. These inputs also are considered fluxes within the ecosystem and will be
discussed in greater detail In the next section. The lead concentration of most vegetation is relatively
low despite evidence that It is probably five-times higher than prior to Industrialization (Shirahata et al.,
1980). In forested sites under a variety of conditions in Europe and North America, the amounts of
lead, cadmium, and nickel in vegetation are somewhat lower than the amounts of the same metals in
soil (Smith and Siccama, 1981; Friedland and Johnson, 1985). Of course, near smelters, vegetation
levels of most trace metals are quite high (Jackson and Watson, 1977).
It is generally accepted that soil and other environmental trace metal concentrations are
substantially greater today than they were hundreds of years ago (Settle and Patterson, 1980). The
most direct evidence of increased trace metal concentrations in this century is available from forest
floor (the organic horizon overlying the mineral soil) samples collected at two points in time from the
same locations in Vermont and Massachusetts (Siccama et al., 1980; Friedland et al., 1984a). These
studies indicate that lead concentration has increased by as much as 40% in less than two decades.
By examining buried, organic horizons at a northern hardwood forest in New Hampshire, Johnson et
al. (1982) were able to approximate lead concentrations from earlier this century. They assumed that
the buried organic horizons were sealed from the atmosphere at the time of the disturbance (tree-falls
of unknown date, but probably in the early 1900s). Concentrations of these buried organic horizons
were 24 /vg/g compared to a modem organic horizon lead concentration of approximately 88 /jg/g.
Regional Patterns In the Northeastern United States
During the 1960s and 1970s, soil samples were collected from a number of locations in the
Northeast (Andresen et al., 1980; Johnson et al., 1982; Friedland et al., 1984b; Friedland et al., 1986a).
All of these samples were analyzed for their trace-metal content (usually lead, copper, zinc, nickel,
and cadmium). In 1978, the researchers collected forest floor samples from 40 sites in the mid-
Atlantic states and southern New England. The survey was extended to include the rest of New
England (excluding Maine) and more of the mid-Atlantic region for a total of more than 400 samples
from 80 sites. Since 1980, forest-floor samples have been collected at random intervals. These
samples have been archived for future reference to samples that will be collected during 1990 and
35
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1991. These data wfll enable the descriptions of spatial patterns of trace metal concentrations and
amounts in forest floors in the northeastern United States.
Known Effects of Trace Metals on Forested Ecosystems
Bioavallablllty and Biological Effects of Trace Metals
There are important differences between the concentration and amount of an element in the
soil. Individuals concerned with the effects on biological organisms including plants will be more
Interested In the concentration of an element than the amount.
Rickard and Nriagu (1978) describe the chemistry of lead in soils as controlled by three
parameters: (1) adsorption at the soil, mineral interface, (2) the formation of relatively stable organo-
lead complexes and insoluble organo-lead particulates and chelates, and (3) the precipitation of
relatively insoluble lead compounds. The relative importance of these three controlling factors varies
depending on the type of ecosystem and ultimately determines the bioavailability of each element
(Cataldo et al, 1987).
There are many methods for determining the speciation and mobility or bioavailability of trace
metals in terrestrial ecosystems. Sequential extractions usually are conducted to determine the
relative difficulty In removing successive fractions of trace metals from a soil or sediment (Miller et al.,
1983). Because the results from sequential extractions are site- and procedure-dependent, it is
difficult to generalize as to the relative fractions of metals in terrestrial ecosystems. In forests with
abundant quantities of organic matter present, many authors have theorized that most of the trace
metals are organically bound and there is relatively little free" lead (Johnson et al., 1982).
A study of the relative fractions of trace metals in forest soil was conducted on a sandy soil
underlying an oak forest in an Industrial area of northwestern Indiana (Miller and McFee, 1983). In
this study, the "0" (organic) horizon was removed and the upper 2.5 cm of mineral soil was used for
the analysis of cadmium, zinc, copper, and lead fractions. The authors found that the largest fraction
of most trace metals was bound to organic-matter, but there were substantial amounts held on soil
exchange sites, and associated with carbonates and iron oxides. Lead had the highest percentage
associated with organic matter (41.2%); cadmium the lowest (21%). It is Important to realize that the
percent organic matter (as estimated by loss on ignition) for the samples in this study ranged from
7.9 to 15.9, and the average was 9.7. In a forest floor, with loss-on-ignltion of 50 to 70%, it is likely
that the fraction of trace metals in the organically bound component would be substantially greater.
Generally, lead complexation Is related to organic matter content and carbonate levels In son.
36
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Regardless of the form and avaBabOity of the trace metals, they will have some effect on soil
microorganisms. Many researchers have found that enzyme activity, CO2 production, nitrogen
mineralization, and numbers of microorganisms are lower in areas dose to smelters or other point
sources of pollution when compared with areas distant from pollution sources. It is known that
extreme trace-metal concentrations (103 mg/kg or greater for lead, zinc, and nickel; 102 mg/kg or
greater for copper, and 10 mg/kg or greater for cadmium) affect microbial respiration and soil
organism populations (Ruhling and Tyler, 1970; Jackson and Watson, 1977). Several attempts have
been made to determine the effects of moderate concentrations of trace metals on decomposition
(Inman and Parker, 1978; Spalding, 1979; Friedland et ai., 1986a). Low-tc-moderate, trace-metal
concentrations are of interest because the metal concentrations found in forests of the northeastern
United States could be considered greater than "background levels" but much lower than
concentrations found near smelters (Johnson et al., 1982; Friedland et al., 1986b).
Effects of Metals on Forest Floor Processes at Camels Hump, VT
Friedland et al. (1986b) reviewed a number of studies that assessed forest floor, litter, and
organic soils and examined the effects of trace metals on soil microfauna, nitrogen mineralization,
carbon dioxide evolution, and enzyme activities (Figure 1). Trace-metal concentrations (in single- or
multiple-metal experiments) that were associated with a significant effect on one of the above
parameters are shown with solid lines (Figure 1a). Those levels associated with no detectable effect
are shown with broken lines (Figure 1b).
The maximum trace metal concentrations found at three sites in the northeastern United
States are identified in Figure 1. For all five metals shown, (lead, copper, zinc, nickel, and cadmium),
the concentrations found at Square Lake, ME, are well below concentrations necessary to significantly
and noticeably alter soil biological activities. The concentrations of trace metals at Camels Hump, a
mountain in the Green Mountains of Vermont, where research has been conducted for 10 years, are
also below the concentrations that probably affect soil activities. Only the concentrations encountered
near a zinc smelter in Palmerton, PA, are high enough to justify the conclusion that trace metal
concentrations affect the activity of soil microorganisms.
However, ft is quite possible that long-term exposure to the concentrations found at Square
Lake, ME, or Camels Hump, VT, could alter species composition, species diversity, or other subtle
measures of health, vigor, or resilience of a soil microorganism population. Current techniques of
assessing chronic effects of trace metals in natural settings (Le., in the forest) are probably not
sensitive enough to detect any immediate adverse effects of metal concentrations such as those
found on Camels Hump.
37
-------
©
Cu
Mi
Cd
Zn
Cu
Ni
Cd
01
*S
•o
o>
to
E
10*
101
10*
10'
"
XT
.5.
10"
Cu
Ni Cd
Zn Cu
I
'«
*
Ni
'
::":
.(L.
*
Cd
FIGURE 1. Trace metal concentrations In organic matter at which CO, production, enzyme
activity, litter accumulation, mineralization rate, or microbial populations were observed. —, an
effect of the metal was observed (a); —, no effect was observed (b). Metal concentrations are either
found at the sampling site or those added by the Investigator. Data In this figure were compBed from
the studies reference in this paper. Metal concentrations for three localities In New England are also
Shown: A, Square Lake, ME, (Fernandez and Czapkowsky. 1985); B, Camels Hump, VT. in 1966
(Friedland, Johnson and Siccama, 1964a); C. Camels Hump, VT, In 1980 (Friedland, Johnson and
Siccama. 19B4b); D, Palmerton, PA, (Buchauer, 1973). Source: Friedland et al., I986b.
38
-------
CONCLUSIONS
Trace metals have accumulated in the organic sols of forests in the northeastern United
States. The concentrations of trace metals present in the forest floor at high-elevation sites such as
Camels Hump. VT, are probably not high enough to interfere with sol organism activities, including
organic matter decomposition. It is possible that trace metals have affected organic sol processes
such as sol organism species composition and diversity.
39
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ATMOSPHERIC DEPOSITION OF TRACE METALS TO UCHENS AND EFFECTS OF
TRACE METALS ON UCHENS
Thomas H. Nash III
Department of Botany
Arizona State University
Tempe, AZ
INTRODUCTION
Uchens are symbiotic organisms composed of a photobiont (alga and/or cyanobacterium)
and mycobiont (mostly Ascomycotina) that commonly occur as epiphytes, or as free-Irving organisms
on rocks and soil. On a world-wide basis they are the dominant plant group on approximately 8% of
the land surface (Larson, 1987) and are particularly prominent in polar regions. For example,
Kershaw (personal communication) estimates that lichens dominate almost 40% of Canada. In
addition, they are important components of many western and northern forests as epiphytes with
biomasses as high as 700 to 3300 kg/ha (Boucher and Nash, 1990). In such forests finely dissected
lichens may constitute over 50% of the canopy surface area, and hence are important in the
interception of atmospheric aerosols.
One must recognize that lichens are fundamentally different from vascular plants and as a
consequence may be better suited for metal deposition studies (Puckett, 1988). First, lichens possess
no roots and have no other known mechanism for absorbing nutrients from soOs. Therefore, they are
largely dependent on atmospheric deposition for their nutrition (see below). Second, lichens possess
no stomates or cuticuiar waxes and thus may take up atmospheric deposition over their entire surface
areas. Third, lichens possess considerable intercellular space [(up to 20% of the volume in a
Xanthoria species (Collins and Ferrer, 1978)], that provides a repository for particulates. Furthermore,
40
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lichens are slow-growing perennials, many species of which may live hundreds of years. Because
lichens do not shed parts (such as leaves for many vascular plants), their morphology remains
uniform across seasons and years. As a consequence they provide uniform long-term receptors for
atmospheric deposition. The fact that many lichens occur over wide geographical areas allows for
both local and regional deposition monitoring. Finally, as discussed below, lichens have several
mechanisms for taking up metals, and hence accumulate metals as the length of exposure to
atmospheric deposition increases.
STATEMENT OF THE PROBLEM
Because of many of the ecological and morphological properties discussed above, lichens
certainly are capable of intercepting atmospheric deposition of toxics. In addition, lichens are capable
of retaining contaminants once intercepted and are therefore capable of providing useful biological
monitors of trace metal deposition. On the basis of many studies of pollution source sites, elevated
patterns near the pollution sources are well documented and therefore the trace metals must be
retained. In addition, mechanisms for uptake and retention are well documented and principally
include: (1) ion exchange, (2) intracellular uptake, and (3) paniculate entrapment (Brown and Beckett,
1983; Nieboer and Richardson, 1981; Nash, 1989). Cell walls of the dominant sites, estimated
variously as having a cation exchange capacity ranging from 6 to 77 /^mol/g (Nash, 1989). These
sites may become saturated in a matter of minutes when exposed to solutions of metals. By
contrast, intracellular absorption occurs gradually over hours and follows Michaelis-Menten kinetic
patterns. The magnitude of such internal absorption is generally much lower than external uptake at
cell wall exchange sites. Even more important is paniculate entrapment in intercellular spaces. This
has been directly demonstrated by using scanning electron microscope (SEM) procedures in concert
'with an electron probe of the X-ray diffraction analysis of the particulates found with the SEM (Garty
et al., 1979). The chemical profile of these particulates closely matches the known profile of
particulates emitted from nearby pollution sources.
DEPOSITION OF TRACE METALS
Studies on metaliferous concentrations in lichens surrounding pollution source sites are
numerous and well reviewed by James (1973), Nieboer and Richardson (1981), Puckett (1988),
Puckett and Burton (1981). In most cases distinctly elevated concentrations are found near pollution
sources as compared to areas farther away. In the absence of multiple pollution sources,
concentrations may decrease logarithmically with distance (Nieboer and Richardson, 1981). Nieboer
41
-------
et al. (1978) summarized much of the early literature and Initially constructed a table comparing
elevated concentrations with background concentrations for a number of elements including metals.
This information Is edited to delete nonmetals and is further expanded to include metals not in the
original tables (Table 1). The original table summarized baseline Information from several locations.
To llustrate the range of background data for Individual regions, the recent investigations of Gough et
al. (I988a) for epiphytes (fypooyrnn/a and Usnea species) from the northern coast area of California,
and of Gough, et al. (1988b) for Parmelia sulcata, an epiphyte common in western North Dakota, and
a variety of sol species from the Arctic (Puckett and Finegan, 1980) are also included.
This table needs to be interpreted with caution. Data for a few metals are poorly defined at
the low end because of limits in detectabfllty (less than symbols). Also, reported background
concentrations may vary regionally with different substrates as Is llustrated by the two data sets from
Gough, et al. (1988 a,b) (Table 1). In addition, the widespread occurrence of industrialization within
temperate latitudes may lead to relatively higher 'background' levels than in areas of the world remote
from industrialization. Interpretation of the elevated levels as concentrations toxic to, or tolerated by,
the species is also problematic for several reasons. First, the viability of the specimens analyzed was
rarely checked. Second, most of these values are for total metal contents and thus do not distinguish
between metals found intracellularly and those found extracellularly (as particuiates or at cell wall
exchange sites). Finally, there are few polluted sites where only one metal is present. Most
situations involve an array of metals and gaseous pollution and this potential toxicity needs to be
assessed within a mutttvariate context and with a regard for potential interactions.
KNOWN EFFECTS OF TRACE METALS TO LICHENS
There are relatively few examples where a trace metal was demonstratty toxic to lichens in
the field. One of the clearest examples, however, is for zinc toxicity, as documented by Nash (1972,
1975, 1989) surrounding a zinc smelter in Palmerton, PA. Zinc, cadmium, and sulfur dioxide were
present near the smelter In sufficient concentrations to be toxic to some species. Because saturation
soB extracts of zinc ions in concentrations toxic to at least some lichen species extended beyond the
zone of detectabfllty for sulfur dioxide, k was Inferred that zinc was the principal toxic agent. At a
control site 84 lichen species were found, but in the zone of extreme Impact only nine species
survived (Nash, 1972). Only two of the survivors, Micarea trisepta on rocks, and Scoliciosporum
chlorococcum on trees, were abundant In the polluted site and both of these were of greater
abundance in the polluted area than in the control area.
42
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TABLE 1. COMPARISON OF
BORDERUNE AND CLASS B
BACKGROUND AND ELEVATED CONCENTRATIONS ^g/g) OF CLASS A,
METALS (NIEBOER AND RICHARDSON 1980).
Element
Class A
Al
Ba
Be
Ca
Ce
Cs
Eu
K
La
U
Lu
Mg
Na
Nd
Sc
Sm
Sr
Background
Metals'
300-400
100-1800
1300-3000
10-110
60-100
2-400
2-400 x 102
12-130 x 108
27-75 x 102
0.18-0.89
0-700
500-5000
620-2700
1375-1660
0.09-0.68
0.09-0.55
100-1000
340-2600
540-1000
350-916
50-1000
120-560
159-233
0.09-.050
0.06-0.46
0.03-0.68
0-700
5.5-40
18-42
References
Nieboer et al., 1978
Gough et al., 1988a
Gough et al., 19885
Gough et al., 1988a
Gough et al., 1988b
Nieboer et at., 1978
Nieboer et al., 1978
Gough et al., I988a
Gough et al., 1988b
Gough et al., 1988a
Nieboer et al., 1978
Nieboer et al., 1978
Gough et al., 1988a
Puckett and Finegan, 1980
Gough et al., 1988a
Gough et al., 1988a
Nieboer et al., 1978
Gough et al., 1988a
Gough et al., 1988b
Puckett and Finegan, 1980
Nieboer et al., 1978
Gough et al., 1988a
Puckett and Finegan, 1980
Gough et al., 1968a
Gough et al., 1988a
Puckett and Finegan, 1980
Nieboer et al., 1978
Gough et al., 1988a
Gough et al., 1988b
Elevated
1300-1900
740-3700
900-4500
367-487 .
400-550 x 102
2.2-7.2
0.03-0.13
5000-9500
1.7-5.6
7.3-11.4
0.01-0.04
1000-12,000
1000-6000
0.26-1.22
0.17-0.70
References
Nieboer et al.. 1978
Olmez et al., 1985
Mueller et al., 1986
Nash and Sommerfeld, 1981
Nieboer et al., 1978
Olmez et al., 1985
Olmez et al., 1985
Nieboer et al., 1978
Olmez et al., 1985
Nash and Sommerfeld, 1981
Olmez et al., 1985
Nieboer et al., 1978
Nieboer et al., 1978
Olmez et al., 1985
Olmez et al., 1985
43
-------
Table I.Jcontd).
Element
Background
References
Elevated
References
Class A Metals (contd.)
Tb
Th
U
Y
Yb
Borderline
As
Cd
Co
Cr
Fe
Ga
Mn
Ni
0.5-1.0
0.06-0.37
1.3-3.8
Metals
0.60-1.5
0.26-0.44
1-30
0.05-0.3
0.12-0.72
0.51-0.68
0-10
0.34-13
5.6-11
1.5-2.3
50-1600
360-1900
1600-3900
50-3560
0.09-0.49
10-130
40-330
57-110
30-85
0-5
3.4-26
4-17
2.5-2.9
Beckett et al., 1982
Gough et al., I988a
Gough et al., 1988b
Gough et al., 1988b
Puckett and Finegan, 1980
Nieboer et al., 1978
Gough et al., 1988a
Gough et al., 1988a
Puckett and Finegan, 1980
Nieboer et al., 1978
Gough et al., 1988a
Gough et al.. 1988b
Puckett and Finegan, 1980
Nieboer et al., 1978
Gough et al., 1988a
Gough et al., I988b
Puckett and Finegan, 1980
Gough et al., 1988a
Nieboer et al.. 1978
Gough et al., 1988a
Gough et al., 1988b
Puckett and Finegan, 1980
Nieboer et al., 1978
Gough et al., 1988a
Gough et al., I988b
Puckett and Finegan, 1980
3.1-21
0.17-0.72
8.0-20.0
3.0-151.0
0.06-0.25
2.6-8.0
30-330
25-130
up to 45
400-90,000
up to 17,000
4875-14,400
0.2-2.2
300-5000
326-466
10-300
Up to 25
Olmez et al., 1985
Olmez et al., 1985
Beckett et al., 1982
BoOeau et al., 1982
Olmez et al., 1985
Olmez et al., 1985
Nieboer et al., 1978
Nieboer et al., 1978
Vestergaard et al., 1986
Nieboer et al., 1978
Vestergaard et al., 1986
Noeske et al., 1970
Olmez et al., 1985
Nieboer et al., 1978
Nash and Sommerfeld, 1981
Nieboer et al., 1978
Vestergaard et al., 1986
44
-------
Table 1. (contd).
Element
Background
References
Elevated
References
Borderline Metals
Sb
Sn
Ti
Zn
0.08-0.13 Puckett and Finegan, 1980
0.9-4.4 Gough et al., 1988a
6-150 Nieboer et al., 1978 150-3800
5.6-88 Gough et al., 1988a 100-240
12-19 Gough et al., 19885
48-68 Puckett and Finegan, 1980
0-10 Nieboer et al., 1978 10-300
0.21-6.3 Gough et al., 1988a 10-20
2.7-5.7 Gough et al., 1988b 10-58
1.0-2.0 Nygard and Harju, 1983
1.2-4.0 Puckett and Finegan, 1980
20-500 Nieboer et al., 1978 1000-25000
9.1-40 Gough et al., 1988a up to 2100
60-320 Gough et al., I988b
16-25 Puckett and Finegan, 1980
Nieboer et al., 1978
Nieboer et al., 1982
Nieboer et al., 1978
Schwartzman et al., 1987
Nygard and Harju, 1983
Nieboer et al., 1978
Vestergaard et al., 1986
Pass B Metals
Cu
Hg
Pb
1-50 Nieboer et al., 1978
1.3-10 Gough et al., I988a
12-120 Gough et al.. 1988b
6.2-8.5 Puckett and Finegan, 1980
0-1 Nieboer et al.. 1978
0.12-0.16 Gough et al.. 1988b
5-100 Nieboer et al.. 1978
3.6-20 Gough et al., 1988a
4.2-5.6 Puckett and Finegan, 1980
15-1100 Nieboer et al., 1978
1000-4900 Alstrup and Hansen, 1977
up to 160 Vestergaard et al., 1986
7.9-29 Lodenius, 1981
2-8 Bargagli et al., 1987
100-12,000 Nieboer et al., 1978
up to 950 Tomassini et al., 1976
up to 600 Schwartzman et al., 1987
up to 3000 Jones et al., 1982
1740-2750 Noeske et al., 1970
These three classes of metals fall into natural groups within the periodic table of chemical elements
and may be divided by a covalent index that reflects different ligand binding affinities. Class A metals
prefer O-N-S donors, whereas Class B prefer S-N-O donors and borderline metals fall in between.
45
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In the Sudbury nickel mining complex In Ontario. Canada, Scoliciosporum chlorococcum was the
only epiphytic lichen found on Populus balsamlfera. a widely planted tree in the zone where the original
forest was destroyed (LeBlanc et al., 1972). Both the Stereocaulon and Cladonia species on soOs and
Umbilicaria species on rocks can stll be found at distances from the smelters. Tomassiniet al. (1976)
measured Ni concentrations up to 220 to 310, Fe up to 17800 to 5200, and Cu up to 250 /;g/g.
Similarly, In the vicinity of a steel complex in England. Seaward (1973) reports concentrations fag/g) In
Pettigera rufescons as high as 90,000 Fe, 5000 Mn, 91 Cu, 127 Cr, 454 Pb, and 38 Ni. A Cladonia and
Co/eocau/on species had somewhat lower concentrations. In the dty of Leeds values up to 35.800 Fe.
349 Mn, 159 Cu. 97 Cr, 3124 Pb, and 183 Ni are reported for Lecanora muralis. Other studies are
Incorporated into Table 1 and could easBy be discussed further here. For example, the numerous
Investigations of the areas surrounding electrical power plants could be summarized, but the release of
metals from power plants is frequently an order of magnitude lower than that which occurs near smelters.
POTENTIAL EFFECTS
Metals may well become toxic If accumulated within the cytoplasm. Cytological effects are not
well documented in lichens, but generally it is known that binding centers in proteins and enzymes satisfy
the reactivity requirements of class A, class B, or borderline ions. Specificity of metals occupying these
binding centers is related to their size and geometry as well as ligand type. Denial (1977) therefore
divided mechanisms of metal-ion toxicity into three categories: (1) blocking of the essential biological
function groups of proteins and enzymes, (2) displacing the essential metal ion in proteins and enzymes,
or (3) modifying the active conformation of proteins and enzymes. There is good evidence that at least
some lichens are tolerant of higher internal concentrations of metal ions than other lichens (see
discussion above), but particular cellular mechanisms that allow such tolerance are not currently known.
Thus, It may be useful in future studies to determine if the tolerant species (ecotypes, etc.) possess some
specific cellular mechanism for preventing toxicity by one or more of the mechanisms discussed by
Ochia! (1977).
RESEARCH RECOMMENDATIONS
There Is a definite need to determine the degree to which toxics, including both trace metals and
synthetic organics. are distributed within the U.S. as well as around the globe. Therefore, It is timely that
major monitoring programs such as the U.S. EPA's Environmental Monitoring and Assessment Program
(EMAP) be initiated. Because there is evidence for long distance transport of toxics to the Arctic where
cooler atmospheric temperatures act as a huge distortion column for the organics in particular, It may
46
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well be appropriate for an initial concentration of studies to be planned there. In polar regions, lichens
should be the organism of choice to provide biological monitoring to supplement instrumented monitoring
stations to achieve greater local and regional detail on deposition patterns. A joint Canada-U.S. program,
perhaps developed cooperatively between Environment Canada and the U.S. EPA. Ideally, the program
should be expanded to include the Soviet Union, and this may be possible through the joint U.S.-U.S.S.R.
program on Air Pollution Effects on Vegetation.
Concurrent with the field studies, there should also be laboratory studies with lichens. There
should be studies for toxics that have not previously been well investigated, to: (1) determine lichen
uptake kinetics and retention characteristics, and (2) define dose-response relationships for toxics. In the
latter case, there is very little information, particularly for synthetic organics.
CONCLUSIONS
An important question raised concerning the need for an air toxics program is: "where are the
dead bodies?' The short answer is, decomposed and disappeared. In the case of lichens, we know that
major declines in the lichen flora of urban and industrial areas have occurred on all continents (Nash and
Wirth, 1988). In addition, regional level declines are known. Gaseous pollutants, such as sulfur dioxide
and hydrogen fluoride, are known to occur in sufficient concentration to be detrimental to many lichen.
Some of the less studied airborne contaminants may also occur in sufficiently high concentrations to be
toxic. One need only reflect that as recently as the early 1950s, ozone, the air pollutant currently
recognized as a regionally important phytotoxic agent, was not even identified as an air pollutant. To
determine which of the air toxics are really ecologically important will require both a monitoring program
and effects research.
47
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FATE AND EFFECTS OF PAHs IN THE TERRESTRIAL ENVIRONMENT:
AN OVERVIEW
Nelson T. Edwards
Environmental Sciences Division
Oak Ridge National Laboratory
Oak Ridge, TN
INTRODUCTION
A recent review of potycydic aromatic hydrocarbons (PAHs) in the terrestrial environment
(Edwards, 1983) pointed to the need for more information and a better understanding of factors
controlling the fate of PAHs. With ever-increasing energy demands by a growing world population,
the amount of PAHs released from energy-producing technologies becomes potentially greater. When
PAHs are released into the atmosphere in the vapor phase, most are adsorbed to particles (e.g., fly
ash). They are also found in solid and liquid wastes from various energy technologies. Polycyclic
aromatic hydrocarbons released to the atmosphere may be carried hundreds of kilometers from the
source before being deposited on vegetation and soil. The well-known PAHs such as benzo[a]pyrene
(B[a]P) are innocuous by themselves, but can be biologically activated by enzymes to form epoxides
that are carcinogenic and mutagenic (Levin et al., 1978). This report summarizes our knowledge
about PAHs In the terrestrial environment with emphasis on their fate and effects on vegetation.
SOURCES OF PAHs
Polycydic aromatic hydrocarbons are found throughout the world in water, air, soil, and the
biota. The ubiquitous nature of PAHs may be a consequence of synthesis in terrestrial vegetation
(Andelman and Suess, 1970). Suess (1976) and Shabad (1980) pointed to microbial synthesis, higher
plant synthesis, and volcanic activity as major contributors to the natural background levels of PAHs,
48
-------
but emphasized that quantities of PAHs formed by natural processes are very small in comparison to
those from anthropogenic sources. Blumer (1961) suggested that PAHs found in rural soil remote
from major highways and industries cannot be attributed totally to air pollution, but are instead
endogenous to soil. However, Lunde and Bjorseth (1977) demonstrated that PAHs are transported
over relatively long distances from industrial areas of England, France, and Scotland. Conversely,
Lyall et al. (1988) presented evidence of very llttie transport of PAHs from urban to rural areas In
Australia. Graf and Diehl (1966) grew plants In a PAH-free nutrient solution and concluded from the
results that plants can synthesize PAHs. However, Grimmer and Duvel (1970) found that plants
grown in filtered air contained no detectable PAHs, whereas their counterparts in unfiitered air did
contain PAHs.
The major anthropogenic sources of PAHs are those resulting from conversion of fossil fuels,
refuse burning, and agricultural burning. The origin of most PAHs is through high-temperature
pyrolysis of various naturally occurring organic materials (e.g., coal). Seuss (1976) estimated global
B[a]P emissions during 1968 to be about 4570 metric tons, mostly from fossil fuel burning (Table 1).
Estimated annual B[a]P emissions in the United States In 1968 ranged from 1000 metric tons (Guerin,
1978) to 1700 metric tons (NAS, 1972). Guerin (1978) estimated 640 to 780 metric tons of B[a]P
emissions in the United States in 1972 (Table 1). Nearly half of this total came from heating and
power, mostly from residential furnaces and fireplaces.
Coal-fired steam plants contributed less than 1 metric ton/y. Gasoline- and diesel-powered
vehicles contributed >2% of the total B[a]P emissions. However, this number can be misleading
because other PAHs (e.g. coronene) are found in higher concentrations in vehicular exhaust than
B(a]P. Lyall et al. (1988) sampled eight PAHs in the Latrobe Valley in Australia and found that the
motor vehicle was the main contributor to PAH contamination of the atmosphere and that domestic
heating added a significant contribution during cold weather. They found that coal-fired power plants
added very little to the atmospheric PAH concentrations. However, if used on a large scale, coal
conversion processes could have a significant effect on anthropogenic Input of PAHs to the
environment. Reports of high releases of potentially hazardous chemicals from coal gasification plants
support this hypothesis. Analysis of gas, liquid, tar, and solid effluents from coal-fueled gastfiers
showed PAHs to be of special concern (Qeland. 1981). For example, B[a]P ranked near the top of
the list as a potentially hazardous chemical in the gas and tar streams, with a relatively high
discharge rate in the liquid stream. Maximum discharge concentration of B[a]P in the gaseous
stream of one of the gasffiers was 5000 mg m*3, and the concentration in tar was 3500 mg/g. Total
B[a]P discharged was 120 mg/g of coal.
49
-------
TABLE 1. ESTIMATED ANNUAL EMISSIONS OF BEN20[a]PYRENE (B[a]P) IN THE UNITED
STATES AND GLOBALLY.
B[a]P Emissions
(metric tons/y)
Sources
U.S."
Global"
Heating and Power
Coal fired residential furnaces
Coal intermediate sized furnaces
Wood, home fireplaces
Coal, steam power plants
OH, residential heating
Gas. residential heating
Total
Industrial Processes
Coke production
Petroleum catalytic cracking
Total
Enclosed Incineration
Coal Refuse Burning
Forest and Agricultural Fires
Other Burning
Vehicle disposal
Other open burning
Total
Vehicles
Diesel powered
Gasoline powered
Rubber tire degradation
Trucks and buses
Automobiles
Total
Grand Total
270
6
23
2
2
304
0.05-153
6
6-153
3
281
10
5
9
14
10
10
21
640-780
c
2360
950
90
620
380
130
30
10
40
4570
* From Guerin (1978)
b From Suess (1976)
c No data available.
Not enough data have been collected to estimate total PAH emissions (other than B[a]P) from
anthropogenic sources. It is generally accepted, however, that B[a]P represents only a small
percentage of the total and that anthropogenic sources are far greater than natural sources.
50
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PAH CONCENTRATIONS IN AIR, SOIL, AND VEGETATION
Concentrations of PAHs In air vary greatly, both spatially and temporally. Sawicki et al.
(1960) examined the B[a]P content of air In 131 urban and nonurban areas of the United States. In
nine large cities the highest levels of 8[a]P occurred during winter months, and the lowest levels
occurred during summer. Concentrations of B[a]P In the air of nonurban areas ranged from 0.01 to
1.9 ng/m3 whereas concentrations in urban areas ranged from 0.1 to 61.0 ng/m9. Gordon (1976)
reported the annual geometric mean concentration of 15 PAHs in air samples collected from 13 areas
in Los Angeles County, CA, to be 10.9 ng/m3. Concentrations of B[a]P only accounted for 4.2% of
the total.
Relatively few studies have quantified PAH concentrations In sol and vegetation at various
distances from known sources. Typical concentrations of B[a]P in soils of the world ranged from
2100 to 1000 ng/g. A typical range for total PAHs is about 10 times the value for B[a]P alone. The
actual measured range of B[a]P concentrations in sofl, including data from very highly polluted areas
and from protected remote regions, is 0.4 ng/g (Shabad et al., 1971)) to 650,000 ng/g (Fritz, 1971).
Although PAH concentrations in the environment have increased, our knowledge of their
movement through terrestrial food chains has remained static. Results from past research on the
uptake, conversion, and concentration of PAHs by vegetation are conflicting. A book by Grimmer
(1983) on PAHs devoted only 4 of 250 pages to contamination of food, with no references to PAH
movement through terrestrial food chains.
Not enough data are available on a sufficient number of different PAHs (most research has
been with B[a]P only) to make reasonable predictions about their food chain transfer and
accumulation rates. Results from previous research do suggest, however, that PAHs can enter the
food chain by contamination of vegetation. Concentrations of 16 PAHs in garden vegetables collected
3 km downwind from a coal-fired power plant in Tennessee ranged from 8 ng of total PAHs/g dry
weight of lettuce roots to 107 ng/g of lettuce leaves (Table 2). The PAHs found in greatest
concentrations were fluoranthene, phenanthrene, and acenaphthylene. B[a]P averaged only 2% of the
total PAH contamination. Previously reported concentrations of B[a]P in vegetation ranged from 0.1
(Kolar et al., 1975) to 150 ng/g (Fritz, 1971), with typical concentrations of 1 to 10 ng/g. Much
greater PAH concentrations were found in vegetation growing adjacent to a tar waste pit located <2
km from a coal gasification plant hi Tennessee (Table 3). This site was expected to have potential for
both high inputs of PAHs from the atmosphere and high uptake via plant roots from the tar waste pit.
Total PAH concentrations in roots ranged from 534 ng/g In Carex sp. to 6401 ng/g in lambs quarters
51
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TABLE 2. PAH CONCENTRATIONS IN GARDEN VEGETABLES COLLECTED FROM 8 KM
SOUTHWEST OF A COAL-FIRED STEAM PLANT IN EASTERN TENNESSEE (SITE A) AND FROM
3 KM NORTHEAST OF THE SAME STEAM PLANT (SITE B)*.
Site A She B
PAH Radish Spinach Lettuce Lettuce Lettuce Lettuce
Leaves Leaves Leaves" Leaves Leaves" Roots"
Naphthylene
Acenaphthyiene
Acenapthene
Ruorene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz[a]anthracene
Chrysene
Benzo[a]fluoranthene
Benzo[k]fluoranthene
Benzo[a] pyrene
Dibenz[a,h]anthracene
Benzo[ghi]perylene
lndeno(1 ,2,3-cd)pyrene
TOTAL
.«
20
-
-
36
13
19
9
2
2
4
2
1
2
<1
110
—
12
-
2
9
-
13
8
1
-
2
1
1
2
1
52
—
11
-
1
4
-
28
3
<1
1
1
<1
<1
<1
-
53
— m
25 5
-
4 <1
6 <1
<1 <1
49 5
6 <1
11 <1
1 <1
2 <1
<1 <1
1 <1
1 <1
-
107 41
4
<1
<1
<1
<1
<1
2
<1
<1
-
<1
<1
<1
-
-
8
* All values are in nanograms per gram.
" Samples were rinsed with cold water.
c Concentrations were not detectable.
(Chenopodium album L), a perennial forb. Concentrations in leaves ranged from 14 ng/g in Johnson
grass (Sorghum halepense L) to 562 ng/g in Care* sp. The average concentrations were 2584 ng/g
In roots and 294 ng/g hi leaves. The B[a]P concentrations ranged from 0.2 to 13.5% of the total
PAH concentration, varying with both plant species and plant organs in no consistent patterns.
Polycydic aromatic hydrocarbons found In greatest concentrations in roots were anthracene,
acenaphthyiene, phenanthrene, and fluoranthene. Those found in greatest concentrations in leaves
were fluoranthene, acenaphthyiene, indeno (1,2,3-cd) pyrene, chrysene, and anthracene. Fluoranthene
and benzo[a]anthracene were found in relatively high concentrations (340 and 459 ng/g, respectively)
In stems of a sycamore tree (Platanus occidentalis L). Concentrations of PAHs In vegetation are
generally less than those in the sol. Bioconcentration factors (concentration in
vegetation/concentration in sofl) reported in the literature range from 0.002 to 0.33 for B[a]P.
52
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TABLE 3. PAH CONCENTRATIONS IN VEGETATION COLLECTED PROXIMALLY TO A COAL TAR
BURIAL TRENCH AND < 3 KM FROM A COAL GASIFICATION PLANT IN EASTERN TENNESSEE.*
Carexb
PAH Roots Leaves
Naphthylene
Acenapthene
Fluorene
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Benz[a]anthracene
Chrysene
Benzo[a]fluoranthene
Benzo[k]fluoranthene
Benzo[a]pyrene
Dibenz[a,h]anthracene
Benzo[ghi]perylene
lndeno(l,2,3-cd pyrene)
Total
c
142
-
56
.
118
84
13
67
9
8
10
18
31
40
534
.
89
29
-
61
45
31
11
44
27
5
13
21
28
81
562
J. Grass
Roots Leaves
.
193 1
44
34 <1
<1
-
94
64
735
100 <1
50 <1
111 <1
59 10
24 <1
<1
817 14
L Quarters
Roots
.
774
25
733
1972
675
530
75
-
216
37
92
168
143
226
6401
Ragweed
Leaves
.
15
-
41
-
98
46
22
23
20
8
17
16
13
9
305
Sycamore
Stems
12
67
-
10
1
340
39
459
-
1
1
2
2
-
3
960
" All values are in nanograms per gram.
bCarex=Carex sp.; J. Grass=Johnson grass; L quarters=Lambs quarters; Ragweed =Ambrosia sp.
c Concentrations were not detectable.
Only one paper reported PAH data (other than just B[a]P data) for both plants and soil from
the same location. Wang and Meresz (1981) analyzed garden vegetables and soil for 17 PAHs,
Including B[a]P, and found most of the PAH contamination in the peels. Their vegetation/soil
concentration ratios ranged from 0.0001 to 0.085 for B[a]P and 0.001 to 0.183 for total PAHs. The
amounts and kinds of PAHs ingested by humans and other animals from vegetation are partially
dependent on whether PAHs are absorbed versus adsorbed, and how easily they are rinsed off with
water. Kveseth et al. (1981) suggested that lower molecular weight PAHs are adsorbed on leaves,
whereas higher molecular weight particulated compounds are washed off by rain. Kolar et al. (1975)
found that washing vegetables removed a maximum of 25% of PAH contamination and generally less.
However, the data presented in Table 2 demonstrate that 87% of the total PAH mass over a wide
range of molecular weights is removed from lettuce by rinsing the leaves in cold water.
53
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PAH UPTAKE, TRANSLOCATION, AND METABOLISM IN VEGETATION
Relatively few experiments have been conducted that address questions relating to the uptake
and translocatlon of PAHs in vegetation. Graf and Nowak (1966) demonstrated growth stimulation of
tobacco, rye, and radishes by a number of PAHs. including B[a]P, and concluded that the
compounds were assimilated through the roots. Gunther et al. (1967) found no translocation of
several PAHs from orange rind (point of application) to other plant parts, and Harms (1975) reported
negligible translocation of 1t>B[a]P from wheat roots to shoots. Edwards et al. (1982) reported both
the uptake of 'ID-anthracene from nutrient solution Into Gtycine max Men. (soybean) roots and
translocation to leaves, and the negligible uptake from air Into leaves and translocation to roots.
However, Ellwardt (1977) concluded, from field experiments with fresh compost containing a number
of PAHs and with several agricultural crops, that little or no uptake by plant roots occurred.
The rate (amount/unit time) of PAH uptake by plants is dependent on a number of factors,
including PAH concentration and plant species. Deubert et al. (1979) found that Zea mays L (corn)
and Triticum aestivum L (wheat) seeds absorbed B[a]P in proportion to B[a]P concentrations in
water used for soaking the seeds. Edwards et al. (1982) found that 11C-anthracene uptake by
soybeans from a nutrient solution was directly proportional to anthracene concentration in the
solution. Several studies (Shabad et al.. 1971; Shabad and Cohan, 1972; Unne and Martens, 1978)
have demonstrated differences in assimilation rates with different plant species.
Other factors that affect PAH uptake rates by plants include the nature of the substrate in
which the plant is growing, PAH solubility, PAH phase (vapor or paniculate), and molecular weight.
D6rr (1970) found no uptake of B[a]P by Triticum spp. (wheat) and Seca/e cerea/e L (rye) from
nutrient solution or soil when B[a]P was applied in Insoluble font) (i.e., not dissolved in oil or some
other solvent). However, uptake and translocation did occur when B[a]P was dissolved in oil before
applying to the substrate. Recent research (NT. Edwards, unpublished data) demonstrated good
correlation between PAH uptake rates by bean plants from nutrient solution and the relative aqueous
solubilities of anthracene (ANTH), B[a]P, and benzo(a]anthracene (B[a]A). Muller (1976) reported
greater uptake of 1t>B[a]P by vegetable plants growing hi sand culture when the 1t>B[a]P was
applied dissolved In benzene rather than when the 't>B[a]P was dissolved in plant oil or detergent.
MOIIer also reported greater uptake of B[a]P by plants growing in sand than those growing in soli and
compost. D6rr (1970) reported greater uptake of B[a]P by rye from nutrient solution than from soD,
but reported no effect of different sol types on uptake rates. Edwards et al. (1982) found that 87% of
the total ANTH dosage was assimilated by soybean plants from a nutrient solution over a four-day
period, but only 7% was assimilated over the same time period from son containing the same ANTH
dosage. In a related five-day experiment, twice as much ANTH and/or Its metabolites was assimilated
54
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by bush bean plants from sol containing ANTH in solution as from sol containing ANTH adsorbed to
fly ash (NT. Edwards, unpublished data). Nearly 90% of the ANTH assimlated from the fly ash soil
was retained in the roots, whereas only 50% of that assimlated from sol solution was retained in the
roots, with the remainder translocated to stems and leaves (Figure 1).
Although relatively little is known about PAH uptake and translocation in terrestrial plants, even
less is known about the chemical fate of these compounds within plants. Dorr (1970) found a decline
in B[a]P concentrations in rye plants after 30 days of growth, following a period (20 days) of
increasing concentrations through uptake from nutrient solution and sol. The dedine in B[a]P
concentration was attributed to degradation or chemical changes of the B[a]P within the plants.
Phaseolus vulgaris L (bush bean) plants grown for 30 days In nutrient solution containing 11C-labeled
ANTH, B[a]A, or B[a]P assimilated, respectively, 64. 66, and 41% of the total 1t dosage
(NT. Edwards, unpublished data). Only 6, 1, and 0.5% of the total '1C dosage associated with ANTH,
B[a]A, and B[a]P, respectively, was translocated to the stems and leaves. Of the 't; extracted from
the plants, these compounds were in the form of polar metabolite compounds of ANTH, B[a]A, or
B[a]P: more than 95% of that from the leaves and 82 to 97% of that from the stems (Figure 2).
Extremely small fractions of the extractable '1C in the stems and leaves were associated with the
parent PAH compounds. In the roots, however. 26, 27, and 65% of the '1C was associated with
B[a]P, ANTH, and B[a]A, respectively, Indicating that B[a]A was the most resistant of the three PAHs
to metabolism by the plant.
Thus, although conclusions vary widely between experiments, and much more research is
needed, a consensus would suggest that PAHs can enter the food chain by contamination of
vegetation. The seriousness of this problem depends in part on molecular species and contamination
levels, and in part on degradation rates and degradation products in air, son, and vegetation.
PAH DEGRADATION
Degradation usually Implies a reduction in chemical complexity of a compound and is often
thought of as a process that wDI change a hazardous compound to an innocuous one. However,
some nonhazardous compounds may be altered chemically to form hazardous compounds. For
example, B[a]P is innocuous but can be biologically activated by enzymes to form epoxkJes that are
carcinogenic and mutagenic. In the discussion that follows, the term degradation means simply a
chemical modification. Besides metabolism by green plants discussed above, the major ways that
PAHs are degraded in the terrestrial environment are by photochemical oxidation and metabolism by
microorganisms.
55
-------
100
75
I
"55
o
01
50
25
Soil Only
Fly Ash
FIGURE 1. A comparison of the distribution of "C-ANTH or Its metabolites in Phased us vulaaris
after 5 days growth in sofl containing fly ash that had been previously vapor phase coated with
14C-ANTH. or In soB containing "C-ANTH In solution.
56
-------
100
2 80
60
40
Ł
3
20
^100
2 80
UJ -
S 20
^100
2 80
«
2 60
20
0
Polar Metabolites
Nonpolar Metabolites
Anthracene
Benz[a]pyrene
Benz[a]anthracene
Roots
Stems Leaves
FIGURE 2. A comparison of the distribution of 'tarbon labeled PAHs and metabolites in Phased us
vulparis L after 30 days growth in nutrient solution containing ANTH, B[a]A, or B[a]P.
57
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Photooxidation may be one of the most important processes In the removal of PAHs from the
atmosphere (MAS, 1972). Yet some of the reaction products may also be carcinogenic, making It
risky to evaluate carclnogenicity solely on PAH content (Fox and Olive, 1979). One of the more
common photooxidation reactions of PAHs is the formation of endoperoxides that ultimately undergo
a series of reactions to form quinones (MAS, 1972). Katz et al. (1979) observed that B[a]P is
photooxidized to B[a]P quinones that are direct-acting mutagens. They found the reactions to occur
faster when B[a]P was irradiated in the presence of ozone. Evidence for faster PAH degradation in
the presence of ozone is supported by LyaJI et al. (1988) and by Blokzijl and Guichertt (1972) who
reported greater photooxidation of PAHs in summer than in winter, presumably as a result of higher
temperatures, but possibly because of the combined effects of higher temperatures and elevated
tropospheric ozone concentrations in the summer months. Evidence exists that discredits the widely
held belief that paniculate association of PAHs wll promote their photooxidation. For example, B[a]P,
pyrene, and ANTH will photooxkJize efficiently In liquid solution, but are highly resistant to photo-
oxidation when adsorbed on fly ash (Korfmacher et al., 1980).
Degradation of PAHs by microorganisms has been demonstrated in a number of
Investigations. Groenewegen and Stolp (1981) isolated microorganisms that can use naphthalene,
anthracene, and phenanthrene as their sole carbon source. However, they could show degradation of
some of the PAHs that are less water-soluble (e.g.. B[a]A and B[a]P) only when the PAHs were mixed
with soil, water, and a substance to stimulate growth of oxygenase-active organisms. Shabad et al.
(1971) discussed several experiments that demonstrated bacterial degradation of B[a]P in soil. They
found that the capacity of bacteria to degrade B[a]P increased with B[a]P content in the soil and that
the microflora of son contaminated with B[a]P were more active in metabolizing B[a]P than those in
'dean' soil. Biochemical pathways for the degradation of a number of PAHs by soil microorganisms
have been proposed (Evans et al., 1965; Gibson et al., 1975).
EFFECTS OF PAHs ON PLANTS
The limited amount of research that has been performed on the effects of PAHs on plants
indicates that PAHs may act as regulators of plant growth and morphogenesis. Graf and Nowak
(1966) observed that B[a]P and other PAHs stimulated the growth of algae and higher plants in
solution culture, and suggested that PAHs synthesized by plants may act as plant growth hormones.
Wettig et al. (1976) germinated several species of vegetable and grain seeds in hydroponic
systems, with and without an added carbon source, and found that the seeds with the added carbon
increased in B[a]P content. Bomeff et al. (1968a; 1968b) demonstrated that algae can synthesize
58
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some PAHs from 'ID-acetate added to their growth medium. Reinert (1952) demonstrated
stimulatory or inhibitory effects of B[a]P on the growth of fern prothallia depending on concentrations
used. Forrest et al. (1989) observed earlier morphological transitions and accelerated cell proliferation
in B[a]P-treated fern prothallia than in untreated controls. However, B[a]P concentrations greater than
0.32 /jg/mL inhibited cell proliferation in the fern prothallia, and concentrations of 10 /sg/mL or higher
decreased spore germination and plant survival. Benzo[a]anthracene stimulated tobacco callus tissue
cell differentiation into roots and shoots (Kochhar et al., 1970). Simlarly,
7-12-dimethylbenzo[a]anthracene stimulated callus tissue of Hawortha variegate to differentiate into
roots and shoots (Majumdar and Newton, 1972).
CONCLUSIONS
A number of conclusions about the fate of PAHs in the terrestrial environment can be drawn
from past research. These conclusions can be generally stated as follows: (1) vegetation can
assimilate PAHs from soU through the root systems, but plant/soil bioaccumulation factors are
generally < 1; (2) assimilation rates vary with substrate, plant species, phase (whether adsorbed to
particles or in solution), concentration, and chemical properties (e.g., solubility in water) of the
individual PAHs; (3) PAHs are scavenged from the air by plant leaves and remain primarily on the
surface, but are available for wash off to the soB; and (4) PAHs are chemically transformed by soil
microorganisms, by vegetation, and by photooxidation especially at warm temperatures and high
ozone concentrations. Reported results on the fate of PAHs are incomplete (most studies have
concentrated on B[a]P only), and there is a need for research in this area. Studies should include
not only PAHs that have been identified as having carcinogenic potential (such as B[a]P) but also all
PAHs frequently found in the environment. Perhaps the most pressing questions concerning the fate
of PAHs in the terrestrial environment are these: (1) what are the products of PAH metabolism in air,
soil, and vegetation, and (2) what is the fete of those metabolic products? Studies of effects Indicate
that PAHs may act as regulators of plant growth and morphogenesis. However, given the paucity of
studies dealing with the effects of PAHs on vegetation (most studies have been conducted on simple
algae or at the cellular and tissue culture level In higher plants), new initiatives in this area of research
are needed and should be given high priority. Such studies should examine the effects of PAHs on
the growth and physiology of genetically simOar higher plants established at sites known to be
contaminated and at selected control sites in order to extrapolate results to the community and
ecosystem level. Mechanistic studies could be conducted under more controlled conditions.
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EFFECTS OF ORGANIC CHEMICALS IN THE ATMOSPHERE ON TERRESTRIAL PLANTS
Jeffrey R. Foster
Holcomb Research Institute
Butler University
Indianapolis. IN
INTRODUCTION
Thousands of organic compounds are released Into the atmosphere each year by human
activity. Of these, the U.S. Environmental Protection Agency (EPA) has classified 328 as Priority
Pollutants, based on scientific evidence of toxlcity In humans and other animals. However, only two,
benzene and vinyl chloride, have National Ambient Air Quality Standards. Until recently, our
knowledge of environmental releases of these substances was sketchy. In 1986, Congress passed the
Emergency Planning and Community Right-to-Know Act, requiring manufacturers of Priority Pollutants
to report the amounts of each that they released, deliberately or accidentally, into air. water, and soil.
The results of the first round of the Toxic Release Inventory (TRI) revealed that 865,000 metric tons of
toxic organic compounds were released to the atmosphere in 1987 (US EPA. 1989). Small
manufacturers (<3,400 kg/year) and end users (e.g., farmers applying pesticides) were not required
to report, and no attempt was made to incorporate emissions from area sources (dry cleaners, motor
vehicle exhausts, etc.). Thus, total emissions were probably several times larger than the TRI data
Indicate.
Following emission, organics are dispersed widely In the environment Long-distance
transport occurs primarily In runoff, groundwater, and the atmosphere (Cohen, 1986). The latter
pathway has spread organics to the most remote regions of the globe (Schroeder and Lane, 1988).
Chlorinated pesticides such as DDT, lindane, dieldrin, and hexachlorobenzene (HCB) are the best-
known examples, having been detected in inorganic media and organisms as distant from emissions
60
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as the Arctic and Antarctic (Alias and Giam, 1961; Tanabe et al., 1983; Gregor and Gummer, 1989).
Long-distance dispersal of nonpesticide organics, including potycydic aromatic hydrocarbons (PAHs),
polychlorinated biphenyls (PCBs), chlorophenols, and alkanes, has also been demonstrated (Lunde
and Bjorseth, 1977; Hett ef at., 1984; PaasMrta et al., 1985; Wickstrom and Tolonen, 1987; Gregor
and Gummer, 1989).
What are the consequences of this global organic contaminant load on the structure and
function of organisms and natural ecosystems?
PROBLEM STATEMENT
A vast amount of literature has been accumulated on the ecotoxicology of organics in natural
environments. However, it is quite dear that the driving force for research has been the concern over
their accumulation in crops for human or livestock consumption, primarily as residues following
pesticide applications, and in accumulation via Womagnlfication up aquatic and terrestrial food chains
to organisms that are consumed by humans (e.g., fish), or organisms that bring us aesthetic pleasure
(e.g., birds). There is, comparatively, a severe dearth of Information concerning the fate and effects
of toxic organics in noncrop, terrestrial, vascular plants.
The only comprehensive assessment of the literature on organic chemicals and terrestrial
plants is the EPA-funded PHYTOTOX computer database (Fletcher et al., 1988). This database
contains 77,825 dose-response (effects) records gleaned from 9,700 bibliographic references published
between 1926 and 1984. Of the 1,569 species represented in PHYTOTOX, only 417 are native
species growing in natural habitats. The latter value Increases to 557 If old-field-succession species
are included. Together, wild-grown and old-field species constitute 4,524 records, or 5.8% of the total
database. Thus, the existing literature is heavily oriented toward agronomic species. Pesticides
constitute 21% of all PHYTOTOX effects records. In fact, of the 20 most often cited chemicals, only
two (both hormones) are not pesticides. And, only 4.7% of records are for native plants growing in
natural environments.
From the lexicological standpoint alone, It is evident that Insufficient data exist for EPA to
undertake risk analyses for the effects of organics on natural terrestrial vegetation. In addition, little is
known concerning deposition rates of organics to terrestrial ecosystems, rates of plant uptake, and
subsequent translocation and metabolism.
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The objectives of this report are to: (1) discuss mechanisms of atmospheric deposition of
organic substances to terrestrial vegetation; (2) describe probable pathways of terrestrial vascular
plant uptake of organics; (3) summarize existing literature on the effects of organics on terrestrial
vascular plant growth and physiological processes; (4) speculate on potential, as yet unstudied,
effects; and (5) recommend areas for future research.
I rely heaviy on the herbicide literature, which greatly exceeds that for all other organics
combined, but make no daim to provide any but the most cursory overview. I wll consider all kinds
of airborne organics, with the exceptions of peroxyacetyl nitrate (PAN), a known gaseous oxidant and
phytotoxin that has received much attention during the last 20 years (see Mudd, 1975, for a review),
and the chlorofluorocarbons. Also, I wll not review the effects of water-borne organics on aquatic
macrophytes.
CONCEPTUAL APPROACH
Figure 1 shows a model of the various fluxes of airborne organic substances onto, into, and
through terrestrial vascular plants. Fluxes specific to the canopy are shown in detail in Figure 2. In
the discussion that follows, I wDI deal with most of these fluxes, with the exception of wtthin-soil
fluxes, which are beyond the scope of this paper. Following that, I wDI describe known effects of
organics on terrestrial plants, and finish with recommendations for future research.
ATMOSPHERIC DEPOSITION
Atmospheric Occurrence
Organics enter the atmosphere primarily in gaseous or paniculate form. Initial mixing in the
air leads to partitioning between the vapor, solid, and liquid phases (Schroeder and Lane, 1988).
Phase distributions are determined by inherent chemical properties, including vapor pressures, Henry's
Law constants (H, equilibrium ratio of the concentration in air to concentration In water), diffusion
coefficients in air and water, and equilibrium constants (pK.; for weak acids) (Tucker and Preston,
1984; Mackay et al., 1986; Bidleman, 1988; Ugocki and Pankow, 1989). Organics enter precipitation
by dissolving directly from the vapor phase into water droplets and by being physically scavenged as
aerosols by falling droplets. The "washout ratio* (W; mass of organics per unit volume of rain divided
by mass of organics per unit volume of air) describes the net effect of these processes (Bidleman,
1988).
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wet & dry
deposition
trans) ocati on
detoxification
chemical
decomposition
biological
degradation
root
exudation
root uptake
leaching
water table
FIGURE 1. Conceptual model of fluxes of airborne organics to and within terrestrial ecosystems.
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wet
deposition
•tomatal
absorption
cuticle
absorption/
adsorption
surface
moisture
absorption
washoff
guttation
volatilization
photodecomposition
FIGURE 2. Same as Fig. 1, but enlarged to show fluxes associated whh leaves.
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Compared to the major air pollutants (ozone, sulfur dioxide, nitrogen dioxide, carbon
monoxide), air concentrations of organics are not well-characterized. A great variety have been
detected whenever and wherever they have been looked for. Urban air contains hundreds of volatile
organic compounds (VOCs) from industrial emissions, motor vehicle exhausts, and other
anthropogenic sources (Atlas and Giam, 1988; Bruckmann et al., 1988; Shah and Singh, 1988;
Edgerton et al., 1989). Some VOCs are secondary products of the photochemical reactions that
produce smog (e.g., nltrophenols [Nojima and Kanno. 1977]). These VOCs are transported downwind
to rural regions (Rice et at., 1986; Rlppen et al., 1987; Shah and Singh, 1988; Schroeder and Lane,
1988). Reverse transport also occurs; although having highest concentrations In agricultural areas,
chlorinated pesticides are regularly found in urban air (Abbott et al., 1966; Stanley et al., 1971;
Kaushik et al., 1987; Bruckmann et al., 1988). Natural vegetation, especially trees, puts out substantial
quantities of VOCs, primarily monoterpenes, which mix with anthropogenic VOCs (Roberts et al., 1985;
Trainer et al., 1987; Kreuzig et al., 1988; Petersson, 1988). Table 1 summarizes the range of
concentrations of selected organics measured in air and rain.
TABLE 1. CONCENTRATIONS OF SOME PESTICIDES, PCBS, PAHS AND OTHER VOLATILE
ORGANIC COMPOUNDS IN AIR AND WATER. UNITED STATES'
Compound Air (ng/m^ Rain (ng/L)
Hexachlorobenzene
a-/THCH
DDT/DDE
Chlordane
Dieldrin
Toxaphene
PCBs
Pthalate esters
Benzo(a)pyrene
Formaldehyde
Acetaldehyde
Benzene
Phenol
Benzo(a)pyrene
0.13-0.29
0.30-5.4
0.02-0.33
0.04-1.3
0.02-0.08
0.6-13.1
0.06-9.3
1.0-290
1.3-500
2375-12250"
0-9900b
1950-10700"
. 39-430b
0.01-61
0.40-1.0
6.4-37
0.50-4.3
trace-2.3
0.8-2.0
7.31-59
0.10-200
>200
2.2-7.3
<0.01-1.1
<0.04-1.1
c
< 10-14000
~
'Sources: Pearson (1982), Nriagu and Simmons (1984). Giam et al. (1984), Chapman et al. (1986),
Atlas and Giam (1988), Shah and Singh (1988), Leuenberger et al. (1988), Dulnker and Bouchertall
(1989). Sawicki et al. (1960)
" lower/upper quartiles
c no data available
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Wet Deposition
The flux of organics to terrestrial ecosystems In precipitation Is simply the volume-weighted
product of the concentration In solution and precipitation volume, summed over the time period of
interest. A growing literature reports concentrations of organic compounds in rain. Most regularly
and widely detected are organic acids, especially aldehydes such as formate and acetate (Chapman
et al., 1986 and references therein; Qaffney et al., 1967). Chlorinated pesticides, PCBs, phenols, and
n-alkanes have all been reported In rain (Eteenretch et al.. 1961; Farmer and Wade, 1986; Agarwal et
al., 1987; Gaffney et al., 1987; Jones et al., 1989; Leuenberger et al., 1988; Alias and Giam, 1988).
Few estimates of wet deposition flux for organics exist Farmer and Wade (1986) observed
22 to 670 pg/m2/day for C12- CK hydrocarbon fluxes at Norfolk, VA. An estimate for bulk
precipitation, which incorporates some aerosols, was 40 and 20 pg/m2/y for PCBs and lindane,
respectively, at Paris, France (Chevreuil et al., 1989). Precipitation falling onto the Great Lakes has
been analyzed for PCBs and pesticides for some years. Deposition rates are 0.6 to 9.6 pg/m2/y for
chlorinated hydrocarbons, 424 yg/m8/y for PAHs, and 120 pg/m*/y for PCBs (Nriagu and Simmons,
1984).
Dry Deposition
Dry deposition of atmospheric substances onto natural vegetation is described by the
equation
where F = flux density, C. = atmospheric concentration at a reference height above the vegetation,
and Vd = deposition velocity. Deposition velocity has four components:
V< - V(r. + fb + r. + re)
where r. = bulk aerodynamic resistance, rb = leaf or sofi boundary layer resistance, r, = stomatal
resistance, and re = chemical resistance.
Aerodynamic resistance represents the turbulent downward movement of parcels of air;
carrying entrained gases and particles. It is affected by wind speed, the surface area, and vertical
distribution of canopy components (foliage, branches, etc.). Boundary layer resistances are related to
wind speed and wind direction, and to collecting surface size and geometry. Stomatal resistance is
66
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directly proportional to stomatal aperture, which is under physiological control by plants. Chemical
resistance comes into play when the concentration of the compound at the absorbing surface is not
zero (i.e.. It is not quantitatively destroyed or sequestered upon deposition). In this case, the
concentration gradient driving flux wRI be less than C..
Deposition of gaseous organics occurs primarly to dry sol and vegetation surfaces and is
controlled largely by solubility in the waxy cuticles of plant surfaces and vapor-sol particle partitioning.
Absorption of gaseous organics into water is important to wet sols, on plant surfaces following rain or
during fog and dew events, In permanent wetlands, and In riparian areas during flooding. Deposition
of organics in aerosol form is achieved by the same mechanisms as for aerosols in general:
sedimentation, diffusion, and impaction. Sedimentation is most important for large aerosols, diffusion
for small aerosols, and impaction for intermediate-sized aerosols (Garland et al.. 1988).
Dry deposition of organics has been measured directly by accumulation on surrogate
surfaces, including silicon oil-coated nylon screens (Sodergren, 1972); glass plates sprayed with
mineral oil or water (McClure and Lagrange, 1977; Murphy, 1984); stainless steel funnels and sinks
(Farmer and Wade, 1986; Chevreuil et al., 1989); and pans filled with water, ethyfene glycol-water, or
glycerin-water (Bidleman and Christensen, 1979; Christensen et al., 1979). Observed fluxes have been
in the range 10''to 102^g/m2/day for PCBs and various pesticides, yielding Vdvalues of less than
0.01 to more than 5 cm/s. Confidence in these fluxes is not high because the efficiency collection of
surrogate surfaces may vary substantially from that of natural surfaces. Bidleman and Christensen
(1979) found that Vd for PCBs, chlordane, and DDT varied geographically and by one to two orders of
magnitude from day-to-day. In addition, aerosol and gaseous deposition cannot be distinguished, and
volatilization of adsorbed gases is not accounted for. Deposition velocities measured by accumulation
have been used to infer dry deposition to the Great Lakes from known concentrations of organics in
air (Eisenrelch et al., 1981).
Fogs. Clouds, and Dew
Cloud water and fog water contain higher concentrations of organic acids and much higher
concentrations of pesticides than does rain (Glotfetty et al., 1987; Igawa et al., 1989; Munger et al.,
1989; Schaefer et al., 1989). Organic adds have also been reported from dew (Mulawa et al., 1986).
Interception of cloud water and of advecttve (seacoast) fog water is modeled as a dry
deposition process, with cloud and fog droplets acting like large aerosols (Lovett, 1984). Deposition
of radiation (valley) fog water is similar to precipitation (i.e.. gravitational settling of water droplets).
Dews form by condensation of water vapor onto surfaces whose temperature drops to the dewpoint;
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deposition rates can be predicted using the Penman-Montelth equation. Organics enter dews by
gaseous absorption and aerosol sedimentation. There appear to be no deposition estimates for
organics in fogs, clouds, and dews.
Total Atmospheric Deposition
Accumulation rates of persistent (slow to degrade) organics in sediments yield an estimate of
net atmospheric deposition (i.e.. not accounting for volatilization). For example, Heit et at. (1981,
1988) and Gschwend and Hites (1981) measured PAHs in lacustrine and marine sediment cores in the
northeastern United States. Fluxes in recent decades ranged from 36 to 4,870 /sg/m2/y. with greater
fluxes near urban areas. Jones et al., (1989) measured PAHs in archived, dry soil samples collected
since 1843 in rural England. Mean deposition rates (1880-1980) were 4,560 pg/m2/y. Wlckstrom and
Tolonen (1987) found that PAH fluxes into sediments of small Finnish lakes had increased
substantially since the early nineteenth century.
Swackhamer and Armstrong (1986) estimated net atmospheric deposition of 2 ^g/m2/y of
PCBs in recent decades to several small seepage lakes in Wisconsin. Rapaport and Eisenreich (1988)
measured PCBs, HCB, DDT, and toxaphene accumulation in peat bog sediments. Maximum
deposition rates in the eastern United States occurred between 1960 and 1978, with reduced rates
since 1980.
The only estimate that could be found for recent atmospheric deposition of an organic to a
terrestrial ecosystem was that of Matzner (1984) for German beech and spruce forests. Assuming no
root uptake or exchange between surface deposits and canopy tissues, he estimated total PAH fluxes
of 0.25 to 0.69
A regional application of mass balance was applied to the Great Lakes by the International
Joint Commission (summarized in Elder et al., 1988). Measurements of PCBs and pesticide
concentrations in dated sediments yielded values for long-term inputs. When known contributions of
tributaries, runoff, and point sources were subtracted from these values, and outflows added, the
difference was ascribable to atmospheric deposition. The atmosphere accounts for 60% of PCB
inputs and 78% of benzo[a]pyrene inputs to the Great Lakes.
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FLUXES OUTSIDE OF PLANTS
Fates In Soil
Once deposited, the disappearance of organics in sofls usually follows first-order kinetics,
provided no additional input occurs, and plant uptake is either constant or absent (Ryan et al., 1988).
Factors contributing to the disappearance of organics include volatilization, photodecomposition,
adsorption to sol particles, chemical and biological (microbial) degradation, unsaturated and saturated
flow, and leaching to deeper sol layers (Figure 1). Consideration of these processes (except
volatilization) is beyond the scope of this paper. However, modeling of disappearance rates is useful
In determining the likelihood of organics to accumulate to concentrations at which plant root uptake
will become significant (Ryan et al., 1988).
Volatilization
In general, volatilization rates of organics from plant canopies increase with increasing vapor
pressures and decreasing Henry's law constants (H). Foliar volatilization rates do not fit simple first-
order kinetics because of diurnal variability in environmental conditions (solar radiation, wind speed,
etc.), and plant uptake (Willis et al., 1986). There is no simple relationship between volatilization rates
of organics from soils and either vapor pressure or H values; such factors as adsorption coefficients
to soil particles and soil moisture content are also important (Kilzer et al., 1979; Mackay et al., 1986;
Spencer et al., 1988). Measured volatilization losses of herbicides in agricultural fields show great
variability; 4.5% per year for dieldrin, but less than 1% per year for carbofuran (Caro et al., 1976);
80% per year for PCBs (Moza et al., 1979); 8, 24, and 34% over a six-day period for toxaphene,
parathion, and fenvalerate, respectively (Willis et al., 1986); less than 1% over seven days for didofop
(Smith et al., 1986); and 21% over five days for 2,4-D (Graver et al., 1985).
A substantial proportion of herbicides volatilized from soils can be redeposited on crop
canopies, especially DDT, PCBs, and the nltrophenol dinoseb (Parker, 1966; Prendeville, 1968; Bead
and Nash, 1971; Kaufman, 1976; Moza et al., 1979; Fries and Marrow, 1981; Bacci and Gaggl, 1985).
Topp et al. (1986) found that the proportion of sol-applied organic chemical uptake by barley foliage
increased as the volatilized fraction of applied chemical increased.
Wash-Off and Resuspenslon
Wash-off occurs when rain physically dislodges particles or carries away dissolved gases and
participates from plant surfaces; these are then deposited to the sol in throughfall (Figures 1 and 2).
The highest proportion of wash-off to amount applied is for pesticides with high solubility in water, for
69
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those applied in particulate form, and when heavy or repeated rains occur shortly after application.
Time-course studies of individual rain events reveal that wash-off is rapid at first, then diminishes
asymptotically toward zero. Wash-off increases with Increasing rain amount, up to some saturating
intensity (Isensee and Jones, 1971; Steffens and Weineke, 1975; Cohen and Steinmetz, 1986; Willis et
al.. 1986).
Resuspension is the reentrainment of particles deposited on plant surfaces in moving air
(Figure 2). Particle adhesion and windspeed are the primary factors influencing this process
(Nicholson, 1988). SoB particles may be entrained In the atmosphere (wind erosion) or ejected into
the air by raindrop splash (Chamberlain, 1975). from whence they may be redeposfted on plant
surfaces (Finder and McLeod, 1988).
FLUXES INVOLVING PLANTS
Organics must penetrate plant tissues before they can exert physiological effects. Thus,
aerosol deposition and nonstomatal deposition of gases are appropriately considered as separate
processes from plant uptake. In the case of stomatal absorption, deposition and uptake are
equivalent because the sink Is the mesophyll.
Uptake rates are determined by external concentration, concentration in plant tissues, and
resistances to uptake. For shoots, the most important physical barrier is the cutide. For roots, the
cell membranes of the endodermis are a significant barrier. Following uptake, if the substance is not
translocated or metabolized to other forms, It may accumulate in tissues, reducing the concentration
gradient across the plant surface, and thus reducing flux. Translocation to other parts of the plant, or
metabolic conversion to other compounds, wOl lower internal concentration near the sites of uptake,
maintaining the concentration gradient.
Foliar Uptake of Vaoor-Phase Oroanlcs
Radioisotope studies with carbon-14 (14C) labeled organics have demonstrated that in most
plant species, uptake of PCBs Is primarily by leaf vapor absorption, even when the PCBs are applied
to the sol (Nash and BeaJI. 1970; Iwata and Gunther, 1976; Weber and Mrozek. 1979; Bacci and
Gaggi. 1985). Nash and Beall (1970) and BeaJI and Nash (1971) found the same to be true of
soybean plants exposed to sol-applied DDT, but sol volatilization was less important for diekJrin,
aJdrin, and heptacNor. PCBs accumulate In both herbaceous and woody species in proportion to
PCB concentration In the surrounding atmosphere (Buckley, 1982). Soybeans absorbed "C-
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anthracene vapors from air (Edwards at al., 1962). Accumulation of PCBs, DDT, penta-chlorophenol
(POP), nltrophenols, dtoxins, furans. and »hexachlorocyclohexane (»HCH) also occurs In conifer
foliage (Eriksson et al., 1989; Hinkel et al., 1989. Reischl at al., 1989a).
Gaseous hydrocarbons are raadyy absorbed by plant foliage. When tomato, barley, and
carrot plants were exposed to hydrocarbon vapors (2 x 10"4 M), foliar damage symptoms appeared
rapidly (Currier, 1951; Currier and Peoples. 1954). Frank and Frank (1989) exposed Norway spruce
saplings to tetrachloroethene in glass chambers. The concentration of tetrachloroethene in the
needles (Cn) increased linearly with increasing concentration in the chamber air. At Cn < 5 /jg/cm3,
the partitioning between air and needles was greater than when Cn > 10 /sg/cm9, suggesting surface
adsorption, followed by cuticular absorption when adsorption capacity was saturated.
Wotverton et al. (1984) reported that house plants absorbed formaldehyde from the air.
However, additional experiments by Godish and Guindon (1989) established that the absorption was
by potting soil rather than the plants. Btoconcentration of vapor-phase organics in leaves is closely
correlated with their octanol-water partition coefficients [(KM (Figure 3); Travis and Hattemer-Frey,
1988; Reischl et al.. 1989b].
Foliar Uptake of Liquid-Phase Oroanics
Penetration of the cuticle by herbicides applied in solutions is positively related to their K^,
values (Shafer and Schonherr, 1985) and is usually linearly related to herbicide concentration
(Bukovac, 1976). Uptake Is generally greater to adaxiaJ than to abaxial leaf surfaces, apparently
because of the higher density of stomata and greater pubescence that occur on adaxial surfaces
(Sargent and Blackman, 1962; Bukovac, 1976). Trichome bases and guard cells are preferred
locations of uptake. Mass movement through stomata occurs if the organic, or Its solvent, are
lipophilic and have low surface tension (Baker, 1970; Bukovac, 1976). Most studies of foliar liquid
uptake by herbicides, using "C labeling, have found substantial uptake (Bukovac, 1976; Steffens and
Wieneke, 1975; van Auken and Hulse, 1979), but 2.4-dichlorophenol and dtoxin do not appear to
penetrate the cuticle (Isensee and Jones, 1971).
When Currier (1951) and Currier and Peoples (1954) exposed plants to pure hydrocarbon
liquids, development of foliar damage was rapid, implying fast uptake across the cuticle. In a
literature review on hydrocarbon ol effects on plants, Baker (1970) concluded that those ois
containing mostly small-molecule, low-volatllty hydrocarbons with low viscosity are most phytotoxic;
the hydrocarbon constituents readly diffuse across cuticles and cell membranes, evaporative losses
are low when applied to sofls or foliage, and penetration through stomata occurs.
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Q-HCH 0 '
.' ' lindane
p,p'-DDT
p,p'-DDE «t -
PCBs
LOGKow
FIGURE 3. The relationship between bioconcentration of organics In plant tissues (Be; ratio of
concentration in plant to concentration in air) and the octanol-water partition coefficient (K.J (after
Travis and Hattemer-Frey, 1988).
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Foliar Uptake of Solid-Phase Oroanlcs
Organics are frequently adsorbed to the surfaces of atmospheric aerosols, but no studies of
particle-associated organic uptake by plants have been conducted. However, cuticular absorption of
lipophilic organics is probably much faster from the vapor phase than from the solid phase (Reischl et
aJ., 19895).
Root Uptake
Many different herbicides are absorbed by the roots of crop plants (Ucrrtenstein and Schuttz,
1965; Uchtensteln et a)., 1967; Shone and Wood 1974; Bukovac 1976; HHton et al., 1976; Briggs et
al., 1982; Wicklrff et al.. 1984; McFartane and WfcMHf, 1985; Scheunert et al., 1986; Topp et al., 1986;
McFariane et al. 1987a,b). A linear relationship is often observed between soil or growth solution
concentrations and rates of uptake, at least when concentrations are low. At higher concentrations,
uptake may be saturated or even reduced (Fletcher et al., 1990), perhaps due to direct toxictty to the
roots. If the original herbicide is not very soluble, root uptake rates may instead be related to
concentrations of soluble soil metabolites (Scheunert et al., 1986).
Root uptake has been documented for the insecticides dieldrin, endrin, aldrin, and lindane, but
does not occur for DDT and heptachlor (Harris and Sans, 1967; Ucrrtenstein et al., 1967; Nash et al.,
1970). Root uptake has also been observed for dioxins, benzene and substituted benzenes, phenol
and substituted phenols, ethanol, naphthol, acetate, tiaryl phosphate esters, pthalate esters, organo-
borates, and trinitrotoluene (Harley and Beevers, 1963; Isensee and Jones, 1971; Kaufman, 1976;
Suzuki et al., 1977; Moza et al.. 1979; Mrozek and LekJy, 1981; Shea et al., 1982; Edwards, 1983,
1986; Casterline et al., 1985; McFariane and Wickltff, 1985; Scheunert et al., 1985, 1986, 1989;
Facchetti et al., 1986; Palazzo and Leggett, 1986; Sacchl et al., 1986; Topp et al., 1986; McFariane et
al., 1987a,b; Adriano et al., 1988; Aranda et al., 1989; Krstich and Schwarz, 1989; Fletcher et al., 1990;
McFariane et al., 1990; O'Connor et al., 1990). Depending on the species, plant roots may or may
not absorb PCBs and PAHs (Edwards 1983, 1986; O'Connor et al., 1990).
During root uptake, organics move by diffusion in the apoplastic water of the conex; then, at
the Casparian strip, cross the cell membranes of the endodermis to reach the xytem. Shone and
Wood (1974), Briggs et al. (1982, 1983), Topp et al. (1986), and McCrady et al. (1987) found a
positive, log-linear relationship between the root concentration factor (RCF; ratio of root tissue
concentration to external solution concentration) and «„, for various organics. However, in a
comparison of nitrobenzene absorption by eight plant species, RCF varied considerably around the
73
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predicted value from Briggs et al. (1962), emphasizing the Importance of species-specific
characteristics (McFarlane et al.. 1990).
Bark Uptake
Penetration of the bark of woody plants by herbicides has been reported (Bukovac, 1976).
Meredith and Hites (1987) found a variety of PCB congeners in the bark of black walnut and tulip-
poplar trees growing near a PCB-contaminated landfill. Very low PCB concentrations in the wood,
much higher PCB concentrations in inner than In outer bark, and a positive correlation between
congener K,., and accumulation in outer bark suggested that uptake occurred from the atmosphere
into the suberin of cork cells.
Translocation
When 14C-labeled herbicides are taken up by roots, most move upward in the transpiration
stream, but some remain in the cortex, apparently unable to cross the endodermis (Hay, 1976; Hilton
et al., 1976). The proportion translocated varies according to length of exposure, plant species, and
type of herbicide (McFarlane et al.. 1987b). What is translocated may be found fairly uniformly in all
the leaves or may be concentrated in young leaves and growing tips (Hay, 1976). Dry soils reduce
rates of translocation, presumably due to lesser transpiration and a lower carbohydrate supply for
phloem transport.
Translocation of root-absorbed, non-herbicide organics has been demonstrated for aldrin,
lindane, bromacil, triaryl phosphate esters, phenol and chlorophenols, dtoxins, pthalate esters, PCBs,
trinitrotoluene, and nltrobenzenes (Uchtenstein et al., 1967; Isensee and Jones, 1971; Mrozek and
Leigy, 1981; Shea et al., 1982; Casterline et al., 1985; Edwards. 1986; Palazzo and Leggett, 1986;
Sacchi et al., 1986; McFariane et al.. 1987a,b, 1990; Aranda et al.. 1989; Retcher et al., 1990). Root-
and leaf-absorbed hydrocarbon ois are transported acropetally and basipetally, respectively (Buckley,
1982). However, there is llttie translocation of root-absorbed PCBs, PAHs, and substituted phenols
(Hilton et al., 1976; Kaufman, 1976; Dorr, 1970; Buckley, 1982; Scheunert et al., 1989), nor do root- or
leaf-absorbed dtoxins and furans appear to be translocated (Reischl et al., 19895).
There is considerable evidence for a positive relationship between transpiration rate and root
uptake (see Figure 4) (Bukovac, 1976; van Oorschot. 1976; McFarlane et al., 1987b). However,
Shone and Wood (1976) discovered that neither external concentration nor transpiration rate had
much effect on uptake and translocation of triazine herbicides by radish.
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1.6
1.2
«
0.8
0.4
0.0
Exposure Time
• 3 days
O 5 days
A 7 days
A 10 days
atrazine in wheat
0.0
0.4 0.8 1.2
Transpiration x Solution Concentration (pg)
1.6
FIGURE 4. The relationship between total uptake of atrazine from solution by wheat and the
expected uptake by mass flow through the xylem (after Walker, 1972).
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Briggs et al. (1983) observed a positive, tog-linear relationship between stem concentration
factor (SCF; calculated In same way as RCF) and herbicide K,,.. rt required 24 to 48 h for SCF to
stabilize, with longer equilibration times for higher K^,. They postulated that stable SCF represented
abalance between upward transport of root-absorbed compounds and their absorption or degradation
by stem tissues.
The herbicide transpiration stream concentration factor (TSCF; ratio of concentration in
transpiration stream to concentration hi external solution) in barley plants was usually less than unity,
implying passive uptake of most herbicides, but greater than unity for 2,4-0, Implying active uptake
(Shone et al., 1973; Shone and Wood, 1974). The mltrobenzene TSCF was consistently less than one
In eight plant species (McFariane et al., 1990). The TSCF was not related in any obvious way to
RCF. However. Shone et al. (1974) demonstrated that the TSCF was correlated with the most readily
diffusible herbicide fraction In barley roots eluted with water following a period of root uptake. This
fraction consisted of the most lipophflic (high K,J herbicides.
Briggs et al. (1982) observed a bell-shaped relationship between TSCF and KMin barley
(Figure 5). They hypothesized that for K^, < 1.8, transport is limited by diffusion across the
endodermal membrane; whereas, for K^, > 1.8, transport is limited by the rate of transport from the
roots to the tops. McCrady et al. (1987) demonstrated that sorption to the xyiem tissue was the main
limitation on transport rates of organics through excised soybean stems. McFariane et al. (1987b)
found differential root uptake and translocation to the shoot of soybean by bromacB, nitrobenzene,
and phenol, although the three chemicals have similar K,^ values. Given the scatter in Briggs et al.
(T982) relationship, this result is not surprising. Boersma et al. (1988) have developed a model of
organic compound transport in the xyiem and phloem. This model is most sensitive to K^,, rates of
detoxification, membrane permeability, and the volume of xyiem and phloem in various organs.
Generally, the bulk of foliar-applied organics is not translocated; that portion that is partitioned
to the stem, little goes to the roots (Hay, 1976; Shone and Wood, 1976; Weber and Mrozek, 1979).
Foliar-applied ureas, triazines, DNOC, PCP, and hexachlorophene are not translocated (Fogg, 1948;
Hay, 1976; Kaufman, 1976; van Auken and Hulse, 1979).
Detoxification
Plants detoxify herbicides primarily by metabolic degradation and to a lesser extent by
adsorption to btomolecules. Adsorption to lecithin and cell membranes has been demonstrated for
2.4-D and 2,4,5-T (Wain and Smith, 1976). The metabolic fates of many different herbicides have
been elucidated by "C labeling studies (Naylor, 1976). In most cases, metabolism converts the
76
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1.0
0.8
0.6
0.4
0.2
barley
0.0
-1
LOGKow
RGURE 5. The relatkxiship between transpiration stream concentratkxi factor (TSCF; ratio of
concentration In transpiration stream:concentration In nutrient solution) of various organics and
octanol-water partition coefficients (K,J in barley (after Briggs et al., 1982).
77
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phytotoxic herbicide to nonphytotoxte or less phytotoxic forms; however, the reverse is sometimes
true (e.g., amitrole, bipyridyliums, 2,4-0 precursors, DNOC esters) (Wain and Smith, 1976; Menn and
Still, 1977). Generally, detoxification proceeds in two steps: (1) an oxidizing, reducing, or hydrolyzing
reaction that modifies the organic toxin by adding reactive groups; and (2) a subsequent conjugation
reaction that further reduces toxfclty and yields a metabolite in a form suitable for storage or transport
(Kaufman, 1976; Menn and StOI, 1977).
The metabolic fates of most nonherbicide organics in plants are not well understood.
Degradation of PAHs, pthalate esters, 4-nltrophenol dichlorobiphenyls, nltrobenzenes and
trinitrotoluene have been reported (Dorr, 1970; Moza et a!.. 1979; Edwards, 1986; Palazzo and
Leggett, 1986; McFarlane et al., 1987a, 1990; Fletcher et al., 1990; Preiss et al., 1989; Schmitzer et at.,
1988), but the phytotoxiclty of the metabolites is unknown.
One major mechanism of plant detoxification of organics is conjugation to glutathione. For
example, com, sorghum, and sugarcane, which possess this metabolic capability, resist atrazine;
whereas broadleaf weeds and grasses, which do not have this capability, are susceptible to atrazine.
Schroder and Rennenberg (1989) reported glutathione conjugation of pentachloronltrobenzene (often
sprayed on crops to protect them from pathogens), but not of atrazine, by Norway spruce.
Plant Efflux
Efflux of organics from plants occurs in three ways: (1) foliar efflux or "leaching,* (2) passive
loss by transpiration, and (3) root exudation (Rovira, 1969; Tukey, 1970). Radiocarbon studies have
shown that root exudation of herbicides is an energy-requiring process, occurs in the zone of active
root elongation, and involves readily translocatable and little-metabolized herbicides (Bukovac, 1976).
The extent to which organics taken up by plants may be subsequently lost by foliar efflux apparently
has not been studied.
Siddaramappa and Watanabe (1979) exposed rice plants to "C-carbofuran in nutrient
solutions sealed at the root collar to prevent volatBized carbofuran from reaching the leaves.
Transpired carbofuran was trapped in ethylene glycd in shoot chambers. Over a 10-day period, 9 to
17% of the absorbed radioactivity was lost via transpiration, apparently carbofuran was carried
passively along in xyiem water. Higher transpiration rates resulted in more carbofuran recovered from
the air. These authors also found significant radioactivity in guttation water. McFarlane et al. (1990)
found that 10-40% of not-absorbed "C-nltrobenzene volatBized from leaves of crops and woody
plants.
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KNOWN EFFECTS ON PLANTS
Visual Injury and Growth
As with plant uptake, there is a dose association between Km values of organics and their
inhibitory effects on plant germination and growth. A good model that fits a wide range of plant data
is a nonlinear equation where toxiclty increases linearly up to a critical value of K^, then falls off at
higher K.. (Hansch et al., 1969).
Herbicides
The mechanisms by which herbicides influence growth include inhibition of mitosis, inhibition
or stimulation of tissue enlargement and elongation, and alteration of patterns of tissue differentiation
(Cartwright, 1976). Herbicides may also alter dry mass distribution among plant organs, prevent
pollination, and cause abscission of leaves and other plant parts (Addicott, 1976; Ratsch et al., 1986);
and, in the case of auxin-tike herbicides such as 2,4-D, abnormal growth of various organs (van Andel
et al., 1976). Effects may be manifested at all stages of growth from germination to reproduction.
For example, Sund and Nomura (1963) observed that some herbicides were most effective in
inhibiting germination, others at inhibiting seedling growth, and stBI others at inhibiting growth of
juvenile plants. Several herbicides, including dinoseb, POP, 2,4-D and 2,4,5-T were inhibitory at two
or more of these life stages.
Substituted phenols vary widely in their effects on germination and seedling growth (Sund and
Nomura, 1963; Amer and All, 1968; Shea et al., 1983). Root growth is stimulated by 2,4-dinitrophenol
PNP) at low concentrations and is inhibited at high concentrations, probably because It is a
respiratory uncoupler (Shea et al., 1983). Chlorosis, wilting, and necrosis of peas is caused by
2-nttrophenol and 2,4-dicNorophenol (Amer and All, 1968).
Van Haul and Prinz (1979) studied the effects of a number of organic vapors on the growth of
several crop species. Over a wide range of concentrations (0.1-7.0 mg m~s), relative growth
reductions were positive log-linear for all chemicals and species studied. Growth reductions relative
to those caused by sulfur dioxide at equivalent concentrations varied between plant species and were
usually higher at lower concentrations. The most phytotoxic substance was ethylene (up to eight
times the growth reduction of sulfur dioxide), followed by formaldehyde and acetic acid.
Dimethyfformamlde and methanol were less phytotoxic than sulfur dioxide. Dichloromethane, toluene,
trichloroethylene, acetone, and xylene caused littie or no growth reduction at concentrations up to
60-160 mg/m3.
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Surfactants are often used to improve VettabHty" of foliar-applied herbicides. These
substances may In themselves Increase, reduce, or have no effect on the growth of crops
(LJchtenstein et al., 1967; Singh and Orsenigo. 1984).
Other Organics
Toxicity of pure hydrocarbon sprays applied to plants increases in the order: straight-chain
paraffins < napthenes and alkenes < aromatics (Crafts and Reiber, 1948). When barley, tomatoes,
and carrots were exposed to hydrocarbon vapors, toxicity Increased In the order benzene < toluene
< xylene < trimethylbenzene (Currier, 1951). The symptoms started with Increased foliage odor,
followed by the appearance of dark areas on the foliage, wBting, and, at higher concentrations, death.
Shoot dry mass gain was positively correlated with vapor concentration for sublethal doses. When
these same compounds were sprayed on foliage as pure liquids, the same order of toxJclty was
observed, but the differences among hydrocarbons were greater, apparently because of differences in
volatilization (vapor pressure decreasing in reverse order to toxicity). Currier and Peoples (1954)
repeated these experiments, finding that toxicity Increased in the order: hexene < hexane <
cydohexane < cydohexene < benzene.
Crude oils, their refined products, and coal liquid cause visual injury and reduce growth of a
variety of weed and crop species (Baker, 1970; Biankenship and Larson, 1978; Warner et al., 1984).
Toxicity is due to the mixture of hydrocarbons in these liquids, including alkanes, alkenes, napthenes,
naphthalenes, phenols, and aromatics. Many hydrocarbons are known to be toxic in the presence of
ultraviolet radiation (Larson and Berenbaum, 1988). For example, the PAH fluoranthene causes foliar
Injury under UV light (Zweig and NachtigaJI, 1975). Creosote, a widely used wood preservative,
contains a variety of PAHs and other aromatics. Concentrations of 18 to 34 mg/L reduced root
growth of onion by 50% (Sundstrom et al., 1986).
Arodor 1254, a mixture of PCBs, Inhibited height growth and fresh mass increase of both
soybean shoots and roots; at higher concentrations, newly formed leaves showed abnormal curling
(Weber and Mrozek. 1979). Pthalate esters caused reduced growth and chlorosis of new leaves and
growing tips of tobacco and com (Buta, 1975; Shea et al., 1982). Trlbutvf phosphate reduced root
growth of rice, radish, and soybean at concentrations of 10 to 100 pg/g sol (Muir, 1984).
Trinitrotoluene (TNT) lowered root and shoot growth of nutsedge (Palazzo and Leggett, 1986).
Sodium tetraphenyfborate (NaTPB) reduced growth of loblolly pine seedlings because of increased
plant boron uptake to toxic levels (Kaplan et al., 1988). NaTPB's major degradation product,
diphenylbortc acid PPBA), did not affect growth. However, both NaTPB and DPBA, as well as
biphenyl, another breakdown product, reduced growth of sorghum, with NaTPB having the greater
influence (Adriano et al., 1988).
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(Many substituted phenols have auxin-like activity. Harper and Wain (1969) found that the
greatest growth promotion in standard auxin activity tests was produced by phenols with substitution
at both ortho-positions. Like phenols, PAHs can stimulate plant growth (Grfif and Nowak. 1966).
Several alcohols promoted growth of excised wheat roots in the light, whereas others had no effect
(Qudjonsdottlr and Burstrom, 1962). Hexadecand and docosanol decreased growth of tobacco
(Bourget and Parups, 1963). Organic adds and aromatic alcohols, aldehydes, and carboxylic acids
Inhibited lettuce germination (Mayer and Evenari, 1953; Reynolds, 1978). The Insecticides parathion,
diazinon, and lindane significantly reduced root growth, but not shoot growth, of peas; whereas,
simazlne reduced both root and shoot growth (Uchtenstein et al., 1967).- Vapor-phase aldehydes at
concentrations > 0.2 ppm caused foliar Injury to petunias (Brennan et al., 1964).
Gas Exchange and Water Relations
Herbicides
For most herbicides studied to date, stimulation, inhibition, or no effect on dark respiration
have been reported (van Oorschot, 1976). In some cases, lower concentrations stimulate; whereas,
higher concentrations inhibit, a typical response for oxidatlve uncouplers such as DNP (Shea et aj.,
1983). Respiration may also be affected differentially by the same herbicide in different plant organs.
Inhibition of net photosynthesis has been observed in a wide variety of herbicides when plants
are exposed to herbicides in soils or nutrient solutions, or applied to foliage (van Oorschot, 1976).
With the exception of acylamkJes, virtually all herbicides tested to date inhibit net photosynthesis and,
when a range of concentrations is studied, inhibition increases with Increasing concentration. Unlike
respiration, stimulation of net photosynthesis has not been observed (van Oorshot, 1976).
All herbicides that reduce net photosynthesis reduce transpiration as well, such as DNP
(Barber and Koontz, 1963; Pemadesa and Korelege; 1977). This phenomenon raises the question as
to whether or not photosynthetic inhibition is primarily due to stomatal closure, or to nonstomatal
(biochemical) effects. Nonstomatal inhibition can be represented as a "mesophyll resistance' (r J in
series with rb and r.. Imbamba and Moss (1971) found that rm for CO2 in com exposed to atrazine
increased without an accompanying change in r,. Most evidence for the relative roles of r, vs. rm is
indirect, however. When photosynthesis and transpiration were measured simultaneously following
exposure to ureas, triazines, diazines, bipyridyiiums, and simeton, the relative reduction in
photosynthesis was greater than that of transpiration, and commenced sooner (van Oorshcot, 1976).
This suggested direct nerbiddal action at the sites of photosynthesis. However, for PMA, DSA,
toxynD, propanD, nitrofen, and fluorodffen. photosynthesis and transpiration were inhibited to a similar
extent and at the same time, Implying that stomatal closure restricted uptake of CO2. Similar
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variations in the relative reduction of transpiration vs. photosynthesis were observed for nitrobenzene
(McFarlane et al., 1990). Herbicides also disrupt the water economy of plants by reductions in root
water uptake, as reported for phenoxy herbicides, ureas, triazines, propanB, and dinoseb (van
Oorschot, 1976).
Herbicides that inhibit the light reactions of photosynthesis often inhibit photorespiration as
well. This has been demonstrated by similar reductions In net photosynthesis of wheat and corn in
air with normal and low 0, concentrations following exposure to diuron (Downton and Tregunna.
1968), and by the absence of the postllumination CO, burst In barley exposed to atrazine (Imbamba
and Moss, 1971). However, when com was exposed to atrazine, CO, evolution in the light was
greater than in darkness, suggesting llttie or no suppression of photorespiration (van Oorschot, 1976).
Other Organics
Organic chemicals other than herbicides have been shown to influence plant growth and
development. Wood et al. (1985) and Wood and Payne (1986) found that pecan leaves sprayed with
the fungicides propiconazole, benomyl, triphenyttin hydroxide, and dodine, or with pyrethrold,
carbamate, and organophosphate Insecticides showed inhibition of net photosynthesis of up to 20%
within one day following application, with full recovery usually occurring within several days.
Hydrocarbon oils universally reduce transpiration and net photosynthesis of crops (Baker, 1970).
Recovery of both processes Is correlated with the dissipation of foliar-applied oils, and raising the C02
concentration of the air reduces Inhibition, suggesting that blockage of stomata is the cause of
Inhibition. However, due to the rapidity with which hydrocarbons penetrate leaf tissues, nonstomatal
effects are likely as well. Hydrocarbon oils either increase or decrease respiration, with the effect
differing among oils and between plant species (Baker, 1970). Nitrobenzene at 8 g/L in growth
solution had varying effects on gas exchange of several species; ranging from no effects on soybean
to complete photosynthetic repression of green ash (McFarlane et al., 1990). Even at 100 g/L,
nitrobenzene did not affect soybean gas exchange (Fletcher et al., 1990).
One hypothesis to account for European forest-tree decline or "Waldsterben* is phytotoxicity
of anthropogenic and biogenic hydrocarbons. When young Norway spruce were exposed to 0-pinene
vapors, photosynthesis was Inhibited and chlorophyll degraded (Gross et al., 1988; Wagner et al.,
1989). ^pinene inhibits the HOI reaction (PaJy, 1981), resorting In uncoupling of oxidatfve
phosphorytation (Wagner et al., 1989). Frank and Frank (1985) were able to reproduce the decline
symptom of needle yellowing In Norway spruce by exposing branches to vapors of two anthropogenic
VOCs, tri- and tetrachloroethene, in the field. They later showed (1986, 1989) that these compounds
were absorbed by spruce needles and, In the presence of sunlight, the compounds caused
photodegradation of photosynthetic pigments. Rippen et al. (1987) have speculated that nltrophenols
injure forests.
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Ultrastructure and Membrane Function
Because most organics are lipophBlc, ft Is not surprising that they often disrupt cell
membranes. At low concentrations, dissolution of organic molecules In membranes forces apart the
fatty acid chains of the phospholipids, Increasing membrane permeability. At higher concentrations,
the biiayer configuration may disintegrate completely. Some organics disrupt ceil membranes
Indirectly through the formation of free radicals, inhibition of lipid synthesis, or changes in the types of
lipWs synthesized. The disruption of membranes probably accounts for many of the ultrastructural
changes attributed to organics.
Herbicides
Contact foliar damage ("burning* or "wetness") by 2.4-D, endothal, and several other
herbicides is believed to result from physical disruption of cell membranes (Morrod, 1976). Paraquat
disrupts membranes because Its metabolism produces hydrogen peroxide (Morrod, 1976). Two
consequences of membrane damage are leakage of cell contents and altered rates of active uptake.
For example, DNP increases plant cell permeability to water, whereas stomatal closure caused by
atrazine, 2,4-D, and DNP may be caused by leakage of K4 from guard cells (Morrod, 1976; Shea et
al., 1983). Ultrastructural changes induced by herbicides in leaves include swelling of thylakokJs and
cNoroplasts, destruction of phloem and vascular cambium, abnormal differentiation of xylem, bark,
and root cells, and mitotic abnormalities (Amer and All, 1969; Unck, 1976; Moreland and Hilton, 1976).
Other organics
Despite rather low K^ values, bulky hydrocarbons are effective at disrupting membranes. For
example, heptene-2, with a bent configuration, causes discoloration and death of cotton hypocotyls,
while straight-chain heptane shows no phytotoxicity (Morrod, 1976). It has also been shown that
l-pinene alters the relative proportions of lipids and causes disappearance of microsomes In Norway
spruce needles (Frosch et al., 1989; KJingler and Wagner, 1989).
Translocation
Root ion uptake is inhibited by many organics. For example, nltrophenols, 2,4-D, simazine,
and propanR inhibit phosphate uptake, DNP inhibits Ca8+ uptake, the organophosphorus insecticide
Systox inhibits CT uptake, and phenolic acids inhibit phosphate and K* uptake (Barber and Koontz,
1963; Glass, 1973, 1974; Oertii and Ahmadi, 1975; Kaufman, 1976; Morrod. 1976). As uptake of these
ions is energy-requiring, the mechanism of reduced uptake may be inhibition of respiration or
uncoupling of oxidative phosphorytation.
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Herbicides alter the qualitative and quantitative nature of solute transport In plants. The major
effects are physical obstruction of the phloem caused by unorganized cell division, blockage by
callose formation, localized xylem/phloem tissue injury or death, stimulation or inhibition of metabolic
sinks, and decreased transpiration (Ashton and Bayer, 1976). The only reference found for a non-
herbicide organic was Wedding and Rtehl (1958), who reported that petroleum o» inhibited phosphate
transport from roots to leaves In lemon seedlings.
Metabolism
Herbicides
Herbicides have various modes of action. Many interfere with mitochondria! metabolism.
Chloro- and nltro-phenols, benzimidazoles, nttrfles, and phenoxy acids are uncouplers of oxidatrve
phosphorylation (i.e.. the synthesis of ATP is stopped even though respiration continues) (Beevers,
1953; Kaufman, 1976; Kirkwood, 1976; Shea et al, 1983). Substituted phenols vary widely in their
uncoupling ability (Gaur and Beevers, 1959). Other herbicides, including amides, benzoic acids,
carbamates, dinitroanilines, halogenated aliphatics, thiocarbamates, triazines, triazoles, and ureas,
reduce oxidath/e phosphorylation by either inhibiting the transfer of energy to intermediates in the
formation of ATP, or Inhibiting the flow of electrons along the electron-transport chain (Kirkwood,
1976). Glycolysis is inhibited by halogenated aliphatics and phenoxy acids; the latter also inhibit the
pentose-phosphate pathway (Kirkwood, 1976).
Many herbicides negatively affect photosynthesis. Inhibition of electron transport in
Photosystems I and II occurs by removal or inactivation of intermediate electron transport carriers, or
by the herbicide acting as an electron acceptor in competition with normal acceptors. This inhibition
is light-dependent. Uncouplers dissociate electron transport from ATP synthesis
(photophosphorytation). Some herbicides are both inhibitors and uncouplers. For a comprehensive
review, see Moreland and Hilton (1976).
Herbicide effects on Intermediate metabolism are numerous. Many inhibit rates of synthesis
or degradation of carbohydrates, lipids, proteins, and nucleic acids, or alter the types of these
compounds formed (Cherry. 1976; Kirkwood, 1976; Morrod, 1976). However, 2.4-D and 2,4,5-T
stimulate lipid synthesis (Morrod. 1976). Some herbicides inhibit nitrite reduction, causing nitrite to
accumulate In plant tissues (Klepper, 1979).
Other Organics
Fungicides (e.g., captan) inhibit respiration and photosynthesis at the cellular level in a similar
manner to herbicides (Budimir et al., 1976). Evidence exists for inhibition of gylcolysis, aerobic
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respiration, and uncoupling of oxkJattve phosphorytation by hydrocartxxi ols (Baker, 1970). Ravanel
et al. (1989) reported the uncoupling ability of chlorophenoi in maple cell suspensions.
Plant Resistance
Plant species differ in their resistance to herbicide effects, and species sensitMty rankings
differ among herbicides. Other factors influencing sensitMty are genotype, growth stage, nutritional
status, physical damage, cuticular permeabtity. temperature, water stress, light intensity, leaf area, leaf
morphology, transpiration rate, and rooting depth (Aberg and Stecko, 1976; Muzik, 1976). It is well
established that the primary causes of "herbicide selectivity* (the pesticide science term for plant
sensitivity) are due primarily to three factors: (1) different rates of uptake by the roots and/or shoots;
(2) different rates of translocation and differences in the organs and tissues to which translocation
occurs; and (3) different capabilities for, or rates of, metabolic detoxification (Sargent, 1976; Wain and
Smith, 1976). At the slte(s) of herbicide action, species differences In sensitivity often disappear. For
example, the HQI reaction in isolated chloropJasts from different plant species is inhibited to the same
degree by the same concentrations of simazine and several other herbicides (van Oorschot, 1976).
Differential resistance of plant species to hydrocarbon ols has also been observed (Baker, 1970). In
the case of the Umbellilerae, cell membranes have Inherent high resistance to penetration by oils
(van Overbeek and Blondeau, 1954).
It is worth noting that plants contain a bewildering variety of organic compounds that are
toxic to animals. Most of these are believed to be evolved, qualitative or quantitative, defenses
against herbrvory. Interestingly, many of these are known phytotoxins, including organic acids and
phenols puke, 1977). This fact clearly indicates that many plants already possess mechanisms to
detoxify some priority air pollutants.
POTENTIAL EFFECTS ON PLANTS
Ecological Effects
When photosynthesis and respiration are inhibited, or ATP synthesis uncoupled from
metabolism, the whole-plant reaction should be reduced growth and reproduction. Energy allocated
to alleviating the stress of chronic toxin uptake may also reduce the plants ability to produce
defensive chemicals. Inhibition of root water or ion uptake may increase allocation to roots at the
expense of shoots. Together, these processes lead to lesser competitive abSlty of affected species.
Except near point sources of toxic organlcs, these effects may occur gradually as small but
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continuous atmospheric inputs accumulate. Ultimately, shifts in species composition and dominance
may become apparent at the plant community level.
Just which species may suffer most is largely a matter of speculation. For those organics
that readily cross the root endodermis, rates of translocation. and hence of bioaccumutation, may be
higher for plants with greater transpiration rates (I.e., Inherently smaller stomatal resistances). These
tend to be early successJonal species wtth fast growth rates and among late-successional species that
are shade intolerant Conifers have the highest stomatal resistances of any major plant life form
(Komer et al., 1979), yet they also have exceptionally thick, waxy cuticles Into which organics may be
absorbed. The major route of entry, foliage vs. roots, of organics could play a pivotal role In
determining the relative resistances of different plant taxa.
Interactions wtth other Pollutants
Many organics are weak acids, and their lipophBiclty is greater in the unionized form.
Penetration of weak-acid herbicides through cuticles Increases as solution pH decreases (Fogg, 1948;
Sargent and Blackman, 1962; Bukovac, 1976). This phenomenon probably accounts for the enhanced
phytotoxiclty of nltrophenols and organic acids at low pH (Simon and Beevers, 1951; Simon et al.,
»
1952). As a result, there is a strong possibility that acidified wet deposition Increases the
phytotoxiclty of organics by enhancing uptake (Mullen, 1986).
Biogenic monoterpenes present an interesting case. Like anthropogenic VOCs, they
contribute to the photochemical formation of ozone and PAN (Lurmann et al., 1983; Trainer et al.,
1987; Chameides et al., 1988) that have been shown unequivocally to reduce photosynthesis and
growth of crop and tree species (Reich, 1987). Wagner et al. (1989) have hypothesized that ozone
penetrates the mesophyll of conifers, increasing membrane permeability and liberating monoterpenes
stored in secretory vesicles. In the presence of ozone, monoterpenes promote the oxidation of sulfur
dioxide to sulfate, possibly causing local enhancement of acid deposition to forests (Stangl et al.,
1988).
Interactions between different types of gaseous pollutants are well documented (Ormrod,
1982). Because terrestrial plant communities are receiving chronic inputs of possibly hundreds of
different organic compounds, It Is possible that the overall effects in combination are additive, more
than additive, or less than additive.
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CONCLUSIONS
Thousands of organic compounds are released into the atmosphere each year by human
activities, including pesticides, aliphatic and aromatic hydrocarbons (both halogenated and
unhalogenated). polychlorinated biphenyls, and phthalate esters. These are transported long distances
and reach natural vegetation via wet and dry deposition. Deposition fluxes and velocities have been
characterized for only a few organics In a few geographic areas.
The potential for plant uptake and btoconcentration of organics is high and correlated with
octanol-water partition coefficients. Greater transpiration rates produce greater root uptake. Foliar
uptake of vapor-phase organics has received little attention. Translocation patterns differ among
organic compounds. Metabolic detoxification plays a major role In plant resistance, but has been well
characterized only for pesticides.
With few exceptions, the modes of action of organics in plants are well-known only for
herbicides. A variety of effects by herbicides, usually but not always Inhibitory, have been observed
on plant growth and morphology; photosynthesis, respiration, and transpiration; cell membranes and
uttrastructure; translocation; and respiratory, photosynthetic, and intermediate metabolism. Many of
these same effects have been observed with nonherbicide organics. I found no studies on
population-, community-, or ecosystem-level effects of airborne organics.
Relevance of Previous Work
As the analysis of the PHYTOTOX database indicated, there is only one class of organics -
herbicides - for which the physiological effects on plants are relatively well known. This is hardly
surprising, because their purpose is to kill plants, especially weeds. Although It would .be desirable to
know more about nonherbicide organics, existing knowledge of herbicide effects is helpful in
predicting the Impacts of organics in general on terrestrial vegetation. For example, herbicides are at
times applied aerially to crops or forests, with consequent downwind drift onto adjacent natural plant
communities. Some herbicide precursors (e.g., 2,4-dichlorophenoi, used to make 2,4-D, and maleic
anhydride, used to synthesize pyridazlnes) are on the EPA's Toxic Release Inventory (U.S. EPA, 1989).
Point releases of these, and of the finished herbicides themselves, from sites of manufacture may
affect natural vegetation. A few herbicides have other human sources besides pesticide manufacture
and use. For example, DNP, DNOC, and other nitrophenois are produced as by-products of smog-
forming reactions. Finally, the observed correlations between plant bioconcentration or modes of
action and the physicochemlcal properties of herbicides permit general predictions to be made
concerning the phytotoxic effects of nonherbicide organics.
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Nevertheless, the literature emphasis on herbicides poses many obstacles to understanding
potential phytotoxtefty of airborne organlcs. One obstacle is that the conditions under which
herbicides are applied to plants - high concentrations over short time periods - are quite different from
those encountered by natural vegetation - low concentrations in a repeated or continuous manner.
Significant differences between acute and chronic exposures to the major airborne pollutants have
been demonstrated (Lefohn and RunecUes, 1967). Thus, extrapolation from acute agricultural doses
to chronic natural vegetation doses may be unwarranted. Another difficulty is the extremely high
concentrations of organics used in most dose-response experiments. For example, solution
concentrations of 10s to 10° ng/L are frequently encountered h phytotoxtelty studies. Compare these
with the concentrations In Table 1.
The best use that we may be able to make of the existing literature is to rank the likelihood of
plant uptake of organics according to the vapor pressures. KM values, and half-lives in soil of the
organics (Ryan et al., 1988; Travis and Hattemer-Frey, 1988).
RECOMMENDATIONS
The most critical areas in need of further knowledge are: (1) spatial and temporal patterns of
atmospheric concentrations and wet/dry deposition rates; (2) rates of uptake by roots from soil, and
of vapors and wet deposition by shoots; (3) translocation patterns; (4) physiological effects from the
whole-plant to the subcellular level; and (5) modes of action. In addition, much more emphasis needs
to be placed on nonpestickje organics and on native, nonagricurtural species growing in natural
environments. In all such work, we must use acute or chronic doses that bridge the range of
exposures actually occurring in natural ecosystems.
Some specific suggestions are as follows.
(1) Upgrade routine air, aerosol, and precipitation monitoring networks (e.g., National Atmospheric
Deposition Program, National Dry Deposition Network) to measure concentrations of at least
the most common and highly concentrated organics.
(2) Characterize dry deposition to various plant communities. Traditional direct measurement
techniques • flux-gradient and eddy correlation (Hicks, 1986) - are difficult due to the very low
concentrations of most organics. However, new advances in laser technology - FTIR, Ikjar,
tunable diode lasers - hold promise for use of the flux-gradient technique in the near future.
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(3) Develop methods to measure short-term uptake of organlcs by plant shoots. These would
complement 14C work by distinguishing between cuticular and stomatal absorption, and would
permit scaling-up of leaf-level uptake to entire canopies in inferential methods of estimating
dry deposition (Hicks. 1986). The technological challenge of generating and measuring low
but stable concentrations is great, but the potential gain In predictive capability should be
worth the effort.
(4) Conduct experiments on dose-response relationships of whole plants growing in native soils,
including growth, allocation, morphology, water relations, carbon assimilation, and nutrition.
Growth chamber, continuously stirred tank reactor (CSTR), and open-top chamber studies, in
order, would provide Increasing realism in terms of environmental conditions.
(5) Determine the cellular/subcellular modes of action of various classes of organics. Previous
work with herbicides suggests that virtually every translocation and metabolic process is
worthy of investigation.
(6) Measure dose-response relationships for species representative of other major life forms of
terrestrial plants besides crops: annual and biennial weeds, perennial herbs, ferns, evergreen
and deciduous shrubs, evergreen and deciduous hardwoods, and conifers.
(7) Develop sensitivity rankings based on absorbed rather than external dose. This approach has
yielded better predictive capability for major gaseous air pollutants (Reich, 1987). Because
organics vary so widely in their lipophilicity, even within similar structural classes, relationships
between external dose and response for one organic are unlikely to apply to other organics.
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EFFECTS OF ATMOSPHERIC POLLUTANTS ON PEATLANDS
Noel R. Urban
Environmental Engineering Sciences
Department of Civil and Mineral Engineering
University of Minnesota, Minneapolis, MN
INTRODUCTION
The phenomenon of long-range transport of atmospheric pollutants is well documented. From
the Arctic (Murozumi et al., 1969; Weiss et al., 1971; Ng and Patterson, 1981), to the Antarctic
(Boutron and Patterson, 1983), and to remote Pacific islands (Atlas and Giam, 1981; Duce et al.,
1983), there exists no location on Earth that does not receive measurable atmospheric anthropogenic
contaminants. Although humans have been disseminating pollutants through the atmosphere for
several thousand years (Patterson, 1980; Kuster et al., 1988), the intensity of air pollution has
increased exponentially since the industrial revolution (Edgington and Robbins, 1976; Ferguson and
Lee, 1983a) both in terms of the absolute masses of pollutant emissions and the number and type of
pollutants (Hett et al., 1981; U.S. EPA, 1989). That air pollutants can cause intense and catastrophic
damage to organisms and ecosystems on a local scale is well documented (Gorham and Gordon,
1960a,b, 1963; Glgnac and Beckett, 1986). Recent regional declines in forest health have indicated
that synergistic effects may occur among pollutants and natural stressors (Schulze, 1989), but our
understanding of the effects of chronic exposure to multiple pollutants is very limited.
Peatiands have been considered ideal environments in which to study atmospheric deposition
of anthropogenic contaminants. First, peatiands classified as ombrotrophic bogs are Isolated from
surface and groundwaters, and thus receive all of their hydrdogic, nutrient, and mineral inputs from
the atmosphere. There is, therefore, no ambiguity about the source of the inputs; substances
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deposited in bogs were derived from the atmosphere. Second, the continuous accumulation of
organic-rich sediments provides an excellent substrate for sorting metallic and organic pollutants, and
sealing them beneath subsequent peat accretions In a fashion that preserves the chronological
sequence of deposition. For these reasons, concentrations of S, N, trace metals, major cations, and
organic contaminants in surface moss have been used to deduce spatial patterns in atmospheric
deposition; and historical changes in deposition of these constituents have been inferred from peat
profiles (Mattson and Koutier-Andersson, 1955; Rapaport and Eisenreich, 1986; Norton and Kahl, 1987;
Kuster et al., 1988).
The factors that render bogs useful monitors of atmospheric deposition also render them
particularly susceptible to damage from atmospheric deposition. The dependence of bog plants on
atmospheric supplies of nutrients, trace metals, and major tons implies that changes in atmospheric
supply will affect the growth of the bog plants. The situation is exacerbated by the very adaptations
of the mosses and lichens to the ombrotrophic conditions. The thin or absent cell walls and the lack
of a vascular transport system facilitate the uptake of wet and dry deposition and its associated
contaminants (Winner and Atkinson, 1987). The anaerobic, water-logged conditions that cause the
rapid sediment accretion also promote acidic, anoxic conditions that enhance the toxicity and
persistence of many pollutants reaching this environment. Furthermore, the geographic distribution of
peatlands coincides with areas that receive extensive deposition of atmospheric pollutants and further
predisposes these sites to damage. Peatlands are prevalent in latitudes north of 40"and thus
coincide with major industrial regions in northern Europe and eastern North America (Kivinen and
Pakarinen, 1981; Gorham et al., 1987; Matthews and Fung, 1987).
The importance of peatlands as wildlife habitat and vegetation preserves is well known (Karns,
1984; Scoullos and Hatzianestis, 1989). Less well known but of equal or greater importance, are the
chemical and hydrologic linkages between peatlands and other ecosystems. Wetlands greatly modify
water quality of precipitation and streams by consuming alkalinity (Bayley and Schindler, 1987),
sequestering trace metals (Eger et al., 1981) and nutrients (van der Valk et al., 1978), neutralizing
mineral acids (Hemond, 1980; Wieder and Lang, 1982), and exporting organic acids (Hemond, 1980;
Urban et al., 1989a). Globally, peatlands are significant reservoirs of carbon (Houghton, 1986; Clymo,
1987; Gorham, 1988a), major sources of methane (Matthews and Fung, 1987) and important sources
and sinks of S and N gases (Adams and Farwell, 1981; Hemond, 1983; Bowden, I986a,b, 1987;
Nriagu et al., 1987; Urban et al., 1989a,b). Clearly, deleterious effects of atmospheric pollutants on
peatlands will have many ramifications extending outside the peatlands themselves.
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EVIDENCE OF ATMOSPHERIC TRANSPORT AND DEPOSITION OF POLLUTANTS
There exists a very large body of literature documenting the deposition of atmospheric
pollutants in peatlands. A wide variety of pollutants have been examined Including components of
acid deposition, trace metals, radionudkJes, synthetic organic contaminants, soot and magnetic
particles from fly ash and dust. Atmospheric deposition has been implicated as the source of
pollutants In three different ways. The mere presence of anthropogenic contaminants in remote
ombrotrophic bogs constitutes the first piece of evidence implicating atmospheric deposition as the
source of contamination. Ombrotrophic bogs receive all hydrologic inputs from the atmosphere; they
have no in-flowing streams, receive no groundwater discharge, and receive little or no runoff from
surrounding mineral softs. Hence, the only possible source of contaminants found In such sites is the
atmosphere. The relatively high radioactivity In mosses in English bogs could only have come from
bomb fallout or local emissions to the atmosphere (Gorham, 1958a, 1959). Aerial drift from local use
of pesticides has been implicated as responsible for the high burdens of toxaphene in bogs in Maine
and Minnesota (Rapaport and Eisenrefch, 1986). The presence of pesticides (toxaphene. mirex,
endosulfan, dieldrin, chlordane, aldrin, lindane, heptachlor, methoxychlor) and industrial chemicals
(polychlorinated biphenyls [PCBs], polycydic aromatic hydrocarbons [PAHs], and hexachlorobenzene
[HCB] in remote bogs in Minnesota, Ontario, and Quebec clearly implicates long-range transport of
atmospheric pollutants (Rapaport et a!., 1985; McCrea and Wickware, 1986; Rapaport and Eisenreich,
1986, 1988). The existence of untransformed DDT. a compound whose use has been banned in
North America for nearly 20 years, In surface peat from sites across northeastern North America can
result only from atmospheric transport and deposition of pesticides used in Central America (Rapaport
et al., 1985; Rapaport and Eisenreich, 1988).
The second piece of evidence linking atmospheric deposition with the presence of pollutants
In peatlands comes from geographic surveys that show correlations between pollutant concentrations
in peatlands (surface waters or surface vegetation) and proximity to pollutant emission sources. In a
large survey of North American peatlands, Gorham et al. (1985) demonstrated that concentrations of
major ions in surface bog waters were controlled by atmospheric deposition of sea spray (Na, Cl, Mg,
SO/), soil dust (Ca, Mg, Na, K), and air pollutants I (SO/) see also Urban et al., 1987a]. At these
same sites, concentrations of Fe and Al in surface waters also were related to rates of soD dustfall,
but concentrations of Pb were explained largely by rates of deposition from anthropogenic sources
(Urban et al., 1987b). Similarly, Blancher and McNichol (1987) demonstrated that, compared to
peatlands out of the plume, peatlands downwind of the Sudbury smelter (Ontario, Canada) had higher
dissolved concentrations of SO/, H*. Ni, Mn, and Cu. Taylor and Crowder (1983) observed that
concentrations of Cu, Nl, Fe, and Mg in marsh sediments decreased with increasing distance from
smelters. Zoltai (1988) found that concentrations of Zn. Fe, Al, and Pb were elevated in the waters of
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peatlands within 5 km of a smelter in Manitoba, but concentrations of Zn, Pb, Cu, and As were
elevated above background concentrations in peat from sites up to 100 km from the same smelter.
Similarly, Gignac and Beckett (1986) observed elevated concentrations of Cu and Ni in bog waters
within 5 to 10 km of the Falconbridge smelter (Ontario, Canada) and elevated concentrations in peat
up to 30 km from the smelter. Concentrations of SO/ and H+ in English bog waters also increased
with proximity to industrial, urban centers (Gorham, 1958b; Gorham and Detenbeck, 1986). Within six
months, Sphagnum moss transplanted from remote peatland sites in England to bogs near industrial
centers showed large increases in concentrations of N, Pb, and Fe in relation to transplants in more
remote areas (Ferguson el al., 1984; Woodin et al., 1985) and were attributed to high rates of
atmospheric deposition of these elements in the industrial regions. A host of studies have
documented correlations between element concentrations in surface moss or peat, and proximity to
industrial centers. Concentrations of S in living Sphagnum moss and lichens in Finnish peatiands are
linearly correlated with rates of S deposition and decrease from south to north (Pakarinen, 1981 a).
Similarly, elevated concentrations of S, Pb, and Hg were noted in Sphagnum from peatiands close to
point sources of these pollutants in the maritime provinces of Canada (Percy, 1983; Percy and
Borland, 1985). Concentrations of trace metals in surface moss from Finland (Pakarinen and Tolonen,
1976a,b), Canada (Glooschenko and Capobianco. 1978; Percy. 1983; Santelmann and Gorham, 1988),
and New England (Furr et al., 1979) have been shown to vary in relation to proximity to industrial or
automotive source areas. Similarly, concentrations of ash in peat from Minnesota bogs were found to
be highly correlated with distance from agricultural areas (Gorham and Tilton, 1978). The ability of
mosses and lichens to accumulate atmospheric pollutants is so well established that they have been
used widely as monitors of atmospheric deposition of trace metals (Ruhling and Tyler, 1970, 1984;
Pakarinen, I981b; Folkeson, 1981), synthetic organic chemicals (Thomas and Herrmann, 1980;
Thomas, 1983), and radionuclides (Gorham, 1958a; Persson, 1970; Daroczy et al., 1988).
Peat profiles that show correspondence between periods of high rates of accumulation of
pollutants and periods of high rates of atmospheric emissions of the pollutant constitute the third type
of evidence linking atmospheric deposition with the presence of contaminants in peatiands. Were
direct dumping or surface water discharges the source of pollutants to peatiands, different sites would
not be expected to show the same historical record of pollutant inputs, and this historical record
would not be expected to match the historical pattern of nation-wide or worldwide usage of the
chemical. Rapaport and Eisenreich (1986, 1988) have demonstrated that the historical patterns of
inputs of DDT, PCBs, toxaphene, HCB, and other pesticides are nearly identical in bogs across
northeastern North America; clearly atmospheric deposition is the common source for all of these
sites. Similarly, historical records of metal smelting are preserved in Pb profiles within English peat
bogs (Lee and Tallis, 1973; Uvett et al., 1979). Long-term records of emissions to the atmosphere
have been inferred from peat profiles in many locations; for example, Pb in Denmark (Aaby and
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Jacobsen. 1979); Pb, Zn, Cu. and magnetic minerals In Canada (Toionen and Otdfield. 1986: Zottai.
1988); Pb. Zn, and V In Maine (Norton, 1983; Norton and Kahl. 1987); Pb and Cd h south Germany
and Poland (Sapek, 1976; Kuster et al.. 1988); and magnetic minerals and trace metals In England.
Minnesota, and Finland (0
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effects Identified at present include changes in hydrology, metal speciation, nutrient retention, and
enhancement of erosion.
One adverse effect of acid deposition is the leaching of cations from surface peat. Urban and
Bayley (1986) and Bayley et al. (1986, 1987) have documented the increase In base cations in surface
bogwaters following experimental spraying of acid to the surface of a bog in the Canadian
Experimental Lakes Area. Approximately half of the acid applied to the minerotrophic. marginal area
of the peatland was neutralized by ion exchange or leaching of base cations from peat. Skiba et al.
(1989) have shown that the base saturation of surface peat throughout Scotland is inversely correlated
with rates of acid deposition and the loss of acid-sensitive lichens. They argue that the geographic
gradient in acid deposition has caused extensive leaching of base cations and acidification of surface
peats in southern Scotland. Bogs are oligotrophic. acidic environments that are vegetated by only the
few plant species capable of obtaining nutrients under acidic conditions. The abundance of
individuals, and the number of animal species is similarly restricted In these harsh environments
(Karns, 1984). In acidic environments the abundance of Ca and Mg is particularly important in
blocking the harmful effects of hydrogen ions and aluminum (Mount et al., 1988; Shortle and Smith,
1988). Thus in acidic bogs, leaching of cations from surface peat will further restrict the number of
species capable of inhabiting these ecosystems. Reductions in productivity, in rates of decomposition
and associated gas release, and in resilience of these species-impoverished systems also may
accompany the leaching of cations from bogs. More drastic are the changes likely to occur in poor
fens - peatlands with slightly higher pH and base saturation than bogs. Such changes have not been
studied yet, and will be discussed below under potential impacts.
Another documented effect of acid deposition on peatlands is the inhibition of plant and
bacteria enzyme systems. Press et al. (1985) observed that arylsulfatase activity of peat decreased
with increasing proximity to urban industrial centers in England and Wales. The low enzyme activity
in peat might be due to a number of components in the peat including sulfur compounds and trace
metals. However, peat transplanted to polluted sites and isolated from underlying peat also quickly
lost the capacity to cleave arylsutfate esters; this Inhibition was reproduced in the laboratory by
application of HSO,' at concentrations measured in rainfall at these sites (Ferguson and Lee, 1983a).
Jarvis et al. (1987) noted a decrease in arylsulfatase activity in surface peat relative to deeper peat
from a Virginia peatland. They also attributed the decline In activity in surface peat to inhibition by
acidic deposition. Changes In arylsulfatase activity in peat may lead to a number of changes In sulfur
cycling Including changes In reduced sulfur gas emissions and changes in suffate reduction and
methane production (Yavttt et ail., 1987; Giblin and Wieder, 1990). More significantly, they may
indicate that other microbial processes have been impacted by acidic deposition.
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High rates of sulfur and nitrogen deposition also cause reduced enzyme activity arid
subsequently reduced growth and loss of sensitive species in peatiand vegetation. It has been
demonstrated by several studies that high rates of N deposition suppress nltrate-reductase activity and
growth rates in Sphagnum (Woodin et al., 1985; Press et al., 1986; Rudolph and Voight, 1986; Woodin
and Lee, 1987). Although the increasing rates of N deposition in North America appear to stimulate
Sphagnum growth (Urban et at., 1989b), the much higher rates of N deposition in the Pennine
mountain region of England are adequate to inhibit growth of all Sphagnum species except
Sphagnum recurvum (Press and Lee, 1982; Ferguson et al., 1984; Press et al., 1986; Woodin and
Lee, 1987). Thus, It is felt that the species composition of English peatlands is determined In part by
rates of NO, emissions.
The deposition of sulfur compounds has had a drastic impact on peatlands in Great Britain
and in Canada. The toxicity of S02 and HS03" to mosses and lichens (Ferguson et al., 1978;
Ferguson and Lee. 1980; Aulio, 1984) has resulted in the disappearance of Sphagnum and lichen
species from large areas of Great Britain (Tallis, 1964; Ferguson and Lee, 1983b; Lee et al., 1987;
Looney and James, 1988). Although emissions of SO, have declined over the past forty years,
combined emissions and toxicity of SO, and NO, are sufficient to prevent recolonization by
Sphagnum (Lee et al., 1987; Woodin and Lee, 1987). Similarly, the combined exposure to SO2 and
trace metals has caused loss of Sphagnum from peatlands within a 15-km radius of the metal
smelters in Sudbury, Ontario (Gignac and Beckett, 1986). The effects of acidic deposition have not
ended with the loss of species. For example, the loss of Sphagnum carpets from blanket bogs in
England has resulted In accelerated erosion, decreased nutrient and water retention, and enhanced
mobilization and toxicity of trace metals (Lee et al., 1987; Helmer et al., 1990). Similarly, loss of
Sphagnum species has caused lowered water tables, accelerated decomposition and subsidence of
peat, and enhanced metal toxicity in peatlands surrounding smelters in Ontario. Acid deposition in
Britain provides the best documented example of the susceptibility and potential for widespread
damage to peatlands from air pollution. Damage has not been confined to changes in species
composition and changes in water quality, but extends to biotic impoverishment of large areas and
serious alteration of the functional and structural attributes of the affected peatlands.
Trace Metals
Documented effects of trace metal deposition in peatlands include plant toxicity, amelioration
of SO 2 toxicity, and Inhibition of enzyme activity in peat Only five species of vascular plants are
capable of growing in peatlands within 2 km of metal smelters in Sudbury, Ontario due to the
combined toxicity of Cu, Ni, and S02 (Gignac and Beckett, 1986). It Is likely that SO, is responsible
for the elimination of bryophytes within 15 km of these same smelters (Ferguson et al., 1978; Baxter
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et al., 1989a). However, the altered hydrology within the sites acts to exacerbate the metaf toxiclty,
and bryophytes are unable to fully colonize peatiands within 30 km of the Sudbury smelters (Gignac
and Beckett, 1986). Trace metals including Fe. Ma and Cu ameliorate the toxteity of bisutfate by
catalyzing the extracellular oxidation to SO4* (Baxter et al., 19895). This interaction among pollutants
enables Sphagnum species to grow in relatively more pollutod sites in central England than would be
possible in the absence of trace metal pollution. However, the trace metals are not entirely beneficial;
Press et al., (1985) demonstrated that concentrations of trace metals in severely polluted English bogs
were high enough to inhibit aryisutfatase activity in peat The catalytic oxidation of SO 2 by trace
metals also may be responsible for the ability of Sphagnum species to grow in moat areas around the
periphery of peatiands located 15 to 20 km from Sudbury smelters; concentrations of Fe and Mn are
highest in the moat and decrease toward the center of these peatiands (Gignac and Beckett, 1986).
Although many other potential effects of trace metals on peatiands may exist (see discussion below),
little research has been performed to document these Impacts.
Potential Effects
Acid Deposition
Research of the past 10 years has highlighted many potential effects of acid deposition on
peatiands in addition to those effects that have been dearly documented. The magnitude of impacts
depends on the intensity, aerial extent, and duration of acid deposition. In many peatiands, the
effects wfll likely be irreversible. Because ecosystems are not wholly self-contained but have
numerous biotic, hydrologic, and chemical links with other ecosystems, effects on peatiands may
result in serious indirect effects on other ecosystems as well. Although these remain conjectural, it is
important to consider such indirect effects in order to fully assess the risks and costs of air pollution.
The results of acid deposition are most easPy understood mechanistically by examining the
effects of the components of acid deposition: nitrogen, sulfur, and hydrogen ions. In small
quantities, addition of N as NO,* to peatiands will fertilize the vegetation (Verhoeven et al., 1988;
Urban et al., 1989b). Because many peatiands in North America are nitrogen-limited (Watt and
Heinselman, 1965; TBton, 1976), increasing N inputs not only may enhance the growth of existing
species, but may allow invasion by other species currently excluded by low N avaDabOity (Verhoeven
et al., 1983, 1988). In higher doses. NO,' wfll inhibit the growth of some plant species (Woodln et al.,
1985; Press et al., 1986; Rudolph and Vokjht, 1986; Woodin and Lee, 1987). When the capacity of
the plants for NO,' uptake is exceeded, increased fractions of NO,' inputs wDI be emitted as N2O or
N2 as a result of denttrtfication (Hemond, 1983; Koerselman et al., 1989; Urban et al., 1989b).
Increased emissions of N2O must be viewed as undesirable because of their contribution to global
warming and stratospheric ozone depletion. High rates of NO,* deposition are toxic to bog
vegetation (Ferguson et al., 1984; Rudolph and Voight, 1986; Lee et al., 1987).
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Effects of Increasing S deposition range from changes in S cycling to plant toxtetty and loss
of plant species. Low rates of S addition are likely to enhance rates of dissimBatory reduction (Behr.
1985; Urban et al., 1989c) resulting In Increased emissions of reduced S gases that become
reoxidized in the atmosphere and extend the geographic and temporal extent of acid deposition
(Nriagu et al,. 1987). The Increased rate of suKate reduction also leads to enhanced accumulation of
labile, reduced organic sulfur that is readly reoxidized In dry months; such reoxidation leads to
flushes of acidic water (Braekke, 1981) or to severe episodes of cation leaching from peatiands
(Bayley et al., 1986; Sheppard and Thibautt, 1988). Higher rates of S deposition are toxic to
bryophytes and lichens (Ferguson and Lee, 1980; Winner and Atkinson, 1987; Looney and James,
1988) and wOl cause the loss of Sphagnum and lichen species (Tallus, 1964; Ferguson et al., 1978;
Looney and James, 1988).
The effects of hydrogen ion deposition depend on the extent to which uptake of S and N
neutralize the acidity. Because uptake of S and N Is neither instantaneous nor 100% efficient
(Hemond. 1980; Urban and Bayley, 1986; Urban et al., 1987b, 1989b; Urban and Eisenreich, 1988),
much of the H+ input from acid deposition is not neutralized and may cause cation leaching, site
acidification, plant toxicity and species changes, trace metal mobilization and toxteity, decreased
decomposition rates, decreased emissions of CO 2 and CH4, changes in carbon storage rates,
changes in production and export of organic acids, and a host of changes in lakes and streams
receiving peatland runoff. Enhanced leaching of cations and decreased base saturation of peat as a
result of acid deposition have been documented (Urban and Bayley, 1986; Skiba et al., 1989). In fens
with only moderate base saturation, loss of nutrient cations is likely to result in a sequence of events
culminating in a large decrease In pH; major changes in plant species composition, nutrient cycling
and retention, rates of gas emissions; and ultimately, changes in hydrology and water quality that will
seriously impact streams and lakes (Gorham et al., 1984, 1987). The pH of fens is determined by the
balance between bicarbonate inputs from ground or surface waters and the production of organic
acids. The pH wDI remain above 5.5 as long as the supply of bicarbonate exceeds production of
organic acids, but wfll fall rapidly to about 4 as organic acids predominate over bicarbonate (Urban
and Eisenreich, 1989). Acid deposition may Induce this large, rapid pH change both by directly
titrating inputs of bicarbonate and by leaching cations from surface peat. The reduced base
saturation of the surface peat wOl favor rapid colonization by addophDic Sphagnum species (Clymo
and Hayward, 1982; Clymo, 1987; Gorham et al., 1987) that further acidify the site by decreasing
decomposition rates (Verhoeven et al., 1990), and thereby promoting accumulation of organic matter
that isolates the vegetation from base inputs in groundwater (Bellamy and Riely, 1967; Glime et al.,
1982; Wiicox et al., 1986). Eventually this buldup of organic matter is likely to result In decreased
water flow through the site and increased concentrations of organic acids. The reduced groundwater
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inputs and Increased organic acid concentrations cause further cation leaching as well as mobilization
of trace metals and phosphorus (Richardson, 1985; Wieder and Lang, 1986; Verhoeven et al., 1988).
The effects of add deposition on poor fens Is thus likely to greatly accelerate their natural conversion
to acidic bogs.
The effects of such acidification are not confined within peatlands. The rapid change in plant
and mterobial species may result In the complete disappearance of species unique to poor fens
(Gorham et al., 1984). The long-term implications of such btotic Impoverishment are not fully
understood, but may represent a loss of potentially valuable resources, and may preclude the
reversibility of the environmental changes (Gtgnac and Beckett, 1966; Gorham, 1988b). Bird, insect,
amphibian, and mammal species would be affected by habitat acidification (Karns, 1984). Fish,
Invertebrates, and algae In streams and lakes would be Impacted by increased concentrations of
organic acids and trace metals as well as altered water retention.
Trace Metals
Potential results of atmospheric deposition of trace metals to peatlands include loss of
sensitive species, declines in rates of primary production and decomposition, facilitation of the
methylation of select metals and uptake by biota, and mobBization of metals into streams and lakes.
The deleterious effects of high rates of trace metal deposition in the vicinity of metal mining and
smelting have been documented for peatlands, lakes, forests, and grasslands (Gorham and Gordon,
1963; Gignac and Beckett, 1986). Two unique attributes of peatlands render them both more
susceptible to harm from trace metal deposition and prone to facilitate the transfer of trace metals
into food chains and adjacent streams and lakes.
The first attribute is the high organic matter content and ion-exchange capacity of surface
plants (mosses and lichens) and soDs. Thus, peatlands have a large capacity to retain and
concentrate cationic metals. Trace metals sequestered from atmospheric deposition or surface waters
can build up to very high concentrations in peat; concentrations up to several percent have been
reported for Cu, Pd, and Fe (Fraser, 1961; Coker and DiLabio. 1979; Crerar et al., 1979). This
increases the likelihood of toxlclty to both plants and herbivorous animals.
The second attribute is the high concentration of organic adds in peatiand waters that act to
mobilize trace metals into the water. This phenomenon is most pronounced in acidic bogs where
concentrations of dissolved organic matter are much higher than in less acidic peatlands. The results
of this mobilization are to render metals more bioavaBable. enhance the toxiclty to plants, and
promote the leaching of trace metals from peatlands into streams and lakes. Although the metals that
are mobilized are largely compiexed by dissolved organic matter, they remain available and toxic to
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plants. Organically bound Al has been shown to be taken up to a greater extent than free AT3 and
to elicit similar toxic effects in tree seedlings (Arp and Ouimet, 1986; Hutchinson et al., 1986).
Sparling (1967) hypothesized that Al concentrations in peatiand waters in the British Isles regulated
the distribution of the rush, Schoenus nlgricans. Significant leaching of Pb. Cd, Al. Fe, and Hg from
peatlands into lakes and streams occurs because of high concentrations of dissolved organic matter
in peatiand runoff (Lodenius et al., 1983, 1987; Turner et al.. 1985; Urban et al., 1987b; Helmer et al.,
1990). The organic acids from bogs may Inhibit demethyiation of Hg in lakes and thus promote the
bioaccumulation of this toxic metal In acidic, brown-water lakes (Xun et at., 1987). Again, linkages
among peatlands and other ecosystems may be disrupted or may transfer the ecological damage of
air pollutants from peatlands to other sites.
Synthetic Organic Chemicals
Potential effects of atmospheric deposition of synthetic organic chemicals in peatlands include
inhibition of microblal activity, Inhibition of plant growth, declines or loss of populations of sensitive
plant and animal species, and incorporation of toxins into food chains resulting in stress to individuals
and populations at the top of the food chain. Although definitive proof of such effects is lacking
because of the dearth of research in this area, preliminary evidence gives reason for concern. The
vast wetlands in the Hudson and James Bay lowlands are a major breeding ground for numerous
species of waterfowl. High concentrations of Industrial and agricultural chemicals have been
measured in the peatlands and rivers of this remote region (McCrea and Wickware, 1986; McCrea and
Fischer, 1986). These chemicals can only have reached these sites through atmospheric transport
and deposition. Deposition is facilitated by the cold temperatures that lead to condensation of the
chemicals onto atmospheric particulates that are subsequently scavenged efficiently by precipitation
(Gregor and Gummer, 1989). High concentrations of organic contaminants In precipitation and wildlife
in the Arctic (Muir et al., 1988; Norstrom et at., 1988) attest to the potential for harm throughout
latitudes north of 50°.
The potential for ecological damage from agrochemicals also is significant in more southerly
regions of intense agricultural activity. Microcosm studies have shown that concentrations of
herbicides In runoff reaching midwestem wetlands are adequate to cause acute toxiclty to invertebrate
populations (Huckins et al., 1986). Nitrification is known to be Inhibited In fens and soils receiving
aerial drift of herbicides (Rajagopal and Rao. 1984). Significant aerial drift of agrochemicals into a
nontidal estuary also has been measured; Its effects remain under investigation (Howe, 1990). The
persistence of many of these compounds in peatiand environments (Rapaport and Eisenreich. 1988)
not only Increases the potential for chronic, long-term effects, but also may prolong the environmental
recycling of these compounds for many years to come (Eisenreich, 1987).
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RadionudkJes
The ability of mosses and lichens to accumulate and concentrate radionudides is well
documented (Gorham, 1958a, 1959; Persson, 1970; Daroczy et al., 1988; Henderson. 1988). Thus,
the potential exists for direct mutational effects to these species as well as for organisms consuming
these plants. The Laplanders and their reindeer are probably the most publicized populations at risk
(Persson, 1970).
CONCLUSIONS
There is ample documentation of damage caused to peatJands by atmospheric pollutants.
There Is an abundance of data suggesting that many other effects may exist, but little research has
been performed to document the impacts of many pollutants. Air pollutants are likely to induce not
only direct effects on peatiands, but also are likely to disrupt the Important linkages between
peatlands and adjacent ecosystems. The ramifications of many changes caused by pollutants are not
understood at present and may not be evident for some time. The probability and potential for
serious consequences of the existing widespread exposure of ecosystems to low levels of multiple
contaminants demands greater research and regulatory attention.
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USE OF THE PHYTOTOX DATABASE TO ESTIMATE THE INFLUENCE
OF HERBICIDE DRIFT ON NATURAL HABITATS IN AGROECOSYSTEMS
James E. Nellessen and John S. Retcher
Dept. of Botany and Microbiology
University of Oklahoma
Norman. OK
INTRODUCTION
In agroecosystems, there exists a patchwork of row crops intermixed with pasture and natural
plant communities. The extensive use of herbicides in various agroecosystems across the United
States poses a potential threat to vegetation growing on adjacent land If the chemicals applied to
cropland Inadvertently drift onto nontarget areas. Adverse consequences of such an event could be:
(1) reduced production of plant biomass (yield), and /or (2) change in the species composition
of the nontarget plant community. Whether or not a plant community is affected in either or both of
these ways is dependent on two factors: (1) the nature of the exposure and (2) the sensitivities of the
plant species within the nontarget community.
Estimating chemical exposure of nontarget plants depends on several variables: (1) the
chemical gradient (concentration vs. distance) extending across the nontarget zone, (2) duration of
exposure, and (3) frequency of exposure. The magnitude of each of these variables depends in turn
on the chemical and physical properties of the compound, mode of application, and prevailing
weather conditions. Because all of these parameters can be quantified, the process of predicting the
amount of drift and subsequent exposure of nontarget vegetation lends Itself to mathematical
modeling. A variety of distinctive models have been developed by various Investigators to deal with
102
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different combinations of pesticide application (plane vs. tractor), chemical features (liquid vs. dust),
and mode of drift (direct from applicator, or Indirect following volatilization from field).
The response of nontarget vegetation to pesticide drift has received less attention than
development of the drift models. Some models appear to have never been validated by
biomonltoring (Thompson, 1983), and others have only been validated by examining cultivated crops
(Yates et al., 1978). There are only a few isolated reports on how herbicides Influenced the
productivity or composition of native plant communities (Malone, 1972; Marrs, 1985; Gillen et al.,
1987; Marrs et al., 1989; Swindell et al., 1989). Only one of the reports was conducted in connection
with drift modeling, one dealt with a tree community, and none considered the influence of indirect
(field volatilization) drift.
In the absence of such studies, an alternative is to use dose-response data taken from the
literature for individual plant species that are known to be present in natural plant communities
frequently found in pesticide treated agroecosystems. The dose-response data compiled in the
PHYTOTOX database (Royce et al., 1984) is ideally suited for predicting potential hazards posed by
pesticide drift to nontarget vegetation. In this pilot study we used a portion of PHYTOTOX to predict
the potential hazard posed by the volatilization and drift of trffluralin and alachlor on forest
communities in Illinois.
MATERIALS AND METHODS
The PHYTOTOX database is a computerized information resource that permits the rapid
retrieval and comparison of data pertaining to the response of terrestrial plants to the application of
organic chemicals (Royce et al., 1984). As of January 1, 1990, PHYTOTOX possessed information on
approximately 8,000 different chemicals, 2,000 species, and 50 plant responses. The data has been
compiled from over 3,500 articles published between 1926 to 1988. The database has two files: a
Bibliographic file and an Effects file. The Bibliographic file possesses information on each paper that
has been used as a source of data for compfling effects records. The Effects file contains
approximately 100,000 records. The Effects file differs from most biological databases, because it
contains quantitative numerical data pertaining to chemical doses, plant responses, and experimental
parameters. Each record in the Effects file contains information concerning the effect(s) of one dose
of a single chemical applied to a particular plant species as reported in one publication. The
Information associated with each record is organized under labels that may be sorted separately
during computer searches.
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Trifluralin and alachlor were examined in this study, because of their high usage orrcorn and
soybeans in Illinois (Gianessi and Puffen, 1988}), and their strong tendency to volatilize from the soil
(Glotfetty et al., 1984, 1989). Attention was focused on the oak-hickory community because small
wood lots meeting this description are interdispersed among com and soybean fields in Illinois
(Kuchler, 1964). A hierarchal search of PHYTOTOX was conducted to recover data pertaining to oak-
hickory community plants treated with comparative doses (kg/ha) of either trifluralin or alachlor.
RESULTS
PHYTOTOX possessed information on 100 species which occur in oak-hickory communities.
This number was reduced to eight when only records pertaining to trifluralin and alachlor were
considered (Table 1). The plant list was expanded to include data on an additional 10 species that
are in the same genera as plants found in oak-hickory communities. Justification for expanding the
list in this manner comes from previous analyses where It was shown that species within the same
genus had a high correlation of response to the same chemical (Fletcher et al., 1990).
Examination of the data in Table 1 shows that different species vary widely in their
sensitivities, and a single species will respond differently to different herbicides. For example, while
Senecio vulgaris experienced 100% control (kill) when treated with 1.9 kg/ha alachlor, It was not
affected at all by a somewhat lower dose of trifluralin. This variable response by different plant
species to chemicals has been capitalized on in selectively killing unwanted plants in cultivated fields.
This feature also has the potential of reducing plant biodiversity in nontarget areas If the amounts of
drifting chemicals reach the inhibitory level for sensitive plants.
Trifluralin is an extremely volatile herbicide. It has been shown that 90% of the trifluralin
applied to moist soil will be lost to the air in 2 to 7 days following application (Glotfelty et al., 1984).
These investigators established the maximum rate of volatilization to be 195 g/ha/h and measured air
concentrations as high as 40 pg/m3 (0.003 ppm) at 50 cm above the son surface. Based on these
data It can be hypothesized that If 2.8 kg/ha is applied to a cultivated field It is possible that moving
an air mass could transport 2.5 kg/ha to adjacent nontarget plants over a 2 to 7 day period. When
this level of exposure is compared to dose-response data in Table 1, it was found that eight different
genera (Acer, Cassia, Dioscorea, Ilex, Rhododendron, Solatium, Thuja, and Urtica) had sensitivities
tow enough to be affected. Among these genera the species Urtica chamaedryoides is on the Illinois
endangered and threatened species list.
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TABLE 1. DOSE-RESPONSE DATA FROM PHYTOTOX FOR TWO HERBICIDES COMMONLY USED
IN ILLINOIS AND TAXA FROM GENERA KNOWN TO OCCUR IN ILLINOIS WOODLOTS.
Herbicide
Trifluralin
Plant Species
Acer palmatum"
Cassia obtusifolia"
Dioscores sp.
Euonymus fortune! b
Ilex cornuta"
Ilex crenata"
Juniperus horizontalis
Pinus echinata
Pinus strobus
Rhododendron obtusumb
Sedum brevifoliumb
Senecio vulgarisb
Solanum nigrum
Solanum sp.
Thuja occidental
Urtica urensb
Dose
(kg/ha)
3.1
0.9
0.2
8.9
2.5
1.1
4.5
2.5
1.1
0.6
0.6
0.6
2.5
Response*
(%)
20 injury
81 control
3 DMD
16 RT FMD
18 LF CHL
None
None
35TRD
None
15 control
35 control
3TRD
100 kill
Alachlor
Dose
(kg/ha)
6.2
1.7
20.0
12.5
13.6
6.0
2.5
1.9
13.6
1.9
Response"
(%)
injury
23 control
49 DMI
6SZI
None
9 injury
15 injury
100 control
None
100 control
* All responses refer to whole plants unless indicated otherwise. Control = similar in meaning to
plant kill, DMD = dry mass decrease, DMI = dry mass increase, LF CHL = leaf chlorosis, RT FMD
= root fresh mass decrease, SZI = size increase, TRD = transpiration decrease.
b Taxa that do not occur in Illinois woodlots but are members of genera that can be found in Illinois
woodlots.
Alachlor has a lower field volatility than trifluralin, but even at the reduced rate of 8.1 g/ha/h it
has been shown that 19% of applied doses (420 g/ha) were lost In 21 days (Glotfelty et a!., 1989). If
weather conditions caused this amount of volatilized alachlor to be redeposlted on vegetation adjacent
to an applied field It could influence the growth of three genera listed In Table 1. This level of
alachlor exposure represents 25% of a dose giving 23% control of Cassia obtusifolia and 22% of a
dose giving 100% control of Urtica urens and Senecio vulgaris (Table 1). Two species of Cassia,
three of Senecio, and the aforementioned U. chamaedryoides growing in Illinois woodlots could be
affected.
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CONCLUSIONS
A comparison between maximum drift values estimated from the literature with dose-response
data taken from PHYTOTOX indicated that some nontarget species growing in oak-hickory
communities could be influenced by the drift of trifluralin and/or alachlor. The most sensitive plants
were Cassia and Urtica. Although we are not aware of any field measurements or biomonltoring data
collected in native plant communities which would substantiate this prediction, there is a report by
Behrens and Lueschen (1979), where 0.28 kg/ha of dicamba applied to a corn field affected soybean
growth 60 m downwind In an adjacent field. Such data certainly Is.cause for concern and suggests
that herbicide drift may have a profound influence on the productivity and composition of natural plant
communities.
The biota of the United States has been described as having approximately 116 different
native plant communities (Kuchler, 1964). The geographical location and species composition of each
of these communities is known. Therefore, It is possible to identify natural plant communities that are
in association with various crops in different agroecosystems located throughout the United States. In
this pilot study with PHYTOTOX we considered only one plant community (the oak-hickory forest) and
only two herbicides (trifluralin and alachlor). Similar analyses could be conducted for all 116 plant
communities in the United States and also for an extended list of herbicides.
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BIOLOGICAL MARKERS IN ANIMALS AND PLANTS TO ESTABLISH
EXPOSURE TO, AND EFFECTS OF, ATMOSPHERIC TOXICANTS
John F. McCarthy and Timothy J. Tschaplinski
Environmental Sciences Division
Oak Ridge National Laboratory
Oak Ridge, TN
ABSTRACT
Evaluation of the potential for exposure to toxicants transported through the atmosphere is
extremely complex due to the diversity of possible routes of exposure, differences in bioavailability,
and the complexity of molecular, biochemical, and lexicological interactions within exposed
organisms. This complexity limits our capabilities to either quantify or assess the significance of
exposure to atmospheric toxicants. Exposure cannot be readily quantified by measuring body
burdens of contaminants because many deleterious chemicals are rapidly metabolized. Furthermore,
the relationship between body burden and toxic response is complex and not fully understood.
Assessing the significance of exposure to complex mixtures of chemicals is even more problematic.
Chemical analyses are difficult and expensive; furthermore, possible synergistic or antagonistic
interactions within biota can invalidate predictions based on toxicity of Individual chemicals.
The overall objective of our research Is to develop and validate the concept of using animals
and plant biomarkers as indicators of bioavailable contaminants. Evidence of exposure in animals or
plants provides a temporally Integrated measure of bioavailable contaminant levels, and is therefore
much more relevant to the potential risk to health or the environment than is the analytically
measurable concentration of contaminants in the soil, water, or air.
Our approach is to measure biomarkers (biochemical or molecular indicators of exposure) in
environmental species as sensitive, biologically relevant indicators of toxicant exposure. Biomarkers
under study include several measures of either specific damage to DNA (e.g., adducts) or non-specific
damage to the integrity of the genetic material (e.g., DNA strand breaks), induction of several
detoxification systems (mixed function oxidase system, metallothionein, and oxyradical scavenging
enzymes), as well as biomarkers of reproductive competence. The qualitative pattern and quantitive
response of a suite of biomarkers offers the potential of indicating the extent of exposure to
atmospheric pollutants, and the magnitude of the toxic effects from that exposure. Challenges that
107
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must be resolved before applying this approach to regional studies include establishing an acceptable
reference level of biomarker response in "unexposed" animals or plants, and understanding the effect
of environmental and physiological variables on biomarker responses of Individual organisms.
INTRODUCTION
The U.S. Environmental Protection Agency is mandated to protect human health and the
ecological integrity of the environment from unreasonable risks associated with exposure to pollutants.
Although existing regulations are targeted at a wide range of possible sources of contaminant input to
the environment, there is an increasing concern about the effects of the plethora of anthropogenic
chemicals capable of being transported long distances through the atmosphere. These chemicals
include trace metals, industrial organic compounds, including polycydic aromatic hydrocarbons
(PAHs) and polychlorinated biphenyls (PCBs), as well as chlorinated pesticides, and gaseous
toxicants; such as SO2, NO,, and 03. The effects of chronic deposition of airborne toxic chemicals
on the integrity of terrestrial and aquatic ecosystems is not known, especially with respect to the
potential interactions of these toxicants with other environmental or physiological stresses.
This paper will discuss an approach that can contribute to evaluating the exposure of
organisms to airborne toxicants, and can be particularly valuable as a tool to evaluate the biological
significance of that exposure (NRC, 1987). The approach is based on biological monitoring of
animals and plants in areas impacted by airborne toxicants, and, more specifically, on measurements
of molecular, biochemical, and physiological biological markers (biomarkers) In target species. In this
context, the measurements of body burdens of persistent compounds (or metabolites) and of
biomarkers of exposure or effects permit the animals and plants to serve as biomarkers indicating the
presence of bioavallable contaminants.
Biological monitoring has a number of advantages that make it an informative and necessary
adjunct to measurements of the concentrations of chemicals in the environment. Chemical
measurements of environmental media are specific, quantitative, and exquisitely sensitive and precise.
However, the biological significance of the chemical concentrations measured in air, water, soil, or
food is not at all dear. We understand the toxic action of only a few of the thousands of chemicals
in the environment and have almost no Information on the toxidty of complex mixtures of chemicals
or on the role of environmental stresses on an organism's susceptibility to toxic exposure.
Furthermore, a chemical survey is a snapshot in time and space. Variations in concentrations over
time resulting from intermittent releases of effluents by industries, or from storm events, changes in
108
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winds, and such, cannot be accounted for without repeated analyses. Spatial patchiness of
contaminant patterns also requires extensive and expensive sampling and chemical analyses.
One approach to addressing these problems is to monitor animals and plants in the
environment as sentinels of environmental contamination. Evidence of exposure in sentinel species
provides a temporally integrated measure of bioavaBable contaminant levels and is therefore much
more relevant to the potential risk to the environment than is the analytically measurable concentration
of contaminants in the soil, water, or air. Measurements of body burdens of persistent compounds
provides a direct measure of compound uptake. Measurements of biomarkers provide the following
information that cannot be obtained from direct measurement of body burdens and that is very
relevant to evaluating the potential biological and ecological effects of toxicants:
o Provide evidence of exposure to compounds that do not bioaccumulate, or are rapidly
metabolized and eliminated.
o Integrate the toxicologies! and pharmacokinetic interactions resulting from exposure to
complex mixtures of contaminants, and present a biologically relevant measure of toxicant
interactions at target tissues and the cumulative adverse effect of the exposure.
o Give quantitative measures of adverse effects in sentinel species, and are therefore
relevant to understanding the relationship between exposure to environmental levels of
airborne contaminants, and the potential for adverse health and ecological effects of air
toxicants.
This paper will describe several biomarkers of exposure and effect as they have been applied
to assessing exposure of animal populations. We will then describe the use and potential application
of several biomarkers of airborne toxicants in plants. Finally, the advantages and difficulties in
applying this approach to assessing the effects of airborne toxicants will be evaluated.
BIOMARKERS OF EXPOSURE AND EFFECTS IN ANIMALS
Selection of specific biomarkers wQI depend on the physical and chemical properties of the
toxic chemical, the depth of current understanding of its mode of toxic action, the metabolic
capabilities of the sentinel species, and the objective of the monitoring effort (e.g., is the emphasis on
assessment of exposure or effects). In all cases, the response of animals from areas of suspected
contamination must be compared to those of organisms from ecologically comparable reference sites
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that are known to be free of anthropogenic toxicants. Only a minimal number of animals are
sacrificed to carry out the chemical analysis. The concept of biomarkers is llustrated in Figure 1.
Several classes of biomarkers have been used for environmental monitoring and offer promise as
sensitive measures of exposure and informative measures of effect.
Biomarkers of Exposure
Biomarkers of exposure are indicators of an exogenous substance within an organism
(including body burden of parent compound or metabolites), interactive products formed between the
chemical and endogenous components, or an event in the biological system that can be directly
related to exposure (NRC, 1987). Several categories of biomarkers of exposure are described
genetically In the following discussion.
Biomarkers of Genetic Damage
These can include both measures of damage from specific chemical agents or can be
nonspecific indicators of damage to the integrity of the DNA. The latter biomarkers are especially
useful because they are a measure not only of agents that can directly damage DNA, but they can
also assay the activity and fidelity of the mechanisms responsible for proof-reading and repairing DNA:
(1) DNA adducts - The metabolism of chemicals may result in the production of highly
reactive electrophllic metabolites that can undergo attack by nudeophilic centers in
macromolecules such as DNA, RNA, or proteins. The amount of the reaction product
(adducts) is proportional to the in vivo concentration of the electrophile and length of the
exposure. Therefore, the amount of metabolite bound to cellular DNA (in vivo dose) provides
a reliable dosimetric basis upon which to assess exposure to a genotoxic compound (Table 1;
Perera et al., 1982; Shugart et al., 1987; Shugart et al., 1989).
(2) DNA strand breaks - Certain genotoxic compounds or agents such as metals,
radionudides, and some chemicals do not covalentiy bind to DNA, but nevertheless induce
damage. If this damage is expressed as single-strand breaks (or the potential for single-
strand breaks), It can be detected by measuring the rate of alkaline-Induced strand separation
(Rydberg, 1975; Shugart, 1988; Shugart et al., 1989). This technique provides a general
measure of DNA damage because sites modified by adducts may also be alkaline labile
(Figure 2; Kanter and Schwartz, 1982; Daniel et al., 1985).
(3) Cytogenetic damage - Chemical and physical agents that are genotoxic can produce
damage to DNA that propagate into morphologically detectable chromosomal aberrations.
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Biomarkers: The Concept
Fish Kill
a
E
"c
Decreased Reproduction
Disease
Cancer
Behavioral Changes
Detoxifying
Systems
Immunological
Changes
DNA/Protein
Modification
Toxic Exposure
Concentration x Time
FIGURE 1. The Biomarker concept. The response of an animal to toxicant exposure is a function
of the concentration of a chemical in the environment and the length of time the animal is exposed.
We would like to avoid the long-term, irreversible adverse effects indicated in the upper right hand
portion of the figure. The btomarker approach seeks to measure early responses indicated in the
boxes at the lower left of the figure. These sensitive biological markers Indicate that the animal has
been exposed to chemicals in the environment and provide an early warning of future effects.
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TABLE 1. Benzo[a]pyrene adduct formation in DNA isolated from brain tissue of beluga whales*. The
whale population from the St. Lawrence River in Canada is declining and there have been several
beached whales with bladder tumors. The area from which the whales were collected is known to
have measurable levels of PCBs and PAHs, including benzo[a]pyrene. The data below quantify
"adducts" [the amount of a carcinogen, such as benzo[a]pyrene (B[a]P), chemically attached to DNA].
DNA from the affected population of beluga whales from the St. Lawrence River has been modified by
B[a]P at levels as high as those observed in mice and fish experimentally exposed to a carcinogenic
dose of the chemical; belugas from a pristine area in Canada's Northwest Territories had no
detectable adducts (Martineau et al., 1988).
B[a]P Adduct Formation
Sample Tissue Binding b Level0
St. Lawrence Estuary
#1
#2
#3
MacKenzie Estuary
#1-4
#1-4
Brain
Brain
Brain
Brain
Liver
206
94
69
2.15
0.98
0.73
None detected
None detected
DNA isolation and quantitation was according to Shugart et al. (1983)
B[a]PDE-DNA adducts expressed as nanograms of tetrd N (resulting from binding to DNA of the
anti-B[a]PDE metabolite of benzo[a]pvrene per gram of DNA (see Shugart et al.,1983).
Level expressed as number of B[a]PDE-DNA adducts per 107 DNA nudeotides.
The cytogenetic assays currently in use represent a highly diverse group of tests which include
measures of DNA repair processes, mitotic recombination, sister-chromatid exchange, and such.
Petrochemical-related DNA damage in wild rodents has been detected by these methods (McBee and
Bickham, 1988).
Protein adducts
Hemoglobin has been proposed as an alternative cellular macromolecule to DNA for
estimating the in vivo dose of chemicals subsequent to exposure (Osterman-Golkar et al., 1976;
Calleman, 1984; Shugart and Kao, 1985; Shugart, 1985a, b; Shugart, 1986) because It has reactive
nudeophilic sites that form stable reaction products with electrophilic agents; no mutagenic or cancer-
initiating compound has failed to produce covalent reaction products with hemoglobin. This approach
has been used to monitor exposure in animals from a floodplain in Oak Ridge, TN, contaminated with
PAH, and has demonstrated that the concentration of hemoglobin adducts of benzo[a]pyrene were
highest in animals with the greatest contact with the contaminated son (Loar et al., 1987b).
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9 4
c
=>
Is
0)
"5
Ł 2
re
Ł
ffi
"c
0*
• Bluegill Sunfish
• Fathead Minnow
J_
8 12
Exposure Period (days)
16
20
FIGURE 2. Time course of genetic damage, measured as the number of single-strand breaks in DNA
through the use of an alkaline unwinding assay, in fish exposed to the carcinogen benzo[a]pyrene
(BaP) in the laboratory. BluegHI sunfish and fathead minnows were exposed under flow-through
conditions to an aqueous solution of BaP (I ug/L) for 20 d. Results are plotted as the number of
breaks per alkaline unwinding unit, in the manner of Rydberg (1975) and Shugart (1988).
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Induction of Detoxication Systems as Biomarkers
Most contaminants stimulate synthesis of protective detoxication systems. Higher levels of
these proteins are diagnostic of a molecular response to toxicant exposure. Several classes of
detoxication systems, each responding to specific classes of toxicants, are available as biomarkers:
(1) Mixed Function OxkJase System - The oxidation of many xenobiotic organic compounds
is catalyzed by a group of hepatic cellular monooxygenases. The biochemical system
responsible for these oxidation and conjugation reactions are slmDar In vertebrates ranging
from fish to mammals (Bend and James, 1979; Chambers and Yarbrough, 1976). Because
chronic exposures to many organic contaminants induce the hepatic MFO system In fish
(Leek et al., 1982; Price-Haughey et al., 1986; Jimenez et al.. 1987; Jemenez and Burtis, 1989;
Loar. I987a,b) and mammals (AJvares et al., 1967; Sladek and Mannering, 1966; Lu and West,
1980), the levels of components of this system provides evidence of exposure and effects of
many compounds that are rapidly biotransformed by the detoxication system (Figure 3;
Payne. 1984).
(2) Concentrations of Metal-Binding Proteins (Metallothionein) - Almost all organisms possess
low molecular weight proteins capable of binding toxic metals such as cadmium, zinc, or
copper, and rendering them unavailable for toxic actions within cells. The levels of these
proteins, including the cysteine-rich metallothionein, are induced in many organisms
chronically exposed to toxic metals in the environment (Kojima and Kagi, 1978; Webb, 1979;
Mitane et al., 1986; Hamilton and Mehrle, 1986; Harrison et at., 1987). Quantification of metal-
binding proteins is a promising biomarker for exposure to a range of heavy metal
contaminants, such as cadmium and zinc (Table 2).
(3) Oxyradical Generation and OxidatK/e Stress • A large number of diverse organic
compounds can undergo redox cycling (Mason and Chignell, 1982; Kappus, 1987) In which
the parent compound first undergoes a one-electron reduction to form an organic radical, with
the electron subsequently donated to molecular oxygen, giving rise to 02'. Superoxide may
have deleterious effects, or may give rise to other toxic oxyradicals - H2O3 and OH - that
have been associated with a number of toxic effects, Including altered redox status, llpld
peroxiolation, protein oxidation, and damage to DMA (Fridovich, 1983; Halliwell and Gutteridge,
1984; Kappus, 1987). Additionally, oxidative stress can invoke adaptive responses, such as
inductions of antioxidant enzymes (e.g., superoxide dismutase, catalase, and glutathione
peroxidase), which can be used as biomarkers. For example, Induction of superoxide
dismutase has been demonstrated in fish exposed to a range of organic compounds
(Washbum and DiGiulio, 1989; Mather-Mihaich and DiGiulio, 1986).
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500
^400
1 300
o
D.
•|<200
I
O
o
Ł 100
1 I i i i i I I I I I I i i i i i r
(5)-
(6)
(3)
(5)
(6)
Combined Mean for Control Streams (9)
i i I I I I I I I I I I I I I I I I t
123456789
11 12 13 14 15 16 17 18 19
•Industrial Outfall
Distance Downstream (km)
Municipal Outfall
FIGURE 3. Induction of detoxication enzymes in fish. Fish in a stream contaminated by industrial
effluents demonstrate significantly higher levels of a detoxication system, the mixed function oxidase
system. The levels of one enzyme activity of this system, ethoxyresofurin-)-deethytase (EROD), is
measured in depatic microsomes of Huegill surrfish collected In East Fork Poplar Creek (Oak Ridge,
Tennessee) at different distances downstream from an Industrial source (New Hope Pond) and
compared to EROD levels in fish collected at the same time from Brushy Fork Creek, an unpolluted
reference stream. Each point represents the mean +/- S.E.M. (Q) for each station. Significant
differences (P <0.01) between the combined mean ( )+• SEM ( ) for the reference
stream f control stream*) and for the stations in the stream receiving the industrial effluents are
indicated by asterisks (see Jimenez et a!., 1988).
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TABLE 2. Metal-binding proteins (MBP) in the livers of biuegill sunfish (Lepomls machrochirus)
exposed to cadmium in the laboratory and in redbreast sunfish collected from streams contaminated
with metals, including zinc. Fish were injected intreperitoneally with 2 mg Cd/kg body weight as
CdCI2 in 0.9% saline on each of three consecutive days and then sacrificed on Day 6. Control fish
were similarly injected with 0.9% saline (50uL/100 g body weight) on each of three consecutive days
and sacrificed on Day 6. White Oak Creek drains the southern boundary of the Oak Ridge National
Laboratory and flows Into White Oak Lake. Brushy Fork Creek is used as a reference site because it
has no inputs of industrial or other point source pollutants. MBP concentrations in livers were
measured by the use of the Cheiex-100/10td procedure (Sloop, personal communication), and are
reported as nanomoles of ictd bound per gram of soluble protein ± S.E.M.
Sample Description MBP Concentration
Laboratory Exposure (cadmium)
Control 247 ± 168 (n = 4)
Exposed 977 ± 133 (Q. = 3)'
Field Collection
Reference Stream
(Brushy Fork Creek) 631 ± 121 (n = 4)
White Oak Creek
(3.03-3.4 km) 1900 ± 306 (n - 10)*
White Oak Creek
(2.5 km) 1732 ± 368 (n = 6)'
White Oak Lake 1472 ± 404 (n = 4)'
* MBP concentrations are significantly different from those in control or reference animals at alpha =
0.05.
Biomarkers of Toxins with Specific Modes of Action
Certain compounds exert their toxic action through specific and well understood mechanisms.
Measurements of these key lexicological endpoints can be interpreted as biomarkers. For example,
neurotoxins such as pesticides act by inhibiting the activity of the enzyme, acetylcholinesterase.
Reduction in the activity of this enzyme In plasma or in brain tissue is a biomarker of exposure to this
class of toxicants. Similarly, the activity of the enzyme aminolevulinic acid dehydrase. an important
component in the porphyrin biosynthetic pathway, is Inhibited by heavy metals, and is particularly
diagnostic of toxic exposure to lead. Many other examples of potential biomarkers can be selected,
depending on the potential toxicants of concern at a particular site.
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Biomarkers of Effects
The division between biomarkers of exposure and of effects is somewhat arbitrary. For
example, depression of acetylcholinesterase levels Is a biomarker of exposure to a specific class of
toxicants, but also Indicates the magnitude of the neurotoxic effect of the exposure. In the context of
this discussion, a measurement of impaired function at some level of biological organization is
considered a "biomarker of effect" that may or may not be attributable to the action of chemical
toxicants. Attributing the adverse effect to toxicant exposure wOl generally require that the response
toxicants be propagated to successively higher levels so as to form a chain of logical, albeit
circumstantial, evidence that the 'effect* Is causally related to the exposure. Several different biomarkers
of effects classes have been examined in animals and found to be reliable for environmental monitoring.
Biomarkers of Impaired Reproductive Competence
Reproductive capacity is the key process linking exposure and effects of toxic materials within
an individual to a population-level consequence. Gonads can be examined to determine parameters
such as abnormalities in sperm, numbers of oocytes recruited, and abnormalities in cocyte
development quantified as numbers of atretic (dying) oocytes. Effects on the reproductive potential of
small mammals can be assessed through counts of corpora lutea, embryos, or placenta! scars.
Effects on avian reproduction can be assessed by nesting success surveys. Reproductive condition
of males can be assessed by examining testicular size, which is Indicative of sexual maturity and
reproductive activity. Reproductive condition of females can be assessed on the basis of whether
they are mature or immature, lactating or not, in estrus or not, and whether they are pregnant.
Biomarkers of Impaired Organ or Tissue Function
A number of easily measured parameters can Indicate in quantitative terms the physiological
and bioenergetic status of the animals. Organ indices (e.g., ratios of the weight of liver and gonads
to body weight) and condition indices (ratios of length and width parameters in animals that indicate
the 'plumpness* of the individual), as well as levels of serum and body lipids and triglycerides all
provide useful information on the energy stores of the animals and mobilization of those stores for
physiological functions. Serum enzyme levels, including transaminases (SGOT, SGPT), lactic
dehydrogenase, and sorbltol dehydrase levels Indicate cytotoxic damage, which results in release of
these enzymes into the blood. Histopathologlcal analyses of tissues, especially liver, provide evidence
of neoplastic, necrotic, or parasitic lesions. These data are not only biomarkers of the effects, but
also can aid in interpretation of biomarkers of exposure. For example, contaminant-induced changes
in liver enzyme activities can be seriously impaired by the hepatotoxic damage resulting from parasitic
or necrotic lesions (McCarthy el al., 1989). Information on serum enzymes or histological evidence of
liver damage can help In interpreting the significance of indicators of exposure.
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Indicators of General Health Status and Population Age Structure
In addition to information from biomarkers that indicate both exposure and the effects of
exposure, a significant body of information can be gathered from observations on the health and age
of animals trapped for the population and community studies. For example, gross anatomical
examination can identify lesions, tumors, necrosis, liver size, and reproductive condition, and will help
determine whether tissues should be subjected to histopathological examination to confirm the
presence of tumors or neoplasia. The age structure of the animal populations from different sites can
be assessed by a number of observations. Growth rings in fish scales are an accurate indicator of
age. In mammals, cranial measurements (degree of fusion of cranial sutures, tooth eruption in young
animals, and amount of tooth wear in older animals) provide an indication of relative and absolute
age. The weight of the eye lens increases with age. The length of the animal (length from snout to
dorsal tip of tail, and length of foot from outside of heel to tip of toes), as well as the overall size of
the animal also can be related to age (less mature animals are smaller). The pelage (coat color and
condition) is another Indication of health. These analyses should provide evidence of premature
mortality in populations exposed to contaminants.
BIOMARKERS OF EXPOSURE AND EFFECTS IN PLANTS
The general conceptual approach described for applying biomarker-based monitoring to
evaluate the effects of toxic chemicals in animals can be readily adapted to plants (Figure 4).
Biomarkers can be defined at different levels of resolution from molecular, cellular, organismal,
species, through ecosystem level responses. As with animal systems, there is particular Interest in
biochemical indicators that may have predictive value In that they precede visual, advanced damage,
including mortality, leaf abscission, visible foliar lesions, factors associated with secondary stresses,
growth decline, and breakdown of cell uttrastructure (Figure 4). A potential biomarker in plant
systems should meet several criteria including (1) rapid and unambiguous response to low levels of
toxicants, (2) specific responses for a particular toxicant, (3) simplicity of the measurement, and
(4) reproducible results (Amdt, 1970; Darrall and Jager, 1984). A major limitation in plant biomarkers
is the lack of specificity In biomarker responses, because several stresses that differ greatly in nature
may elicit similar responses.
A suite of markers may, however, prove useful in pinpointing the nature and impact of stress
at an affected site. For example, biomarkers that may prove useful as indicators of exposure include
metabolite pool size and turnover, detoxifying systems, and DNA/protein modification (Figure 4). The
numerous photosynthetic processes, with the exception of pigment concentrations (considered under
the metabolite section), may be useful as biomarkers of general effects, but these processes are
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Biomarkers: The Concept
ID
Premature Leaf Abscission <
Decreased Growth •
Ultrastructure Breakdown
Decreased
Photosynthesis
Toxic Exposure
Tree Mortality
Visible
Foliar Lesions
Secondary Stress
Pathogens
Alteration in
Metabolites
Membrane
Modification
DNA/Protein
Modification
Concentration x Time
FIGURE 4. The biomarker concept as It pertains to plants. The concept is analogous to that
described in Figure 1, but reflects key responses of plants to toxicant exposure.
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affected by such a variety of environmental and physiological factors that they may be of only
marginal utility as indicators of effects that can be associated with pollutants. Photosynthetic
processes that could be considered include changes in electron transport, photophosphorylation,
fluorescence characteristics, carbon assimilation, partitioning, and allocation. The same lack of
specificity also limits the use of membrane modifications as biomarkers, including Ion transport,
oxidative electron transport, and breakdown of compartmentation itself.
Most data collected on biomarkers of stress relate to heavy metals and gaseous pollutants,
including O, NO, S0a HF, and peroxyacyl nitrate (PAN). Because there is a better understanding of
plant responses to toxic gases, this discussion wOl largely focus on the biomarkers of gaseous
pollutants, particularly those that are often used, or that appear to offer the greatest potential as
indicators of exposure and effects in plants. It must be noted, however, that environmental factors
such as temperature, nutrient, and water stress can confound the Interpretation of biomarkers.
Several detailed reviews report a broad array of biochemical variables that have been correlated with
stress (e.g., Koziol and Whatley, 1984; Malhotra and Khan, 1984).
DNA/Protein Modification
Numerous enzymes are impacted directly by environmental toxicants. The primary effect can
be a decrease in enzyme activity, but in time protein degradation and synthesis are also impacted by
altered DNA transcription.
(1) Ribulose-l,5-Bisphosphate Carboxylase/Oxygenase (RUBISCO) - The activity of the main
carboxylating enzyme in C3 plants, (RUBISCO), has often been targeted as a biomarker. The
inhibition of the rate of net photosynthesis (Pn) by S02 has been attributed to reduced activity
of RUBISCO in Spinacia oleracea I. chloroplasts, due to the competitive binding of SO2 and
C02 binding sites (Zeigler, 1972). RUBISCO activity also declines with 0, stress in Oryza
sativa L (Nakamura and Saka, 1978). The use of RUBISCO as a biomarker is limited
because its activity can be readily altered by growing conditions and nutrient regimes.
(2) L-Phenylalanine Ammonia Lyase (PAL) • PAL activity increases with exposure to 03
(Tingey et a!., 1976a). Several gaseous toxicants, such as O,. NO., and SO2, stimulate free
radical formation (Malhotra and Khan, 1984, Richardson et al., 1989). L-Phenyialanine
Ammonia Lyase may increase in response to any oxidizing toxicant as a consequence of
tissue disruption that stimulates secondary metabolism, including synthesis of phenolic
compounds required in cell wall repair. L-Phenytalanine Ammonia Lyase, and phenolic
metabolites may therefore serve as broad-based indicators of oxidant stress, including other
stresses that generate free radicals, such as ionizing UV-B radiation.
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(3) Glutathione reductase - Given that glutathione is a free radical scavenger, glutathione
reductase activity may increase in response to oxidant stress (Bennett et al., 1984) and hence
be a useful indicator.
(4) 'Metallothionein-Jike* Complexes - Although a number of gaseous stresses may induce
similar responses, recent reports on heavy metal toxicants suggests the production and
accumulation of "metallothionein-like' peptide-metal complexes that may be specific for a
particular toxicant (Robinson and Jackson, 1986). Such complexes are found in animal
systems, as discussed above. Homologous complexes remain to be fully characterized in
anglosperms. Robinson and Jackson (1986) isolated a small Cd complex from Cd-resistant
Datura innoria Mill, cells containing glutamate, cysteine, and glycine in a ratio between 2:2:1
and 3:3:1, which resembles metallothionein. These peptides are collectively known as poly(-
glutamylcysteinyl)glycine (EC)nG. It has been hypothesized that these proteins are involved in
metal Ion homeostasis, protection from ionizing radiation, and detoxification of free radicals
(Karih, 1985; Robinson and Jackson, 1986).
Induction of Detoxifying Systems
Toxicants such as NO, and S02 can be initially detoxified by normal metabolic pathways by
increased activity of enzymes involved in assimilation.
(1) Increased Assimilation (e.g., Nitrate Reductase) - Low concentrations of NO and NOS are
readily converted into N03" and NO,' upon dissolution in aqueous solutions, and assimilated
by nitrate reductase and nitrite reductase, and then incorporated into nitrogen metabolism in
higher plants via glutamine synthetase and glutamate synthase. Given that nitrate reductase
activity (NRA) is substrate-inducible, this suggests the possible use of NRA as a biomarker of
NO, impact at nitrogen-poor sites. This has been suggested by Norby (1989), working with
high-altitude red spruce (P/cea rubens). At highly fertile sites the level of constitutive NRA
would already be too high to distinguish NO, impact, and there may be considerable diurnal
and seasonal variation in NRA, limiting its use as a biomarker. If plant tissues are exposed to
acute levels of NO .the plant's capacity to metabolize these toxicants via normal metabolic
pathways is exceeded and membrane damage occurs due to free radical generation and
membrane oxidation.
(2) Free Radical Scavengers - Plants rely on free radical scavenger systems to prevent
extensive membrane disruption, similar to animal systems described above. Enzymes in
plants that function in scavenging of free radicals and their products include superoxide
dismutase, peroxidase (Castillo et at., 1987; Richardson et al., 1989), and catalase (Richardson
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et al., 1989). The lack of specificity of the induction of these systems limits their use as
specific biomarkers, but they are useful as general Indicators of oxidant stress. Nonenzymatic
scavengers, including o-tocopherol (vitamin E), ascorbic acid, ^carotenes, and glutathione
(Richardson et al., 1989) may also have the same limitation. Metabolites that are indicative of
tissue disruption or repair processes may prove more useful.
Specific Metabolites as Biomarkers
(1) Malonaldehyde - A number of studies demonstrate a decrease in lipid content following
SO2, O3, and NO2, often due to Inhibited synthesis (Mudd et al., 1971 a, b; Malhotra and
Khan, 1978, 1984). A common finding is the accumulation of malonaldehyde, a product of
free-radical peroxidation of unsaturated fatty acids (Mudd et al., 1971b; Khan and Malhotra,
1977; Richardson et al.. 1989). Although malonaldehyde concentrations may be a useful
general indicator of oxidants that generate free-radicals, It lacks specificity.
(2) Carotenoids - Pigments have recently received attention as possible biomarkers.
Photosynthettc pigment concentrations are often reported to decline following exposure to
toxicants; however, pigment concentrations may also decline In response to mineral nutrition
status, light, temperature, and water stress (Darrall and Jager, 1984). Mehlhom et al. (1988)
have reported that the ratio of violaxanthin/antheraxanthin (two carotenoids) in two-year-old
needles, the ratio of dry weight/fresh weight differences between current and two-year-old
needles, and ethylene emissions in two-year-old needles were independent of site effects and
correlated with Norway spruce (P/'cea ab/es L) damage in areas affected by forest decline. A
strong correlation between ethylene production, and violaxanthin/antheraxanthin ratio was
reported by Wolfenden et al. (1988), suggesting the carotenoid ratio, a specific indicator,
correlated with the more general stress indicator. The application of the carotenoid ratio test
to controlled exposures of various toxicants in other species is required to determine the
specificity of this test.
(3) Ethylene Emissions - Ethylene production increased in response to S02 exposure.
(Bressan et al., 1979), and ozone exposure (Tingey et al., 1976b). The use of volatile
emissions, such as ethylene. as biomarkers of exposure to toxicants is limited by a lack of
specificity; however, ethylene production is recognized as a good general indicator of stress
(Abeles. 1973; Kimmerer and Kozlowskl, 1982).
(4) Phenol Metabolism - Another general dass of compounds that are receiving attention as
potential biomarkers are phenols and related metabolites. It is not surprising that
concentrations of phenolic compounds increase with oxidant stress, given that the activity of
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PAL and pdyphenol oxkJase (PRO) also increase (Tingey et al., 1976a). Oxidation of phenols
by PPO, laccase, and phenolases is involved in the production of polyphenols, flavinoids, and
alkaloids (Goodwin and Mercer, 1983). Keen and Taylor (1975) reported that ozone induced
large accumulations of isoflavonoids in soybean (Gtycine max (L) Merr. including daidzein,
coumestrol, and sojagol. Hurwltz et al.. (1979) reported the accumulation of 4',7-
dihydroxyflavone in ozone-stressed alfalfa (Medicago sativa L). Tingey et al. (1976b)
subjected Ponderosa pine (Pinue ponderosa Laws.) to chronic ozone exposures and found
that levels of soluble sugars, starch and phenols tended to increase in the shoots and decline
in the roots in response to ozone. These studies suggest that there are likely to be several
phenols and related metabolites that can be used as indicators of ozone stress.
APPLICATION OF BIOMARKER-BASED MONITORING TO EVALUATE AIRBORNE TOXICANTS
The concept of using biomarker measurements in animals and plants for assessing exposure
and effects due to toxic chemicals in the environment has generally been limited to localized "hot
spots" of contamination, such as polluted streams, rivers, or harbors. In general, results have been
encouraging; biomarker responses have correlated with the perceived degree of contamination, and
the relative ranking of sites on the basis of molecular and biochemical responses agrees well with
community level measures of ecosystem integrity (Loar, 1987a, b). In these applications, as in
applications for evaluating airborne toxicants, the biomarker approach offers valuable information that
links exposure to effects and that offers the potential of predicting the potential for long-term
ecological effects from rapidly responding biomarkers. However, airborne toxicants pose new
challenges not encountered in typical applications of localized pollution. For example, what is an
appropriate reference site (le., non-polluted site) for regionally dispersed pollutants transported
through the atmosphere? Also, how can a pollutant exposure of effect be specifically attributed to an
atmospheric source? Each of these considerations will be discussed briefly.
Attributing Exposure to Airborne Toxicants
Although airborne toxicants warrant concern because of the large scale dispersion that is
possible with atmospherically transported material, the pollutants can enter specific ecosystems
through a plethora of routes, including deposition of chemicals in water and son. by entry into food
chains, or by respiration of the air. Thus, identification of a pollutant at a site as atmospherically
derived will probably depend on (1) chemical transport models and/or atmospheric tracer
experiments; (2) direct measurement of chemical concentrations in environmental media, with an
evaluation of whether the presence or concentration could be accounted for by other sources of
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environmental contamination (e.g., an nearby industrial source); and/or (3) identification of marker
compounds or a "fingerprint" consortia of chemicals that can be linked to an atmospheric Input.
Clearly, the identification of specific btomarkers of toxicants In plants has been elusive.
Virtually ail variables Identified as potential Indicators have limitations and do not meet the criteria of
Amdt (1970) and Darrall and Jager (1984) detailed above. The critical shortcoming Is that of
specificity, with various toxicants eliciting similar responses, and with many responses also triggered
by environmental variables. There are, however, several variables that may be related, that can
provide a dear indication of oxidant perturbation. Ethyiene production Increases following stress, With
oxkJants constituting a high degree of stress, via generation of free radicals. Ethyiene stimulates
phenol metabolism, by stimulating PAL, and PPO. Oxidant stress stimulates free radical scavenging
systems, leading to an increase in activities of superoxide dismutase, catalase and peroxldase. A
combination of these variables hold the most potential as general indicators of oxidant stress, but
pinpointing the specific oxidant requires simultaneous environmental monitoring.
Although these methods are needed to identify the potential for a deleterious exposure, It is
the biomarker monitoring that is capable of determining if the toxicants are bioavailable, and if
exposure is sufficient to account for ecologically relevant effects.
Selection of a Suitable Reference Site
Ideally, the reference site should be ecologically identical to the suspect site in all
characteristics other than the potential for exposure to toxicants. In the more typical application of
biomarkers in monitoring (Loar et al., 1987a,b), biomarker responses of animals collected from a site
impacted by an industrial discharge are compared to responses of animals from a similar stream with
no known source of pollution. In the case of widely dispersed atmospheric toxicants, true reference
sites may not exist in all cases.
Perhaps the most reasonable alternative is to define reference sites as those in which
transport models and chemical measurements suggest low impact due to atmospheric or other
sources of pollutant input, and that are characterized by stable, diverse ecosystems. Given the
difficulty in defining an appropriate reference site, It is particularly important to consider Indices of
habitat quality, such as sofl type, percent ground cover, and such, to provide data to statistically test
the alternate hypotheses that habitat or other environmental variables are responsible for any
observed differences in biomarker responses between sentinel populations for two sites.
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Linking Exposure and Effects
Conceptually, the biomarker-based monitoring approach develops evidence, and can
statistically test hypotheses about the linkage between exposure to toxic chemicals and ecologically
relevant effects. The rationale for the approach is indicated in Figure 5, which yiustrates the
relationship between responses at different levels of biological organization and the relevance and time
scales of the responses. Responses at the population and community level are highly relevant to
ecological concerns, but are slow to respond, and are difficult to attribute unequivocally to toxicants.
In contrast, responses at lower levels of organization occur in shorter time frames and can be more
clearly linked to toxic exposure; however, It is difficult to relate these responses to effects at the
population and community level. Our approach is to measure responses at several different levels of
biological organization, including metrics both of exposure to toxicants (generally responses in the
upper left quadrant of Figure 5, but also including tissue burdens of chemicals) and of effects
(generally the lower right quadrant). The goal in examining responses at these different levels of
organization is to answer the following two critical questions.
1. Are organisms exposed to levels of toxicants that exceed the capacity of normal
detoxication and repair systems?
2. If there is evidence of exposure, then is the chemical stress impacting the integrity of the
populations or communities?
- Evidence of exposure from body burdens or from responses at lower levels of biological
organization provide an answer to the first question. In particular, the biomarkers of exposure
indicate the biological significance of chemicals that may have entered the animal or plant (i.e., did
the chemical reach molecular and biochemical targets and cause detectable damage, or induce a
protective response?). The second question can be addressed by determining whether the responses
to the toxicants is propagated up through successively higher levels of biological organization
(biomarkers of effects and population parameters). If chemical exposure is responsible for a high
level ecological effect, responses should be apparent at intermediate levels of organization.
Alternately, If the data do not indicate any evidence of exposure, or If biomarker responses
indicate only minor effects in the most sensitive and responsive exposure parameters (e.g., genetic
damage), but not at any higher levels of biological organization (e.g., histopathological evidence of
neoplasia or tumors, or reduction in photosynthesis), community and population level effects could
not be reasonably attributed to chemical agents. For this reason, any assessment approach also
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Molecular
Physiological
Short-Term
Response
Immunological
High
lexicological
"Relevance
Detox Systems
Bioenergetics
Reproductive
Competence
High
Ecological
Relevance
Population &
Community
FIGURE 5. The relationship between responses at levels of biological organization and the relevance
and time scales of responses.
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needs to examine alternate hypotheses for ecological effects. For example, indices of habitat quality
can be correlated with the population monitoring parameters to determine whether physical
disturbance, or other measures of habitat quality, are better predictors of ecological response than are
measurements of environmental concentrations of known toxicants, or the exposure parameters. If
the biomarker data indicate that some level of toxic effect is occurring, but population parameters are
better predicted by physical use or habitat quality, then multivariate statistics (Adams et al., 1985;
Johnson, 1988) can be used to determine whether effects of chemicals are significantly contributing to
the observed population response.
If a relationship is established between the chemical exposure and ecological effects, then the
statistical models relating atmospheric sources to exposure and effects should be useful for estimating
the extent and magnitude of existing effects, and to project the effect of continued releases.
ACKNOWLEDGEMENTS
We thank G. E. Taylor, N. Edwards and T. Moser for reviewing this manuscript. This work
was supported by the Exploratory Studies Program, Oak Ridge National Laboratory, by the Oak Ridge
Y-12 Plant, Division of Environmental Management, Health and Safety, and by the Biofuels and
Municipal Waste Technology Division, U.S. Dept. Energy. TJT was supported by an appointment to
the U.S. Dept. of Energy Postgraduate Research Program administered by Oak Ridge Associated
Universities. The Oak Ridge Y-12 Plant and the Oak Ridge National Laboratory are operated by
Martin Marietta Energy Systems, Inc., under contract DE-AC05-840R21400 with the U.S. Dept. of
Energy. Publication No. 3661, Environmental Sciences Division, Oak Ridge National Laboratory.
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AIR TOXICS ECOLOGICAL EFFECTS RESEARCH RECOMMENDATIONS
Noel R. Urban1, and Jerry R. Barker2
'Environmental Engineering Sciences. University of Minnesota. Minneapolis. MN
2ManTech Environmental Technology, Inc., U.S. EPA Environmental Research Lab., Corvallis, OR
Nearly 65,000 chemicals are commonly used worldwide for Industrial, agricultural, and
domestic purposes (Schroeder and Lane, 1988). Many of these chemicals eventually are emitted
directly or Indirectly Into the atmosphere as waste products. The airborne chemicals may then be
transported and/or transformed through atmospheric processes and deposited into remote natural
ecosystems as well as rural and urban environments. Airborne toxic chemicals can produce adverse
effects ranging from the biochemical level to ecosystem structure and function. However, not all
airborne chemicals pose the same magnitude of threat to biota and ecosystems. Therefore, research
is needed to prioritize and rank pollutants according to their toxicity, ecological risk, and potential for
human food-chain contamination. Research is also needed to document the spatial extent and
magnitude of air toxics deposition, exposure to biota, and the resulting ecological effects.
Any research directed at assessment of the impacts of air pollution on ecosystems is
confronted with several major problems. These include (but are not limited to): (1) an absence of
background data to define the baseline from which change may be detected, and to determine the
natural variability in ecosystem properties; (2) the existence of geographic gradients in ecosystem
properties that confound efforts to compare impacted and nonimpacted regions; (3) the absence of
systems free from contamination by air pollution, and the Impracticality of maintaining ecosystem-scale
exdosures; and (4) the existence .of natural processes that mimic the effects of pollutants.
There are approaches to identifying changes Induced by air toxics in any type of ecosystem
that minimize the above stated problems. Integral components of a comprehensive approach are
long-term monitoring at a few selected sites, periodic surveys, and process-oriented studies of
contaminant fate and effects. Long-term monitoring and effects research are required to obtain
baseline data on rates of important processes, on sizes and dynamics of populations, and
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environmental reservoirs. Paleoecological studies, such as chemical markers in ice and peat cores,
provide a valuable extension of this monitoring into the past to document natural background
concentrations. Because of the tremendous temporal variability in many ecosystem properties,
periodic surveys are inadequate, and long-term monitoring is critically important for examining
temporal trends. It is particularly important to monitor the rate of atmospheric deposition of
contaminants. Simple measurements of input rates and pool sizes can give important indications of
the mobility and persistence of contaminants. Since single sites are never adequately representative
of a given ecosystem type, periodic surveys are required to evaluate variabllty in ecosystem
properties and to examine spatial trends. It is critically Important to consider natural gradients in
ecosystem characteristics when regional surveys are used to compare the ecological responses of
polluted and nonpolluted areas. Analysis of ecosystem responses across the short-distance gradients
of pollutant exposure that surround many point-sources of air pollutants has been under-utilized in the
past (Gorham and Gordon, 1963; Gignac and Beckett, 1986). At best, however, monitoring and
surveys can only reveal correlations; they cannot prove causality or reveal mechanisms. Therefore,
process-oriented studies that include laboratory and field manipulations are essential for proof of
causality, elucidation of transport and transformation pathways, and anything more than empirical
prediction of contaminant effects.
It is not a lack of vision into research needs that impedes progress, for these
recommendations are far from new. Nevertheless, there is no ongoing national monitoring of
atmospheric deposition of trace metals, synthetic organic chemicals, and other air toxics. There is
also no ongoing monitoring of contaminant cycling in terrestrial and aquatic ecosystems and there is
very little baseline data on fundamental ecosystem processes (Kelly and Harwell, 1982). Very little of
the ongoing process-oriented research in terrestrial and aquatic ecosystems is directed at the
ecological risk assessment of air toxics effects.
129
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