NNAS
RNAE
CIOM
Piiblication-on-Demand Program
Polycyclic Aromatic Hycarbons
-'BRARY. AWBERC, CINCINNA-,
U. S. EPA
26 W. MARTIN LUTHER KING DRIV
CINCINNATI, OHIO 45268
.N/itfionaf Academy Press
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POLYCYCLIC AROMATIC HYDROCARBONS:
EVALUATION OF SOURCES AND EFFECTS
COMMITTEE ON PYRENE AND SELECTED ANALOGUES
BOARD ON TOXICOLOGY AND ENVIRONMENTAL HEALTH HAZARDS
COMMISSION ON LIFE SCIENCES
NATIONAL RESEARCH COUNCIL
National Academy Press
Washington, D.C.
1983
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NOTICE: The project that is the subject of this report was approved by
the Governing Board of the National Research Council, whose members are
drawn from the Councils of the National Academy of Sciences, the
National Academy of Engineering, and the Institute of Medicine. The
members of the committee responsible for the report were chosen for
their special competence and with regard for appropriate balance.
This report has been reviewed by a group other than the authors
according to procedures approved by a Report Review Committee consisting
of members of the National Academy of Sciences, the National Academy of
Engineering, and the Institute of Medicine.
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Academy of Sciences in 1916 to associate the broad community of science
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the Institute of Medicine were established in 1964 and 1970, respec-
tively, under the charter of the National Academy of Sciences.
The work on which this publication is based was performed pursuant
to Contract 68-01-4655 with the Office of Research and Development of
the Environmental Protection Agency.
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BOARD ON TOXICOLOGY AND ENVIRONMENTAL HEALTH HAZARDS
RONALD ESTABROOK, University of Texas Medical School, Dallas, Texas,
Chairman
PHILIP LANDRIGAN, National Institute for Occupational Safety and Health,
Cincinnati, Ohio, Vice Chairman
EDWARD BRESNICK, University of Vermont School of Medicine, Burlington,
Vermont
VICTOR COHN, George Washington University Medical Center,
Washington, D.C.
A. MYRICK FREEMAN, University of Washington, Seattle, Washington
DAVID G. HOEL, National Institute of Environmental Health Sciences,
Research Triangle Park, North Carolina
MICHAEL LIEBERMAN, Washington University School of Medicine, St. Louis,
Missouri
RICHARD MERRILL, University of Virginia, Charlottesville, Virginia
VAUN NEWILL, Exxon Corporation, New York, New York
JOHN PETERS, University of Southern California School of Medicine,
Los Angeles, California
JOSEPH V. RODRICKS, Environ Corporation, Washington, D.C.
LIANE B. RUSSELL, Oak Ridge National Laboratory, Oak Ridge, Tennessee
CHARLES R. SCHUSTER, JR., University of Chicago, Chicago, Illinois
Ex Officio Members
LESTER BRESLOW, School of Public Health, University of California, Los
Angeles, California
GARY P- CARLSON, Purdue University, West Lafayette, Indiana
JAMES F. CROW, University of Wisconsin, Madison, Wisconsin
BERNARD GOLDSTEIN, University of Medicine and Dentistry of New Jersey/
Rutgers Medical School, Piscataway, New Jersey
ROGER 0. McCLELLAN, Lovelace Biomedical and Environmental Research
Institute, Albuquerque, New Mexico
SHELDON MURPHY, University of Texas, Houston, Texas
NORTON NELSON, New York University Medical Center, New York, New York
JAMES L. WHITTENBERGER, Harvard University, Boston, Massachusetts
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COMMITTEE ON PYRENE AND SELECTED ANALOGUES
EDWARD BRESNICK, University of Vermont School of Medicine, Burlington,
Vermont, Chairman
MARSHALL W. ANDERSON, National Institute of Environmental Health
Sciences, Research Triangle Park, North Carolina
ROBERT A. GORSE, JR., Ford Motor Company, Dearborn, Michigan ^
DANIEL GROSJEAN, Environmental Research and Technology, Inc., Westia
Village, California
RONALD A. HITES, Indiana University, Bloomington, Indiana
ATTALLAH KAPPAS, The Rockefeller University, New York, New York
RICHARD E. KOURI, Microbiological Associates, Bethesda, Maryland^ ^
MALCOLM C. PIKE, University of Southern California School of Medicine,
Los Angeles, California
JAMES K. SELKIRK, Oak Ridge National Laboratory, Oak Ridge, Tennessee
LAWRENCE J. WHITE, New York University, New York, New York
JAMES A. FRAZIER, National Research Council, Washington, D.C., Staff
NORMAN GROSSBLATT, National Research Council, Washington, D.C. ,
JEAN E. PERRIN, National Research Council, Washington, D.C., Secretary
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ACKNOWLEDGMENTS
This document is the result of individual and coordinated efforts by
the members of the Committee on Pyrene and Selected Analogues. Although
individual members were responsible for specific sections, the entire
report was reviewed by the full Committee. The summary (Chapter 8) and
the recommendations (Chapter 9) represent a consensus of the Committee
members.
The executive summary was prepared by the chairman, Dr. Edward
Bresnick. Chapters 1, 2, and 3, on sources and atmospheric transforma-
tions and persistence, represent a joint effort of Drs. Robert A. Gorse,
Jr., Daniel Grosjean, and Ronald A. Kites and Mr. James A. Frazier.
Chapter 4, on biologic effects, was written by Dr. Bresnick. Chapter 5,
on pharmacokinetics and effective biologic dose, was prepared by Drs.
Marshall W. Anderson and James K. Selkirk. Chapter 6, concerning human
exposure to and metabolism of the compounds in question, was written by
Dr. Attallah Kappas. Chapter 7, on populations of "hypersensitive"
persons, was written by Dr. Richard E. Kouri. Appendix C, dealing with
human-cancer risk assessment, was prepared by Dr. Malcolm C. Pike.
Appendix D, on public decision-making with respect to source and
emission control, was prepared by Dr. Lawrence J. White.
We acknowledge the special contributions of Dr. Stanley Blacker of
the Environmental Protection Agency, who made a presentation to the
Committee at its first meeting, on May 11, 1981, and provided resource
material for the Committee's use in preparing its report, and to Dr. Roy
Albert of the New York University Medical Center's Institute of
Environmental Medicine, who addressed the Committee at its second
meeting, on May 29.
We express our gratitude to the following persons for providing
resource material and other information:
• Dr. Kent Berry, Environmental Protection Agency
Dr. William J. Blot, National Cancer Institute
Dr. Robert M. Bruce, Environmental Protection Agency
Dr. Marcus Cooke, Battelle Columbus Laboratory
Mr. John Cuttica, Department of Energy
Dr. Gregory J. D'Alessio, Department of Energy
Dr. Jack H. Dean, Chemical Industry Institute of Toxicology
Dr. John W. Farrington, Woods Hole Oceanographic Institution
Dr. Wayne H. Griest, Oak Ridge National Laboratory
Dr. Robert Hall, Environmental Protection Agency
Dr. Ronald W. Hart, National Center for Toxicological Research
Dr. Frederick T. Hatch, Lawrence Livermore National Laboratory
Dr. Dietrich Hoffman, Naylor Dana Institute for Disease
Prev ntion, American Health Foundation
Dr- Gary L. Johnson, Environmental Protection Agency
Dr. Ronald 0. Kagel, Dow Chemical Co.
Dr. Daniel W. Nebert, National Institutes of Health
Dr. Douglas E. Rickert, Chemical Industry Institute of Toxicology
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to:
Special thanks for providing printout8 of the literature are given
• The National Agricultural Library in Beltsville, Md. (AGRICOLA)
• The National Institute for Occupational Safety and Health in
Cincinnati, Ohio (NIOSHTIC)
We acknowledge the contributions of the following in the National
Research Council for providing resource material:
• Dr- Scott R. Baker, Board on Toxicology and Environmental Health
Hazards
• Dr- Robert J. Golden, Board on Toxicology and Environmental
Health Hazards
• Mrs. Barbara Jaffe and the Toxicology Information Center staff
• Dr. Sushma Palmer, Commission on Life Sciences
• Mr. Richard C. Vetter, Ocean Sciences Board
The Committee wishes to commend the excellent assistance of
Mr. James A. Frazier, the staff officer; Mr. Norman Grossblatt, the
editor; Mrs. Jean E. Perrin, secretary; and Mrs. Eileen G. Brown,
manuscript typist.
Extensive use was made of the resources of the Library of the
National Academy of Sciences, the Toxicology Information Center of the
Board on Toxicology and Environmental Health Hazards, the National
Library of Medicine, the National Agricultural Library, the Library of
Congress, and the Air Pollution Technical Information Center of the
Environmental Protection Agency.
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CONTENTS
Executive Summary
Introduction
1 Polycyclic Aromatic Hydrocarbons from Mobile Sources
and Their Atmospheric Concentrations
2 Polycyclic Aromatic Hydrocarbons from Natural and
Stationary Sources and Their Atmospheric Concentrations
3 Atmospheric Transformations of Polycyclic Aromatic
Hydrocarbons
4 Biologic Effects of Smoke, Emission, and Some of Their
PAH Components
5 Effective Biologic Dose
6 Polycyclic Aromatic Hydrocarbons in Food and Water and
Their Metabolism by Human Tissues
7 Some Factors that Affect Susceptibility of Humans to
Polycyclic Aromatic Hydrocarbons
8 Summary
9 Recommendations
Appendix A Lists of Polycyclic Aromatic Hydrocarbons
Appendix B Polycyclic Aromatic Hydrocarbons in the Ambient Atmosphere
Appendix C Human-Cancer Risk Assessment, by Malcolm C. Pike
Appendix D Public Decision-Making with Respect to Atmospheric PAH
Sources and Emissions, by Lawrence J. White
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EXECUTIVE SUMMARY
The Clean Air Act stipulates that from time to time the Administrator
of the Environmental Protection Agency (EPA) shall revise a list that
includes pollutants that may be anticipated to endanger public health or
welfare and for which air-quality criteria have not been issued.
As part of a continuing contract with the National Academy of Sciences
to prepare scientific and technical assessment reports on selected
pollutants, the EPA asked for an evaluation of selected and representative
pyrene compounds and their analogues as they occur as pollutants in the
ambient air, especially those from mobile sources.
The Committee on Pyrene and Selected Analogues, appointed by the
National Research Council, selected representative pyrenes and close
chemical relatives for study. Great difficulties necessarily are
encountered when a study covers a large number of compounds. It is
extremely difficult to be comprehensive and discuss every compound in
detail. The Committee found that there were far more sources of human
exposure to pyrenes than vehicle exhaust—for instance, cigarette-smoking,
coke ovens, wood-burning, and some foods. The Committee is aware that
some of its interpretations are founded on data that are neither clear-cut
nor complete. This is true of its efforts to extrapolate risks, to
identify susceptible groups in the population, and to assess economic
alternatives for control or abatement of the pollutants in question.
The polycyclic aromatic hydrocarbons (PAHs) have been reviewed pre-
viously as components of atmospheric pollution and as potential human-
health hazards. This document attempts to make current the information on
the sources, formation, atmospheric persistence and transformations,
biologic effects, and toxicokinetics of a select group of PAHs and on the
identification of populations hypersensitive to them. The document also
presents material on human risk assessment and develops an approximate
estimate of the societal value of reducing environmental emission of
benzo[a]pyrene. Benzo[a]pyrene is used as a surrogate PAH. It may not be
the best indicator of the biologic effects of other PAHs in soots and
smokes. However, the literature on benzo[a]pyrene is considerably more
voluminous than that on other PAHs. It should also be recognized that the
benzo[a]pyrene concentrations in soots and smokes is small and that other
PAHs present in smokes have greater biologic activity, such as
nitro-PAHs. The specific PAHs discussed in this report were selected on
the basis of their relative concentrations in various emission or
combustion products or because they are pharmacologically active. The
structures of the selected compounds are presented in Appendix A.
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SOURCES. ATMOSPHERIC PERSISTENCE, AND TRANSFORMATIONS OF PAHs
The total annual release of PAHs from mobile sources in the United
States has been estimated. In the case of benzo[a]pyrene (BaP), all
mobile sources produced about 43 metric tons in 1979, including 27 tons
from motor vehicles; 63Z and 37Z of the BaP emission from motor vehicles
occurred in urban and rural areas, respectively. It is projected that the
total motor-vehicle BaP emission will decrease by the year 2000 to 24
metric tons, of which only about 402 will be in urban regions. This
implies a shift toward emission in the rural areas. This projection is
based on the adoption of no further changes in emission standards beyond
those which have been in effect since 1980. If more rigorous emission
standards are adopted in 1985 or later, BaP emission could be considerably
lower.
On the basis of motor-vehicle emission values for 1979, the average
daily BaP concentration in the urban atmosphere has been calculated as
1.3 ng/m3; this value is in excellent agreement with the findings in Los
Angeles, where atmospheric BaP comes largely from motor-vehicle exhaust.
Lower-density urban areas would have lower atmospheric concentrations of
motor-vehicle-contributed BaP. In the year 2000, atmospheric BaP
concentrations in cities with the traffic conditions of Los Angeles are
expected to be approximately 0.6 ng/nr*—a large decrease from 1979.
So-called "hot spots" of BaP, as in severe roadway tunnel congestion, can
lead to motor-vehicle-generated BaP concentrations of approximately 50
ng/m , to which people would be exposed for very brief periods.
Atmospheric BaP concentrations as high as 74 ng/nr* have been measured in
U.S. urban areas (such as Birmingham, Alabama); much of this BaP is
contributed by stationary sources.
The high atmospheric concentrations arise mostly from stationary
combustion, especially that of coal, wood, and oil. National annual BaP
atmospheric emission from all combustion sources, including both mobile
and stationary, is estimated at between 300 and 1,300 metric tons. The
total is decreasing because of emission controls and changes in fuel
use—e.g., less coal is used in residential furnaces, but more coal is
used in power generation. Wood-burning stoves and fireplaces, currently
nonregulated sources of PAHs, are ubiquitous and are important and
increasing sources of atmospheric PAHs in urban areas, as well as in rural
areas with restricted air flow.
A survey of the literature reveals large uncertainties with respect to
the persistence of PAHs, their chemical transformations, and their
transport and fate in the atmosphere. There is evidence from long-range
transport studies and analyses of sediments that PAHs may be transported
over long distances in the atmosphere (e.g., 1,000 km) without substantial
degradation; and laboratory studies have shown that many PAHs react
chemically with atmospheric components in a matter of hours.
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High chemical reactivity and long-range transport of unreacted PAHs
are not necessarily in disagreement. There is a need for more information
on this extremely broad and dynamic issue. The major chemical processes
include photooxidation, reaction with ozone, and reaction with nitrogen
dioxide—the latter can yield potent, direct-acting mutagenic nitro
derivatives. However, these processes appear to be significantly slower
for PAHs adsorbed on atmospherically relevant substrates, such as soot and
fly ash, than for the same PAHs deposited on filters in pure form. It is
still uncertain whether nitro-PAHs occur in exhaust or are artifacts of
the filter-sampling procedures.
BIOLOGIC EFFECTS
Particles from diesel engines were tested for toxicity in intact
animals. Only minimal effects on pulmonary function, reproductive
capacity, glandular or hepatic function, and general neonatal health were
observed. There was some indication of abnormal development of central
nervous system function in newborn rats that were chronically exposed to
diesel-engine exhaust. There was not enough information to ascribe these
alterations to the PAH content of the exhaust.
Organic-solvent extracts of particles from spark-ignition-engine and
diesel-engine exhaust were mutagenic in Salmonella typhimurium in forward-
and backward-mutation assays and were mutagenic in several animal-cell
model systems. In bacterial assays, diesel and spark-ignition combustion
products were directly active; emission from coke ovens and roofing tar,
cigarette-smoke condensate, wood combustion products, and the positive
control 3aP required metabolic activation before they demonstrated any
rautagenic action. The mutagenic efficacy of the combustion products was
reduced by the inclusion of alcohol in the fuel base (a suggestion of
additional advantage to its presence in fuel). The direct-acting
mutagenic property appeared to be caused in part by nitro-PAHs. The
latter were constituents of the automobile-exhaust particles, but were not
found in wood combustion products. The nitro-PAHs tested proved much more
active mutagenically than the parent compounds. Indeed, 1,8-dinitro-
pyrene, a constituent of particles and, in the past, of xerographic
toners, was the most highly mutagenic of all compounds subjected to the
Salmonella/microsome assay. The nitro-PAHs have not been tested
consistently in animal-cell mutagenesis models or whole animals, however.
Because of their mutagenic potency observed in the Salmonella/microsome
assay, it is essential to learn whether they are formed in exhaust
products or are artifacts produced during the course of the sampling
procedure.
The extracts of particles from mobile and stationary sources have been
tested for carcinogenicity, mainly by topical application to mouse skin.
The condensates from spark-ignition-engine exhaust were carcinogenic in
this model, and diesel-exhaust preparations were less active. The exhaust
preparations exhibited both initiation and promotion activities in the
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skin-tumor model. Whether the tumorigenicity reflected the additive
activity of a number of PAHs present as components of the condensates
depended on the experimental protocol used. The cocarcinogenic activity
of mixtures of PAHs must be studied further and the activity relationship
of individual chemicals measured. The compounds of the condensates were
not very active in the inhalation or intratracheal-instillation test
systems used for tumorigenicity, and the major routes of entry of emission
and of PAHs are ingestion and inhalation. Additional effort should be
expended to develop test models that better approximate human
carcinogenesis.
The nitro-PAHs have not been tested for tumorigenicity. Because they
are potent mutagens, they should be tested in several tumorigenicity model
systems. Although there was a moderate correlation between the
mutagenicity and the carcinogenicity of the PAHs, a battery of tests—
including assays for mutation, clastogenesis, primary DNA damage, and
morphologic transformation—would serve better as a monitoring mechanism.
EFFECTIVE BIOLOGIC DOSE
PAHs are readily absorbed after administration to laboratory animals
by various routes and are distributed to a number of tissues.
Nonmetabolized PAHs accumulate and persist in fat to a much greater extent
than in other tissues. This phenomenon may be used to monitor the chronic
exposure of populations to sources of PAHs. PAHs that are present on
particles are retained in the lungs of animals to various degrees as a
function of particle size and composition. The particle-bound PAHs are
desorbed in the lung and distributed systemically to various tissues.
The clearance of PAHs from animals is a function of the "reservoir" of
nonmetabolized material in the fat, metabolism, biliary excretion, fecal
excretion, and, to a smaller degree, urinary excretion. The excreted
metabolites consist of glucuronides, sulfates, and unconjugated
hydroxylated and phenol derivatives.
The metabolism of many of the PAHs has been studied in in vitro
systems. Preparations from virtually all tissues are able to metabolize
PAHs, although liver is the most efficacious in this regard. The initial
metabolism is catalyzed by membrane-bound cytochrome P-450-dependent
monooxygenases. The epoxide product may spontaneously rearrange to a
phenol, which may give rise to conjugated phenols. The epoxides serve as
substrates for another membrane-bound enzyme, epoxide hydratase, which
catalyzes the formation of trans diol derivatives; these may also be
excreted in conjugated form (as glucuronides) or in unconjugated form.
The activation of PAHs to at least some of the ultimate carcinogenic
forms requires recycling of the trans diol derivatives through the
cytochrome P-450-dependent monooxygenases to yield very reactive
diol-epoxides , which can spontaneously form electrophiles capable of
interacting with macromolecular components, such as DNA. The amounts of
monooxygensases and epoxide hydratase in tissues are genetically
determined; and the enzymes are inducible. Indeed, PAH exposure can
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increase dramatically the amount of monooxygenase; the extent of induction
is also genetically determined.
The activation reaction can also be fulfilled through an arachidonic
acid-dependent cooxygenation step involving prostaglandin peroxidase. In
this step, the trans diol is enzymatically transformed to the diol-epoxide
at the expense of prostaglandin 62- Hence, prostaglandin synthesis may
be intimately involved in the elaboration of carcinogenic agents from PAHs.
PAH metabolites, such as diol-epoxides, interact in covalent fashion
with DNA bases to form adducts. The adducts of BaP diol-epoxide with
DNA—BP DE-DNA adducts—form readily in lung, liver, forestomach (of
mice), colon, kidney, brain, and muscle after oral administration of the
PAH to laboratory animals. Human tissues also have this capacity. From
in vivo studies, the BP DE-DNA adduct profiles appear representative for a
particular tissue. The major adduct known to be formed is between BP DE I
and the 2-position of deoxyguanosine in DNA, and'a minor adduct is formed
at the 6-position of deoxyadenosine. The adduct profile is species-
dependent. The quantitative aspects of these reactions do not appear to
be correlated with the susceptibility of a tissue to PAH-induced
carcinogenesis. For example, hepatic tissue is not under ordinary
circumstances a target organ for PAHs, but it can easily biosynthesize
these adducts.
The PAH-DNA adducts have various turnover rates in different tissues.
The relative contributions of a hitherto unknown DNA enzymatic repair
system and cell turnover have not been established under in vivo
conditions. It is apparent, however, that different adducts are removed
from the DNA at different rates.
A linear dose-response relationship has been observed (with BaP) for
PAH metabolite-DNA adducts even at low doses. There appears to be no
threshold dose below which binding of activated PAH metabolites to DNA
does not occur. The administration of several inducers of the cytochrome
P-450-dependent monooxygenases and of various conjugating enzyme systems,
as well as the administration of several antioxidants, dramatically
reduces the formation in vivo of adducts. That suggests that these
substances (i.e., antioxidants and monooxygenase inducers) may decrease
PAH induction of neoplasia as a result of their ability to affect
synthesis of adducts. It is proposed that, given a "susceptible" tissue
(i.e., one that is neoplastically transformed by a PAH), the adduct
concentration can be an appropriate measure of the "effective biologic
dose" of the PAH. With current sensitive radioinmunoassay methods at our
disposal for the determination of these adducts, it should be relatively
easy to determine that concentration, e.g., in human lymphocytes.
The formation of adducts may have three important implications for
toxic effects in humans: (1) Even at low environmental concentrations of
a PAH, continuous exposure could result in persistent formation of the
adducts, leading to a higher incidence of neoplasia. The persistence of
the adduct in a particular tissue will be determined by the extent of
repair or cell turnover. (2) The presence of an adduct, even at a low
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concentration, may influence replication, transcription, and transposition
substantially. In any case, the expression of the genome will be
affected. The extent of this problem will depend on the site of PAH-DNA
adduct formation. (3) Nutrition and other exogenous factors may
influence the activation of a PAH, the extent of conjugation to a
detoxified product, the formation of adducts, and the relative turnover of
these adducts in a particular tissue.
HUMAN EXPOSURE TO AND METABOLISM OF PAHs
Human exposure to PAHs is almost exclusively via the gastrointestinal
and respiratory tracts—and approximately 992 of these substances is
ingested in the diet. PAHs are ubiquitous in foodstuffs. The PAH content
of most foods before preparation is quite low, but some have surprisingly
high concentrations, presumably as a result of pollution from soils,
irrigation waters, and atmospheric fallout and perhaps from the initial
phases of food-processing. The contaminants include 100 or even more
PAHs. The mode of cooking, especially broiling, also affects the
composition and quantity of PAHs in foods.
The extent to which PAHs gain access to the circulation is not known.
Both cellular and humoral routes of entry appear to be involved. The
serum lipoproteins constitute a substantial circulatory pool of PAHs.
PAHs presumably translocate from these serum proteins into cells by a
non-receptor-mediated mechanism. According to data obtained with lower
animals, the PAHs will enter cells or accumulate in fatty tissues and
then slowly re-enter the circulation, undergo a variety of biotransform-
ations, and are excreted via the biliary or the urinary system. There is
a dearth of information on the human toxicokinetics of PAHs other than BaP.
Human normal and malignant tissues have the metabolic capacity to
effect oxidative transformations of PAHs, especially BaP, to form
products—including ultimate carcinogens—comparable with those formed in
experimental animals. There is a very large individual variation in these
enzymatic activities, and the rate of oxidative metabolism of a PAH can
vary considerably in different sites in the same organ from the same
organism. The PAHs are present in various human tissues to a limited
extent, and some can be sporadically, or even regularly, identified. It
is important to note this observation, although the relation of the
finding of specific pathologic conditions to known PAH pollution has not
been established. Nor has a biochemical "marker" involving PAHa been
established by which a patient population with specific enzymatic
characteristics can be distinguished in relation to a discrete pathologic
condition. The question of high inducibility of aryl hydrocarbon
hydroxylase (AHH) activity in lymphocytes and monocytes of lung-cancer
patients continues to be provocative, as well as ambiguous, and the use of
this activity as a marker of high-risk populations deserves further study.
There is very little information to implicate diet-derived PAHs in any
form of clinical pathologic condition, despite the high concentrations of
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these compounds to which humans may be exposed through food contamination.
The lack of information suggests that the gastrointestinal system
(including the liver) may be relatively "resistant" to the toxicity of
PAHS or that this system can biochemically adapt to PAH exposure. The
capacity of the body's detoxification system to be "resistant" is worth
exploring.
POPULATIONS OF HYPERSENSITIVE PERSONS
The toxicity—including mutagenesis, carcinogenesis, and terato-
genesis—of PAHs results from multistage processes, and variations in any
of the intermediate stages can influence susceptibility to the effects.
Sensitivity to PAH-induced biologic effects is probably controlled at the
level of uptake into particular cells, metabolic activation or
inactivation of the parent PAH chemical, capacity of cells to repair PAH
metabolite-DNA adducts, capacity of cells to express DNA damage and allow
progression to the phenotype of a mutant or tumor cell, and iramuno-
competence of the host. Compilation of data from humans and animal-model
systems has demonstrated a degree of genetic regulation of each of these
stages, but the information is far too sketchy for specific conclusions to
be drawn on the role of PAHs.
PAHs in both human and animal systems are taken up and metabolized by
microsomal monooxygenases that are under some sort of genetic regulation.
In murine-model systems, susceptibility to carcinogenesis induced by PAHs
is genetically linked to the capacity to respond to and metabolize these
chemicals. In humans, development of cigarette-smoke-associated lung
cancers also may be linked to the capacity to respond to and metabolize
PAHs. Natural variations in DNA-repair capacity do occur among humans
with specific genetic disorders, and these persons are more susceptible to
cancer; but whether PAHs play a role in such susceptibility is not known.
Variations in capacity to promote (or allow for progression of) carcino-
genesis occur in animal-model systems, but there are virtually no data or
similar variations among humans. Genetically controlled variations in
immunocompetence are observed in humans, and persons with these
alterations are usually more susceptible to carcinogenesis; but, again, no
active role for PAHs has been suggested. This lack of information on the
impact of genetic variation is striking. Factors that tend to make these
genetic differences less distinct include the physical state of the PAHs
and the nutritional and developmental states of the host.
DISCUSSIONS OF RISK ASSESSMENT AND PUBLIC-POLICY COST-BENEFIT
CONSIDERATIONS
The Committee draws conclusions about neither anticipated health risks
associated with exposure to PAHs in ambient air nor the societal costs of
reducing PAH emission. However, some examples of how these considerations
could be addressed are described in Appendixes C and D, respectively.
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INTRODUCTION
Benzo[a]pyrene is a chemical commonly found in the emission products
from most types of fuel combustion, whether it occurs in the engine of an
automobile, the fireplace of a home, or an industrial installation. It has
been found to be one of the chemicals that cause cancer in humans exposed to
them. Because this chemical and other PAH compounds emitted in automobile
exhaust are commonly found in the ambient air (see lists in Appendixes A and
B), the U.S. Environmental Protection Agency asked the National Academy of
Sciences-National Research Council to assess the health risks of humans
exposed to the compounds that can be identified and characterized in the
atmosphere, to identify those persons most susceptible to the toxic effects
of the compounds, and to characterize the other major sources of human
exposure, with emphasis on emission from mobile sources. The NRC Committee
on Pyrene and Selected Analogues, which prepared this report, was formed in
the Commission of Life Sciences, where it was under the oversight of the
Board on Toxicology and Environmental Health Hazards.
The relatively recent oil crisis in the United States has focused
attention on efforts to reduce the amount of crude oil that must be imported
to maintain our standard of living. As part of the move to conserve fuels,
there has been an increase in the use of diesel engines. Unfortunately,
diesel-engine exhaust contains more particulate matter than exhaust from
spark-ignition engines and therefore may contribute heavily to environmental
pollution. The major organic chemical constituents attached to the
particulate matter in diesel exhaust include the polycyclic aromatic
hydrocarbons (PAHs). Benzo[a]pyrene is commonly used in the literature as a
surrogate for the whole class of PAHs, although it may not be the best
indicator for the biologic effects of complex chemical mixtures containing
PAHs. (See discussions of tracer chemicals and surrogates in concluding
section of Chapter 3.)
One of the major objectives of the Committee was to evaluate the
relative contributions of the various emission sources—mobile and
stationary—to the PAH pollution burden and to establish the health risks to
those exposed to the emission. The Committee selected the following
compounds and close chemical relatives as the subjects of published
information to use in preparing its report: acridine, benz[a]anthracene,
benzo[a]pyrene, benzo[e]pyrene, chrysene, coronene, cyclopenta[cd]pyrene,
dibenzothiophene, 1-fluoranthene, fluorenone, methylfluorenone, nitropyrene,
phenanthrene, phenanthrene carboxaldehyde, and pyrene. (See Appendix A for
formulas and CAS numbers of these and other PAHs.) The term "polycyclic
aromatic hydrocarbon" (PAH) has been used throughout the report to refer
broadly to all these chemicals.
The Committee met for the first time in May 1981. In its review of the
published literature then being assembled, it became clear that PAHs have
numerous sources and are ubiquitous in the environment. For example, the
published information revealed that the human diet was a major source of
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exposure. Thus, if the adverse human health effects attributable to
exposure to PAHs in the emission of automobiles and other vehicles were to
be appropriately assessed, the Committee felt that information on exposure
associated with other than mobile sources was needed, so that the relative
amounts from each could be described. The Committee discussed mechanisms
and principles for identifying population subgroups that seem to be more
susceptible to the effects of exposure to PAHs. They discussed the
mechanisms of chemical intoxication, the metabolism of individual PAHs, and
the formation of adducts to DNA.
Chapters 1-3 of the report discuss the mobile and stationary sources of
PAHs emitted to the atmosphere, their atmospheric persistence and
transformations, and their deposition. Chapter 4 gives an overview of
published findings on PAH toxicity and biologic effects. Chapter 5
describes the phannacokinetics of PAH and their role in the formation of DNA
adducts. Chapter 6 discusses human PAH metabolism, the modes and extent of
human exposure to PAHs in the diet (via various foodstuffs and cooking
methods) and from other sources, and deposition in various body tissues.
Chapter 7 discusses enzyme systems and genetic and other anomalies that
can be used to identify or characterize persons who are hypersensitive to
PAHs.
Chapter 8 summarizes the Committee's findings. In Chapter 9, the
Committee presents recommendations for research that it feels will advance
the understanding of the effects of PAH.
In trying to assess the adverse health effects, the Committee discovered
a lack of epidemiologic data on exposure to PAHs from mobile or stationary
sources and effects directly attributable to them. Perhaps the best
epidemiologic work was related to cigarette-smoking, and smoking was
therefore used as a model in the discussions of risk in Appendix C. However
unsatisfactory this may appear at first glance, cigarette smoke is a complex
mixture of PAHs and other chemicals, and the results provide some foundation
for judging the relation of PAH exposure to adverse health effects observed
in cigarette-smokers. The lack of data made it difficult (and perhaps it is
impossible) to characterize cost-benefit relations with respect to the
control or abatement of PAH emission from various sources and the attendant
policy choices (see Appendix D). The cutoff date for literature cited in
the report was June 1982.
1-2
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POLYCYCLIC AROMATIC HYDROCARBONS FROM MOBILE SOURCES
AND THEIR ATMOSPHERIC CONCENTRATIONS
Exhaust products of fuel combustion from stationary or mobile sources
that have been identified as hazardous to humans are commonly targeted for
abatement or regulatory control. A variety of control techniques—e.g.,
particle collectors, gaseous-emission scrubbing devices, catalytically
equipped exhaust systems, and "scavenger" fuel additives—have been used to
convert the unburned and partially burned hydrocarbons, including polycyclic
aromatic hydrocarbons (PAHs), in exhaust to less hazardous chemicals.
This chapter discusses the annual consumption of fuels in various types
of vehicles, sampling, PAH emission from mobile sources, and future control
technologies.
FUEL CONSUMPTION IN THE UNITED STATES
The important fuels consumed in this country are listed in Table 1-1 with
estimates of annual consumption figures for 1979, the latest year for which
all data are available. 0»°1 The major energy source, of course, is crude
oil. AXable 1-2 lists the uses of the major crude-oil fractions for
1979. 0»«1 The U.S. consumption of crude oil is decreasing. In addition,
important changes in how oil is used are possible within the next two
decades. For example, gasoline consumption currently far exceeds the con-
sumption of diesel fuel. Owing to the increased fuel mileage of gasoline-
fueled vehicles, the increasing use of diesel-fueled vehicles, and overall
efforts at energy conservation, it is possible that diesel-fuel consumption
could outstrip gasoline consumption in two decades.
TYPES OF MOBILE SOURCES AND THEIR RELATIVE IMPORTANCE
The term "mobile source" represents a broad range of vehicle classi-
fications with considerable differences in miles traveled, amount and type of
fuel consumed, exhaust emission rate, and location of fuel use. In addition,
the emission from any particular category may change considerably from one
year to the next with technologic advances in engine design and
emission-control techniques. Current estimates of miles traveled and fuel
consumption for each mobile-source category are listed in Table
1-3. 4»46»60»61'67»82'83 The present status and projected changes in
relative importance of each of the categories are discussed below.
Light-duty passenger cars with spark-ignition engines account for most of
the motor-vehicle mileage accumulated in this country. To meet the
exhaust-emission standards for gaseous hydrocarbon (HC) and carbon
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monoxide (CO), most 1975 and later model-year spark-ignition passenger
cars have been equipped with oxidation catalysts on the exhaust system.
The catalysts are poisoned by lead in the fuel and therefore require
unleaded fuel. The use of unleaded fuel in catalyst-equipped cars and
lower lead concentrations in leaded fuel have resulted in considerable
decreases in rates of emission of lead, HC, CO, and particulate material
from passenger cars. »^1 The transition to catalyst-equipped cars has
continued, and older non-catalyst-equipped cars are continually being
removed from service, so the present (1982) mileage of catalyst-equipped
cars is now greater than the mileage of noncatalyst cars. By the early
1990s, more than 952 of the gasoline-fueled passenger-car mileage will be
attributed to catalyst-equipped vehicles. Beginning with the 1981 model
year, most new passenger cars have three-way catalysts capable of reducing
emission of HC, CO, and nitrogen oxides (NOX). More than 502 of the
catalyst-equipped passenger cars will have three-way catalysts by the
mid-1990s. The exact mix will depend heavily on future NOX emission
standards. Three-way catalysts result in significantly lower emission of
CO, NOX, gaseous HC, and particulate material than the original
oxidation catalysts.
Increasingly stringent federal fuel-economy standards (Figure 1-1) are
in effect through the 1985 model year for passenger cars. Coupled
with oil shortages and the goal of decreased U.S. dependence on foreign
oil, the fuel-economy standards will result in an approximate doubling of
new-car fuel economy between 1974 and 1985. The goal of improved fuel
economy is being attained by a decrease in vehicle weight, the use of more
fuel-efficient engines, and an increase in the use of diesel engines in
light-duty vehicles (both in passenger cars and in light- and medium-duty
trucks). Diesel-engine vehicles achieve about 252 higher fuel mileage
than their counterparts among spark-ignition-engine vehicles (a somewhat
smaller improvement if the volume of crude oil or the energy content of
the fuel is used as a basis). The cost advantage enjoyed by diesel fuel
over gasoline has largely dissipated in the last few years and could even
turn into a cost penalty as the demand for diesel fuel increases and the
demand for gasoline decreases.^ It has been projected that 252 of the
passenger-car fleet could be diesel-powered by the mid-1990s, but many
factors will affect the actual rate of approach to that proportion and the
percentage ultimately attained.2,25 ,28,44
Buyers' demand for diesel-powered passenger cars and light-duty trucks
has been strong since about 1979, but there is considerable concern about
possible health effects and urban visibility degradation associated with
emission of particles in diesel exhaust. Diesel particulate-emission
rates are about two orders of magnitude greater than those associated with
catalyst-equipped spark-ignition vehicles. ^>*1»*1
Comparative emission factors are discussed later, but it is evident
from Table 1-3 that there is already extensive use of diesel-engine
vehicles in this country, and it will be well into the 1990s, if ever,
before light-duty diesel particulate emission becomes equivalent in
tonnage to the particulate material from heavy-duty diesel-engine vehicles
nationwide.12,13,44
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Medium-duty trucks (gross vehicle weight, 8,500-33,000 Ib) are now
equipped with spark-ignition engines, but are also undergoing diesel-
ization rapidly. Heavy-duty trucks (>33,000 Ib) are already more than 90%
diesels. In 1980, about 15% of new trucks sold were diesel-engine
vehicles, and this figure could grow to 50% by the year 2000. In
addition, the total number of trucks in the United States is expected to
increase by 5% per year until beyond 2000.44,46,60
Approximately 90% of all commercial buses are powered by diesel
engines; school buses are still powered by spark-ignition engines. Trains
and ships are powered by diesel engines, as is most industrial equipment.
Private boats and planes are powered predominantly by spark-ignition
engines. Commercial aircraft are powered by gas-turbine engines that use
jet fuel.
Although information on engine types used in military vehicles is not
readily available, there is some information in Table 1-3 on military
fuel consumption. Emission from military vehicles is not a major source
of atmospheric PAHs, except possibly in particular areas.
PAH EMISSION FROM MOBILE SOURCES
Internal-combustion engines emit gases, liquids, and solids from the
exhaust system as products of the incomplete combustion of the fuel and as
noncombusted fuel, lubricants, and fuel additives. Chemical processes
also occur in the exhaust system, especially in the catalytic emission-
control devices. Some reactions continue after the exhaust is released
into the atmosphere. The temperatures in the combustion chamber and in
the exhaust system and the volume flow rates depend directly on engine
design, size, operating speed, and working load. These factors are
important in the formation of PAHs and in the amounts of PAHs that are
emitted into the atmosphere.
The combustion process in a spark-ignition engine takes place with
near-stoichiometric amounts of oxygen at temperatures in the vicinity of
3500°C. In diesel engines, there is an excess of oxygen with combustion
temperatures in the vicinity of 2000°C. Exhaust temperatures for
spark-ignition engines are commonly between 400 and 600°C, but diesel
exhaust is typically at 200-400°C (except at high load factors).
Oxidation catalysts typically must be at 400°C or higher before becoming
active, so current types of oxidation or three-way catalysts do not
function efficiently on diesel vehicles.29,43
Many of the PAHs have boiling points of 200-300°C and are
sufficiently volatile to exist predominantly in the gas phase at
temperatures above 200°C. Even at room temperature, some of the more
volatile PAHs are distributed between the vapor and particle-adsorbed
phases.52
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In the open air, vehicular exhaust is diluted by a factor of about
1,000 in the first few seconds, so cooling to near-ambient temperature is
quite rapid. But condensation of PAHs, by adsorption on existing
particles, can occur many feet behind a vehicle, thus allowing some mixing
of exhaust plumes from different emission sources. Normally, black
"elemental carbon" particles, also products of incomplete fuel
combustion, act as condensation nuclei for the condensation of vapor-phase
organic chemicals, such as aliphatic compounds, aromatic compounds
(including the PAHs), aldehydes, ketones, acids, and heterocycles.
The exhaust is the major source of the PAHs; another possible source
of PAHs is engine oil, because it can act as a sink for them. It has been
estimated that crankcase oil collects 10 times as much PAH per mile
traveled as is released from the exhaust system.^^ In the case of
vehicles in which volatile emission from the crankcase is not controlled,
it could be a significant source of PAHs in the atmosphere, but
quantitative assessment is not now possible.
Studies of particle-size distribution of spark-ignition and diesel
exhaust particulate material show mass-median aerodynamic diameters of
0.1-0.25 ym.20>36'63 More than 90% of the mass is in particles less than
1 ym in diameter.'2 Larger particles presumably result from deposition
of particulate material on and later release from the walls of the exhaust
system. Resuspended road dust, roadbed material, and tire particles
result in particle sizes of about 8 ym in median diameter, which can
account for as much as 10% of the measured vehicular respirable
ftfl 7 9
particulate mass in near-road measurements.D0»'*•
Spark-ignition vehicles with oxidation catalysts emit particulate
material that is mostly aqueous sulfuric acid droplets with organic
compounds presumably adsorbed on droplet surfaces. Particle median
diameters tend to be somewhat less than 0.1 yra.^*
Diesel particulate material is mostly elemental carbon. The primary
particles are spherules 0.015-0.03 ym in diameter that agglomerate at high
temperatures to irregular clusters and chains. These clusters, about
0.15 ym in diameter and containing up to 4,000 spherules, act as carriers
of the PAHs and other adsorbed species.^6 The PAHs are adsorbed on the
surface of the carbon and into the minute pores between the spherules.
The small particle size results in long atmospheric residence times and in
deposition in alveolar regions of the lung.3» 13»^,58,64 (See
discussions in Chapter 3 on particles in the atmosphere and in Chapter 5
on relations of deposition of PAHs and particles.)
*This material is not truly elemental carbon, nor is it graphitic
carbon. No term has been found or accepted that properly describes the
material.*2
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Exhaust-emission standards from EPA for HC, CO, and NOX have resulted
in 84, 79, and 562 reductions, respectively, in 50,000-mi emission from
spark-ignition passenger cars, as shown in Table 1-4.^ Particulate
emission and lead emission have also decreased as a result of the use of
catalysts and the decrease in lead concentrations in leaded fuel. Figure
1-2 shows the decrease in urban CO concentrations since 1973,30 and Figure
1-3 shows the decrease in traffic-average lead emission rates as measured in
highway tunnels.'^
SAMPLING OF EXHAUST FROM MOBILE SOURCES
Techniques for sampling exhaust from mobile sources have been thoroughly
described elsewhere and are not reviewed here except as pertinent to the
analysis of PAHs.4»21»53>77
Exhaust-particle sampling in this country commonly involves the use of
dilution tunnels. The dilution tunnel represents a laboratory attempt to
simulate the normal atmospheric dilution and cooling of the exhaust.
Atmospheric dilution is by about 1,000:1 in the first few seconds, whereas
typical laboratory dilutions are between 5:1 and 20:1.19,24 Exhaust from
a vehicle tailpipe is mixed with particle-free, temperature- and humidity-
controlled air in a tunnel that is typically 8-16 in. (20-40 cm) in
diameter. Downstream from the exhaust inlet, a constant fraction of the
diluted exhaust is pumped through a high-efficiency filter to collect
exhaust particles. The weight gain of the filter is a measure of total
particulate emission. Adsorbed organic matter, including PAHs, is isolated
from the carbon particles by solvent extraction or other techniques. The
organic extract material can then be analyzed in many ways, including
high-performance liquid chromatography (HPLC). gas chromatography (GO, mass
spectrometry (MS or GC/MS), and bioassays.5»^.53,77,79 The Salmonella
assay has become a commonly used test in almost all laboratories working
with vehicle-exhaust particulate material.
A technique applied more commonly in Europe uses low-temperature
condensers (as many as three in series, at successively lower temperatures)
followed by filtration.35,54,55 This approach is used either on the
undiluted exhaust or on the exhaust from a dilution tunnel. As mentioned
previously, PAHs with high volatility (molecular weight <250) can be
distributed in the vapor and condensed phases. Therefore, filter-only
sampling from a dilution tunnel misses some of the more volatile PAHs. By
combining the PAHs collected in each of the condensers with the PAHs
collected on the filter, one obtains a better quantitative assessment of PAH
emission. Results from filter-only dilution-tunnel studies are used to
provide a qualitative description of PAHs in exhaust, and condenser-study
results are used to provide quantitative emission rates.
In any attempt to sample a chemical system, it is necessary to show that
the sampling process does not alter the chemical concentrations of the
1-5
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mixture. The production and destruction of chemicals during sampling are
called "artifacts" of the sampling process. Sampling artifacts are
important in a discussion of the production of the nitro-PAHs.
QUALITATIVE DESCRIPTION OF EMISSION FROM MOBILE SOURCES
The combustion of gasoline or diesel fuel in air yields water and carbon
dioxide as the principal combustion products. Nitrogen oxides result from
the high-temperature reaction of nitrogen in the air and from combustion of
nitrogen-containing compounds in the fuel and lubricant. » Carbon
monoxide, gas-phase hydrocarbons, elemental carbon, and particle-adsorbed
organic material are formed as products of the incomplete combustion pro-
cess. Fuel and lubricant additives and impurities and their combustion
products are also found in exhaust. For example, sulfur-containing organic
compounds in the fuel are combusted to gaseous sulfur dioxide, some of which
can be further oxidized to sulfuric acid in the combustion chamber or in the
oxidation catalyst and give rise to sulfuric acid in the particulate
material.
The components detected as gas-phase hydrocarbons are listed in Table
1-5 (from a study of on-road gaseous organic-compound emission). The
quantitative emission rates have not been determined. »
Diesel-exhaust particulate material has been the subject of extensive
study in the last 5 yr. It is typically about 25% extractable into organic
solvents, although different vehicles may have extractable fractions of
10-90%, depending to some extent on operating conditions. More than half
the extractable material is aliphatic hydrocarbons of 14-35 carbon atoms and
alkyl-substituted benzenes and naphthalenes.4,42,78 ^g remaining
extractable mass is PAHs and oxidized derivatives of the PAHs, such as
ketones, carboxaldehydes, acid anhydrides, hydroxy compounds, quinones,
nitrates, and carboxylic acids. There are also heterocyclic compounds
containing sulfur, nitrogen, and oxygen atoms within the aromatic ring. The
alkyl-substituted PAHs and PAH derivatives tend to be more abundant than the
parent PAH compounds.
The particulate-extract HPLC eluent can be separated into nonpolar,
moderately polar, and highly polar fractions. The fractions can then be
further analyzed by GC/MS. Table 1-6 lists the results of such an analysis
of the nonpolar and moderately polar fractions of a particulate extract from
an Oldsmobile- diesel vehicle, including the approximate extract concentra-
tions for this particular vehicle.78 The highly polar fraction has not
been fully characterized. It contains the PAH carboxylic acids, acid
anhydrides, and probably sulfonates and other highly polar species.?8,94
Most (75%) of the direct bacterial mutagenicity resides in the
moderately polar fraction (see discussions of Salmonella strains in Chapter
4).10,71,73,74,76,80,81,88 ^ remaining direct mutagenicity is in the
highly polar fraction. These aspects are discussed further in Chapter 4.
1-6
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Over 50 chromatographic peaks of nitro-PAH compounds have been
identified in diesel particulate extracts, as listed in Table
1-7.22,40,47,65,75,93 i-Nitropyrene is the most abundant of the
nitro-PAHs, ranging from 25 to 2,000 ppra in the vehicle extracts studied.
The other nitro-PAHs are present at concentrations from below the parts-
per-million range to a few parts per million. The nitropyrenes have been
studied in greater detail. They are released in diesel and gasoline exhaust
(according to particulate extracts) at approximately 8.0 and 0.30 ug/mi,
respectively. The latter value was obtained with leaded gasoline; with
unleaded fuel, the rate was 0.20
Gibson^" has determined the concentration of 1-nitropyrene in ambient
particulate extracts obtained from suburban areas in Michigan to be
0.016-0.030 ng/ra-* of air (corresponding to 0.2-0.6 ng/rag of particles.
Gibson has also observed that catalytic converters greatly reduce the
concentration of nitropyrenes. The nitropyrene concentration in extracts
from particles obtained from the emission of a wood-burning fireplace was
less than 0.1 ng/mg of particles. 26
1-Nitropyrene has been the only nitro-PAH detected in spark-ignition
particulate extracts. 51,95 Qn-road heavy-duty diesel and light-duty
spark-ignition vehicles have recently been found to have very low
1-nitropyrene particulate extract concentrations, which thus account for
very small fractions of the on-road direct bacterial mutagenicity associated
with these vehicle categories. *
QUANTITATIVE DESCRIPTION OF PAH EMISSION FROM MOBILE SOURCES
The work of Grimmer and co-workers,33,34 of Kraft and Lies,48 and
* e-i ne '
more recently of Zweidinger and colleagues > can be used to derive
typical rates of emission of many of the PAHs from the different categories
of mobile sources. Because a relatively small number of vehicles have been
used to measure these emission rates, the uncertainty in the derived
vehicle-category emission factors is quite large—probably at least a factor
of 2 and possibly even larger. Table 1-8 lists the best current
measurements of rates of emission of numerous PAHs and their derivatives for
spark-ignition vehicles (light-duty with and without catalysts and heavy-
duty) and for light-duty and heavy-duty diesel vehicles.6.16-18,39,90,91
For the other categories of mobile sources, the estimate of total PAH
emission can be based on the heavy-duty spark-ignition or heavy-duty diesel
emission rate per gallon of fuel consumed and the total fuel consumption of
the category (railroads, aircraft, etc.). When more than one value for a
particular PAH emission rate is available for a source category, the
micrograms-per-gallon-of-fuel figures are averaged. Kraft and Lies found a
very similar distribution of the PAHs for diesel and gasoline vehicles.
Owing to the paucity of emission-rate measurements, we used this observation
to derive emission factors for vehicle categories when measurements are
lacking. Table 1-9 lists the resulting emission-rate estimates for the
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different vehicle categories. The assumptions made in deriving these
estimates were:
• That the PAH distributions for both spark-ignition and diesel
vehicles are the same as the distributions for average light-duty spark-
ignition noncatalyst vehicles.
• That the measured BaP emission rates for the oxidation-catalyst
spark-ignition vehicles and for the three-way-catalyst classes represent the
reductions in all PAH emission rates, compared with the noncatalyst values.
• That the heavy-duty spark-ignition class has the same fuel-
specific emission rates (in micrograms per gallon) as the light-duty spark-
ignition noncatalyst vehicles.
• That the heavy-duty diesel class has the same fuel-specific emission
rates as the light-duty diesel class.
Within the limits of those assumptions, we have a complete list of PAH
emission factors for each of the vehicle categories in terms of micrograms
per gallon of fuel consumed. With typical fuel-economy values for each
class, one can calculate the micrograms per mile for each class. These
results are also listed in Table 1-9. The 1-nitropyrene values are those
from actual experimental measurements, unless a derived value was higher
than the measured value. Therefore, the resulting 1-nitropyrene emission
rate should be considered an upper limit.
Fuel-specific PAH emission rates can be combined with the total
fuel-consumption values in Table 1-3. That yields a total emission tonnage
for each mobile-source category for each PAH and PAH derivative in Table
1-9. The PAHs released from mobile sources in 1979 according to these
estimates are listed in Table 1-10. The total BaP emission from all mobile
sources is estimated to be 43 metric tons. This encompasses all mobile
sources, whereas the motor-vehicle contribution is 27 metric tons (about 63Z
of the total) exclusive of the railroad, aircraft, ship, farm, military, and
other contributions. Motor-vehicle BaP emission was estimated in 1972 at
about 20 metric tons/yr. The calculated mobile-source emission of
1-nitropyrene is 17 metric tons, of which 30Z is calculated to be
contributed by motor vehicles. The non-motor-vehicle categories tend to be
less relevant to polluted-air concentrations, because they are used away
from urban areas (railroads, ships, farm machinery) or because their
emission is dispersed above the boundary layer (aircraft). In addition, the
various motor-vehicle categories are used to various extents in urban
areas. Passenger-car use is 60Z urban, light-truck use 55Z urban, and
heavy-truck use only 20Z urban.™ The urban fraction of the total
motor-vehicle PAH emission is calculated to be 63Z for 1979, owing mainly to
the dominance of noncatalyst-passenger-car emission of PAHs. The relative
contributions of each of the mobile-source categories to PAH and 1-nitro-
pyrene emission are listed in Table 1-11. The 70Z contribution of the
non-motor-vehicle sources to the 1-nitropyrene emission may be an artifact
of the method used for calculating emission rates, inasmuch as 1-nitro-
pyrene from sources other than passenger cars and trucks has not been
investigated.
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By far the largest single contribution to PAH emission (from mobile
sources) is that from noncatalyst spark-ignition passenger cars, which will
soon be supplanted by vehicles equipped with oxidation catalysts and
three-way catalysts. The second most important category for motor-vehicle
PAH emission consists of spark-ignition light trucks. As much as half the
spark-ignition light trucks (all those under 8,500 Ib) will have catalysts
by the year 2000. The relative importance of the heavy-truck diesels can be
expected to increase with that of the light-truck diesel and passenger-car
diesel categories.
Comparison of the estimated BaP emission factors in Table 1-9 for each
of the motor-vehicle categories with values reported in the literature
indicates that the present estimates tend to be on the high side of what
might be expected for fleet-average values. (For the purpose of the
estimates in this report, overestimates are obviously preferable to
underestimates.) For example, light-duty diesel BaP emission rates range
from less than 1 yg/mi to more than 20 yg/mi, with mean values reported in
the vicinity of 3-4 yg/mi. The present estimate is 13 yg/mi for light-duty
diesels. The few measurements of BaP emission rates for heavy-duty diesels
that have been reported18 indicate that the 54-yg/mi value in Table 1-9
may be too high by as much as an order of magnitude. The reason for this
discrepancy is not apparent, but it may reflect a real difference between
the four-stroke indirect-injection light-duty diesel and the two- or
four-stroke direct-injection heavy-duty diesel. If the lower emission rates
are correct, the role of heavy-duty diesel emission is considerably less
than portrayed in later sections of this report.
It is now possible to use the emission rates in Table 1-9 and the
projections previously described to estimate future rates of emission from
motor vehicles. Using the current BaP emission rates, we have calculated
the motor-vehicle BaP emission for the year 2000 and listed the results in
Table 1-12. The 24 metric tons of BaP represents an 11% decrease from the
1979 value of 27 metric tons and reflects the benefit of catalyst-equipped
spark-ignition passenger cars over their noncatalyst counterparts, which is
partially offset by the incursion of diesel vehicles. In the year 2000,
without further particulate-emission controls, diesel vehicles will account
for 40% of the mileage, 50% of the fuel consumption, and 80% of the total
motor-vehicle BaP emission, according to this estimate. If the present
distribution of motor-vehicle use between urban and rural areas will still
be valid in the year 2000, it can be estimated that about 40% of the BaP
from motor vehicles will be released in urban areas in the year 2000,
compared with 63% in 1979. Thus, the BaP tonnage nationwide will decrease
slightly and there will be a shift to more rural emission and away from
urban areas.
The total-tonnage estimates just described do not assess directly the
problem of human exposure to air pollutants. In this regard, emission rates
are not the sole important quantities. What is needed is an estimate of
atmospheric concentrations in the air inhaled by people. The results of
atmospheric-dispersion modeling by Ingalls and Garbe can be used to
1-9
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calculate atmospheric concentrations resulting from motor-vehicle emission
in many of the typical urban-exposure situations.^ ihe model was
constructed on the basis of a hypothetical 1-g/mi traffic-weighted emission
rate for 1980 vehicle distributions. To calculate atmospheric concentra-
tions for an emission component with other than a 1-g/mi emission rate, one
need only multiply the Ingalls and Garbe exposure-concentration factor by
the actual traffic-weighted emission rate in grams per mile. Table 1-13
lists the exposure conditions modeled by Ingalls and Garbe and the 1-g/mi
exposure-concentration factors derived. One can calculate exposure to BaP
on the basis of the data in Tables 1-13 through 1-15. In the calculation,
the effect of binding of BaP to particles on deposition and absorption is
not considered. Because retention of particle-bound BaP, and thus
absorption, depends heavily on particle size and because particle size
varies widely, we have characterized exposure on the assumption of complete
retention. We assume that 90% of the BaP is bound to particles less than
1 ym in diameter.'^ To use these results for calculation of BaP
exposures, one uses Table 1-14 to derive the traffic-weighted BaP emission
rate for 1979. The same data for the year 2000 are listed in Table 1-15.
The 1979 exposure concentration of BaP in a typical roadway tunnel is
0.017 vg/m3 [(15.3 Ug/mi)(10~6 g/yg)(l,123 g/m3 per g/mi)]. A person
exposed to a concentration of 0.017 _ug/m3 for 2 min while breathing at the
rate of 15 m3/d would inhale 0.4 ng of BaP [(0.017 pg/m3)(15 m3/d)
(1/24)(1/60)(2 min)(103 ng/ug)]. The total daily exposure of a person can
be calculated by summing over each of the exposure situations experienced in
the course of the day. The result of these calculations is a degree of
exposure by inhalation. The dose of BaP to the body would be less and would
depend on the fraction of the BaP-laden particles that is deposited in the
body. This fraction is highly uncertain and depends on particle size,
shape, and hygroscopicity and on BaP loading per particle, which also
depends on particle size. The fraction of BaP deposited is probably about
20-50% of the BaP inhaled. This has been done in Table 1-16 for a person
living in a suburb (1,000 m from an expressway) with a 1-h commute to a job
at street level in a central-city street canyon. These are rather severe
conditions and result in higher exposures than would be expected for the
average urban dweller. These conditions lead to a calculated inhalation of
20 ng of BaP. Had the person stayed home all day, the exposure would have
resulted in a 3.0-ng inhalation. For comparison, the BaP inhalation from
one cigarette is 20 ng. Smoking 1 cigarette/d has the effect of being
exposed to over 15 ng/m3 for the entire day (see discussion on BaP
exposure from smoking in Appendix C)—an inhalation of over 330 ng of BaP.
For the traffic composition projected to exist in the year 2000, the
calculated workday exposure of the person is 9.1 ng of BaP, a 55% decrease
from the 1979 value. This shows again the decrease in urban exposure at the
expense of an increase in rural exposure. The 9.1-ng BaP inhalation when
combined with the 15-m3/d inhalation rate gives a calculated average daily
atmospheric concentration of 0.6 ng/m3 in the year 2000.
Calculations of BaP total motor-vehicle emission tonnage, urban
fractions, and inhalation exposures for the year 2000 have assumed no
changes from the present fuel-specific emission rates for the various
motor-vehicle categories. Even modest particulate-emission controls (50%
1-10
-------
reductions) for diesel cars and trucks and for spark-ignition light trucks
would result in significantly greater reductions in BaP (and other PAH)
exposures than those calculated here. This of course assumes that
compliance (i.e., not removing or poisoning catalytic converters) approaches
100%. The benefit of such controls on the basis of BaP exposures will need
to be assessed.
FUTURE CONTROL TECHNOLOGIES
The need for control of emission from light-duty spark-ignition vehicles
seems to be moot, with all new vehicles being sold now having catalyst
systems. Use of catalysts on heavy-duty spark-ignition vehicles, if
feasible, would be expected to result in PAH reductions comparable with
those observed for light-duty spark-ignition vehicles. Control of diesel
particulate material has received much attention recently. The light-duty
diesel particulate-emission standard of 0.6 g/mi that went into effect for
the 1982 model year was achieved by most diesel manufacturers through engine
modifications. The 0.2-g/mi standard proposed for 1985 would not be as
readily attained, at least for the larger vehicles. Currently, only diesels
of less than 2,600 Ib could meet a 0.2-g/mi particulate-emission standard
without exceeding the gaseous-emission standards. Variations in diesel
fuel appear to be inadequate to allow attainment of the standard for all but
the smaller diesel vehicles.
A number of diesel-particle control techniques are under investigation.
In general, these entail after-treatment devices designed to collect
particles from the exhaust stream and to oxidize the collected material
periodically. Diesel-particle control devices are being developed by
several companies and are being tested by automobile manufacturers. Texaco
has reported results on an alumina-coated metal-wool diesel-particle filter
that achieved a collection efficiency of about 70% without increasing the
backpressure enough to sacrifice fuel economy and performance. The col-
lected particulate material must then be removed by combustion." »",87
Johnson-Matthey is developing a wire-mesh particle trap that is coated with
a catalytic material to initiate the combustion of the collected soot.
Efficiencies exceeding 50% have been achieved during 50,000-mi accumulation
with regeneration every 300-1,000 mi. Corning has developed a ceramic-
honeycomb monolithic particle filter that can be coated with a catalyst
material to assist in soot combustion. ' Gorse et^ _£!• and
Williams recently reported emission characterization studies that used
some of the above-mentioned control devices. The ceramic trap removes more
than 90% of the elemental carbon particles and about 50% (with the catalyst
coating) of the particulate organic material and can result in an order-of-
magnitude decrease in the emission of bacterial mutagens per mile of travel
The wire-mesh catalyst trap removes more than 90% of the particulate organic
material and 30% of the elemental carbon. Some of the catalyst-coated traps
can produce very high sulfate emission rates, especially during regeneration.
In general, there appears to be some hope of success for diesel-particle
control, but the devices tested so far need to be tested for durability,
1-11
-------
packaging, and on-road reliability. None of the devices has been evaluated
for use with heavy-duty diesel particle emission. If diesel-particle
control devices are successfully developed and used on light- and heavy-duty
diesel vehicles, reductions in the PAH emission factor by at least a factor
of 2, and conceivably a factor of 10 or better, could be realized.
Cost estimates of the diesel-particle control techniques are premature;
the technology of choice has not yet been determined. Johnson-Matthey
claims that the cost of its device could be as low as $150 for light-duty
diesels. ° General Motors estimates that the cost of light-duty
diesel-particle control could be as high as $830/vehicle to meet the
0.2-g/mi particle standard and the 1.0-g/mi NOX standard at low and high
altitude.^3 The control device itself would represent about half the
total cost, and the control modifications needed to ensure the functioning
of the trap would represent the other half.
1-12
-------
Energy Source
Oil
Coal
Natural gas
Hydrogeneration
Nuclear power
Wood
Other
TABLE 1-1
U.S. Energy Consumption, 1979a
Energy Content,
Consumption
104 BTU/unit
9.8/gal
1.3/lb
O.I/ft3
—
—
1.0/lb
— _
1015 BTU
34.2
15.6
20.4
3.1
2.7
0.17
0.12
Quantity
3.5 x
1.2 x
2.0 x
—
—
1.7 x
— —
11
1011 gal
1012 Ib
1013 ft3
1010 Ib
Total
76
aData from Motor Vehicle Manufacturers Association.^
1-13
-------
TABLE 1-2
Uses of Major Crude-Oil Fractions in the United States, 1979a
Consumption, 1Q9 gal
Destination
Residential
Commercial
Industrial
Electric utility
Oil company
Farm
Military
Rail
Marine
Highway vehicles
Off-highway vehicles
Other
Totals e
Gasoline
_^ _
—
—
—
—
1.3
—
—
—
104.2
1.6
5.5
112.6
141.9
lanufacture:
Distillate
Oil
17
5.9
6.2
2.1
1.1
4.3
0.9
4.9
2.1
8.6d
3.4
1.8
58.3
66.4
rs Association*^
Residual
Oil
_ _
5.1C
11
22
3.3
—
0.2
0.01
7.7
—
—
1.0
50.3
56.7
and National
Other
0.8
0.1
0.1
—
—
—
—
—
—
—
—
100.0
101.0
106.5
Petroleum News.
"Includes jet fuels, kerosene, lubricants, asphalt, etc.
cCombined total for residential and commercial categories.
dFederal Highway Administration data show 18.3 x 10^ gal for highway
use, which is compatible with Table 1-3.
eCompiled from Department of Energy data.
from University of Houston Downtown College, Energy Information
Services, U.S. Annual Energy Facts.
1-14
-------
Fuel User
Highway vehicles:
Passenger cars
Trucks
<33,000 Ib
Trucks
>33,000 Ib
Buses
Motorcycles
Other:
Railroads
Ships
Aircraft
Farm vehicles
Military vehicles
Other
Totals
TABLE 1-3
Mileage and Fuel Consumption, 1979s
Fuel
Gasoline
Diesel
Gasoline
Diesel
Gasoline
Diesel
Gasoline
Diesel
Gasoline
Diesel
Gasoline
Diesel
Gasoline
Jet
Gasoline
Diesel
Diesel
Gasoline
Diesel
Gasoline
Diesel
Jet
Mileage,
1010 mi/yr
112.9
1.1
28.8
0.59
0.34
6.4
0.30
0.31
2.2
152.9
Fuel Consumption,
109 gal/yr
79.0
0.81
23.2
0.74
1.1
15.9
0.41
0.62
0.44
4.4
0.93
8.7
0.77
22.0
1.3
4.3
0.85
5.4
3.5
112.6
39.8
22.0
aData from Jambekar and Johnson,Motor Vehicle Manufacturers Associa-
tion,60 National Petroleum News.61 and Shelton.82'83
1-15
-------
TABLE 1-4
Exhaust Emission Rates for Light-Duty Gasoline-Powered Vehicles3
Emission
Component
HC
CO
NO,
Model
Year
Pre-1968
1968-1969
1970-1971
1972-1974
1975-1979
1980
1981
1982+
Pre-1968
1968-1969
1970-1971
1972-1974
1975-1979
1980
1981
1982
1983+
Pre-1968
1968-1972
1973-1974
1975-1976
1977-1979
1980
1981+
Zero-Mile
Emission Rate,
g/mi
7.25
4.43
3.00
3.36
1.29
0.29
0.39
0.39
78.27
56.34
42.17
40.78
20.16
6.14
60
21
5.00
3.44
4.35
.87
.43
.69
.56
2.
2.
1.
1,
0.75
50,000-Mile
Emission Rate,"
g/mi
8.
5,
4.
4.
2.
1.
1.
15
68
85
21
74
74
34
1.34
89.52
69.09
57.82
52.98
34.46
20.44
19.35
19.01
18.80
3.44
4.35
3.07
2,
2,
2,
63
,19
06
1.50
"Data from U.S. Environmental Protection Agency.^ Emission rates are for
low-altitude 49-state vehicles. High-altitude and California emission rates
are different.
The 50,000-mile emission rates are calculated from zero-mile rate by
addition of term that takes account of EPA-projected deterioration rate
of vehicle combustion and emission-control systems.
1-16
-------
TABLE 1-5
Summary of Gaseous Hydrocarbons Emitted from Vehicles
All ti-alkanes from n-butane through ji-hexacosane
Four methyl-substituted butanes
Ten methyl- and ethyl-substituted pentanes and 11 cyclopentanes
Eleven methyl- and ethyl-substituted hexanes and 35 cyclohexanes
Fifteen methyl- and ethyl-substituted heptanes
Five methyl-substituted octanes
One methyl-substituted nonane
One methyl-substituted decane
One methyl-substituted undecane
Decalin and two methyl-substituted decalins
Two CIQ alkanes
Eleven C^ alkanes
Nine C^ alkanes
Thirteen Cjj alkanes
Eleven Cj^ alkanes
Eight C}5 alkanes
Eight C^g alkanes
Five C^7 alkanes
Three C^g alkanes
Seven methyl-substituted butenes and two methyl butadienes
Eighteen pentenes and pentadiene
Fourteen hexenes
Six heptenes
Four octenes
Decene and dodecene through heneicosene
Seven cyclic olefins
Seventy-one alkyl-substituted benzenes
Eight styrenes and the three xylenes
Fourteen indans and three indenes
Twenty-eight alkyl-substituted naphthalenes
Three alkylthiophenes and two benzothiophenes
Two alkylsulfides and one alkylamine
Six nonaromatic alcohols and eight aromatic alcohols
Eighteen aliphatic and aromatic aldehydes
Six furans, 17 ketones, and six organic acids
1-17
-------
TABLE 1-6
Qualitative Analysis of Nonpolar and Moderately Polar Fractions
of Diesel Particulate Extract
Approximate Concentration
Compounds in Oldsmobile Extract, ppm
Nonpolar fractions:
Phenanthrenes and anthracenes 600
Methylphenanthrenes and methylanthracenes 1,400
Dimethylphenanthrenes and dimethyl-
anthracenes 3,000
Pyrene 1,700
Fluoranthene 1,400
Methylpyrenes and methylfluoranthenes 800
Chrysene 100
Cyclopenta[cd]pyrene 20
Benzo[ghi]fluoranthene 100
Benz[alanthracene 500
Benzo[a]pyrene 40
Other PAHs, heterocyclics 30,000
Hydrocarbons and alkylbenzenes 500,000
Total, nonpolar fractions 539,700
Moderately polar fractions:
PAH ketones:
Fluorenones
Methylfluorenones
DimethyIfluorenones
Anthrones and phenanthrones
Methylanthrones and methylphenanthrones
Dimethylanthrones and dimethylphenanthrones
Fluoranthones and pyrones
Benzanthrones
Xanthones
Methylxanthones
Thioxanthones
Methylthioxanthones
13,500
1-18
-------
TABLE 1-6 (continued)
Compound
PAH carboxaldehydes;
Fluorene carboxaldehydes
Methyl fluorene carboxaldehydes
Phenanthrene and anthracene carboxaldehydes
Methylanthracene and methylphenanthrene
carboxaldehydes
Dimethylanthracene and dimethylphenanthrene
carboxaldehydes
BaA, chrysene, and triphenylene
carboxaldehydes
Naphthalene dicarboxaldehydes
DimethyInaphthalene carboxaldehydes
Trimethylnaphthalene carboxaldehydes
Pyrene and fluoranthene carboxaldehydes
Xanthene carboxaldehydes
Dibenzofuran carboxaldehydes
PAH acid anhydrides:
Naphthalene dicarboxylic acid anhydrides
Methylnaphthalene dicarboxylic acid
anhydrides
Dimethylnaphthalene dicarboxylic acid
anhydrides
Anthracene and phenanthrene dicarboxylic
acid anhydrides
Hydroxy PAHs:
Hydroxyfluorene
Methylhydroxyfluorene
DimethyIhydroxyfluorene
Hydroxyanthracenes and hydroxyphenanthrenes
Hydroxymethylanthracenes and hydroxy-
methylphenanthrenes
Hydroxydimethylanthracenes and hydroxy-
dimethylphenanthrenes
Hydroxyfluorenone
Hydroxyxanthone
Hydroxyxanthene
Approximate Concentration
in Oldsmobile Extract, ppm
1,600
400
2,600
1,600
400
400
300
300
1,000
1,600
600
400
11,200
3,000
1,000
500
600
5,100
1,400
400
1,500
600
900
1,300
000
300
1,000
10,400
1-19
-------
TABLE 1-6 (continued)
Approximate Concentration
Compound in Oldsmobile Extract, ppm
PAH quinones:
Fluorene quinones 700
Methylfluorene quinones 600
Dimethylfluorene quinones 500
Anthracene and phenanthrene quinones 1,900
Methylanthracene and methylphenanthrene
quinones 2,000
Fluoranthene and pyrene quinones 200
Naphtho[l,8-cd]pyrene 1,3-dione 600
6,500
Nitro PAHa:
Nitrofluorenes 30
Nitroanthracenes and nitrophenanthrenes 70
Nitrofluoranthenes 10
Nitropyrenes 150
Methylnitropyrenes and methylnitro-
fluoranthenes 20
300
Other oxygenated PAHs: 8,000 8,000
PAH carryover from nonpolar fraction; 6,000 6,000
Phthalates, HC contaminants: 30,000 30,000
Total, moderately polar fractions 91,000
1-20
-------
TABLE 1-7
Nitroarenes Indicated in Diesel-Exhaust Particulate Extracts
Mononitroarenes:
Nitroindene
Nitroacenaphthylene
Nitroacenaphthene
Nitrobiphenyl
Nitrofluorene
Nitroraethylacenaphthylene
Nitromethylacenaphthene
Nitromethylbiphenyl
Nitroanthracene
Nitrophenanthrene
NitromethyIfluorene
Nitromethylanthracene
NitromethyIphenanthrene
Nitrotrimethylnaphthalene
Nitrofluoranthene
Nitropyrene
Nitro(C2-alkyl)anthracene
Nitro(C2~alkyl)phenanthrene
Nitrobenzofluorene
NitromethyIfluoranthene
NitromethyIpyrene
Nitro(C3~alkyl)anthracene
Nitro(C3~alkyl)phenanthrene
Nitrochrysene
Nitrobenzoanthracene
Nitronaphthacene
Nitrotriphenylene
Nitromethylnaphthacene or
Nitromethylchrysene
NitromethyIbenzanthracene
Nitromethyltriphenylene
Nitrobenzopyrene
Nitroperylene
Nitrobenzofluoranthene
Polynitroarenes:
DinitromethyInaphthaiene
Dinitrofluorene
DinitromethyIbiphenyl
Dinitrophenanthrene
Dinitropyrene
Trinitropyrene
Trinitro(C5-alkyl)fluorene
Dinitro(Cg-alkyl)fluorene
Dinitro(C^-alkyl)pyrene
Nitro-oxyarenes:
Nitronaphthaquinone
Nitrodihydroxynaphthalene
Nitronaphthalic acid
Ni trofluorenone
Nitroanthrone
Nitrophenanthrone
Nitroanthraquinone
NitrohydroxymethyIfluorene
Nitrofluoranthone
Nitropyrone
Nitrofluoranthenequinone
Nitropyrenequinone
Nitrodiraethylanthracene
carboxaldehyde
NitrodimethyIphenanthrene
carboxaldehyde
Other nitrogen compounds:
Benzocinnoline
Methylbenzocinnoline
PhenyInaphthylamine
(C2-Alkyl)phenylnaphthyl-
amine
1-21
-------
TABLE 1-8
Meaaured PAH Eaiaaioa Ratea for Habile Source*
PAH
Anthracene
Pheoanthrene
Methylphenanthrenc
DIM thy If luorene
DiBethylphenanthrcne
Fluoranthene
Pyreae
Benzofluorene
Benzoanthracane
Triphenylene
Cyclopentapyreae
Chryaene
Indeoofluoranthene
Indenopyrene
Methylchryieae
1-Hitropyrene
Benzofluoranthene
Benzolejpyrene
Benzol a]pyreae
Perylene
Cyclopentabenzopyrene
Benzochryaen*
Anthracene
Dibenzanthracene
Benzoperylene
Coroaene
Cyclopen t abeniope rylene
Eaiaaion Rate, \ig/g
al of fuel conauned
Spark Ignition
Light Duty
Leaded Fuel
Noncatalyat
• be
2.251 2,226
8.163 10.005
5.678
2.972
2.014
4,221 5,152
5,066 9,528 331
750
231
189
1,653 3.290
1,689 394
83
281 223
38
Unleaded Fuel
Oxidation Three-Way
Catalyat Catalyat
d Avg. a a
2.239
9.084
5,678
2,972
2,014
4,900 4,758 96 12
3,500 6.031
750
231
189
2.472
735 939
83
70 191
38
Heavy Duty
Leaded Fuel Dieiel, Noncatalyat
Honcatalyat Light Duty Heavy Duty
e f g d Avg. g
f 2.31l] 4,400 4,400
] 13. 436] 4.320 6.689
1.160 1.160
152 152
3.5
3.5
4.3 0.4
422
338
281
197
79
1,126
704
920
182
250 261
42
163
4
83
848
715
114
665 669
455 325
298 273
50
5.9
198
11
196
88
27
163
4
140
79
723
502
114
222
135
840
360
300
16
271 48
840
360
321 133
16
340 340
84 84
-------
Footnotes to Table 1-8:
aData from Grimmer.^3
^Data from Grimmer e_£ al.^
cData from Lang e_£ al.^l
dData from Kraft and Lies.48
eData from Dietzmann et_ al.16~18
fData from Braddock.6
SData from Williams and Chock,90 Hare and Baines,39 and Williams and Swarin.91
i
to
OJ
-------
TABU 1-9
Derived PAH Emission Rate* for Mobile Source*
Eaiaaion Rate*
Spark Ignition
Light Duty
Leaded
Fuel
Honcatalyat
PAH
Anthracene
Phenanthrene
Mechylphenanthrene
Dive thy If luorene
Dioethy Iphenanthrene
Fluoranthene
Pyrene
Be nzof luorene
Benzoanthracene
Triphenylene
Cyc lopentapyrene
Chryaene
I ndenof luoranthene
Indenopyrene
Nethylchryaene
1-Nitropyrene
Be nzof luoranthene
Benzo ( e ) pyrene
Benzol a] pyrene
Perylene
Cyclopentabenzopyrene
Bencochryaene
Anthanthrene
Dtbenzanthracene
Benzoperylena
Coronene
Cyc lopentaben«o-
perylene
Mg/g«l
2,239
9,084
5,678
2,972
2,014
4,758
6,031
750
231
189
2,472
939
83
191
38
3.5
699
325
273
27
163
4
140
79
723
502
114
ug/"i
160
649
406
212
144
340
431
54
17
14
177
67
6
14
3
0.3
48
23
20
2
12
0.3
10
6
52
36
8
Unleaded Fuel
Oxidation
Catalyat
jig/gal
410
1,664
1,040
544
369
871
1.105
137
42
35
453
172
15
35
7
4
123
60
50
5
30
1
26
14
132
92
21
pg/au
23
92
58
30
21
48
61
8
2
2
25
10
0.8
2
1.4
0.2
7
3
3
0.3
2
0.1
1
0.8
7
5
1
Three-Way
Catalyat
pg/gal
48
196
123
64
44
103
130
16
5
4
53
20
2
4
1
0.4
15
7
6
1
4
0.1
3
2
16
11
2
Mg/«u
3
10
6
3
2
5
7
0.8
0.3
0.2
3
1
0.1
0.2
0.1
0.0
0.8
0.4
0.3
0.0
0.2
0.0
0.2
0.1
0.8
0.6
0.1
Heavy Duty
Leaded
Fuel
Noncatalyat
pg/gal
2,239
9,084
5,678
2,972
2,014
4,758
6,031
750
231
189
2,472
939
83
191
38
4
669
325
273
27
163
4
140
79
723
502
114
pg/ai
448
1.817
1.136
594
403
952
1,206
150
46
38
494
188
17
38
8
0.8
134
65
55
5
33
0.8
28
16
145
100
23
Dieael,
Noncatalyat
Light-Duty
Mg/g«l
2,633
10,681
6,676
3,495
2,368
5,595
7,091
882
272
222
2,907
1,104
98
225
45
271
840
382
321
32
192
5
165
93
850
590
134
pg/au
105
427
267
140
95
224
284
35
11
9
116
44
4
9
2
11
34
15
13
1
8
0.2
7
4
34
24
5
Heavy Duty
pg/g«l
2,633
10,681
6,676
3,495
2,368
5,595
7,091
882
272
222
2.907
1.104
98
225
45
271
840
382
321
32
192
5
165
93
850
590
134
pg/ai
439
1.780
1.113
583
395
933
1.182
147
45
37
485
184
16
38
8
45
140
64
54
5
32
0.8
28
16
142
98
22
Motorcycles ,
Leaded Fuel
pg/gal
2,239
9,084
1,113
2,972
2,014
4,758
6,031
750
231
189
2.472
939
83
191
38
3.5
669
325
273
27
163
4
140
79
723
502
114
Mg/"
45
182
114
59
40
95
121
15
5
4
49
19
2
4
0.
0.
14
7
5
0.
3
0.
3
2
14
10
2
,i
8
1
5
1
*Wg/I«l • VI of PAH per gallon of fuel conauawd. Mg/aii • (ug/g«U/(«i/gal).
-------
TABLE 1-10
Estimated PAH Emission from Mobile Sources, 1979
Total Emission,
PAH metric tons
Anthracene 350
Phenanthrene 1,400
Methylphenanthrene 900
Dimethylfluorene 470
Dimethylphenanthrene 320
Fluoranthene 750
Pyrene 950
Benzofluorene 120
Benzanthracene 37
Triphenylene 30
Cyclopentapyrene 390
Chrysene 150
Indenofluoranthene 19
Indenopyrene 30
Methylchrysene 6
1-Nitropyrene 17
Benzofluoranthene 110
Benzo[e]pyrene 52
Benzo[a]pyrene 43
Perylene 4
Cyclopentabenzopyrene 26
Benzochrysene 1
Anthanthrene 22
Dibenzanthracene 13
Benzoperylene 110
Coronene 80
Cyclopentabenzoperylene 18
Totala 6,400
aTotal for all PAHs: about (3)(6,400) - about 19,000 metric tons.
Benzo[a]pyrene from mobile sources is therefore about 0.2Z of total
PAHs.
1-25
-------
TABLE 1-11
Contributions of Mobile-Source Categories to PAH
and 1-Nitropyrene Emission, 1979
PAH Emission, 1-Nitropyrene Emission,
Category X 2
Motor vehicles;
Passenger cars:
Noncatalyst 29.1 0.9
Oxidation catalyst 3.8 0.6
Diesel 0.6 1.3
Trucks, <33,000 Ib:
Spark-ignition 14.7 0.5
Diesel 0.6 1.2
Trucks, >33,000 Ib:
Spark-ignition 0.7 0.02
Diesel 11.8 25.1
Buses:
Spark-ignition 0.3 0.01
Diesel 0.6 1.0
Motorcycles 0.3 62.5 0.01 30.6
Other mobile sources:
Railroads 3.3 6.9
Ships 7.1
Aircraft 16.8
Farm 4.0
Military 0.6
Miscellaneous 6.0 37.8 5.6 69.0
100.3 99.6
1-26
-------
TABLE 1-12
Projected Vehicle Use and Benzo[a]pyrene Emission for the Year 2000
Vehicle Type
Totals
Fuel
Fuel BenzofaJpyrene
Mileage, Fuel Economy, Consumption, Emitted,
mi/yr mi/gal 10^ gal/yr metric tons
Passenger Cars
Trucks,
< 33, 000 Ib
Trucks,
> 33, 000 Ib
Buses
Motorcycles
Gasoline
Diesel
Gasoline
Diesel
Diesel
Gasoline
Diesel
Gasoline
101.2s
39.0
50. Ob
50.0
18.8
0.4
1.7
3.0
30
40
20
30
6
7
10
50
33.7
9.8
25.0
16.7
31.3
0.6
1.7
0.6
1.0
3.1
3.9
5.4
10.0
0.2
0.5
0.2
264
119
24.3
a50Z with oxidation catalysts, 50% with three-way catalysts.
b50Z with oxidation catalysts, 50% without catalysts.
Assumptions:
1.5%/yr increase in passenger-car and motorcycle mileage.
6%/yr increase in light-truck and bus mileage.
5%/yr increase in heavy-truck mileage.
Fuel-economy improvements, as listed in column 4.
Benzo[a]pyrene emission rates (pg/gal) unchanged from values in
Table 1-9.
-------
TABLE 1-13
Summary of Microscale Exposure-Concentration Factors3
Exppsure Situation
1.
6.
Residential garage:
Typical (30-s run time)
Severe (5-min run time)
Parking garage:
Typical (parking level)
Severe: inlet-air component
exhaust-emission component
Roadway tunnel:
Typical
Severe
Street canyon (sidewalk receptor,
includes background):
Typical: 800 vehicles/h
1,600 vehicles/h
Severe: 1,200 vehicles/h
2,400 vehicles/h
On expressway (wind:
2.2 mph):
Typical
Severe
Beside expressway:
Severe: 1 m
10 m
100 m
1,000 m
315 deg relative,
Exposure-Concentration Factor,'5
per 1 g/mi
7,900
67,000
3,900
9,600
46,100
1,123
2,856
42
85
141
282
124
506
Short-Term
397
334
105
13.6
Annual
61
48
14
1.6
aAdapted from Ingalls and
''For 1 g/vehicle-mile (1 g/vehicle-minute for idle conditions). Assumes no
background concentrations except as noted. To use these values with emis-
sion factors other than 1 g/mi (or 1 g/min), multiply the concentration
factor by the actual emission factor in grams per mile (or grams per minute
for idle conditions).
1-28
-------
TABLE 1-14
Urban Motor-Vehicle Data, 1979
Vehicle Type
Fuel
Urban Mileage Fuel Economy, Consumption, Benzofajpyrene,
Fraction mi/gal 1Q9 gal metric tons
Passenger cars:
Noncatalyst, spark-ignition 0.6
Oxidation catalyst,
ignition
Diesel
Trucks, <33,000 Ib:
Spark-igni t ion
Diesel
Trucks, >33,000 Ib:
Spark-ignition
Diesel
Buses :
Spark-ignition
Diesel
Motorcycles
Totals
Urban-traffic weighted
13.46 metric tons
spark-
0.6
0.6
0.55
0.55
0.2
0.2
0.6
0.6
0.8
-
benzofa] pyrene emission
13.5
15.4
16.3
12.4
8.0
3.1
4.0
7.3
5.0
50.0
rate:
26.8
20.4
0.4
12.7
0.4
0.3
3.3
0.3
0.4
0.4
65.4
7.3
1.0
0.1
3.5
0.1
0.08
1.1
0.08
0.1
0.1
13.46
88 x 101U total urban miles "15.3 pg/mi.
-------
Vehicle Type
TABLE 1-15
Urban Motor Vehicle Data, 2000
Fuel Fuel
Urban Mileage Economy, Consumption, Benzo[a]pyrene,
Fraction
mi/gal 1Q9 gal
metric tons
Passenger cars:
Noncatalyat, spark-ignition
Oxidation catalyst, spark-
ignition
Diesel
Trucks, <33,000 Ib:
Noncatalyst, spark-ignition
Oxidation catalyst, spark-
ignition
i Diesel
o
Trucks, > 33, 000 Ib:
Spark-ignition
Diesel
Buses :
Spark-ignition
Diesel
Motorcycles
Totals
Urban-traffic weighted benzol a :
10.14 metric tons
0.6
0.6
0.6
0.55
0.55
0.55
0
0.2
0.6
0.6
0.8
I pyrene emission
30
30
40
20
20
30
—
6
7
10
50
rate:
10.1
10.1
5.9
6.9
6.9
9.2
0
6.3
0.3
1.0
0.5
57.2
0.5
0.06
1.9
0.3
1.9
3.0
—
2.0
0.08
0.3
0.1
10.14
-------
TABLE 1-16
Calculation of Benzo[a]pyrene Inhalation Exposure for 1979
Exposure Situation
Exposure Time
Benzofajpyrene
Concentration,
ng/m3
Benzo[a]pyrene
Inhalation,3
Typical roadway 2 min 17.2
tunnel
Severe roadway 2 min 43.7
tunnel
Typical street 8 h 1.3
canyon, 1,600 veh/h
Severe street 20 min 4.3
canyon, 2,400 veh/h
Severe expressway 2 h 7.7
Beside expressway 14 h 0.2
Total exposure
0.4
0.9
6.5
0.9
9.6
1.8
20. lb
aBased on inhalation rate of 15 a3/d.
^Corresponds to daily average atmospheric BaP concentration of
1.3 ng/m3 [(20.1 ng/d)/(15 m3/d)].
1-31
-------
28
26 -
O
22
o 20
in
18
16
14
12
'IT.*
FUEL ECONOMY / V.O
STANDARDS
•OAO NEW CAA
fUCL CCOMOMT
AVER**! FLEET
FUEL ECONOMY
1» •
1J.1
I I
1966 1970 1974 1978 1982
YEAR
1986
FIGURE 1-1. Automotive fuel economy standards, 1967-1985.
49
Adapted from Kulp et a_l.**7 Based on manufacturers' sales
projections (on-road fuel economy based on actual sales).
1-32
-------
10
2°
I
OJ
Q
UJ
rr
or.
UJ
a.
30
40
50
EPA MOBILE? Passenger Car
S Percenl Reduction
16 Highest US Stations
Air Quality
Percent Reduction
SO U S Slahons
Air Ou.ilily
Perr.enl Reduction
1
1973
1974
1975
1976 1977
YEAR
1978
1979
1980
FIGURE 1-2. CO air-quality and emission-factor trend. Base year, 1973. Air-quality reductions are
reductions in highest 8-h CO yearly concentration averaged over 50 and 16 important U.S. locations.
Adapted from Chang e_t al."
-------
30
UJ
_J
o
X
UJ
o
uj
ec
LU
£
O
a.
i
Hi
2
_J
O
CO
a:
LU
GL
20
10
o>
0
liter
I
I
1
I
I
I
I
I
1
0.7
0.6
0.5
cc
UJ
.o
a.
o»
UJ
0.4 o
0.3
0.2
O.I
LU
O
<
or
LU
1970 1971 1972 1973 1974 1975 1976 1977 1978 1979
YEAR
FIGURE 1-3. Decrease in traffic-average lead emission rates as measured in highway tunnels.
Reprinted with permission from Pierson and Bracharzek.^2
-------
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73. Pitts, J. N., Jr., D. M. Lokensgard, W. Harger, T. S. Fisher, V.
Mejia, J. J. Schuler, G. M. Scorziell, and Y. A. Katzenstein.
Mutagens in diesel exhaust particulate. Identification and direct
activities of 6-nitrobenzo[a]pyrene, 9-nitroanthracene, 1-nitro-
pyrene and 5H-phenanthro[4,5-bcd]pyran-5-one. Mutat. Res.
103:241-249, 1982.
74. Rappaport, S. M., Y. Y. Wang, E. T. Wei, R. Sawyer, B. E. Watkins,
and H. Rapoport. Isolation and identification of a direct-acting
mutagen in diesel-exhaust particulates. Environ. Sci. Technol. 14:
1505-1509, 1980.
75. Riley, T., T. Prater, D. Schuetzle, T. M. Harvey, and D. Hunt.
The analysis of nitrated polynuclear aromatic hydrocarbons in diesel
exhaust particulates by mass spectrometry/mass spectrometry tech-
niques, pp. 115-118. In EPA Diesel Emissions Symposium, October
5-7. Raleigh, N.C.: U.S. Environmental Protection Agency. 1981.
' [Also in Anal. Chem. 54:265-271, 1982.]
76. Salmeen, I., A. M. Durisin, T. J. Prater, T. Riley, and D.
Schuetzle. Contribution of 1-nitropyrene to direct-acting Ames
assay mutagenicities of diesel particulate extracts. Mutat.
Res. 104:17-23, 1982.
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77. Schuetzle, D. Air pollutants, pp. 969-1005. In G. R. Weller and
0. C. Denser, Eds. Biochemical Applications of Mass Spectro-
tnetry. New York: John Wiley & Sons, 1980.
78. Schuetzle, D. Sampling of vehicle emissions for chemical analysis
and biological testing. Environ. Health Perspec. J., 1982. (in
press)
79. Schuetzle, D., and C. V. Hampton. GC/MS in air pollution studies.
In S. Safe, F. Karasek, and 0. Hutzinger, Eds. Mass Spectrometry
in the Environmental Sciences. Oxford, England: Pergamon Press,
Ltd. (in press)
80. Schuetzle, D., F. S.-C. Lee, T. J. Prater, and S. B. Tejada. The
identification of polynuclear aromatic hydrocarbon (PAH) derivatives
in mutagenic fractions of diesel particulate extracts. Int. J.
Environ. Anal. Chera. 9:93-144, 1981.
81. Schuetzle, D., T. L. Riley, T. J. Prater, I. Salmeen, and T. M.
Harvey. The identification of mutagenic chemical species in air
particulate samples, pp. 259-280. In J. Al.baiges, Ed. Analytical
Techniques in Environmental Chemistry. II. Oxford, England:
Pergamon Press, Inc., 1982.
82. Shelton, E. M. Diesel Fuel Oils, 1980. Report DOE/BETC/PPS-80/5.
Bartlesville, Okla.: U.S. Department of Energy, Bartlesville
Energy Technology Center, 1980. 15 pp.
83. Shelton, E. M. Motor Gasolines, Winter 1980-81. Report
DOE/BETC/PPS-81/3. Bartlesville, Okla.: U.S. Department of
Energy, Bartlesville Energy Technology Center, 1981. 67 pp.
84. U.S. Environmental Protection Agency. Compilation of Air Pollution
Emission Factors: Highway Mobile Sources. 3rd ed. Supplement
11. Report No. AP-42. Research Triangle Park, N.C.: U.S.
Environmental Protection Agency, Office of Air Quality Planning and
Standards, 1981. 85 pp.
85. U.S. Environmental Protection Agency. Control of Air Pollution
from New Motor Vehicles and New Motor Vehicle Engines; Federal
Certification Test Results for 1982 Model Year. Washington, D.C.:
U.S. Environmental Protection Agency, (handout)
86. Vuk, C. T., M. A. Jones, and J. H. Johnson. The measurement and
analysis of the physical character of diesel particulate emissions.
SAE Technical Paper 760131. SAE Trans. 85:556-597, 1976.
87. Wade, W. R., J. E. White, an«i J. J. Florek. Diesel Particulate Trap
Regeneration Techniques. SAE Paper 810118. Warrendale, Pa.:
Society of Automotive Engineers, 1981. 22 pp.
88. Wang, Y. Y., S. M. Rappaport, R. F. Sawyer, R. E. Talcott, and E. T.
Wei. Direct-acting mutagens in automobile exhaust. Cancer Lett.
5:39-47, 1978.
89. Williams, R. L. Diesel particulate emissions: Composition, con-
centrations and control, pp. 14-31. In EPA Diesel Emissions
Symposium, October 5-7. Raleigh, N.C.: U.S. Environmental
Protection Agency, 1981.
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90. Williams, R. L., and D. P. Chock. Characterization of diesel
particulate exposure, pp. 3-33. In W. E. Pepelko, R. M. Banner,
and N. A. Clarke, Eds. Health Effects of Diesel Engine Emissions:
Proceedings of an International Symposium. EPA-600/9-80-057a.
Cincinnati, Ohio: U.S. Environmental Protection Agency, Office of
Research and Development, 1980.
91. Williams, R. L., and S. J. Swarin. Benzo(a)pyrene Emissions from
Gasoline and Diesel Automobiles. SAE Technical Paper 790419.
Warrendale, Pa.: Society of Automotive Engineers, Inc., 1979.
8 pp.
92. Wolff, G. T., and R. L. Klimisch, Eds. Particulate Carbon: Atmos-
pheric Life Cycle, pp. v-vi. New York: Plenum Press, 1982.
93. Xu, X. B., J. P. Nachtman, Z. L. Jin, E. T. Wei, S. Rappaport, and
A. L. Burlingame. Isolation and identification of mutagenic nitro-
arenes in diesel-exhaust particulates, pp. 556-558. In EPA Diesel
Emissions Symposium, October 5-7. Raleigh, N.C.: U.S. Environ-
mental Protection Agency, 1981.
94. Yu, M-L., and R. A. Hites. Identification of organic compounds on
diesel engine soot. Anal. Chem. 3:951-954, 1981.
95. Zweidinger; R. B. Emission factors from diesel and gasoline powered
vehicles; correlation with the Ames test, pp. 95-108. In EPA
Diesel Emissions Symposium, October 5-7. Raleigh, N.C.: U.S.
Environmental Protection Agency, 1981.
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POLYCYCLIC AROMATIC HYDROCARBONS FROM NATURAL AND STATIONARY
ANTHROPOGENIC SOURCES AND THEIR ATMOSPHERIC CONCENTRATIONS
Although the emphasis of this report is on the identification of the
polycyclic aromatic hydrocarbons (PAHs) emitted from motor vehicles, PAHs
are ubiquitous substances. They are found in terrestrial and aquatic
plants, in soils and bottom sediments, in fresh and marine waters, in
emission from volcanoes and naturally occurring forest fires, and in the
products of numerous human activities. The anthropogenic sources vary
widely—major oil spills and the inestimable minor spills of petroleum
products, emission from coal- and gas-fired boilers and electric-power
generating plants, space heaters (especially in individual residences),
municipal and industrial incinerators, and all sorts of industrial
processes. It is not possible to list all the sources or to count or
measure the PAHs produced by them. The various PAH compounds and the
amounts emitted into the environment from each of the sources result in a
complexity that makes it difficult to trace and identify the major
contributing sources.
PAH COMPOUNDS IN PETROLEUM AND FOSSIL-FUEL PRODUCTS
The carcinogenic potential of petroleum hydrocarbons was examined in a
critical review of world literature in the period 1960-1978 by Bingham e_t
al. Although the carcinogenic potential of some samples of petroleum and
other fossil-fuel material tested in experimental animals could be
associated with the presence of benzo[a]pyrene (BaP), others without
benzofajpyrene were also carcinogenic. Thus, the authors suggested that
benzo[a]pyrene may not be the most prevalent or important component in the
samples and recommended further chemical analyses of a variety of petroleum
samples to determine the profile of PAH compounds in them. The review
included references to the carcinogenicity of high-boiling-point (above
260°C) petroleum fractions, residues, and products and to the occurrence
of cancer in workers in refineries and industries in which these materials
are used. The review did not discuss the carcinogenicity of pure PAHs or
studies of environmental pollution from general sources, as are covered in
this report.
PAH COMPOUNDS IN CRUDE OILS. COAL. AND OIL-SHALE DERIVATIVES
Mutagenicity testing (with standard j>. typhimurium procedures) by Guerin
e£ £!• of several crude oils and shale- and coal-derived petroleum
substitutes showed the petroleum-substitute rautagenicities to be equal to or
10-100 times greater than those of petroleum products.
The nonpolar neutral constituents generally were found to contribute
over half the mutagenicity. Those findings give added incentive for
identifing and characterizing the individual PAHs in these products.
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Crude Oil
In their review of the carcinogenicity of petroleum hydrocarbons,
Bingham et^ al. cited several of the early researchers' work (1931) on the
carcinogenicity of crude-oil fractions from various sources when applied to
the skin of mice. The distillate fractions in the 300-400°C range were
later found to be more carcinogenic than whole crude oil in mice and
rabbits. Although the review cited studies by Hueper in 1965 that failed to
produce cancer within the normal life span when undiluted crude oil was
applied to C57BL mice, it cited Barr-Nea and Wolman in 1972, who found
papilloraas after 5-7 mo of topical application of acetone-extracted crude
oil. Numerous other studies cited in the review described the carcinogenic
potential of crude oils from various locations when tested in animals.
The review by Bingham e_t al. cited references that gave the content of
BaP: 40, 1,320, and 1,660 yg/L, respectively, in Persian Gulf, Libyan, and
Venezuelan petroleum (Graf and Winter, 1968) and 1,000 and 2,800 yg/kg in
South Louisiana and Kuwait crude oils, respectively (Panceron and Brown,
1975).
Several studies cited by McKay and Latham22 reported qualitative
findings of anthracenes, phenanthrenes, benzophenanthrenes, fluorenes,
chrysenes, pyrenes, perylenes, and coronene in virgin petroleum. In the
process of cracking of petroleum distillates, the high-temperature
hydroconversions formed ring systems, such as benzocoronenes,
dibenzocoronenes, and tribenzocoronenes. The authors identified seven
polynuclear aromatic compounds not previously found in virgin petroleum
distillates (temperature, 335-550°C): 1.12,2.3- dibenzoperylene,
l,12-£-phenyleneperylene, pyreno[1.3:10'.2'jpyrene, 2.3,10.11-dibenzo-
perylene, 1,2,4,5-dibenzopyrene, benzo[e]pyrene (BeP), and
benzo[g]chrysene. Similar quantitative evidence of the presence of pyrene,
BaP, BeP, chrysene, and 1,23-benzoperylene was reported by Coleman et al.
in fluorescence emission and fluorescence excitation analysis of Prudhoe Bay
crude oil.
Coal and Oil-Shale Derivatives
Coal gasification has been used to produce clean fuels in many countries
since 1880. The measurement of individual PAHs in the synthetic oils
produced by coal liquefaction and in natural crude oil remains a difficult
problem. The samples often are from small-scale processes with questionable
resemblance to the products of eventual commercial-scale operations. Guerin
e_t j^. analyzed fractions of two coal-derived crude oils (synthoil from
catalytic hydrogenation of coal, synthoil C, and syncrude from pyrolysis of
coal, syncrude D), shale-derived crude oil shale B, and a petroleum mix
(crude-oil mixture from California, Canada, Alaska, Iran, Louisiana-
Mississippi, and Arabian Light). The results of the chromatographic
analyses are shown in Table 2-1. The coal-derived crudes had larger
quantities and a greater variety of PAHs than the petroleum sample, and
2-2
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shale B had less than either of them. The summation of the PAHs produced
the following totals: synthoil C, 135 mg/g; syncrude D, 132 mg/g; shale B,
36 mg/g; and petroleum mix A, 58 mg/g.
The increased interest in greater use of petroleum substitutes from oil
shale and coal has raised concern about the health hazards associated with
these fuels. This concern led Buchanan et_ a_l. to study the mutagenicity
of subtractions from several fossil fuels, principally the primary aromatic
amines (PAAs). Aminofluorene and aminoanthracene were among the PAAs
identifed in the subfractions, and azabenzofluorene, azabenzopyrene, and
azaanthanthrene were identified in the azaarene group. They mentioned that
the rautagenic activity of PAAs is greater than that of azaarenes. Katz and
Ogan1 identified chromatographic peaks of benz[a]anthracene, dibenzfah]-
anthracene, and benzo[ghi]perylene in coal liquid (EPA Chemical Repository
Samples CRM-1-3). Earlier, White et^^l..,49 during the process of
development of gas-chromatographic analysis, used a coal-liquefaction sample
from the Synthoil Process Development Unit, Bruceton, Pennsylvania, and
identified fluorene, 9,10-dihydrophenanthrene, 1,2,3 ,4,5,6,7,8-octahydro-
phenanthrene, 1,2,3,4-tetrahydrophenanthrene, phenanthrene, anthracene,
fluoranthene, and pyrene.
Nichols e£ a_1.34 studied raw gases from fixed-bed reactors fueled by
different coal or vegetative fuels and collected in a stainless-steel
cooler-condenser during the run. The gas-chromatographic analyses of the
condensate are reported in Table 2-2 as micrograms per gram of solid feed to
the gasifier unit.
USED ENGINE OIL
The 1972 NRC study^^ on particulate polycyclic organic matter (POM)
reported findings in the literature that BaP emission increased as the
vehicle aged and oil consumption increased from 1,600 mi/qt to 200 mi/qt and
that the BaP preferentially concentrated in the crankcase. The report did
not list any other POM in used oil.
There are numerous analytic problems in isolating and analyzing for PAHs
in used engine oil. Lee e_t £l.« sampled oil taken from the oil pans of
four randomly selected 4-, 6-, and 8-cylinder automobiles. The qualitative
results of high-resolution (capillary) gas chromatography showed peaks for
fluorene, phenanthrene, anthracene, 4-5-raethylene, 9-methylphenanthrene,
fluoranthene, pyrene, 1-methy1pyrene, triphenylene, chrysene, BaP, BeP,
perylene, and dibenz[ac]anthracene.
Peake and Parker^ estimated that 500 million gallons of used motor
oil is not reclaimed, but is haphazardly discharged into sewers or onto
wasteland each year. In analysis of motor oil by gas chromatography-mass
spectroraetry, they found a predominance of alkyl-substituted aromatic
compounds and 11 alkylfluorene isomers. Table 2-3 lists identified
compounds and amounts per milliliter of oil, based on the detector response
to perdeuteroanthracene.
2-3
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PETROLEUM AND OTHER FOSSIL-FUEL COMBUSTION FOR HEAT AND POWER GENERATION
The data presented below show the individual PAHs present in emission
from several sources of combustion effluents. Compounds known to be
carcinogenic are identified in the tabulations of some of the data. A 1980
NRC study-^O recommended that future research should continue to monitor
such emission and measure (in a mass-balance study) the contribution of
known carcinogenic or mutagenic compounds to the environment. There is a
need for specific-site and broad-scale mass-balance studies of the release
of PAHs into the atmosphere. The highest PAH emission rates in heat- and
power-generation categories given in the 1967 review by Hangebrauck et^
£l.l^ were associated with small, domestic, coal-fired furnaces used to
heat single-family homes. The emission from oil-burning was generally much
lower than that from coal-burning and slightly higher than that from
gas-fired units. The emission rates for 10 PAH compounds from under-feed
stokers and hand-stoked coal furnaces were higher by several orders of
magnitude than those from coal-fired power-plant units, as shown in Tables
2-4 and 2-5, respectively. In Table 2-6, the emission rates for
intermediate-sized coal, oil, or gas units using different firing methods
show that the under-feed coal-stoker units emit all 10 PAHs at the higher
rates.
The EPA Industrial Environment Research Laboratory, Research Triangle
Park, N.C. (IERL-RTP), developed the source-assessment sampling system
(SASS) train for collection of gaseous, particulate, and volatile exhaust
matter. The SASS was used to determine emission data on 74 inorganic trace
elements and 21 PAHs in the effluent of 11 industrial coal-stoker-fired
boilers. These 11 units represent a wide range of designs, which reportedly
have changed very little over the last 20 yr. Three units were spreader
stokers with reinjection from the dust collector (Table 2-7), three were
without reinjection (Table 2-7), and five were mass-fired over-feed
stokers. The study was done by the American Boiler Manufacturers'
Association under the joint sponsorship of the U.S. Department of Energy and
the EPA by Burlingame e£ al.6 There were 23 SASS tests conducted, and the
emission data were presented in three units: nanograms per joule of energy
input, micrograms per dry standard cubic meter of flue-gas samples, and
micrograms per kilogram of fuel input. The PAH totals of all the SASS tests
were reported, but only the average emission in micrograms per standard
cubic meter of effluent and in micrograms per kilogram of fuel is listed in
the table.
PAHs FROM COKE PRODUCTION
The EPA's July 1981 report^** on background information for a proposed
standard on coke-oven emission gives a comprehensive overview of emission
and locations of coke plants, as well as a cancer risk assessment.
Only some of the literature showing the PAH compounds found in the
emission from coke ovens and the relationship to particle size is discussed
here.
2-4
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Lao £t al^. 19 reported the results of sampling emission from coke ovens
in the steel industry. Two samples were collected on glass-fiber filters
and two on O.S-y m-pore silver membrane filters, extracted for 24 h in
Soxhlet extractors, and measured in GC/MS and GC/FID systems. The results
are given in Table 2-8 in micrograms per gram of extract.
Some of the early studies on carcinogenicity of coke-oven effluent (as
early as 1875) were reviewed by Hoffmann and Wynder15 in 1976. The
relative risk of developing lung cancer for men employed 5 yr (1951-1955) or
more at the full topside of the ovens from exposure to coke-oven effluent
was 6.9 times the expected (the prediction was 2.5 times greater than for
the general population); an unexpected finding of kidney-cancer incidence
7.5 times greater than in the general population was also reported.
Several reports have discussed the PAH emission from coke ovens in other
countries: Norway,4 Finland,41 Czechoslovakia,21 Canada,37 and
Brazil.24
In studying concentrations of PAHs on particles of the various sizes,
Bjorseth found the greatest amount on the particles of 0.9-3 ym and only
about 1% of the total on particles larger than 7 ym (see Table 2-9). The
particles were collected at the top of a coke-oven battery, fractionated
according to particle size by a Lundgren impactor, and analyzed in a
glass-capillary GC/MS system. Miguel and Rubenich24 gave the
concentrations of BaP found on particles in various size ranges (collected
with an eight-stage low-pressure impactor that separated by aerodynamic
diameter) from an urban automobile traffic tunnel, from ambient air, and
from the bench (push side) of a steel-mill coke oven (see Figure 2-1).
COAL MINING
The PAHs found in respirable coal-dust samples were identified as
phenanthrene, pyrene, benzo[ghi]fluoranthene, chrysene, perylene,
benzoperylene, benzochrysene, and dibenzoperylene. These compounds were
more common in coal dust in the mine than in dust from other locations. The
OTA report3^ of 1979 stated that the long-term exposure of coal miners to
these suspected carcinogenic compounds had not been analyzed. Shultz et^
al.4^ reported finding 13 PAHs in the respirable fraction of mine dusts.
WOOD-BURNING FOR HEAT AND POWER GENERATION
As the prices for home heating with electricity, gas, and oil increase
and fuel availability is threatened, many home owners are using wood or coal
stoves as a supplement for space heating. Duncan e£ At* estimated an
increase of 40,000 wood-burning stoves in 201 counties of the TVA
power-distribution area from 1974 to 1976. According to DeAngelis et
al.,1^ the U.S. Bureau of the Census data showed that 452,000 new homes
had fireplaces and that 550,000 wood-burning stoves were shipped by
manufacturers in 1975. Owing to the difficulty in achieving controlled
2-5
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combustion in fireplaces and wood- and coal-burning stoves, there is often
not an efficient burn; consequently, there is a need for more frequent
cleaning of chimneys. Chimney-cleaning equipment is being sold for use by
individual homeowners, and the occupation of chimneysweep has become
prominent once again. Hazards of exposure to the particulate matter in
chimney-cleaning are recognized to be associated with not only skin
exposure, but also inhalation. The carcinogenic health hazards associated
with chimney-cleaning were reviewed by Bagchi and Zimmerman.'- The number
of housing units burning wood was estimated by DeAngelis et_ aj^. ,* using
1970 U.S. Census of Housing data in conjunction with the 1976 Housing
Survey. The state-by-state tabulation showed totals of 912,000 units
burning wood as the primary source of heat and 35,467,900 burning wood for
auxiliary or aesthetic purposes, with an estimated consumption of 5,122,000
metric tons a year for the primary units and 11,500,000 metric tons for
auxiliary or aesthetic units. The range of POM emission from wood stoves
and fireplaces was 0.01-0.4 and 0.02-0.04 g/kg, respectively. In 1981,
Peters, a coauthor of the above work, estimated the annual emission of POM
into the ambient air from primary heating units at 1,383 metric tons, from
auxiliary units at 2,376 metric tons, and from fireplaces at 78 metric tons,
for a total of 3,837 metric tons.
The NRC report Indoor Pollutants*° assessed some of the sources of
pollutants indoors and their effect on air quality. When wood stoves were
in use, the BaP concentration monitored over 24 h indoors was 5 times higher
than when stoves were not in use.
In assessing the impact of wood-combustion emission on the environment,
the U.S. Department of Energy,^ in 1979, stated that the major pollutants
of concern from residential wood-combustion devices were unburned
combustibles, carbon monoxide, particles, and hydrocarbons. Owing to the
inefficient combustion in many home-heating units, large quantities of all
of them are emitted. According to the 1980 Department of Energy report^
on health effects of residential wood combustion, the emission from such
combustion is a major environmental problem affecting local air quality.
POM is the most important group of organic compounds among the noncriteria
substances emitted. There is no federal regulation of atmospheric emission
applicable to residential space-heating units.
In the preliminary EPA assessment of wood-fired residential combustion
equipment, it was stated that the emission of organic substances, including
POM, is relatively high, owing to the use of large pieces of fuel, highly
resinous fuel, uneven fuel distribution, and hand-feeding in batches. The
emphasis on use of wood-burning, air-tight stoves may greatly increase the
magnitude of the emission problem. The PAHs emitted from wood-burning in
fireplaces and baffled and nonbaffled stoves are shown in Table 2-10.
In a study of modifications of combustion to reduce emission from
residential wood-burning, EPA^7 suggested several techniques to reduce the
gaseous components in the emission. The following compounds were identified
in the emission from wood-burning stoves that burned oak or green pine in
either the up-, down-, side-, or high-turbulence-draft delivery systems.
The up-draft system had the lowest total-particle and gaseous emission rates
2-6
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Naphthalene
Acenap&thene
Acenaphthyleae
Fiaorene
Phenanthrene
Anthracene
Methylanthracene/aethylfluorantfaenes
FTuoranthene
Pyrene
Me thy1pyrene s/aethy1fluoran thene 3
Benzla]anthracene
Chrysene
Methylehrysenes
Diaethylbenz Is.} aathracesie
Senzofluoranthenes
Benzo { e ] pyr ene
Perylene
Indeno 11,2, 3-cd ] pyrene
Be nz o :I gh i ] pe ry leae
Coronene
Dibeozo [ ah ] pyrene
PAH emission rates under controlled burn conditions were characterized by
Hubble _et_ _al_., and their results are shown in Table 2-11. The
relationship of ambient air concentrations of PASs to the sources of
wood-banning has been studied by several investigators. In sose
circumstances, wood—burning can be the largest combustion source for various
ataospiieric pollutants in urban areas. For example, the contribution of
wood-burning emission products to atsospheric PAHs in Telluride, Colorado,
has been estimated by ?§urpfay et .si.. Telluride is a ssall coasuaity in a
valley with poor ventilation and with large temperature inversions. It
depends heavily on wood-burning as a residential heatiag source and has only
light automobile traffic. In this cosssunity in 1980, the 3aP concentration
in the air reached 7.i ng/a , which exceeds several tisses over that which
is fouad in a nuaber of U.S. metropolitan areas, such as Los Angeles, and
Telluride is not unique in this respect.
The aecaanisms of PAH foraation during the combustion of wood are poorly
understood. It is known that wood contains substantial asounts of
alkylbenzene derivatives, which contribute to the formation of the ?AHs.
The latter reactions depend markedly on teaperature: no iap-ortant forsatioa
of hydrocarbons occurs below 450°C. Approximately 75 organic compounds
have been identified in flue-gas samples of the identified organic
substances, and the PASs sake up about 351 of the aass. The PAHs that are
produced during the pyrolysis of wood and are found in the saoke include
anthracene, phenanthrene, dibenzfaj3anthracene, dibenzfah]anthracene,
fluoranthene, benzo[ghi]fluoranthene, benzo[b]fluoranthene, benzo[c3phen-
anthrene, benzolghilperylene, pyrene, BaP, BeP, 3-aethylcholanthrene,
dibenzo[cg]carbazole, dibenzo[ai]carbazole, cyclopentaicdjpyrene, and sosae
aethylated substances.
A detailed study of the effects of wood type, degree of seasoning (e.g.,
the aoistare content), and type of wood stove on the emission of PAHs in
saoke has been conducted.^^ A fireplace, a baffled and a nonbaffled wood
stove, and seasoned (4-5J moisture) and green (27-301 ooisture) oak and pine
were used. The results are indicated in Table 2-13. Mood saoke generated
in a fireplace contained such less PAH than that foraed in wood stoves. So
difference in relative distribution of particular PAHs was noted when green
and seasoned wood were burned in the fireplace (data not shown).
Furtheraore, neither wood type nor extent of baffling significantly
influenced the pattern of PAH emission in the wood saoke.
2-7
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Peters has compared the emission of PAHs from several residential
combustion sources as a function of thermal efficiency (Table 2-14).
Wood-fired heating resulted in a much higher output of PAHs than did
coal-, oil-, or gas-fired furnaces; i.e., the mass of PAHs emitted per
joule was 10, 5,000, and 30,000 times greater, respectively.
REFUSE BURNING AND INCINERATION
Before the passage of the Solid Waste Disposal Act in 1965, numerous
municipal dumps practiced open or uncontrolled burning throughout the
United States. Owing to the standards established under this Act, most
of the large incinerators then in operation were shut down, largely
because upgrading them would have been expensive. Congress passed the
Resource Recovery Act in 1970 and the Resource Conservation and Recovery
Act in 1976. As a consequence of the passage of these acts, numerous
demonstration projects for solid-waste disposal and energy and resource
recovery were initiated. The concluding remarks of the 1981 NRC report
The Recovery of Energy and Materials from Solid Waste^' stated "that
the technologies for energy recovery were still under development and
that the most highly developed and least risky was mass burning, but that
other technologies were being tested."
The solid waste from residential, commercial, and institutional
sources amounted to 130 million metric tons in 1976 and is projected to
increase to 180 million tons by 1985.^
In EPA's 1971 review^ of the literature on municipal incineration,
the findings of Hangebrauck et .§_!• (Table 2-15) showed the emission
of 10 PAHs from municipal and commercial units burning wastes from
households, grocery stores, and restaurants. In 1976, Device et al.^
reported that the measured emission of PAHs was similar to that found by
Hangebrauck. They used a modern, continuous-feed municipal refuse
incinerator rated at 9.14 tons of refuse per hour with a water-spray
cooling tower (cooled to 250-300°C) and an electrostatic precipitator
in the flue-gas stream. The analytic results on samples collected after
the electrostatic precipitator are given in Table 2-16.
The findings by several investigators on the emission of chlorinated
dibenzo-£-dioxins in waste incineration have caused some concern about
the human-health implications of this group of compounds, but discussion
of dioxin emission is beyond the scope of this report.
METAL PROCESSING
The stack gases from a smelter processing lead from batteries were
sampled. The PAHs and their concentrations found in four samples are
shown in Table 2-17- The polymeric organic battery casings were included
in the process and presumed to be the contributing source of the organic
emission. Lao and Thomas'-" identified several PAH compounds in the
2-8
-------
particles collected on glass-fiber filters in the exhaust flue from the
"pot room" of a nonferrous-metal production room. Particles from iron
foundries in Finland were analyzed for PAHs by Schimberg,3** and the
following were identified: phenanthrene, anthracene, fluoranthene,
pyrene, benzo[a]fluorene, benzo[c]phenanthrene, benzofluoranthenes, BeP,
BaP, perylene, c>-phenylenepyrene, dibenzanthracenes, benzochrysenes, and
benzo[ghi]perylene.
NATURAL SOURCES
The mechanism whereby complex mixtures of PAHs occur in natural
deposits of peat, coal, crude oil, and shale oil is unknown. Neff33
discussed the sources of PAHs in the aquatic environment and their
possible biosynthesis by bacteria, yeasts, and higher plants. Although
some publications apparently provide evidence of biosynthesis, others
refute it. Until relatively recently, it was assumed that PAHs are
formed only by pyrolysis of organic material. However, the finding of
l-methyl-7-isopropylphenanthrene (retene) in pine tar and the isolation
of PAHs directly from plant material have helped to keep alive the
uncertainty of PAH biosynthesis in plants.
FOREST FIRES
Forest fires are sporadic and sometimes uncontrollable occurrences
that apparently contribute significantly to the PAHs found in the
atmosphere. Some laboratory data are given below on the amount of PAH
emission. The discussion in Chapter 3 of physical removal of PAHs from
the atmosphere mentions forest fires as a source of atmospheric PAHs.
This source is mentioned here, because it has the potential for being a
significant contributor; however, data are insufficient for an assessment
of the impact on atmospheric quality. The 1976 NRC report Air Quality
and Smoke from Urban and Forest Fires^ described "prescribed forest
fires" as fires set to reduce the amount of secondary plant or roughage
undergrowth in the hope of reducing the incidence of wild fires, for
disease control, and for other management purposes* In the 1976 report,
it was noted that wild forest fires consume about 3 times as ouch fuel as
prescribed fires and produce about 3 times as much particulate matter per
ton of fuel burned as prescribed fires. On the average, prescribed fires
consume 3 tons of fuel per acre (range, 1-10 tons/acre) and wild fires 9
tons/acre (range, 1-50 tons/acre).
Emission from forest fires varies widely, owing to the variety of
fuels, fir* typ* (heading fires, with the fire line moving with the wind,
or backing fires, with the fire line moving against the wind), fire
intensity, and combustion phase (flaming vs. smoldering). The results of
a screening experiment reported by McMahon and Tsoukalas,23 using slash
pine needle litter as fuel, are shown in Table 2-18. The PAHs present in
the particles show that heading fires produce higher total amounts of
particles, but smaller total amounts of PAHs.
2-9
-------
AQUATIC ENVIRONMENTS (FRESHWATER AND MARINE)
In its twenty-fifth report to Congress, the Committee on Governmental
Operations^3 considered groundwater destruction one of the most serious
environmental problems of the 1980s. It estimated that 50 billion
gallons of water are placed in industrial surface-water impoundments
throughout the country each year. The report stated that tens of
millions of Americans obtain their water from private wells, because no
public water supply is available.
A high percentage of PAHs is removed from municipal water by
flocculation and sedimentation of suspended particulate matter. The
chlorination process removes a large percentage of residual PAH
contamination (the amounts removed by this process are discussed
elsewhere in this report). We do not have national survey information on
the incidence of well-water contamination by PAHs, and this route of
human exposure has not been assessed. But it is of potential importance,
in that most well-owners do not treat the water before use, on the
assumption that it is uncontaminated. Owing to the large number of
industrial wastewater impoundments that are unlined (no plastic liner)
and are directly over groundwater sources, there is a need for
investigation to determine whether there is PAH contamination. (See the
section of this report on human exposure for discussion of PAHs in
drinking water.)
In the comprehensive review by Neff33 on PAHs in the aquatic
environment, the following sources of pollution were listed: industrial
and domestic sewage effluent, surface runoff from land, deposition of
airborne particles, and spillage of petroleum and petroleum products into
water bodies. Specific PAHs from these sources are discussed earlier in
this section; but Neff's work includes more detail on the effects on the
aquatic environment. The 1975 NRC report Assessing Potential Ocean
Pollutants32 discussed the effects of several sources of pollution,
e.g., ocean discharge of dredge spoil, municipal sewage sludge, petroleum
discharge, and spills. Although that report did not discuss the effects
specifically of PAHs, the sources of pollution described in it have been
identified as sources of PAHs. An NRC committee is in the process of
reassessing the pollutants and their effects on the marine environment.
In 1974, owing to the increasing amounts of petroleum being
transported over the world's waterways, the need for a comprehensive,
international marine environmental monitoring assessment program led to a
symposium and workshop on marine pollution monitoring held at the
National Bureau of Standards in Gaithersburg, Maryland. At the
symposium, deep concern was expressed about the amount of petroleum
pollution observed in the marine environment and its effects and ultimate
fate. Among those expressing concern were representative of several
national and international organizations with interests and oversight for
pollution problems in the marine environment, such as the Special
2-10
-------
Committee on Problems of the Environment of the International Council of
Scientific Unions, the International Oceanographic Commission, the
Intergovernmental Working Group on the Global Environmental Monitoring
System (the EARTHWATCH program), the Integrated Global Ocean Station
System, and the Experts on Scientific Aspects of Marine Pollution
(supported by the United Nations Environment Program). Out of concern
like that expressed at the 1974 symposium, there arose the Mussel Watch
that uses mollusks (mussels, clams, and oysters) as biologic monitors of
aquatic pollution. The 1980 NRC report The International Mussel
Watch^l described the program and gave the two general aims of the
"watch": to produce information on the contamination of coastal
ecosystems and food resources and global data on the abundance of
anthropogenic contaminants.
2-11
-------
TABLE 2-1
Polycyclic Aromatic Hydrocarbons in Coal, Oil-Shale, and
Petroleum Isolates8
Relative Peak Values**
PAH
Fluorene
9-Methylf luorene
1-Methylf luorene
Phenanthrene
Anthracene
2-Methylanthracene
1-Methylphen-
Coal
Synthoil C
5.3
2.0
6.0
20.4
—
2.4
<1.2
Coal
Syncrude D
9.9
1.4
<4.7
12.0
4.1
3.0
<5.1
Shale B
2.2
0.7
3.2
4.0
1.8
1.1
3.6
Petroleum
Mix A
3.9
2.6
5.6
5.3
T
—
8.1
anthrene
9-Methy1anthracene
Fluoranthene <1.9
Pyrene 35.0
Benzo[a]fluorene 2.5
Benzofb]fluorene 3.4
l-Methylpyrene <8.0
Benzo[cIphenanthrene <0.6
Benzo[ghi]fluor- 3.2
anthene
Benz[a]anthracene <2.2
Chrysene 2.5
Benzo[b, j, and/or <1.3
k]fluoranthene
Benzo[e]pyrene 1.3
Benzo[a]pyrene <1.2
Perylene T
Dibenzfac and/or
ah]anthracene,
indeno[l,2,3-cd]-
pyrenec
Picenec
3-Methylchol- <1.2
anthrene
£-Phenylenepyrene 2.6
Benzo[ghi]perylene 6.6
Anthanthrene <0.8
—
<3.7
14.2
2.1
<1.5
<6.0
<2.2
—
T
<1.5
<0.5
a. 2
<0.5
<0.6
T
1.6
3.7
1.1
1.2
1.6
1.3
—
1.0
1.5
0.5
0.3
0.4
T
—
0.9
4.3
2.1
2.2
1.6
3.1
0.3
T
2.6
0.2
0.3
0.2
0.7
0.70
1.1
4.3
T
aAdapted from Guerin et al.^2
"Relative peak values show a relationship of amounts present. Dash
indicates none detected. T indicates trace quantity detected. < indicates
known to contain co-eluting species.
cPeaks identified, but relative values not sought.
2-12
-------
TABLE 2-2
PAHs Resulting from Gasification of Western Coal, Coal Blend,
Wood, and Peat3
Concentration, ug/g of feed stock
Compound
Naphthalene
Acenaphthylene
Fluorene
Phenanthrene
Anthracene
Pyrene
Benz [a]-
anthracene
Chrysene
Benzo[b]-
fluoranthene
Benzo[k]-
f luoranthene
Benzo[a]pyrene
Indeno[l ,2,3-
cd]pyrene
Dibenz [ah]-
anthrgeene
Benae[ghi]=
perylgnt
North
Dakota
Lignite
1,520
322
137
443
81.4
221
35.7
18.6
22.9
7.1
17.1
4.7
5,2
2,9
Wyoming
Smith-
Roland15
899
182
82.2
394
65.9
253
48.3
32.1
32.1
7.9
23.7
2.2
11,0
4, a
Montana
Rosebud
Coal
1,430
234
45.5
375
49.0
405
49.0
30.9
54.3
1.2
52.5
4.4
33,1
34,0
111. #6
& CaCO
Blendc
878
278
117
412
224
230
98.2
69.0
43.9
23.0
37.6
4.2
16,7
10,4
111. #6
& Wood
Pelletsd
236
53.6
28.9
94.0
32.7
31.3
6.3
3.8
3.1
2.1
3.8
0.7
0,7
0,7
N. C.
Peat
Pellets6
2,480
343
101
542
108
395
67.0
33.9
35.8
17.9
34.9
0.9
11,3
7,3
ea«l,
e70I §§sl, 30* limeatene,
Carolina coastal peat (191 moisture),
2-13
-------
TABLE 2-3
Polycyclic Aromatic Hydrocarbons in Used Motor Oil Identified
by Gas Chromatography-Mass Spectrometry3
Compound ug/ml oil
Methylbiphenyl 0.74
Methylbiphenyl 0.36
Methylbiphenyl 0.26
Fluorene 1.47
Methylbiphenyl 0.42
Methylbiphenyl 0.18
Methylbiphenyl 0.09
Methylfluorene 0.10
Methylfluorene 1.19
Methylfluorene 0.08
Phenanthrene 7.80
Deuterated anthracene0 0.50
Methylfluorene 0.08
Dimethylfluorened 0.58
0.10
Anthracene 0.33
Dimethylfluorened 0.61
Methylphenanthrene 2.63
Methylphenanthrene 3.62
Methylphenanthrene 2.95
Triraethylfluorene6 0.12
Methylphenanthrene 2.44
Trimethylfluorene6 0.29
Triraethylfluorene6 0.36
Phenylnaphthalene 0.90
Trimethylfluorene6 0.15
Trimethylfluorene6 0.18
Dimethylphenanthrened 0.22
Diraethylphenanthrened 0.16
Dimethylphenanthrened 0.75
Methylanthracene 0.09
Diraethylphenanthrened 2.45
Dimethylphenanthrened 4.21
Methylanthracene 0.28
Dimethylphenanthrened 2.80
Fluoranthene 4.36
Methylanthracene 0.21
0.08
Ethylcyclopenta[def)phenanthrenec 0.79
Ethylcyclopenta[def]phenanthrenec 0.46
0.10
Trimethylphenanthrene6 0.10
Pyrene 6.69
2-14
-------
Table 2-3 (cont.)
Compound yg/ml oil"
Ethylcyclopenta[def]phenanthrened 0.17
0.39
Trimethylphenanthrene6 0.08
Triraethylphenanthrene6 0.29
Terphenyl 0.12
Triraethylphenanthrene6 1.32
Dimethylanthracened 0.08
Dimethylanthracened 0.10
Trirnethylphenanthrene6 2.72
Triraethylphenanthrene6 0.48
Dihydroraethylpyrene* 0.13
Triraethylphenanthrene6 1.16
Benzo[a]fluorene 0.93
Benzo[b]fluorene 1.38
Benzo[c]fluorene 0.44
Methylpyrene' 1.19
Trimethylanthracene6 0.51
Dihydromethylpyrene^ 0.32
Methylpyrene^ 1.14
Methylpyrene 1.14
Methylpyrene^ 0.78
DiethylphenanthreneS 0.18
Dimethylpyrene^ 0.13
DiethylphenanthreneS 0.14
Dimethylpyrened 0.70
Diethylphenanthrene? 0.34
Dimethylpyrene 0.28
Dimethylpyrene 0.33
Benzo[c]phenanthrene 0.12
Diraethylpyrened»f 0.27
Ethylraethylpyrened»f 0.14
Benzo[a]anthracene 0.87
0.22
Chrysene + triphenylene 2.48
Cyclopenta[cd]pyrene 0.78
Methylbenzo[a]anthracene" 1.68
Methylbenzo[rano]fluoranthene1 0.15
Methylbenzo[a]anthracenen 0.26
Methylbenzo[a]anthraceneh 0.23
Methylbenzofmno]fluoranthene1 0.15
Methylbenzo [a] anthracene*1 0.26
Methylbenzo [a] anthracene*1 0.15
Methylbenzo[a]anthraceneh 0.28
Ethylbenzo[a]anthracene 0.44
Ethylbenzo[a]anthracene 0.21
Benzo[k]fluoranthene 1.44
Benzo[e]pyrene 1.74
2-15
-------
Table 2-3 (cont.)
Compound yg/ml oil
Benzo[a]pyrene 0.36
Perylene 0.13
Methylbenzofluoranthene 0.18
Methylbenzopyrene-' 0.41
0.32
Benzo[ghi]perylene 1.67
aReprinted with permission from White et_ _al_.;^9 copyright Ann Arbor
Science Publishers, Inc.
"yg/ml of oil based on the response of deuterated anthracene.
clnternal standard.
dCould be ethyl- or dimethyl-.
eCould be ethylmethyl-, trimethyl-, or propyl-.
fCould be a pyrene or fluoranthene.
SCould be diethyl-, ethyldimethyl-, tetramethyl-, raethylpropyl-, or butyl-
"Could be a derivative of chrysene, triphenylene, benzo[c]phenanthrene or
benz fajanthracene.
^ould be a derivative of benzo[mno]fluoranthene or cyclopenta[cd]pyrene.
^Compounds with molecular weight 276 can be any of the following:
indeno[1,2,3-cd]pyrene, indeno[1,2,3-cd]fluoranthene, cyclopenta[cd ]-
perylene, phenanthro[10,-l,2,3-cdef]fluorene, acenaphth[l,2-a]-
acenaphthylene, dibenzo[b,mno]fluoranthene, dibenzo[e,mno]fluoranthene,
and dibenzo[f,mno]fluoranthene. Further possibilities are the benzo
derivatives of cyclopenta[cd]pyrene and cyclopenta[cd]fluoranthene.
2-16
-------
TABLE 2-4
FAH Eeiieaioo Suaaary: Heat Generation by Co*1-fired Residential Furnace**
Sanple
No.
1SS-19
34
36
59
60
20
57
58
Firing
Method
Under-
feed
atokera
Hand-
atoked
Beniene-
aoluble
Organica
• l*r
10* Itu
2.4
47
73
20
8.7
91
170
350
Croup 1
•aP
MS/g
t aol.
1,600
1,400
1.100
3.400
990
3,900
10,400
9,400 '
Mg per
1.000 •*=
3,400
40,000
44,000
18.000
2.200
340,000
690.000
1,500,000
Mg per
Ib fuel
52
900
1.200
930
120
6.000
25,000
46,000
Bap
P
leP
Per
BuhiP Anth
Cor
Group 2
A
Phen
3,800
65,000
81.000
67,000
8,600
400,000
1,700,000
3,300,000
7,700
300,000
190.000
160,000
45,000
600.000
2,700,000
9,100,000
5.400
39,000
59,000
55,000
7,700
100,000
870,000
1,500,000
7,900
4,800
5,500
430
60,000
220,000
350,000
580
61,000 6,100
58,000 3,000
59.000 1,300
6,300
300,000 90,000
1,400.000 270,000
2.200,000 490,000
1,200
4,100
3.400
30,000
49.000
97,000
70,000
48,000
14,000
1,300
400.000
1.100,000
2 , 900 , 000
29,000
610,000
350,000
170,000
51.000
1,000,000
2,300,000
7,500.000
47.000
330.000
150.000
320.000
76,000
1,000,000
4,300,000
11,000,000
Heprinted I ram Hangebrauck e£ a_K '* A blank indicate* that Che compound vae not detected in the aaaple.
bHicroa.rau per |raa> of benaene-eoluble organic eubatancea.
cHicro|raB)a per 1,000 •* of flue fa* at atandard condition* (7O°F. 1 eta).
-------
TABLE 2-}
PAH (aiaaion Sueeury
8en*aoe-
aoluble Croup 1
K>
1
H-"
00
Settle
1SS-22
24
25
27
28
29
10
11
42
44
41 4 62
61
64
65
Type of Unit
Pulveriaad coel
(vertically fired.
dry-kottoe> furnace)
Pulverised coal
(front-wall- fired.
dry-botcoa furnace)
Pulvericed coal
( taagentially fired.
dry-bottoai furnace)
Pulveri«ed coal
( oppoaed- . dovnvard-
inc lined burner*;
wet-bottoai furnece)
Point'
§
,
A
A
8
A
8
A
A
A
A
A
A
A
Organic*
it | per
106 Itu
1.0
0.99
1.7
2.0
1.4
1.4
1.1
1.2
0.14
0.62
0.65
1.1
1.1
1.0
B.P"
: Heat Generation by Coal-Pired Power Planta*
Croup 2
M 8/8 M« P«r
1 aoluble* 1.000 ••
16
22
11
9.1
19
18
110
110
48
11
220
110
21
21
110
50
42
42
110
120
910
270
19
48
120
110
57
46
Ml P*r
H Ib fuel
0.58
0.26
0.21
0.22
0.66
0.61
5.1
1.5
0.21
0.28
1.7
1.9
0.29
0.28
8aP
P
BeP Per BghiP Anth Cor A Phen
Fluor
yf per Bullion Btu heat input
49
22
19
19
56
55
440
110
17
21
140
140
22
21
150
110
190
120
180
210
840
74
200
160
140
110
51
19
45 16
11 19
41
250 66 160 15 4.7 110 820
79 71 81
55 14 200
84 71 150 4.9 7.1 12
420 1.100 91
110 190 19
72 150 8.1
190
210
190
120
410
1.700
84
160
11
190
210
65
55
-------
T«bl« 2-5 (continued)
Bencene-
aoluble Croup 1
Organica BaP"
Sample
No.b Type of Unit
69 Cruahad coal
(cyclone-fired.
70 wet-bottom furnace)
71 Spreader atoker
(traveling grate)
72
73
•Reprinted from Hangebrauck ejt
Additional filter uaed before
CB: lampling point before fly
Sampling
Point0
A
A
A
A
A
.1.1*
the bubblera
aah collector
1 ''"Leaa than" valuea for benxo|a|pyrene were
( per
10 Btu ,
2.1
0.92
1.1
1.8
1.4
in the aampling
i A: after fly
calculated for
Ug/g Mg per pg per
( soluble* 1,000 m3f |b fuel
175
81
22
8.5
11
train.
aah collector
samples having
710 5.2
170 l.l
58 0.11
16 0.19
33 0.19
.
concentrations below
BaP
Mg per
370
76
24
15
15
the lim
Group 1
t BeP Per BghiP Anth Cor A
million Btu heat input
1.800 680 14 160
250 110 16 11
59 61 9.5
12
21
it of quantitative determination
Phen Fluor
110
370
44
59
32
21
(approximately 0.6 „ g per aaaiple). Similar calculation! were not included for the other PAHa (indicated by blanka in the
table).
eMicro(ra»* per gra» of benxene-aoluble organic aubatancea.
'Hicrograaa per 1,000 m3 of flue gaa at atandard conditiona (70°F, 1 atn).
-------
TABLE 2-6
PAH Eaieiion SuaaMrjr: Heat Generation by Intermediate Coal-Pired Unit*
and Intermediate «nd Siull Oil- and Caa-Fired Unit**
Sample
Ho.
ISS-6
7
5
4
14
8
12
10
17
13
15
11
9
18
16
47
Firing
Fuel Net hod
Coal Pulverised]
Chain-grate
atokar
Spreader
• toker
Onderfeed
•toker*
Oil Steear-
atomised
Low-praaaura
air-atomised
Preaaure-
atomiied
Vaporised
Gaa Pram is
burner*
Bensene
•oluble
Organic*
"8 P«r
106 Btu
2.9
1.9
5.4
3.0
4.0
1.4
3.3
14
8.1
3.6
3.5
1.1
1.2
0.95
0.65
5.2
Group 1
BeP
US/8
4 *olc
11
19
4.7
3.400
29
<11
IS
65
<4.6
<17
<33
<17
170
<21
<35
51
B*P»
US P«r
1.000 «3d
75
71
49
7.900
61
<38
40
1.900
<26
<27
<34
<29
350
<23
<30
71
Mt P«r
Ib fuel
0.43
0.44
0.35
140
1.6
<0.3
0.89
18
<0.9
-------
•Reprinted from Hangebrauck e_t ml.
14
''"Lei* Chan" value* for benzo[a]pyrene were calculated for (ample* having concentration* below the liait of quantitative determination
(approximately 0.6 u( per laaple). Similar calculations were not included for the other PAH* (indicated by blank* in the table).
cHicrograa* per gran of bensene-ioluble organic *ub*tance*.
dMicrograu per 1,000 a3 of flue ga* at standard condition* (70°F, 1 at*).
I
to
-------
Compound
TABLE 2-7
Summary of Average PAH Emission from Coal-Stoker-Fired Boilers
With and Without Reinjection from the Stoker Dust Collector8
Concentration"
Reinjected Without Reinjection
Mg/dry SCM0 yg/kg of fuel ug/dry SCMC Ug/kg of fuel
Phenanthrene 9.60
Anthracene 0.29
Methylanthracenes/ 0.66
phenanthrenes
Fluoranthene 2.50
Pyrene 0.86
Methylpyrene/ 0.21
fluoranthene
Benzo[c]phenan- 0.069
threne
Benz[a]anthracene 0.10
Chrysene 0.15
Methylchrysenes 0.051
DimethyIbenz- 0.028
anthracenes
Benzofluoranthenes 0.13
Benzo[e]pyrene 0.044
Benzo[a]pyrene 0.039
Perylene 0.036
Indeno[l,2,3-cd]- 0.055
pyrene
Benzo[ghi]perylene 0.025
3-Methylcholanthrene —
Dibenzopyrene
Coronene —
66.0
2.0
9.40
8.50
0.84
8.60
0.37
0.12
0.093
0.07
0.12
0.059
0.09
0.047
0.043
0.043
0.061
0.093
0.028
0.047
0.029
76.0
68.0
aAdapted from Burlingame e_t a_l.*> Average of all emission from cyclone
filters, condensates, impingers, and XAD-2 absorbent resins. Note: The
analytic results are accurate to within a factor of 3.
^Dashes indicate not calculated.
cug of emission component per standard cubic meter (SCM) of effluent.
2-22
-------
N>
K
Saaple Ho. lb
TABLE 2-8
PAH Concentration! in Coke-Oven
Saaple Ho. 2C
Staple Ho. 4C
Compound
Octahydrophenanthrene
Oc t «hyd roan t h r acene
Dihydrof luorene
Dihydrof luorene
Benxlndene
F luorene
Dihydrophenanthrene
Dihydroanthracane
2-Hathylfluorene
l-Methylfluorene
9-Hethylfluoreoe
Hechy If luorene
Benzoquinol in*
Acridinc
Phenanthrene
Anthracene
Fluorine c«rbonitril«
Methylphenanthrene
He thy 1 anthracene
Bthylphenanthrene
Ethylanthracene
Octahydrof luoranthrene
and octahydropyrene
Dihydrof luoranthene
Dihydropyrene
Fluoranthene
Dihydrobenzo(a) f luorene
Dihydrobenxo(b| f luorene
and dihydrobenco(c)-
t luorene
Peak Ho.
1
la
2
2a
3
4, 4a
4b. 5
6
7
a
9
10
11
12, 12a
13
14
15
16
17
16
19
20
21
22
23
24
25
Concentration,
ua/£ of aaaple
31.85
29.89
30.31
18.76
106.73
271.52
586.98
168.88
98.71
73.46
44.32
87.84
77.74
85.98
2,828.54
942.85
180.29
1,023.41
1,692.26
1,578.60
1,096.71
280.42
115.07
575.06
5,979.74
791.41
213.53
Peak No.
„ ._
1
2
—
2«
3
4
5
6
7
a
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
Concentration,
UB/B ot lanple
8.77
13.62
—
6.72
20.86
79.55
21.92
9.06
24.35
11.81
10.77
8.34
32.44
163.53
46.44
16.76
44.67
85.30
58.04
49.19
11.06
73.57
36.65
269.74
24.03
28.77
Peak No.
1
la
—
2
3
4
5
Sa
6
7
—
8
9
11
lie
lib
12
13
14
15
16
17
18
19
—
20
Concentration,
MB/8 of aaeiple
__
29.98
15.68
—
58.15
19.29
316.49
101.68
8.87
44.01
6.71
—
31.41
74.98
458.80
305.89
32.55
130.22
258.93
202.97
249.86
64.36
64.87
52.29
451.95
—
99.07
Peak No.
_ —
2
2a
2b
2c
2d
3
4
5
6
6a
7
8
9
9a
9b
10
11
12
13
13a
13b
14
15
—
16
Concentration,
Wl/g of aaaple
70.31
18.26
23.20
20.81
15.37
325.83
232.54
40.28
102.83
10.38
79.77
172.79
636.98
500.36
11.47
283.92
573.25
197.39
1,473.68
44.53
31.86
404.25
1,097.82
46.38
-------
Table 2-8 (continued)
ro
I
Compound
Pyrene
Benzo(a) f luorene
Benzo(b) f luorene
Benzole | f luorene
Me thy If luoranthene
Me thy If luoranthene
Methytpyrene
Hethylpyrena
Benzofc Iphenanthrene
Benzo ||hi ] f luoranthene
Dihydrobenz [ a ) anthracene ,
dihydrochryaene, and
dihydrotriphenylene
Bens ( a J anthracene
Chryaene and triphenylene
Oihydrovethylbenz(a)-
anthracene, dihydro-
awthylchryaene, and
dihydroMthyltri-
phenylene
Methylbenz [a (anthracene
Methyl triphenylene
Methylchryaene
DihydroMthylbencofk and
bjf luoranthenea and
dihydro»ethylbenio(a
and elpyrcnea
Diaethylbenz(a)-
anthracene, di«ethyl-
triphenylene. and
d late thy Ichryiene
Benzol j] f luoranthene
Benzol k]f luoranthene and
benzo(b) f luoranthene
Saaiple Ho.
Peak Ho.
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44, 44a
45
46
lb
Concentration,
MI/I of aaaple
4.627.33
971.18
109.45
627.02
1.817.37
390.94
1.016.76
856.91
220.45
677.35
383.03
2.740.45
4,202.02
841.67
159.33
463.99
1,151.61
434.38
246.35
176.92
3,930.34
Saaple Ho. 2C
Peak No.
25
26
27
28
29
30
31
32
33
34
34a
35
36
37
38
39
40
40a
Concentration,
u8/K of aaaple
206.35
87.42")
16.70J
38.96
124.73
21.87
31.12
106.34
82.70
164.25
54.75
105.15)
119.041
93.30
22.36
40.81
107.50
8.63
Sample Ho. 3d
42
43
46.73
18.66
155.03
Peak No.
21
22
23
24
25
26
27
28
29
29a
30
30a
31
32
33
34. 35
36
37
38
39
40
Concentration,
UK/K of aaaple
472.09
62.97
290.03
126.03
179.04
97.80
233.37
1.510.34
201.74
101.78
5.509.43
247.80
1.015.70
371.97
193.68
185.08
174.71
62.25
30.16
80.04
2,170.92
Saaple Ho. 4e
Peak No.
17
17a
17b
17c
18
18a
19
20
21
21a
22
23
23a
24
25
26
26a
27
28
29
30
Concentration,
MB/g of aaaple
1.446.64
37.21
22.63
98.47
101.56
34.83
120.66
175.42
2.156.14
151.88
271.38
2,673.65
86.35
1.669.72
369.10
449.19
137.26
160.89
451.70
285.34
2.556.98
-------
Table 2-8 (continued)
Sample Mo. lb
Sample No. 2C
Sample Ho. 3d
Sample Ho. 4e
Compound
Me thylbeaco(k) fluor-
anthene and me thy 1-
benzo(b) fluoranthene
Benzole ) pyrene
Benzo [ a ] pyrene
Perylene
Methylbenzol a 1 pyrene
Dime thy Ibenco JkJ-
fluoranthene and
dimethylbenzo(b)-
f luoranthene
Dime thy Ibenco [ a ) pyrene
Dibenzanthracene
Is0 o-Phenylenepyrene
fo Benzol ghilperylene «nd
*•" anthanthrene
Me thy Idibens anthracene
Me thy Ibenco [ghilperylene
Coronene -
Dibencopyrene
Peak No.
47
48
48a
49
50
51
52
53
54
55
56
57
58
59
Concentration,
iig/g of sample
735.95
103.86
2,630.92
702.12
330.85
116.74
82.68
123.66
101.54
72.35
89.04
36.79
864.55
693.21
Concentration,
Peak No. pg/g of sample
44 33.05
—
45 122.15
46 22.07
47 5.88
—
—
—
—
—
—
—
—
__ _«
Peak No.
41
—
42
43
44
45
46
47
—
—
—
—
51
52
Concentration,
pg/g of sample
430.05
—
2,007.75
616.85
344.12
73.13
70.18
84.52
—
—
—
—
833.30
587.05
Concentration,
Peak Ho. M^/> of sample
31 492.13
—
32 2,297.83
33 698.44
34 247.96
—
—
—
—
—
—
—
35 766.58
36 493.27
•Reprinted with permission from Lao e_t al. *9
bClass-fiber filter No. 1: total weight of material collected, 12.59 mg; total volume of solvent extract, 1 ml; injected sample size, 10 pi.
'Glass-fiber filter No. 2: total weight of material collected, 39.67 mg; total volume of solvent extract, 1.0 ml; injected sample size, 10 pi
^Silver membrane filter No. 1: total weight of material collected, 5.75 mg; total volume of solvent extract, 0.6 ml; injected sample size,
6 pi.
'Silver membrane filter No. 2: total weight of material collected, 1.96 mg; total volume of solvent extract, 1.0 ml; injected sample site,
10 UL.
-------
TABLE 2-9
PAHs on Particles from Coke Ovens as a Function of Particle Size3
Size
PAH
Phenanthrene
Anthracene
Methylphenanthrene/
methylanthracenec
Fluoranthene
Dihydrobenzo [ab]-
f luorene
Pyrene
Benzo [a] f luorene
Benzo [ b 1 f luorene
Benzo [c ] phenanthrene
Benz [a Janthracene
Chrysene/triphenylene
Benzo [bjk]f luoranthene
Benzo [e]pyrene
Benzo[a]pyrene
Perylene
o-Phenylenepyrene
Benzo [ghi Jperylene
Anthanthrene
Coronene
> 15 ym 7-15 ym 3-7 pm
vg
0.5
0.3
d
d
d
d
d
d
d
0.03
0.06
0.05
0.03
0.02
d
d
d
d
d
yg
0.5
0.2
d
d
d
d
d
d
d
0.08
0.10
0.40
0.14
0.12
d
d
d
d
d
yg
0.8
0.6
0.4
4.2
0.6
4.3
3.5
2.6
2.0
3.3
5.8
4.8
3.4
5.2
1.3
3.8
5.2
1.2
5.9
0.9-3 ym <0.9 ym
lig
2.0
0.7
1.3
10.7
2.1
12.3
9.1
7.2
5.2
9.0
13.2
10.6
8.9
12.1
2,9
7.2
7.8
3.2
6.3
yg
b
b
d
2.0
d
1.3
0.5
0.3
0.8
4.0
1.8
7.2
0.5
1.4
d
0.5
1.3
0.3
0.5
Dibenzopyrene d d 0.9 1.5 0.3
Reprinted with permission from Bjorseth; copyright Ann Arbor Science
Publishers, Inc.
^Detectable.
clsomer not determined.
dNot detected.
2-26
-------
POM
TABLE 2-10
Eniaaion (g/kg)*
ro
i
N>
COMpdunda
Ant;..acene/phenanthrene
Methyl-anthracenei/
-phenanthrenea
Ca-alkyl-anthracenea/
-phenanthrenea
Cyc lopenta-anthracenea/
-phenanthrenea
Fluoranthene
Pyrene
Methyl-fluoranthenea/
— pyrenea
Benzol ghi ] fluoranthene
Cyclopenta(cd)pyrene
Benzo[c Iphenanthrene
Benc [ a] anthracene/
chryaene
Methyl-benzanthracenea
-benzphenanthrenea/
-chryaenea
Ca-alkyl-bencanthracenea/
-benzophenanthrenea/
-chryaenea
Benzof luoranthenea
Benzopyrenea/perylene
Hethylcholanthrene
Indeno [1,2, 3-cd ] py rene
Be nzo [ gh i ] per y lene
An than th rene
Dibenzanthracenea/
-o-phenanthreoea
Dibenzocarbazolea
Dibeozopyrenea
Total1- (• column turn)
Fireplace
Seaaoned
POM
Train
0.0082
0.0027
0.0014b
0.0014b
0.00l4b
0.0014b
0.0014b
—
0.0014b
—
0.0014b
0.0014b
—
0.0014b
0.00146
—
—
—
—
—
—
0.0249
Baffled Stove
Oak
SASS
Train
0.0114
0.0034
0.0011
0.0004
0.0026
0.0026
0.0023
0.0009
0.0010
0.0004
0.0020
0.0013
0.0009
0.0022
0.0017
—
0.0013
—
0.0003
—
0.0007
0.036S
Green Pine
POM Train
0.0069
0.0083
0.0014
0.0014
0.0016
0.0016
0.0016
0.0014
0.0014
0.0013
0.0014
0.0016
0.0014
0.0016
0.0014
—
0.0015
—
0.00005
—
0.0001
0.0360
Seaaoned Oak Seaaoned Pine
POM Train
0.0745
0.0211
0.0040
0.0032
0.0180
0.0156
0.0128
0.0048
0.0048
0.0016
0.0125
0.0062
0.0055
0.0128
0.0083
0.00007
0.0045
—
0.0007
—
0.0011
0.2121
POM Train
0.1463
0.0510
0.0070
0.0086
0.0316
0.0240
0.0167
0.0067
0.0089
0.0023
0.0138
0.0104
0.0044
0.0159
0.0116
--
0.0099
—
0.0014
—
0.0010
0.3715
0.0037
0.0112
0.0084
0.0043
0.0010
0.0007
0.18851
Green Pine
POM
Train
Honbaffled Stove
Seaaoned Oak
POM Train
0.0618 0.1034
0.0167 0.0513
0.0045 0.0094
0.0030 0.0047
0.0208 0.0188
0.0169 0.0188
0.0103 0.0142
0.0047 0.0047
0.0051 0.0138
0.0016 0.0046
0.0076 0.0371
0.0062 0.0048
0.0047
0.0141
0.0094
0.0048
0.00005
0.00002
0.3187
0.0104
0.0028
0.0008
0.0002
0.0012
0.0013
0.0016
0.0004
0.0005
0.0002
0.0013
0.0009
0.0005
0.0015
0.0011
0.0011
0.0002
0.0005
0.0265
-------
•Reprinted fro* DeAngelis e£ a_U10 POM train used an XAD-2 reain and cooler (reduction to 2l°C) to trap organic gaaes. SASS
dource aaaeaaaent •••pling ayatea) train uaed three cyclonea, a filter for particle-aize fractionation, an XAD-2 trap, and a trace
inorganic iapinger trap. Blank* indicate not detected.
bCo*pound waa identified, but not quantified, because of the detection liaiita of the analytic method.
cTh« detection liaiit waa taken aa the eaiaaioo factor for compounds that were identified, but not quantified.
KJ
i
ho
ot>
-------
TABLE 2-11
POM Emission Factors3
Compounds
Phenanthrene/ anthracene
Ci-Phenanthrenes/anthracenes
C2~Phenanthrenes/ anthracene a
Cyclopentaphenanthrenes/
anthracenes
Pyrenes
Fluoranthene
Benzo[a]f luorene
Unidentified POMs
C^-Fluoranthenes/pyrenes
Benzofghi] f luorene
9-Phenylanthracene
C3~Phenanthrenes /anthracenes
Benzo [ ghi ] f luoranthene
Cyclopenta[cd]pyrene
Benzo [ c ] phenanthrene
Benz[a]anthracene/chrysene
Higher-molecular-weight POMs
Total
Emission Factor,
0.12-m
Logs,
0.82-kg/h
Burn Rate
0.88
0.42
0.11
ND
0.33
0.25
0.26
0.70
0.10
0.04
0.04
0.20
ND
0.08
ND
ND
0.44
0.25
mg/kgb
0.06-m
Logs,
7.73-kg/h
Burn Rate
2.30
0.18
0.04
ND
1.39
0.10
0.26
1.31
ND
0.10
0.20
0.03
ND
ND
ND
ND
1.35
1.26
4.10
8.52
aAdapted from Hubble £t £K
bND - not detected.
2-29
-------
TABLE 2-12
Estimation of Annual Emission of POM Contributed
by Various Sources*
Source
Residential heating:
Wood-fired (total):b
Coal-fired
Oil-fired
Gas-fired
Annual Emission of Total POM
Metric tons %
3,837
102
7.4
9.8
34.8
0.9
0.1
0.1
Open burning:
Agriculture open burning
Prescribed burning
Forest wild fires
Coal-refuse fires
Land-clearing waste burning
Structural fires
Coke production
Mobile sources:
Auto—gasoline
Auto—diesel
Trucks—diesel
Industrial boilers:
Coal
Gas
Oil
Wood/bark
Bagasse
1,190
,071
,478
28.5
171
86
632
2,160.8
1.2
103.5
69.0
2.1
1.3
1.2
0.3
10.8
9.7
13.4
0.3
1.6
0.8
5.7
19.6
0.1
0.9
6.3
0.1
tabulated from Peters36 for the early to middle 1970s.
"Primary heating, auxiliary heating, and fireplaces responsible
for 1,383, 2,376, and 78 metric tons, respectively.
2-30
-------
TABLE 2-13
Emission of PAHs in Wood Smoke3
Emission, mg/kg of wood burned
Fireplace
Seasoned
PAH Oak
Anthracene/ 8
phenanthrene
MethyKanthracenes/ 3
phenanthrenes)
Fluoranthene <1
Pyrene < 1
Me thyKf luoranthenes/ <1
pyrenes)
Benzol ghi ]-
f luoranthene
Cyclopenta[cd]pyrene <1
Benzo [c ] phenanthrene
Benz [a] anthracene/ <1
chrysene
Benzof luoranthenes <1
Benzopyrenes/perylene <1
Benzo [ ghi ] perylene
Baffled
Seasoned
Oak
75
21
18
16
13
5
5
1
13
13
8
1
Stove
Seasoned
Pine
146
51
32
24
17
7
9
2
14
16
12
1
Nonbaffled
Seasoned
Oak
62
17
21
17
10
5
5
2
8
11
8
1
Stove
Green
Pine
103
51
19
19
14
5
14
5
37
14
9
0
aData from Peters.
36
2-31
-------
TABLE 2-14
Emission Factors for PAHs from Residential Combustion
Sources as a Function of Thermal Efficiency3
Combustion Source
Wood-burning stove
Wood-burning fireplace
Automatic coal furnace
Oil furnace
Gas furnace
Approximate
POM Emission
Factor, pg/J
(fuel input)
17,000
2,000
1,900
4
1
Average Thermal
Efficiency, I
55
10
60
70
85
Approximate PCM
Emission Factor,
pg/J (thermal
output)
31,000
20,000
3,200
5.7
1.2
aData from Peters. °
2-32
-------
TA1LB 2-15
Polycyclic Aromatic Hydrocarbon Eaiaiion Sueaury by Incineration Source*
Croup 1
Croup 2
rO
1
U>
U)
Type ot Unit
Municipal:
250-ton/d aultiple
chaaber
50-ton/d aultiple
chaaber
Coaaerciel:
5.3-ton/d aingla
chaabar
3-ton/d aultiple
chaabar
Benzol a Ipyrene,
1*1/1,000 Pyrene
Sampling Point a' ug/lb of refuae charged
•reaching (ahead of 19 0.075 8.0
eat t ling chaaber)
Breeching (ahead of 2.700 6.1 52
acrubbar)
Stack (behind 17 0.089 2.1
acrubber)
Stack 11.000 S3 320
Stack 52.000 260 4.200
Benzole)- lenzo(ghi)- Anthan- Anthra- Phenan- Fluoran- Benz[a]-
pyrene Perylena pcrylene threne Coronena cene threne thane anthracene
Ht/lb of refuae charged
0.34 -- — -- 0.24
— — 9.8 0.37
12 — 34 — 15
— 18 4.6 —
0.50 — 0.63 -- 0.63
~ — 3.3 0.15
45 3.1 90 6.6 21
** 140 220 4.6
260 60 870 79 210
86 59 3,900 290
•leprinted from Bj»irona*ntal Protection Agency. Review of Literature. Cited in Uangebrauck tt_ a_l. 1* — • not detected in
the lamp la.
^ticrograma par 1.000 a3 of flue gae at atandard condition! (70°r. 1 ata).
-------
TABLE 2-16
N>
I
U>
Emission of Polycycllc Aromatic Hydrocarbons3
Stack Cases Residues
Compound
Fluoranthene
Pyrene
Benz [a janthracene +
chrysene
Benzofbjf luoranthene +
benzo[k]f luoranthene +
benzo f J ] f luoranthene
Benzo[a Ipyrene +
benzo [e] pyrene
Perylene
Benzo [ghl ] perylene
Indeno [1,2, 3-cd ] pyrene
Coronene
rag/ 1,000 m
0.58
1.58
0.72
0.32
0.02
0.18
0.42
0.18
0.04
mg/db
274
745
340
151
9
85
198
85
19
Pg/kg
58
49
171
292
147
82
47
10
20
mg/dc
1,360
1,150
4,010
6,850
3,450
1,920
1,100
230
470
Water
Input ,
Mg/L
0.08
0.08
0.03
0.03
0.03
0.02
0.01
0.01
0.01
Output,
Ug/L
0.62
0.54
0.64
0.14
0.14
0.13
0.03
0.01
0.01
Output,
mg/dd
15.5
13.5
16.0
3.5
3.5
3.3
0.8
0.3
0.3
aReprlnted with permission from Davles et_a±.;^ copyright 1976 American Chemical Society.
^Assuming 12 h dally operation; levels measured at flow 655 mVmln corrected to dry, 20°C,
760 mm Hg.
cAssuming 23Z average moisture content and 30-metrlc-ton dally output.
^Assuming a discharge of 25,000 L/d from the quench trough.
-------
TABLE 2-17
PAH Emission from a Secondary Lead Smelter Processing Batteries3
Compound
Emission Concentration,*3
ng/Nm3
Anthracene/ phenanthrene
Methylanthracenes
Fluoranthene
Pyrene
Methylpyrenes/fluoranthenes
Benzo [ c ] phenanthrene
Chrysene/benz [ a. ] anthracene
Benzo[a]pyrene
aData from Bennett e al.^
H * corrected to 70°F and 1 atm. Sampling runs 1 and 2 were taken
during separate visits to the smelter. Gas stream was split, and two
simultaneous samples, A and B, were taken.
1A
600
25
160
31
2
10
25
1
IB
740
34
170
22
2
11
23
1
2A_
770
41
330
28
3
17
28
1
2B
940
33
310
30
2
13
25
1
2-35
-------
TABLE 2-18
PAHs from Burning of Pine Needles, by Fire Type3
Concentration, ng/g of fuel burned (dry weight basis)
to
I
PAH
Anthracene/phenanthrene
Methylanthracene
Fluoranthene
Pyrene
Methylpyrene/fluoranthene
Benzo[c]phenanthrene
Chrysene/benz[a janthracene
Methylchrysene
Benzofluoranthenes
Benzo[a]pyrene
Benzo[e)pyrene
Perylene
Methylbenzopyrenes
Indeno[l,2,3-cd Jpyrene
Benzofghi]perylene
Total
Total suspended partic-
ulate matter (TSP)
Benzene-soluble organic
substances
Backing fires
0.1 Ib/ft
12,181
9,400
14,563
20,407
18,580
8,845
28,724
17,753
12,835
3,454
5,836
2,128
6,582
4,282
6,181
171,750
21 Ib/ton
55%
0.3 Ib/ft
2,189
1,147
2,140
3,102
2,466
1,808
5,228
1,891
1,216
555
1,172
198
963
655
1,009
25,735
9 Ib/ton
50%
0.5 Ib/ft
584
449
687
1,084
1,229
468
2,033
877
818
238
680
134
384
169
419
10,249
5 Ib/ton
45%
Heading Fires
0.1 Ib/ft
2,525
1,057
733
1,121
730
244
581
282
164
38
61
33
65
—
—
7,632
20 Ib/ton
44%
0.3 Ib/ft
5,542
4,965
974
979
1,648
142
543
1,287
129
40
78
24
198
—
—
16,549
73 Ib/ton
73%
0.5 Ib/ft
6,768
7,611
1,051
1,133
2,453
175
836
1,559
241
97
152
46
665
—
—
22,787
118 Ib/ton
75?;
aReprinted with permission from C. K. McMahon and S. N. Tsoukalas. "Polynuclear aromatic hydro-
carbons in forest fire smoke," pp. 61-73. In P. W. Jongs and R. I. Freudenthal, Eds. Polynuclear
Aromatic Hydrocarbons. Vol. ^. Carclnogenests. New York: Raven Press, 1978. Moisture content
for all fires ranged between 18% and 27%.
-------
TABLE 2-19
Estimated Total Annual Benzofajpyrene Emission
for 1975 and 1985
Benzo[a]pyrene Emission,
metric tons
Estimate 1975 1985
Minimum 346 67
Intermediate 588 358
Maximum 1,676 885
2-37
-------
3 »-
I
Q
Cok«
So*
AEROOYNAMC DIAMETER, dp,
FIGURE 2-1. Distribution of B«P and BghiP as function
of aerodynamic diameter. Total PAH concentrations, in
ng/nr: tunnel, 42.8 BaP; coke oven, 58.5 BaP and
44.2 BghiP; and ambient air, 1.01 BaP. Reprinted with
permission from Miguel and Rubenich.^^
2-38
-------
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2-43
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ATMOSPHERIC TRANSFORMATIONS OF POLYCYCLIC AROMATIC
HYDROCARBONS
GENERAL CONSIDERATIONS:
PERSISTENCE AND TRANSFORMATIONS OF PAHs
The atmospheric persistence of PAHs has received considerable
attention in recent years and continues to be actively investigated. Two
extreme situations can be envisioned. In the absence of any chemical
interaction, the lifetime of PAHs adsorbed onto particles will depend
solely on physical characteristics—the size of the carrier particle and
scavenging processes, including wet and dry deposition. In addition,
carrier-particle size is also critical with respect to the rate of
deposition in (and clearance from) the human respiratory system and the
rate of elution from the carrier particle by the lung tissue. Because
submicrometer particles have atmospheric residence times of several days,
experimental evidence on long-range atmospheric transport of PAHs and
their distribution in sediments appears to support a hypothesis of
negligible chemical transformation of PAHs in the atmosphere. Given the
same carrier-particle residence time, even relatively slow chemical
reactions could compete effectively with physical processes, with respect
to PAH removal from the atmosphere. A substantial body of experimental
evidence has been accumulated on chemical reactions between PAHs and
pollutant gases under laboratory conditions, with reaction times as short
as a few hours. The products of these reactions are in some cases much
more potent rautagens than the parent PAHs, thus warranting concern about
the implications of these chemical transformations with respect to human
exposure.
The foregoing considerations suggest the format of this chapter.
Pertinent information concerning the chemical and physical processes
governing atmospheric persistence of PAHs is summarized first, followed by
chemical reactions of PAHs, with an attempt to organize the somewhat
conflicting published data according to reactant species and substrate,
i.e., carrier particles and other substrates, including filters.
PAH FORMATION: CHEMISTRY AND PHYSICS
CHEMISTRY
The exact synthetic chemistry that produces PAHs in a fuel-rich flame
is not well known, but PAHs can be produced from almost any fuel burned
under oxygen-deficient conditions.2 As an example of the PAH assemblage
produced by combustion systems, Figure 3-1 (top) shows identified gas-
chromatographic mass-spectrometry (GC/MS) peaks on PAHs produced by the
3-1
-------
combustion of kerosene.-*' Note that fluoranthene (peak 22) and pyrene
(peak 25) are present in about equal abundances; that the abundance of
phenanthrene (peak 14) far exceeds that of anthracene (peak 15), a less
stable compound; and that benzo[a]pyrene (peak 39) is always found with
its noncarcinogenic isomer benzo[eIpyrene (peak 38).
A particularly interesting group of compounds in combustion effluents
are those with a vinylic bridge, such as acenaphthylene (peak 4) and
cyclopenteno[cd]pyrene (peak 32). Peak 23, although not labeled, has been
positively identified as acephenanthrylene,** which also has a vinylic
bridge. We emphasize this structural feature because of its chemical
reactivity (compared with that of the fully aromatic portions). This
reactivity is important in considering the fate of PAHs in the atmosphere.
The PAHs shown in Figure 3-1 (top) are typical of those produced from
the combustion of various fuels. The combustion of almost any fuel will
produce the mixture of compounds shown. The relative abundances, however,
can be substantially different, depending on the temperature of combus-
tion. In fact, the relative abundances of the alkyl horaologues of PAHs
depend heavily on the temperature at which the fuel is burned. Although
Figure 3-1 shows very modest amounts of alkyl homologues (see the region
between peaks 25 and 30), other fuels burned under other conditions can
show considerably greater abundances of alkyl PAHs.
PHYSICS
Adsorption
PAHs are formed in almost all combustion processes. As the effluent
temperature decreases, PAHs initially present largely as vapors become
adsorbed on condensing carriers, such as soot and fly ash.-^8, 67 It .is
generally accepted that the adsorption process is virtually completed at
or near the point of emission into the ambient air, and that PAHs in
ambient air are adsorbed on carrier particles. Studies of the distribu-
tion of selected PAHs between the gaseous and particulate phases in
ambient air"»^ have shown that, even though the smaller PAHs (e.g., with
three and four aromatic rings) may have measurable gas-phase
concentrations, these are, as a result of adsorption, lower by several
orders of magnitude than those expected on the basis of the corresponding
vapor pressures (see Table 3-1).
Particle Size Distribution
Once PAHs are adsorbed onto carrier particles, their size distribu-
tion in the atmosphere is governed by aerosol dynamics, including co-
agulation and condensation processes. Thus, carrier particles may evolve
into substantially different "stable" size distributions. In many com-
bustion processes, PAHs are emitted in the so-called nucleation mode,
i.e., adsorbed on particles less than 0.1 pra in diameter. In diesel-
engine exhaust, the carrier-particle distribution has mass median
diameters of about 0.1-0.25 urn (National Research Council, unpublished
3-2
-------
manuscript). The contributions from various anthropogenic emission
sources may have significant effects on the size distribution of airborne
PAHs.
In early studies of the PAH size distribution in urban air, DeMaio and
Corn1^ reported that most of the benzo[a]pyrene (BaP) was found to be
associated with small particles (less than 2.3 yra in diameter).
Kertesz-Saringer and co-workers^7 reported that 50% of the BaP in
Budapest air was found in particles smaller than 0.3 um. Size distribu-
tion measurements were later extended to other PAHs, including benzofk]-
fluoranthene, *• 8-PAH and 2-PAH quinones,7^ 4-azaarenes, and 3-alkyl-
substituted PAHS.**7 More recently, the application of new size-
segregating sampling devices, such as the low-pressure impactor, has given
more detailed information on the distribution of PAHs in the submicrometer
range. Miguel and Friedlander6^ reported on the distribution of BaP and
coronene in Pasadena, California, ambient air (Table 3-2). The largest
concentration of both PAHs was found in particles with aerodynamic
diameters between 0.075 and 0.12 ym.
The influence of particle size on human respiratory uptake has been
the subject of a number of theoretical and experimental studies.^»16,68
In the specific case of PAHs, it has been conclusively shown that the rate
of uptake by lung membranes is much higher for PAHs adsorbed on physio-
logically inert carrier particles than for the same PAHs inhaled in Che
crystal state. '-^ In addition, simultaneous measurements of carrier-
particle and BaP clearance from the respiratory tract of mice for small
(0.5-1.0 ym) and large (15-30 ym) carbon particles^ showed that, even
though smaller particles were cleared from the respiratory tract faster
than larger ones, more BaP was eluted from small particles than from the
large ones. These results are consistent with findings in Cumorigenesis
studies of Farrell and Davis, in which the BaP-carbon combination of
the smallest particles (0.5-1.0 ym) was the most carcinogenic, and
underline the importance of PAH size distribution for human toxicity.
However, most respiratory deposition-clearance studies have been limited
to two sizes (i.e., about 1 ym and over 10 ym), and no information is
available on the effect of carrier-particle size on PAH retention in the
range of interest, i.e., less than 0.25 jjm. Such size resolution would
provide valuable information not only on PAH retention from ambient
particles, but also on the relative contribution of various emission
sources to PAH uptake.
PHYSICAL REMOVAL PROCESSES FOR ATMOSPHERIC PAHs
Once PAHs are released from the combustion system and adsorbed on soot
or fly ash, they are exposed to potential atmospheric degradation. In the
absence of major photodecomposition or other chemical transformations,
PAHs would be removed from the atmosphere by dry and wet deposition.7^
Dry deposition involves sedimentation, turbulence-induced collision with
surface electrostatic deposition, and inertial impaction. Although
settling velocities have apparently not been determined for PAHs, it is
generally accepted that they are controlled by those of the carrier
particle.
3-3
-------
Carrier-particle settling velocities can be estimated from Stokes's
law, i.e., assuming that the settling velocity is proportional to the
square of the particle diameter, to a term that includes the particle-
to-fluid (air) density ratio, and to the reciprocal of the fluid
viscosity. Thus, for a 1-ym particle with a density of about 2 g/cnr in
air at 20°C, the settling velocity is about 6 x 10"^ m/s,"^ in
agreement with experimentally determined velocities of about 10 x 10"^
m/s for 1-um particles.°^ For such a particle suspended in air at a
height of 20 ra and with an average wind speed of 4 m/s (about 9 mph), it
would take 4 d to settle to the surface. Assuming a constant wind speed
of 4 m/s and constant wind direction over the 4-d period, this atmospheric
residence time is equivalent to atmospheric transport over a distance of
1,400 km. Experimental evidence of such regional- and subcontinental-
scale transport of PAHs in the atmosphere is discussed below.
A simple way in which to note the relative degradation suscepti-
bility of the various PAHs is to compare the GC/MS data on PAHs coming
from a combustion system (see Figure 3-1, top) with the PAH profile of
atmospheric particles (Figure 3-1, middle). PAHs without vinylic bridges
are still prevalent, the ratio of fluoranthene to pyrene is still about
1:1, and the ratio of phenanthrene to anthracene is about 10:1. Compounds
with vinylic bridges (acenaphthylene, peak 14; acephenanthrylene, peak 23;
and cyclopenteno[cdjpyrene, peak 32) have completely vanished from the PAH
mixture found in the atmosphere. The increased chemical reactivity of the
relatively localized double bond found in these compounds apparently makes
them susceptible to photolytic oxidation.
Assuming that most PAHs are stable in the atmosphere, what happens to
these compounds after they are released from combustion systems throughout
the world? Two types of data address this question: data on PAHs in
marine and lacustrine sediments, presumably the ultimate environmental
sinks of atmospheric PAHs; and data on PAHs in air sampled at remote
locations.
PAHs IN MARINE AND LACUSTRINE SEDIMENTS
Many workers have observed significant concentrations of PAHs in
aquatic sediments. For example, Figure 3-1 (bottom) shows a GC/MS
analysis of PAHs in the sediment of the Charles River. In comparing the
bottom and the middle of Figure 3-1, one sees considerable resemblance.
The ratios of the major groups of compounds are the same; the PAHs with
vinylic bridges are missing, as they were in the atmosphere; and the alkyl
homologues are about as abundant as one might expect. Similar data have
been obtained, but in a more quantitative fashion, on over 50 sediment
samples from around the world." These data indicate that PAHs are
ubiquitous and that they are found in almost all samples both near and
remote from urban areas. The PAH pattern in all these samples, even the
most remote, is similar to that shown in Figure 3-1 (bottom).
3-4
-------
Even though the relative distribution remains constant, the total
amount of PAHs decreases dramatically with distance from urban centers.
Figure 3-2 shows a plot of the total PAH abundance in five marine-sediment
samples taken from Massachusetts Bay as a function of distance from
Boston. There is a decrease by 3 orders of magnitude in the total
abundance of PAHs within 100 km of Boston. At that point, the total PAH
concentration is about 100 ppb; remarkably, that is what is seen in almost
all other samples from areas remote from urban centers.
On the basis of these and other measurements of PAH concentrations,
the following scenario is suggested for the transport of PAHs. The
various fuels that are burned in metropolitan areas produce airborne
particulate matter (soot and fly ash) on which PAHs are adsorbed. These
particles are transported by the prevailing wind for distances that depend
heavily on particle diameter. The long-range airborne transport of small
particles may account for the presence of PAHs in deep-ocean sediments.
Larger airborne particles will settle back onto the urban area; rain
then washes them from streets and buildings. The PAHs in this urban
runoff eventually accumulate in local sinks. These highly contaminated
sediments could be slowly transported by resuspension and currents to
seaward locations, where the sediments accumulate in basins or in the deep
ocean. The rapid decrease in PAH concentration to 100 ppb within 100 km
of Boston (see Figure 3-2) indicates that this transport mode is a rather
short-range effect.
The stability of PAHs is also apparent when one examines sediment
samples taken in such a way as to preserve the historical record. This
can be done by carefully coring sediments, segmenting the core into 2-
4-cm sections, and analyzing each section for PAHs quantitatively. An
example of such data is shown in Figure 3-3; this represents a core from
the Pettaquamscutt River in Rhode Island, an anoxic basin. ^ The total
PAH concentrations range from 14,000 ppb near the sediment surface to less
than 120 ppb at the core bottom. Despite the range of concentrations, the
relative distribution of the PAHs (excluding the natural products retene
and perylene) is indicative of combustion. For example, the ratio of the
isomers (nonalkylated) to their monoalkyl homologues
i-3 3*0 JL 0.4:1. In no case does this ratio become less than
unity, which would be expected if the source were direct fossil-fuel
contamination. The ratio of the C^gH^Q isomers to the G^gH-^
isomers is 2.7 +_ 0.3:1, and the ratio of the C^^H^g isomers to the
^20^12 isomers is 0.46 +_ 0.08:1. These ratios are consistent
throughout the core and are indicative of combustion sources." Com-
bustion seems to have been the source of the PAHs in all sections of the
core.
With a reported deposition rate of 3 mm/yr, total PAHs (excluding
retene and perylene) in the Pettaquamscutt core were plotted against year
of deposition (see Figure 3-3). For comparison, the BaP data reported by
Grimmer and Bohnke^' for a core from the Grosser Ploner Sea are also
plotted in Figure 3-3. The similarity between these two core profiles is
3-5
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remarkable. Both show rapid increases in PAH concentrations beginning
around 1900. The increases could be due to the heavy industrialization
that occurred at the turn of the century and the combustion associated
with it.
A slight decrease in total PAHs around 1930 is present in both cores
(see Figure 3-3). It is intriguing to speculate that this reflects an
event chat occurred in both Europe and New England at this time. The
Depression could be such an event. During the Depression, total U.S.
energy consumption decreased from 25 x 10^5 BTU (in 1929) to 18 x 1015
BTU (in 1932) before resuming its increasing trend.^°
The Pettaquamscutt data are from a core deep enough to allow the
assessment of the PAH burden before 1900. The PAH concentrations are low
and constant (about 200 ppb) for the 50 yr before the turn of the
century. That may be indicative of PAHs from natural combustion
processes, such as forest fires. Contributions from natural processes
appear to be insignificant in areas or periods of high anthropogenic
activity.
The decrease in PAHs after 1950 is interesting. It may reflect the
change from coal to oil and natural gas as home heating fuels that
occurred in the 1950s. During the period 1944-1961, the use of coal in
the United States decreased by 40X, and the use of oil and gas increased
by 20%. ° Combustion of coal usually produces more PAHs than oil and
gas, so the change in fuel would result in a decrease in PAH production
during the same period. A return to coal as a major energy source without
stringent emission controls might therefore have an important effect on
man's input of PAHs into the atmosphere and the sedimentary environment.
In an effort to measure the deposition rates for PAHs from the
atmosphere in both remote and urban locales, PAH concentrations in
sediment cores from water bodies in several areas in the northeastern
United States have been determined, and the corresponding atmospheric PAH
fluxes to these sites have been calculated. In assessing flux information
(rather than concentrations), many of the differences between sites are
taken into account, thereby allowing useful comparisons. PAH fluxes
calculated for lakes on islands and for remote high-altitude lakes were
particularly interesting, in that these sites should reflect most
accurately the atmospheric deposition of these combustion-derived
pollutants. This background flux could then be compared with PAH inputs
found nearer to urban centers, thereby showing the relative importance of
long-range airborne vs. short-range runoff delivery of PAHs.
With the observed PAH concentrations and information on the sedimen-
tation rates and in situ dry densities, the fluxes of individual PAHs to
five remote and three urban sites were calculated. ^ Table 3-3 (top)
shows the results of these flux calculations for core subsections
reflecting PAH deposition in remote sites at present, in the interval
including 1950, and at the turn of the century (1900). The first point to
notice is that the average fluxes for most individual PAHs (except anthra-
cene) to remote northeastern U.S. sites are 0.8-3 ng/cra^ at present.
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Where core subsampling resolution permits, it can be seen that atmospheric
PAH fluxes approximately 30 yr ago were 2-3 times greater than and those
around 1900 were one-tenth to one-fifth of the present flux rates. This
historical PAH record clearly shows that man's activities over the last
century resulted in an influx of PAHs to the environment and that
coal-derived energy was a much greater source of polluting PAHs than
energy derived from oil and gas.
For comparison, similar flux estimates for three sites much closer to
urban centers were calculated. The results are shown in Table 3-3. These
locations all have much greater PAH fluxes than the remote locations. As
suggested above, such locations probably receive most of their PAH
contamination via water runoff from the watershed. This source of PAHs in
sediment overwhelms the background atmospheric deposition rates seen at
remote sites.
In summary, several things are now apparent about the physical
transport of PAHs from source to depot:
o PAHs (except retene and perylene) in continental aquatic sediments
originate largely in anthropogenic combustion.
o The watershed runoff resuspension mechanism is of short range
(about 100 km) and delivers a near-shore flux of about 35 ng/cm^ per yr
for an individual PAH.
o The airborne transport mechanism is of long range and delivers a
flux of about 1 ng/cnr per yr for an individual PAH.
o Anthropogenic sources of PAHs were first observed in sediments at
concentrations significantly higher than natural background in around
1900, and the maximal deposition was in about 1950.
We need more information about the mechanisms of PAH transport to
remote sediments. For example, what fraction of the PAH flux is delivered
by aquatic transport mechanisms and what fraction by atmospheric fallout?
How much, if any, is lost to the water column?
PAHs IN AIR SAMPLED AT REMOTE LOCATIONS
Experimental evidence of long-range transport of PAHs has been
presented by Lunde, Bjorseth, and co-workers. .61,62 xhey analyzed the
PAH content of particulate samples collected in Norway with respect to air
trajectory. As seen in Table 3-4, PAH concentrations in air masses
originating in industrialized areas in western Europe were 20 times higher
than those measured in air masses originating in Norway and were as high
as those typically measured in urban and industrial areas. These results
support the concept, at least for the 20 PAHs listed in Table 3-4 and
collected during the winter (i.e., low temperature and low light
intensity, resulting in little, if any, photochemical activity), of
atmospheric transport of PAHs over long distances from anthropogenic
sources.
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CHEMICAL REMOVAL PROCESSES FOR ATMOSPHERIC PAHs
Chemical reactions of PAHs in the atmosphere have received steady
attention for about 30 yr, as have their implications for human health.
In their classic work demonstrating the presence of BaP and other PAHs in
the Los Angeles atmosphere and the carcinogenicity of atmospheric organic
particulate matter in mice, Kotin e_t £!_• investigated the interactions
between BaP deposited on filters and pollutant gases, including nitrogen
dioxide (NC^) and ozone (03). In this early study of the currently
much investigated interactions between PAHs and oxides of nitrogen (NO )
and the health implications of nitro-PAH compounds, Kotin e_t a_l. reported
a 60% loss of BaP deposited on filter paper when it was exposed to N02-
Later research has focused on BaP and a number of other PAHs; on photo-
lysis and photooxidation, as well as on thermal reactions of PAHs with
63, NOX, and sulfur dioxide (802); and on the influence of the
physical and chemical nature of the substrate on the reactivity of
adsorbed PAHs. The corresponding literature is somewhat conflicting,
owing in part to the large number of characteristics that influence these
complex and still only partially understood heterogeneous reactions.
Thus, it is not surprising to note, even in the recent literature,
statements to the effect that PAHs are not chemically reactive and are
removed from the atmosphere by rain and sedimentation (e.g.,
Fishbein^O). AS discussed below, chemical reactions—including
photooxidation, reactions with S02 and NOX, and reactions with 03
and other oxidants—may, in fact, constitute major pathways for removal of
PAHs from the atmosphere. This discussion focuses on studies of the
reactions of PAHs deposited or adsorbed on a variety of substrates (e.g.,
soot, silica gel, alumina, and glass-fiber filters). The corresponding
literature concerning PAH chemistry in the bulk liquid phase (e.g.,
National Research Council^) is not included here, except for a few
studies directly relevant to the chemistry of adsorbed PAHs.
REACTION OF PAHs WITH OZONE
Kotin £t aJU*2 first reported on the reaction of pure BaP deposited
on a filter and exposed to various pollutants and mixtures of pollutants,
including 03, N02, and 03 plus N02. More recently. Lane and
Katz,57 Pitts et a_l.,74»75 and Katz et_ aj..45 -have reported on the
chemical half-lives of PAHs exposed to 03 and on the nature and
mutagenic activity of the products.
In experiments conducted with BaP, benzo[b]fluoranthene (BbF), and
benzofk]fluoranthene (BkF) exposed to OT (at 0.19-2.28 ppm) in air with
and without irradiation, Lane and Katz^' reported half-lives of about 40
min for BaP exposed to 03 at 0.19 ppm in the dark. For the three PAHs
studied, half-lives decreased with increasing 03 concentration and were
further reduced by irradiation with quartzline lamps (Table 3-5). Sub-
stantial differences in reactivity were observed, with BbF and BkF being
some 10 times more resistant than BaP to ozonolysis, both in the dark and
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under irradiation. Katz et al_.^^ extended this study to nine PAHs
deposited on cellulose thin-layer chromotography (TLC) plates and exposed
to 03 at 0.2 ppra in the dark, simulated sunlight, and both. The
corresponding results (Table 3-5) show significant ozonolysis of some PAHs
in the dark, with half-lives ranging from about 0.6 h for BaP and 1.2 h
for anthracene to 7.6 h for BeP. Pyrene (half-life, about 16 h), BkF (35
h), and BbF (53 h) were more resistant to dark ozonolysis. For seven of
the nine PAHs studied, half-lives were further reduced by irradiation.
Pitts ej^£l.7^»75 also determined a half-life of about 1 h for BaP
deposited on a glass-fiber filter and exposed to 03 at about 0.2 ppm.
The results, including the direct comparison between adsorbed and liquid-
phase data reported by Katz e_£ a_l. ,^ clearly demonstrate that the
reactivity of PAHs with 03 is much greater for PAHs deposited on solid
substrates than for PAHs in the bulk liquid phase.
Products of the reactions between BaP and 03 have been analyzed.
Katz e_t a_1.^5 identified the 1,6-, 3,6-, and 6,12-diones as major
products and noted that all three BaP diones had been identified by Pierce
and Katz'l in ambient Toronto air. Pitts e_t £l.^ also reported the
epoxide BaP 3,4-oxide as a reaction product. Van Vaeck et al. ^
identified a variety of reaction pathways and tentative structures of
oxygenated reaction products of the gas-phase ozonolysis of BaP as shown
in Figure 3-4. With respect to health implications, Katz et al.^5
stated that the BaP diones are direct-acting mutagens, but Pitts e_t
a_U'* found these products inactive in the Salmonella/microsome assay.
BaP 4,5-oxide, a DNA-binding metabolite of BaP, is a strong, direct-acting
REACTIONS OF PAHs WITH OXIDES OF NITROGEN
Kotin e_t a_1^.52 reported substantial (60%) loss of BaP deposited on a
filter and exposed to N02- The high activity of nitro derivatives of
PAHs—many of which are potent, direct-acting rautagens63,85—nag
prompted renewed interest in the possible formation of these compounds in
the atmosphere by reaction of adsorbed PAHs with coemitted NOX. Recent
studies discussed here include those of Jager, *• Gundel et al.,
Pitts e_£ al.,7* Hughes e_t al-i^7 Jager and Hanus,^2 Butler and
Crossley,^~~and Tokiwa e_t £l« With the exception of Butler and
Crossley,7 who used a mixture of nitric oxide (NO) and N02, all
studies have focused on N02« Hughes e_t^ a^.^' reported no reaction
between NO and PAHs adsorbed on coal fly ash, alumina, and silica gel. A
list of the 14 PAHs studied to date is given in Table 3-6. The molecular
structures of corresponding nitro-PAH products are shown in Figure 3-5.
In all product studies cited above, exposure of adsorbed PAHs to N©2
at parts-per-million concentrations resulted in the formation of nitro-PAH
derivatives. These are also listed in Table 3-6 and include mononitro as
well as dinitro derivatives, the latter identified as nitration products
of BaP^2 and pyrene.-*7 In the study of Jager and Hanus,^2 dinitro-
BaP was readily produced under conditions relevant to air pollution, i.e.,
by exposure of BaP adsorbed on fly ash to N02 at 1.33 ppm for 4 h at
20°C. Tokiwa et_ a^.^5 have reported extremely high mutagenic activity
3-9
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for dinitro-PAHs as direct mutagens in the Salmonella/microsome test
(e.g., 192 x IQ revertants/nmol for 1 ,6-dinitropyrene with strain TA 98
without metabolic activation). Mononitro-PAHs , although not as potent
rautagens as their dinitro homologues, also exhibit substantial activity as
direct mutagens.63'77'85 Two nitro-PAHs, 1-nitropyrene and
3-nitrof luoranthene, are carcinogenic in rats.°'
The effect of substrate on the nitration of adsorbed PAHs has recently
been investigated. Hughes ej: a_l. compared silica gel, alumina, and
coal fly ash and noted that the extent of nitration of pyrene depends on
the acidity of the substrate. Jager and Hanus^ discussed the effect of
PAH structure, substrate chemical and physical characteristics, N02
concentration, temperature, and exposure time on the yields of nitro
products of pyrene and BaP. For both PAHs, the yields of nitro products
were substrate-dependent according to the sequence silica gel > fly ash >
alumina > carbon (soot), with silica gel-to-carbon yield ratios as large
as a factor of 280 for nitropyrene (Table 3-7). As expected, nitration
yields increased as a function of NC>2 concentration and exposure time,
but not necessarily in a straightforward manner, owing to such complex
factors as the adsorption-desorption behavior of N02 on the substrate.
In view of the complex heterogeneous interactions involved, it is not
surprising to note large differences in the nitro-PAH yields reported by
several investigators. Tokiwa et _§_!• prepared nitro derivatives by
exposure of pyrene, phenanthrene, fluorene, chrysene, and f luoranthene
deposited on Toyo #2 paper filters to N02 for 24 h at 30°C in the
dark. Large yields were obtained with N02 at 10 ppm, but yields of only
a few percent were obtained at 1 ppm. These low yields are consistent
with those reported by Jager and Hanus^ for BaP and pyrene exposure to
NC>2 at 1.33 ppm for 4 h at 20°C with carbon as substrate. In con-
trast, Pitts £t a_l.75 reported 40% conversion of BaP deposited on
glass-fiber filters and exposed to NC>2 at 1 ppm for 8 h at ambient
temperature, and a yield of 18Z after exposure to N02 at 0.25 ppm under
the same conditions. The higher yields obtained on glass-fiber filters
may be due to a greater catalytic effect of the glass-fiber filter than of
soot (carbon) substrates.
In the same way, reported PAH half-lives due to reaction with N02
vary considerably with experimental conditions. From the above results,
one can derive a half-life of 10 h for BaP in the study of Pitts et_
al_. , 5 as opposed to half-lives of several days (or weeks) for several
PAHs, including BaP, as investigated by Tokiwa £t a_l.^ and Jager and
^ Butler and Crossley7 recently determined half-lives for 10
PAHs adsorbed on carbon (soot from a burner) and exposed to N(>2 at 10
ppm for up to 50 d. Their results, listed in Table 3-6, indicate PAH
half-lives ranging from 4-7 d for the more reactive PAHs (anthanthrene,
BaP, and benzofghi] perylene) to about a month for the least reactive
compounds (phenanthrene, fluoranthene, coronene, and chrysene). These
half-lives are consistent with those derived from the work of Ja'ger and
Hanus^^ and Tokiwa e_t al_. In view of the substrate used
(combustion-generated soot), the results of Butler and Crossley7 and
Jager and Hanus are probably applicable to heterogeneous nitration of
PAHs in the atmosphere.
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Given the high mutagenic potency of nitro-PAHs, it appears appropriate
to speculate on the fate of these compounds in ambient air. Four
nitro-PAHs have been reported ig urban particulate matter—6-nitro-
BaP,^ 3-nitrofluoranthene,*1'85 1-nitropyrene, and 5-nitroacenaph-
thene"—and indirect evidence of the presence of nitro-PAHs in Wayne
County, Michigan, air has been presented on the basis of mutagenicity
assays conducted with nitroreductase-deficient strains.^8,89 Photolysis
of nitro-PAHs, such as 9-nitroanthracene, yields the corresponding diones
(e.g., 9,10-anthraquinone), both in solution' and on silica gel.7' On
exposure of pyrene to N02, Jtfger and Hanus*2 noted the appearance of
new products after 4 h, and the nitropyrene yield decreased substan-
tially. However, the retention times of these as yet unidentified com-
pounds were different from those of the pyrene diones. The atmospheric
relevance of these and other pathways should be investigated further.
REACTIONS OF PAHa WITH SULFUR DIOXIDE
Ja*ger and Rakovic^ have reported the formation of sulfonic acids
and other sulfur-containing compounds on exposure of BaP and pyrene
(adsorbed on fly ash and on alumina) to 862 at a high concentration (10%
in air). These sulfonic acids are also easily prepared in the liquid
phase by reaction of PAHs with sulfuric acid at room temperature.88
Tebbens ejt ai. ^ observed significant consumption of BaP adsorbed on
soot from a propane burner and exposed to S02 at 50-80 ppra for 4 h, both
in the dark and under irradiation.
At lower S(>2 concentrations more relevant to ambient pollution, PAHs
do not appear to react readily with S02- Hughes e£ al_.^ observed no
reaction between SC>2 at parts-per-million concentrations and BaP or
pyrene adsorbed on silica gel, alumina, and coal fly ash. Butler and
Crossley' exposed 20 PAHs adsorbed on carbon (soot) to S(>2 at 5 ppm in
air for up to 100 d. The matrix air contained water vapor, and thus some
sulfuric acid was presumably present. Within the stated analytic preci-
sion, no significant reaction was observed for phenanthrene, coronene,
fluoranthene, chrysen«, BaP, pyrene, benz[a]anthracene, benzo(ghi)-
perylene, and anthanthrene. Because the 20 PAHs studied include both
highly reactive and essentially inert compounds with respect to reaction
with, for example, 03 and NC>2, the conclusions of Butler and
Crossley' as to the absence of significant reaction with SC>2 can
probably be extended to many of the PAHs present in polluted air.
However, 302 and sulfuric acid may play a role as catalysts for other
PAH reaction*, including nitration, and this possible catalytic role
should be investigated.
REACTIONS OF PAHs WITH OTHER OXIDIZING SPECIES
Reactions of adsorbed PAHs with atmospheric pollutants other than
03, §02, and N02 have received very little attention. Pitts e±
al. " exposed BaP deposited on a glass filter to peroxyacetylnitrate
3-11
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[CH3CO(00)N02, or PAN] at about 1 ppm for 16 h and observed the
formation of BaP diones and other oxidation products. Ambient concen-
trations of PAN in the Los Angeles atmosphere often reach 30 ppb during
episodes of photochemical smog, * so PAN may contribute, with 0^, to
the oxidative degradation of PAHs in photochemically polluted air.
Reactions of PAHs with free radicals, including the hydroxyl (OH) and
hydroperoxyl (H02) radicals (well documented in the bulk liquid phase),
have not been studied in the context of atmospheric pollution. On the
basis of studies conducted with aromatic compounds, such as toluene, the
OH photooxidation products in the presence of NOX include particulate-
phase hydroxynitrotoluene and dihydroxynitrotoluene as major pro-
ducts. 1 It is possible that atmospheric oxidation of PAHs initiated by
reaction with the OH radical results in the formation of nitro, hydroxy,
and hydroxynitro derivatives.
PHOTOCHEMICAL REACTIONS OF PAHs
The mechanisms involved in photochemical reactions of PAHs with
singlet oxygen in the bulk-liquid phase have received considerable atten-
tion and have been the object of several comprehensive reviews.24,66 ^
is illustrated in Figure 3-6 for anthracene, products of these reactions
include the PAH diones, formed either directly or by further reaction of
primary endoperoxide products, as well as other oxygenated compounds.
Photomodifications of BaP and other PAHs in the adsorbed state have
received significant attention with respect to both product distribu-
tion and influence of substrate. Product studies are in good agree-
ment, and the chemical distribution of PAH phototransformations in the
adsorbed state closely resembles that obtained in the bulk-liquid phase.
However, reactivity reportedly varies widely as a function of substrate,
and that makes it difficult to extrapolate laboratory studies to the
ambient atmosphere.
Falk, Markul, and Kotini? first reported on the photodecomposi-
tion of 10 PAHs, including BaP, deposited on Whatman #1 paper filters and
exposed to air in the dark, to air under irradiation, and to synthetic
smog prepared by the reaction of 0-j with gasoline. Their experiments
were conducted with pure PAHs, as well as with PAHs adsorbed on
gasoline-engine exhaust soot. On exposure of light in air, all PAHs
adsorbed on soot were more resistant to photodecomposition than the same
A1 R9
compounds in the pure form. Tebbens et_ *JL* 1,0* investigated photo-
transformations of BaP, perylene, pyrene, and fluoranthene adsorbed on
soot or deposited on paper, acetylated, and glass-fiber filters. Losses
of BaP of up to 40% were observed on irradiation for some 45 min in air;
the major reaction products were the three BaP diones and a carboxylic
acid derivative. Thomas e_t a^_.°^ reported similar results for BaP.
Phototransformations of BaP and other PAHs have also been observed on
a variety of substrates, including alumina, *>51 silica gel.^j^O.Sl
cellulose,39,45 acetylated cellulose, soil,'-' carbon micro-
3-12
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needles,3 atmospheric particulate matter,21 and coal fly ash.^O A
summary of the products of heterogeneous photooxidation of BaP on various
substrates is given in Figure 3-7.
For comparison, Table 3-8 lists half-lives and percent losses
determined for a number of PAHs deposited on cellulose TLC plates^ and
on Whatman #1 paper^' and adsorbed on soot*-' and on fly ash.^0
Although the four sets of data are directly comparable for only two
compounds, perylene and BaP, the effect of substrate on PAH phototrans-
formations is evident. PAHs deposited in the pure form on cellulose TLC
plates exhibit short half-lives, from 23 rain for anthracene to about 20 h
for BaP. On Whatman paper, pure PAHs appear more resistant to
photooxidation; e.g., the half-life for perylene is about 2 d, compared
with only 4 h on cellulose TLC plates. In the adsorbed state, PAHs appear
to be much more resistant to photooxidation, with losses of only 10% on
soot after 48 h of irradiation.^-' On fly ash, only modest photodecom-
position rates (up to 20%) are observed, in striking contrast with rapid
photooxidation in the liquid phase and on silica gel.^0 if one neglects
in a first approximation the important differences in experimental
conditions, it appears from the data listed in Table 3-8 that PAHs
adsorbed on atmospheric particles may be somewhat resistant to photo-
oxidation, with half-lives ranging from several days to several weeks,
depending on the reactivity of each compound.
Korfmacher e££^.^0 have discussed the possible physical and
chemical factors involved in the resistance of PAHs to photooxidation when
adsorbed on fly ash. Resistance to photooxidation on soot, although even
more relevant to urban pollution, where submicroraeter particles contain a
substantial fraction of carbonaceous material,^ has not been fully
investigated. In addition, specific PAH-substrate interactions have to be
considered. For example, Korfmacher et_ al_.^ and Kotin et al.^3
observed rapid decomposition in the dark of some PAHs adsorbed on fly ash
and on soot.
Until more data become available, caution must be exercised in
extrapolating laboratory results to PAH photooxidation in the atmosphere.
INTERACTIONS OF DEPOSITED PAHs WITH AMBIENT POLLUTANTS
It is somewhat surprising, in view of the critical need to obtain
overall PAH chemical deposition rates over a range of ambient condi-
tions, that only a few studies have investigated interactions of PAHs with
ambient polluted air. Pitts e_t £l/ exposed pure BaP deposited on a
glass filter to particle-free ambient Riverside, California, air for 3 d.
BaP was partially oxidized under these conditions, yielding BaP diones and
a variety of oxygenated (but not nitro) derivatives. In contrast, Fox and
Olive^ found only trace amounts of anthraquinone from anthracene (a
reactive PAH in the studies discussed above) adsorbed on ambient parti-
culate matter (suburban location near Austin, Texas) and exposed to atmos-
pheric gases for 4 d in the dark. Comparison of the results of these two
studies suggests that, as noted for photooxidation, PAHs appear more
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resistant to degradation in the adsorbed state than in the pure form.
Peters and Seifert^0 exposed glass-fiber filters impregnated with
l^C-labeled BaP to ambient air in Berlin, Germany, and noted substantial
losses of BaP, typically 75% over 24 h. Simultaneous determination of
14C activity (only 10% loss in 24 h) established that BaP losses were
due to chemical reaction, rather than to BaP evaporation from the filter.
In addition, a relationship was noted between BaP reaction rates and
ambient 03 concentrations.
A recent investigation of PAH concentrations in the plume of a
coal-fired power plant as a function of distance from the stack has been
reported by Kalkwarf and Garcia.^ Fluoranthene, BaP, pyrene, and BeP
in the plume were found to be 50% reacted 3, 6, 8, and 12 km from the
stack, respectively (with'correction for plume dilution and dispersion).
The loss of the PAHs was attributed to their reaction with coemitted N02
and S02.
PAHs IN AMBIENT AIR
Source identification is a key problem in the development of a
pollution abatement or control program. In 1973, Friedlander22 des-
cribed a technique to identify the sources of air pollutants in emission
inventories for particulate matter- Many chemical elements—such as
sodium, chlorine, silicon, and aluminum—are found in natural back-
ground aerosols of the atmospheres of urban and industrial basins, such as
Los Angeles. These are differentiated from other chemical tracers—such
as lead, vanadium, zinc, and barium—which are attributable to human
activities (see Figure 3-8). Thus, if some of the major sources are known
in a given area, the source contributors to the atmosphere can be iden-
tified and calculated by measuring the elemental concentrations at a given
point and fitting the data into a mathematical model. One of the major
problems in using this technique has been the need for instrumentation for
real-time measurement of the tracer elements. The use of trace metals for
identification of sources of particles was examined by Moyers et_ al. in
1977.65 with these tracers, several sources of particulate species in
desert, rural, and urban atmospheres could be determined.
In 1979, Daisey e_t_ a_l_.-^ described three methods for source identi-
fication for the PAHs in the complex mix of the atmosphere. Although the
evaluations of these methods are in the early stages, it was found that
statistical modeling does not depend on source emission data, if the
ambient-air measurement data base is large. In 1981, Daisey and Kneip^
reported that it was possible to use multivariate regression models of
ambient-air data for apportioning the contributions of emission sources to
airborne particulate organic matter. The contributing sources of respir-
able particles were determined by analysis of the ambient-air measurement
data taken in New York City: 19% were from automobiles and related
sources, 40% were from oil-burning, and 15% were soil-like particles.
Although this study using tracer chemicals had good results, the methods
should be validated for predictive use by testing in other locations.
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A comprehensive discussion and critique of environmental sampling and
analytic methods used for polycyclic organic matter are in the EPA
report.7^ Lee et al., in a book on the analytic chemistry of PAHs,
discussed sampling of mobile and stationary sources, ambient air, water,
food, soils, and the aquatic environment. The cleanup and separation pro-
cesses for the various collection media include solvent partitioning for
analysis by column, paper; thin-layer, gas, and high-pressure liquid
chromatography. The percentages of recovery with the analytic methods for
the various PAHs were described by Lee e£ al., but are not discussed here.
In 1967, Hangebrauck £t al.,3^ in a review of known sources of PAHs,
gave the results of a survey made by the National Center for Air Pollution
Control to screen thevprocesses likely to produce emission in urban air.
Although the survey was not intended to establish statistically the
average emission from the sources, the data from it do characterize and
classify the rate of emission of several PAHs from four major source
categories: heat generation, refuse-burning-, industrial processes, and
motor vehicles. The 1972 NRC report Particulate Polycyclic Organic
Matter"" estimated that heat and power generation produced BaP at 500
metric tons/yr; refuse-burning, at 600 tons/yr; coke production, at 200
tons/yr; and motor vehicles, at 20 tons/yr. BaP has often been used as a
surrogate in estimating source contributions of complex mixtures of PAHs.
Surrogate chemicals have been used commonly in monitoring environ-
mental quality, for various reasons—e.g., analytic methods are often
available only for the surrogate, and it costs less to monitor only one
chemical. However, a PAH surrogate may not be useful unless studies have
been conducted to characterize the PAH profile and percentage relation-
ships for each type of environment. For example, in 1979. Bjorseth e_t
al. showed that the relative distribution of PAHs is not the same in
all environments. Figure 3-9 shows comparisons of the percentages of PAHs
found in the particulate matter from an aluminum plant and a Soderberg
paste plant. He recommended that a parent PAH profile (PPP) be estab-
lished before a surrogate compound was chosen. In 1981, Gammage and
Bjorseth^ stated that there are no established techniques for real-time
monitoring of selected PAHs and that BaP is not a universally accepted
proxy or surrogate for PAHs. It is known that the numerous PAHs found in
the outdoor air can be radically different, qualitatively and quanti-
tatively, from those in the workplace environment and that monitoring one
compound as a surrogate for others is unreliable. The recommendation was
made again that a PPP be determined before a proxy or surrogate compound
was chosen.
Considerable information is available to show the profile of the
various PAHs from various sources. Comparison of the rate of PAH emission
between different categories is complicated, owing to the different units
used to measure them: ug/BTU, ug/barrel of oil, Vg/g of particles, ug/lb
of material processed, ug/m3, ug/n»i. Qualitative comparisons of PAH
emission sources do appear feasible and can identify the various contri-
butors to the overall pollution burden. Lists of the PAHs found in the
3-15
-------
following five broad categories have characterized the variety of sources
and identified some of the major contributors: heat and power generation,
refuse-burning, industrial processes, motor vehicles, and natural sources,
In the 1978 review of the sources of PAHs, Baum, using data
assembled by EPA in 1974, estimated that 97% of the BaP emitted in the
United States could be attributed to stationary fuel combustion. The
major contributors were the inefficient combustion of coal in residential
furnaces, coke ovens, and refuse fires. This is in close agreement with
the NRG report Particulate Polycyclic Organic Matter, which stated that
90Z of the annual nationwide BaP emission was attributable to coal- and
wood-fired residential furnaces, coal-refuse fires, and coke production.
A wide range in concentrations of BaP (0.1-388 ng/rn^) has been
reported by Colucci and Begeman^ for U.S. and foreign cities (see Table
3-9). These results are for measurements taken between 1952 and 1966. The
objective of the authors was to study BaP concentrations in the atmosphere
in relation to automobile traffic. They used tracers to identify auto-
motive and nonautomotive sources and calculated correlation coefficients
of BaP with CO (a motor-vehicle tracer) as 0.65, with lead (a gasoline-
vehicle tracer) as 0.74, and with vanadium (an oil tracer) as 0.54.*
The seasonal variations show that the concentrations were highest in fall
and winter and lowest in spring and summer. The winter vanadium concen-
trations were twice the summer concentrations; that indicates that the
higher amount of BaP was attributable to combustion of residual fuels used
for heating or to the lower inversion heights prevalent during cold
weather.
The concentrations of BaP in England, Italy, Norway, Sweden, and
Germany, as shown in Table 3-10 for 1953-1964, were given in the report by
Louw,°0 for the purpose of comparing the findings in South Africa. The
concentrations of BaP ranged between 5 and 49 ng/m . One sample, taken
near a road-tarring operation, was extremely high, 1,113 ng/m .
In Ontario, Canada, five locations were sampled for BaP by Katz e_t
a_l. from April 1975 through March 1976. The highest concentration was
observed in Hamilton (3,498 ng/nr), and the lowest in Sudbury (111
ng/m ); the latter was attributed to the electrostatic precipitators in
use at the nickel-copper smelter 5 mi away. The concentrations are given
in Table 3-11.
A study was conducted in Karlsruhe, Germany, to determine the relative
amounts of BaP from residential heating systems and automobile traffic
(see Figure 3-10). The concentrations ranged from 0.1 ng/m^ (at the
low-traffic Municipal Garden) to 28 ng/m^ (at a railroad underpass).
With lead as the tracer, it was determined that the highest concentration
caused by automobiles was in the underpass. The low concentration in the
The correlations indicate that 42-55% of the variation in BaP concen-
trations is related to motor-vehicle tracers and 292 to stationary
oil-burning.
3-16
-------
Municipal Garden during summer was attributed to deposition of airborne
particles on leaves, trees, and shrubs. During the winter, the increase
in the concentration of BaP was attributed to increased residential heat-
ing. The air samples taken at Karlsruhe Nuclear Research Center, 11 km
(by air) north of the city, had the lowest concentrations, except for
those in the Garden during May and June.
BaP was determined in four locations around the industrial city of
Essen, Germany, by Grimmer £t a±. ,28 from October 1978 to March 1979.
There were four sampling sites at each of five locations. The authors
concluded that the concentrations of BaP in ambient air varied by a factor
of more than 10 from one station to another during the cold-weather heat-
ing period. Thus, they did not give any average values; the approximate
ranges of concentration at each location were as follows: 1-75 ng/ra in
an area that used hand-stoked coal-heating in residences, 1.5-21 ng/nr
in an area with oil-heating only, 10-100 ng/nH in a tunnel with car
traffic, 15-210 ng/nr in an area with coke ovens, and 1-75 ng/ra^ in an
area described as rural, outside the city.
Two very thorough studies of the PAH content of Los Angeles air have
been made by Gordon and Bryan^° and Gordon. ^ The earlier study was
of four locations in the Los Angeles basin (see Figure 3-11), and the
latter included 13 sampling locations (see Figure 3-12). Analyses were
performed for 14 PAHs, including BaP, sampled over the course of a year.
From the relationship between meteorology, traffic density, and PAH
concentrations, the authors concluded that most (at least 60%) of the PAHs
was contributed by automobile traffic, but that the concentrations were
lower than in many other cities. This result was expected, because of the
extensive use of natural gas and hydrothennal energy in the West and the
nonuse of coal in Los Angeles. The warm climate also limits wood-burning
in fireplaces. The Colucci and Begeman^ results demonstrated much
higher ambient BaP concentrations in urban areas that depend extensively
on coal, oil, and wood combustion. They determined that the automotive
contribution to Detroit ambient BaP was only 5-42%, with typical BaP
concentrations 3 times as high as in Los Angeles.
3-17
-------
TABLE 3-1
The Effect of Adsorption on Carrier Particles on the Distribution of PAHs
between Gaseous and Particulate Phases
Equilibrium Gas
Measured Ambient Air Concen-
traction, ng
PAH (Number of Aromatic Rings)
Naphthalene (2)
Anthracene (3)
Phenanthrene (3)
Benz( ajanthracene (4)
^ Pyrene (4)
i
f~~*
00 Benzo[k ]f luoranthene
Benzo[a]pyrene (5)
Benzole Jpyrene (5)
Perylene (5)
Benzolghi Jpery lene (6)
Coronene (7)
Vapor Pressure Phase Concentration, Gas Particulate P/G
at 25°C, torra at 25°C, ng/m3b Phased Phase Ratio
8.2 x 10~2 — All None 0
2 x 10~4
44.7 1.21 0.03
6.8 x 10 ~4
1.1 x 10~7 1.03 x 103 3.87e 12. 2e 3.15*
6.8 x 10~7 74 x 103 3.36 1.64 0.49
16 2.0f 23. lf 11. 5f
5.5 x 10~9 85
5.5 x 10~9 85 2.69 20.1 7.5
—
1.0 x 10~10 1.6
1.5 x 10~12 0.024 None8 A118
aData from Santodonato et al.78
bData from Pupp ^t al ,^~
cData from Cautreels and Van Cauwenberghe.8 Samples collected in Antwerp, Belgium, Oct. 1976; average
temperature, 10.2°C.
^Includes particulate PAH loss from filter at sampling linear airflow rate of 0.43 m/s.
eSum of benz[«? Janthracene and chryaene.
^ Sum of benzo I a ) f luoranthene and benzo [ b ] f 1 uor an thent" .
KD^tla from Mi.a
-------
U)
I
TABLE 3-2
Size Distribution of Benzola]pyrene and Coronene in
Pasadena, California, in Ambient Aira
Concentration,
Aerodynamic
diameter, p m
0.05-0.075
0.075-0.12
0.12-0.26
0.26-0.50
0.5-1.0
1.0-2.0
2.0-4.0
>4
Oct.
BaP
30
196
79
28
36
16
16
21
25-28, 1976
Cor
431
1,390
460
205
166
107
118
136
Dec.
BaP
83
523
208
67
81
44
30
23
14-17, 1976
Cor
1,200
4,220
907
113
218
<45
<45
<45
Feb.
BaP
72
431
238
90
83
39
32
28
1-4, 1977
Cor
800
2,850
920
307
227
93
93
80
Mar.
BaP
69
139
65
28
26
16
14
12
1-6, 1977
Cor
387
593
206
65
70
50
45
45
aData from Miguel and Friedlander.
64
-------
TABLE 3-3a
PAH fluxes (in ng* cm~2- yr~M to sediments from five remote sites in the northeastern United States for
three age intervals: present, approximately 1950, and 1900; and to sediments from three urban sites for 1940 to
the present.
Remote Sites
Lake Superior,
0.02 cm/yr
Isle Royale,
0.09 cm/yr
Somes Sound,
0. 1 cm/yr
Had lock Lower Pond,
0.07 cm/yr
Coburn Mtn. Pond,
0.3 cm/yr
Average
Interval
1955-now
1930-1955
1870-1920
1974-now
1951-1955
1960-now
1940-1960
1880-1940
1950-now
1920-1950
1975-now
1943-1947
1898-1901
present
1950
1900
In Situ
Density
0.55
0.55
0.55
0.33
0.32
0.43
0.52
0.51
0.12
0.11
0.036
0.057
0.057
—
—
— —
GI Chry.+
Phen.
0.3
0.2
0.06
0.4
1
2
4
0.4
2
0.3
2
8
0.8
I
3
0.4
Anth.
0.03
0.02
0.004
0.01
0.05
0.2
0.4
<0.02
0.1
0.04
0.2
0.5
0.07
0.1
0.2
0.03
Phen.
0.3
0.2
0.08
0.5
2.5
1
4
<0.2
2
—
4
12
1
1.5
4.5
0.4
Fluo.
1
0.9
0.2
0.4
1
5
8
0.4
4
0.3
4
11
0.5
3
4
0.4
Pyr.
0.6
0.5
O.I
0.3
0.7
4
5
0.4
3
0.2
3
8
0.3
2
3
0.3
B[a]A
0.3
0.3
0.07
0.2
0.2
2
3
< 0.05
0.6
0.04
0.7
3
0.2
0.8
1.5
0.1
Tri.
I
1
0.3
0.8
0.7
2
4
< 0.2
1
0.1
2
6
0.9
1.5
2.5
0.2
B[e]P
0.9
0.9
0.2
0.8
0.9
2
2
0.2
0.8
0.1
2
7
0.4
1.5
3
0.5
B[a]P
0.4
0.3
0.07
0.2
0.2
2
2
0.2
0.6
0.06
0.7
4
0.1
0.8
1.5
0.1
-------
Table 3-3 (continued)
In Situ
Urban Sites Interval Density
Boston Outer Harbor, 1900-now 0.93
0.1 cm/yr
Buzzards Bay, Mass., 1940- now 0.3
0.3 cm/yr
Pettaquamscutt 1940- now 0.16
River, 0.3 cm/yr
Average —
Phen. Anth
24 5.6
18 2
46 5
30 4
*~* aReprinted with permission from Gschwend and Kites.32
^Includes all €20^12 iCornera except perylene.
Phen. Fluo. Pyr.
17 37 39
Chry.+
B[a]A Tri. B[e]P B[a]P
36
25
53
93
55
48
93
55
19
37
42
30
14
23
37
42
35 ~ 25
17
140b
130b
/"30
-------
OJ
I
ho
ho
TABLE 3-4
Concentration of Polycyclic Aromatic Hydrocarbons in Norway Aerosols3
PAH
Phenanthrene
Anthracene
Me thy Iphenanthrene/
anthracene
Fluoranthene
Dihydrobenzo [ a&b ] -
f luorenes
Pyrene
Benzofa] fluorene
Benzo [ b J fluorene
l-Methylpyrene
Benzofc Iphenanthrene
Benz [ajanthracene
Chrysene/triphenylene
Benzo[b&k] f luoranthene
Benzo [e] pyrene
Benzol a ] pyrene
Perylene
Ideno [ 1 , 2 , 3-cd ] pyrene
Benzofghi ] perylene
Anthanthrene
Coronene
Total
Concentration
England,
France ,
Feb. 20-21,
1976
*)4.725
3
0.661
6.637
0.874
4.864
0.815
0.571
0.147
1.021
0.585
1.756
4.312
1.191
0.965
0.090
1.306
1.142
0.225
0.212
32.099
in Aerosol in Air
Northern Eng. ,
Scotland,
Nov. 25-26,
1975
1.216
0.278
0.216
3.965
0.363
3.293
0.318
0.149
0.099
0.957
0.740
3.269
4.013
2.635
2.053
0.191
1.920
1.971
0.423
0.183
28.252
from:
Northern
Norway,
Jan. 25-27,
1976
0.036
0.038
—
0.171
0.032
0.135
0.021
0.117
0.009
0.038
0.041
0.099
0.083
0.066
0.059
Trace
0.062
0.064
0.007
— —
1.078
Stationary air,
Southern Norway,
Feb. 1, 1976
0.146
—
0.052
0.324
0.032
0.286
0.026
0.148
0.009
0.108
0.073
0.194
0.464
0.135
0.098
0.011
0.144
0.140
0.022
0.020
2.432
aData from Lunde and Bjorseth"* and Bjorseth et
-------
TABLE 3-5
Heterogeneous Photo-oxidation and Ozonolysis Half-Lives of PAHs
on TLC Plates3
Half-Life, h
PAH
Anthracene
Benz [ a ] anthracene
Dibenz [ah ] anthracene
Dibenz [ ac ] anthracene
Pyrene
Benzofa] pyrene
Benzole] pyrene
Benzo[b] fluoranthene
Benzo[k] fluoranthene
Ozonolysis
in Dark
(0, - 0.2 ppm)
1.23
2.88
2.71
3.82
15.72
0.62
0.4b
0.3C
7.6
52.7
10. 8b
2.9<=
34.9
13. 8b
3.3C
Photo-oxidation
(quartz-lamp
irradiation in
air)
0.2
4.2
9.6
9.2
4.2
5.3
—
—
21.1
8.7
—
—
14.1
—
—
Photo-oxidation
and
Ozonolysis
0.15
1.35
4.8
4.6
2.75
0.58
0.2b
0.08C
5.38
4.2
3.6*
1.9C
3.9
3.1*
0.9C
aData from Lane and Katz57 and Katz
b03 »0.7 ppm.
- 2.3 ppm.
3-23
-------
TABLE 3-6
Reaction of Adsorbed PAHs with Nitrogen Dioxide
PAH
Phenanthrene
Anthracene
Fluoranthene
Chrysene
Pyrene
PAH-N02
Reaction
Half-life, a
d
30
—
27
26
14
Nitro
Derivatives
Identified
Mononitro,
isomer not
specif iedb
9-Nitroc?d
3-Nitroband
8-nitrob
6-Nitrob'c
l-Nitrob«c>e
Nitro
Derivative
Yield
Measured
b
—
b
b
b,c,e
Effect of
Substrate
Investigated
c
—
c
c,e
Benzo[a]pyrene 7
Benzo[e]pyrene 24
Perylene
Benz[a]anthracene 11
Benzofghilperylene 8
Anthanthrene 3.7
Fluorene
Coronene 29
Carbazole
and dinitro6
6-Nitro,c'f
1-nitro,
3-nitro,^
and dinitro0
3-Nitrod«f
c,f
2-Nitrob
Two
unspecified
isomersb
aData from Butler and Crossley.7 dData from Gundel £t al.
bData from Tokiwa e^ al.8^
cData from Jager and Hanus
^
eData from Hughes e£ al
^Data from Pitts e_t al.
37
3-24
-------
TABLE 3-7
Yields of Nitro PAHs as Function of Substrate3
Nitro-PAH yield, ug/100 ugof PAH
1-Nitropyrene
Substrate
Carbon
Alumina, deactivated
Alumina, activated
Fly ash
Silica gel
Dark
0.45
2.4
1.9
36.8
—
UV Light
0.38
2.3
2.0
41.8
—
Daylight
0.40
2.6
2.8
57.3
112.5
6-Nitro-BaP,
Daylight
Trace
7.9
8.2
15.8
25.5
aData from J'ager and Hanus.*2
3-25
-------
TABLE 3-8
Influence of Substrate on Photo-Oxidation of PAH
Half-life, h, I Destruction in 48 hb
PAH
Anthracene
Benz [a] anthracene
Dibenz [ab] anthracene
Dibenz [ ac ] anthracene
Pyrene
Benzo[a] pyrene
Benzofe] pyrene
Benzo(b] fluoranthene
Benzo[k] fluoranthene
Anthanthrene
Phenanthrene
Fluoranthene
Benzo[ghi] perylene
Coronene
Chrysene
Pure PAH on
Cellulose
TLC Plate3
0.2
4.2
9.6
9.2
4.2
5.3
21.1
8.7
14.1
—
—
—
—
—
—
Pure PAH Loss, I,
on Whatman Adsorbed Adsorbed on
Paper on Soot Fly Ashc
17-26
—
—
—
42 1 5-13
22 10 9-17
7
—
—
44 5
60 -- 0
24 40
0 0 —
0
0
aData from Katz e_t al..45
bData from Falk £t al_.17
cData from Korfmacher e_t al.50 Different light sources (xenon,
quartz, etc.). Times up to 100 h.
3-26
-------
TABLE 3-9
Benzo[a]pyrene Content of Urban Air3
Benzo[a]pyrene Content, ng/m^
Location
New York:
Commercial
Freeway
Residential
Detroit:
Commercial
Freeway
Residential
Atlanta
Birmingham
Detroit
Los Angeles
Nashville
New Orleans
Philadelphia
Pittsburgh
San Francisco
Hamburg, Germany
London, England
Sheffield, England
Cannock, England
London:
Traffic
Background
Milan, Italy
Copenhagen, Denmark
Prague, Czechoslovakia
Budapest, Hungary
South Africa:
Pretoria
Johannesburg
Durban
Osaka, Japan
Commercial
Residential
Sidney, Australia
Spring
0.5-8.1
0.1-0.8
0.1-0.6
7.2
—
~—
2.1-3.6
6.3-18
3.4-12
0.4-0.8
2.1-9.0
2.6-5.6
2.5-3.4
—
0.8-0.9
14.72
25-48
20-44
4-16
20
11
12
6
—
—
—
—
— —
5.7
3.3
0.6-2.4
Summer
0.7-3.9
0.1-0.7
0.1-0.3
—
4.0-6.0
0.2
1.6-4.0
6.1-10
4.1-6.0
0.4-1.2
1.4-6.6
2.0-4.1
3.5-19
0-23
0.2-1.1
10-26
12-21
21-33
6-11
11
1
3
5
13-36
17-32
10
—
— —
1.7
1.4
0.6-1.8
Fall
1.5-6.0
3.3-3.5
0.6-0.8
—
3.4-7.3
— ~
12-15
20-74
18-20
1.2-13
30-55
3.6-3.9
7.1-12
2.9-37
3.0-7.5
66-296
44-122
56-63
27-31
57
38
25
14
—
—
—
—
• ~
9.4
3.8
2.5-7.4
Winter
0.5-9.4
0.7-1.3
0.5-0.7
5.0-17.0
9.2-13.7
0.9-1.8
2.1-9.9
23-34
16-31
1.1-6.6
25
2.6-6.0
6.4-8.8
8.2
1.3-2.4
94-388
95-147
64-78
27-32
68
42
150
15
53-145
72-141
22-28
22-49
5-28
14
6.7
3.8-8.2
aReprinted with permission from Colucci and Begeman;10 copyright
1971 American Chemical Society.
3-27
-------
TABLE 3-10
Benzo[a]pyrene Concentrations in Pretoria,
Johannesburg, Durban, and Other Large Cities3
Benzo(a]pyrene Concentration, ng/w
Country
S. Afr.
England
Italy
Norway
Sweden
Germany
City
Pretoria
Johannes-
burg
Durban
Merseyside
and other
northern
localities
Salford
Sheffield
Cannock
Milan
Oslo
Stockholm
Hamburg
Period of Sampling
26-27 Aug. 1963b
23-24 Sept. 1963b
20-21 Jan. 1964C
27-28 Apr. 1964b
4-5 May 1964b
18-19 May 1964b
12 May 1964b«d
10-11 June 1964b
11-12 June 1964b
16-17 June 1964b
(i) Jan. -Dec. 1958
1954-1957b
(ii) 1954-1957b
Nov. 1952-Mar. 1953
June 1949-Apr. 1950
July 1949-June 1950
Jan. -Oct. 1958
Feb. -Dec. 1955
Mar- -July 1960
Sept. 1961-Apr. 1963
Uncorrected
67
83
31
65
141
146
3,340
42
16
83
11
17
6
197
20
4
3
0.86
1.1
10.1
Corrected for ~~
Benzofk] fluorantheng
22
28
10
22
47
49
1,113
14
5
28
108
166
37
290
28
32
231
15.2
10.0
388
aAdapted from Louw.60
bWinter.
cSummer; determined by direct chromatography of the cyclohexane-soluble
fraction of sample on thin-layered alumina.
"Road-tarring operation.
3-28
-------
ro
vo
Table 3-11
Seasonal Concentrations of Benzo[a]pyrene in Air of Ontario Cities
April 1975-March 1976a
Apr.-Jun. 1975 Jul.-Sept. 1975 Oct.-Dec. 1975 Jan.-Mar. 1976
Location
South Sarnia
Hamilton
Toronto (Kennedy
ng/mj
338
1,404
657
Ug/gb
5.5
9.6
8.7
ng/m3
114
2,351
408
ug/gb
2.4
16.9
6.2
ng/m3
596
3,498
729
ug/gb
11.4
50.6
11.7
ng/m3
190
1,934
814
Ug/gb
7.0
23.1
5.9
at Lawrence,
suburban)
Toronto (Bathur 789 8.8 1,047 11.0 1,674 20.2 720 9.2
Street at 401)
Sudbury (5 mi 175 5.4 111 2.6 342 15.3 444 19.0
from nickel-
copper smelter)
aAdapted from Katz et_ al.46
yg of PAH per gram of total particles—same as parts per million.
-------
III 11
\ 69
»1
Kerosene Soot
30 I
10
Air Particulates
Charles River Sediment
FIGURE 3-1. Gas chrooaCograms of PAH mixtures obtained from (top) soot
from kerosene flame, (middle) urban air particles, and (bottom) sediment
of Charles River in Boston.
3-30
-------
Peak Identifications:
2 Biphenyl
4 Acenaphthylene
8 Fluorene
10 C14H8
14 Phenanthrene
15 Anthracene
18 Methylphenanthrene
19 4H-Cyclopenta[def]-
phenanthrene
22 Fluoranthene
23 Benz(e]acenaphthylene
25 Pyrene
27 Methylfluoranthene
30 Benzofghi]fluoranthene
31 G (unknown)
32 Cyclopenta[cd]pyrene
33 Benz[a]anthracene
34 Chrysene
35 Methylchrysene
37 Benzofluoranthene
38 Benzo[eIpyrene
39 Benzo(a]pyrene
40 Perylene
42 £21^12 (unknown)
43 ^21^12 (unknown)
44 Indenof1,2,3-cdIpyrene
46 Dibenz[acJanthracene
47 Benzo[ghijperylene
48 Anthanthrene
3-31
-------
- o
V
5
x
<
o.
S 10"
•o
10 20 90 40 SO GO 70 BO 90 100
OtST FROM BOSTON llml
FIGURE 3-2. Total PAH concentrations vs. distance from Boston
for Massachusetts Bay samples. Reprinted with permission from
Windsor and Hites; copyright 1979, Pergamon Press, Ltd.
14
_ 13
I12
0 "
5 10
= 9
p 8
*
| 7
I 6
s
e 5
Z 4
JJ
_ 3
o 2
^
1
; /v
• >
i i i
/ \
\
\ » -
_ «^ \ ;
U V -
/•'
r1:
I i
/ i'
/<
H — r— ^ — '—-r— T-1 i . i . i . i .
2.4
2.2 -
2.0
1.8'
1.6 '
1.4
1.2 i
1.0
0.8
o
O.6 *
0.4*
0.2 •
IB?0
I6CO 1830 i«OO I9ZO 1940 i960 1980
tcet e>
FIGURE 3-3. Total PAH abundance in the various Pettaquamscutt
River sediment core sections vs. date of deposition (solid line,
left scale); BaP abundance in Gosser Ploner Sea vs. date of
deposition (dotted line, right scale). Reprinted with permission
from Hites et al.;^ copyright 1980, Pergamon Press, Ltd.
3-32
-------
COOM
OII-M
0.5-4 k
FIGURE 3-4. Major reaction pathways and tentative structure of
products of gas-phase ozonolysis of BaP. Most structures given
as examples of possible isomers. Reprinted with permission from
Van Vaeck ^t £l.;86 copyright 1980, John Wiley & Sons Ltd.
3-33
-------
Nitrocarbazole (2 unspecified isomers)
9-Nitroanthrscene
Nitrophenanthrene (iaomer not specifi«
1-Nitropyrene (also dinitropyrene,
iaomer not specified)
(B)
3-Nitrofluoranthene (A) and
8-nitrofluoranthene (B)
6-Nitrochrysene
FIGURE 3-5. Nitro products identified in heterogeneous reactions
of PAH with nitrogen dioxide.
3-34
-------
3-Nitroperylene
NO,(A)
N02 (B)
l-Nitrobenzoja]pyrene (A),
3-nitrobenzo[a]pyrene (B),
and 6-nitrobenzo[a]pyrene
(C)--also dinitro-BaP,
isomer not specified
2-Nitrofluorene
FIGURE 3-5. (continued)
3-35
-------
9,10-Endoperoxide
(yield, 2-7Z)
9,10-Anthraquinone
(12-192)
OH
l-Hydroxy-9,10-
anthraquinone
(2-5Z)
Dione dimer (3-13Z)
FIGURE 3-6. Product* of heterogeneous photooxidation of anthracene
on atmospheric particulate matter. Reprinted with permission from
M. A. Fox and S. Olive, Science 205:582-583, 1979;2* copyright 1979
by the American Association for the Advancement of Science.
3-36
-------
1,6-Dione
3,6-Dione
6,12-Dione
7H-Benz gle] -
anthracene-7-one
3,4-dicarboxylic
acid
Anhydride
FIGURE.3-7. Products identified in heterogeneous photooxidation of
BaP. Reprinted with permission from Tebbens et al.
81
3-37
-------
- HCI
HBr
GAS PHASE
CHEMICAL ELEMENTS IN POLLUTED ATMOSPHERES
FIGURE 3-8. Inner ring encloses elements present
in natural background (soil dust and marine aerosol);
second ring, primary particulate matter introduced
by man; outermost ring, secondary material formed in
atmosphere. Elemental carbon added to second ring.
Reprinted with permission from Friedlander;^ copy-
right 1973 American Chemical Society.
3-38
-------
•/.of Totol
100
FIGURE 3-9. Parent PAH profile of PAH in particulate matter from
aluminum plant (dashed line) and Soderberg paste plant (solid line).
Soderberg paste plant is one in which electrodes used in production
of aluminum are made of anthracite or of anthracite and petroleum
coke. During baking of these electrodes, volatile components are
produced from anthracite ore base. Reprinted with permission from
Bjorseth;" copyright Ann Arbor Science Publishers, Inc.
3-39
-------
0,1 0.5 I 2 5 10 20 30 SO 70 80 90 95 9« 99 99,5 99,9
Probability V.
FIGURE 3-10. Frequency distribution of airborne BaP concentra-
tion at different measuring sites in area of Karlsruhe. 1, nuclear
research center, Karlsruhe, November 1974-March 1975; 2, municipal
garden, May-June 1975; 3, railroad underpass, May-June 1975; 4, muni-
cipal garden, October 1975-March 1976; 5, railroad underpass, October
1975-March 1976.
3-40
-------
Location
Component
Total participate
muc, MO/1"1
Benzene solubles.
Load,
Traffic density
x 10" 3. vehicle
mi/mi*/day
PAH,ng/m3
Coronene
Pyrene
Fluoranthene
Benz(a)-
anthracene
Chrysana
Banzo(a) pyrana
Benzo(«) pyrana
Benzo<6)>
floor an trtene
BenzoU)-
fluoranthana
Ban20-
fluoranthana
Anthanthrana
Baflzo(0n/)-
parytana
indano(1.2.3-orf)-
pyrana
215
21.7
6.35
200
6.4
2.0
1.9
1.1
2.6
3.0
1.1
1.6
0.6
0.8
0.5
0.4
9.2
1.2
2
131
13.2
2.50
130
3.2
1.4
0.8
0.8
1.6
1.8
0.5
0.9
0.3
0.3
0.3
0.2
4.2
0.4
3
102
8.3
1.97
95
2.8
3.6
3.4
3.1
3.8
3.2
3.5
1.8
0.8
1.3
1.2
1.1
7.1
0.3
40
2.6
0.50
0.20
0.18
0.12
0,04
0.04
0.09
0.03
0.09
0.01
0.03
0.01
0.01
0.21
0.03
FIGURE 3-11. Components in Los Angeles airborne particles.
Composite June 1971-June 1972. Map shows approximate location
of sampling sites. Reprinted with permission from Gordon and
Bryan;26 copyright 1973 American Chemical Society.
3-41
-------
Area
PAH
PYR
FLT
BAA
CHY
SEP
BAP
BJF
BKF
ANT
GEE
IMP
COR
1
0.41
0.36
0.28
0.23
0.18
0.13
0.62
0.44
0.81
0.75
0.47
0.32
0.17
0.13
0.16
0.14
0.25
0.18
2.86
2.73
1.10
1.09
1.83
2
0.46
0.50
0.32
0.32
0.18
0.18
0.68
0.60
1.01
1.04
0.63
0.45
0.25
0.19
0.26
0.19
0.35
0.25
3.78
3.79
1.22
1.51
2.54
3
0.37
0.40
0.20
0.26
0.15
0.15
0.36
0.49
0.73
0.85
0.41
0.36
0.17
0.15
0.16
0.15
0.28
0.21
3.02
3.08
1.11
1.22
2.06
4
0.49
0.60
0.33
0.39
0.26
0.22
0.65
0.72
1.06
1.26
0.56
0.54
0.18
0.22
0.21
0.23
0.29
0.31
4.33
4.57
1.89
1.82
3.06
5
0.47
0.48
0.30
0.31
0.21
0.18
0.76
0.59
1.00
1.02
0.54
0.44
0.23
0.18
0.23
0.18
6.26
0.25
3.84
3.69
1.55
1.47
2.47
6
0.76
0.61
0.50
0.40
0.44
0.23
0.92
0.75
1.34
1.30
0.77
0.56
0.28
0.23
0.27
0.23
0.42
0.32
5.07
4.72
2.05
1.88
3.16
7
0.84
0.46
0.61
0.30
0.23
0.17
7.02
0.57
1.22
0.98
0.76
0.42
0.26
0.17
0.29
0.18
0.38
0.24
4.02
3.56
1.96
1.41
2.38
8
0.67
0.33
0.55
0.21
0.24
0.12
0.68
0.40
0.88
0.69
0.53
0.30
0.14
0.12
0.25
0.13
0.25
0.17
2.67
2.52
1.18
1.00
1.69
9
0.34
0.39
0.25
0.25
0.10
0.15
0.53
0.48
0.92
0.83
0.41
0.36
0.14
0.15
0.14
0.15
0.18
0.20
2.99
3.02
1.33
1.20
2.02
10
0.34
0.38
0.27
0.25
0.12
0.14
0.42
0.46
0.77
0.80
0.35
0.35
0.14
0.14
0.16
0.14
0.15
0.20
3.05
2.91
1.18
1.16
1.95
11
0.33
0.30
0.21
0.19
0.11
0.11
0.38
0.36
0.65
0.63
0.24
0.27
0.10
0.11
0.11
0.11
0.13
0.15
2.31
2.29
0.90
0.91
1.53
12
0.42
0.44'
0.30
0.28
0.17
0.16
0.66
0.53
0.89
0.92
0.38
0.40
0.15
0.16
0.19
0.17
0.17
0.22
3.41
3.35
1.48
1.33
2.24
1
0..
0..
0..
0..
o.:
o.:
O.t
o.:
o.;
o.e
O.i
o.;
0.1
0.1
0.1
0.1
0.1
0.1
2.2
2.3
l.C
o.s
1.5
Upper value in each pair - observed; lower value calculated using the average PAH/COR ratio for areas 3,11, and 13.
Italicized observed values exceed calculated values by at least three times the coefficient of variance among 3, 11, and 13
Area
10
11
12
0.75 1.2
Automobile Traffic Density (AID), 10" Mi/Da/Mi' (5)
1.5 l.S 2.0 1.2 0.65 0.4 0.8
1.05 0.95 0.95
13
0.9
Sample Yields, M9/m'
Suspended Participate Matter
Goom.mean.
79.0
Geom. fnean,
year 7.7
Benzene-soluble Particulate Matter
9.6 6.5 9.0 8.4 11.3 10.8 9.2 7.9 7.3 5.3 6.8 5.9
FIGURE 3-12. Observed PAH annual geometric mean concentrations,
ng/m3, and calculated on basis of patterns in coastal areas.
Map shows approximate location of sampling sites. Jointed with
permission from Gordon;« copyright 1976 American Chemical Society.
3-42
-------
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4
BIOLOGIC EFFECTS OF SMOKE, EMISSION, AND SOME OF THEIR PAH COMPONENTS
The environment is a major contributor to the development of a
variety of pathologic conditions in humans. Indeed, Doll and Peto^
have estimated that as much as 13% of all human deaths from cancer may
be attributed to exposure to harmful polluting substances in our
environment. The purposes of this chapter are to describe the biologic
activity of various kinds of polluting emission and some of their PAH
components and to attempt to relate the toxic impact of such emission
to its content of specific PAHs. (See Chapter 3 for discussion of
particle size and respiratory uptake, Chapter 6 for discussion of PAH
transfer in tissues, and Chapter 9 for recommendations.) It considers
biologic activity in bacteria, animal-cell systems, and intact animals,
as well as the nature and advantages of some biologic models used in
emission toxicology.
Hilado and colleagues^^-68 have reported considerable morbidity
in experimental animals that were exposed to the products of combustion
of hard woods, such as birch and oak, or soft woods, such as fir and
pine; they noted no difference in toxicity between the products of
these hard and soft woods. The problem of interpreting results related
to wood is compounded by the presence of preservatives and other
additives in the wood. It is often difficult to establish whether any
observed toxicity is caused by the combustion products of the wood
itself or of a contaminating additive. And it has not been established
whether the PAHa generated during combustion contribute more to the
observed toxicity than the gaseous products. Considerable additional
work with subacute and chronic exposure is required to characterize
toxicity, particularly in view of the current increase in such emission,
Toxicity has been measured in rats and mice intermittently exposed
to diesel exhaust for periods up to 308 h.^-0 The total cumulative
particulate exposure varied from 7.75 to 1,310 mg/nH-h. However,
only minimal changes from normal were observed. Glutathione reductase
and lactic acid dehydrogenase activities, which might serve as
indicators of lung-cell damage, were increased in lavage fluid after 3
wk of exposure at the high dosage; although exposure was continued,
they returned to normal by 6 wk. Neutral protease activity was
increased in lavage fluid after 1 wk of exposure at the medium dosage
(30.6 mg/m^-h) and the high dosage, but returned to normal by the
twelfth week of continued exposure. It is of interest that no
alteration in cytochrome ?450 activity was observed in either mouse
or rat liver at any time in any group. After 12 wk of exposure at the
highest dosage, an increase in the number of macrophages was seen in
the lavage fluid.
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TOXICITY TO SPECIFIC ORGANS AND ORGAN SYSTEMS IN ANIMALS
Manifestations of toxicity to specific organs and organ systems
were detected in animals that were exposed to various kinds of emission.
PULMONARY FUNCTION
Many studies have been conducted in which animals were exposed to
diesel-exhaus t particles (generally 0.1-0.2 pm in size)- Abraham et
al. reported little change in pulmonary resistance or in airway
reactivity to a carbachol aerosol in conscious sheep exposed for 30 min
to diesel-exhaus t particles. Battigelli' exposed human volunteers
for up to 1 h to diesel exhaust at total hydrocarbon concentrations of
2-6 ppm (comparable with the environment in railroad shops) and then
measured pulmonary resistance as an index of function. With this
rather insensitive assay of only relatively short duration, no changes
in function were observed.
Mauderly e_t a_1. measured tracheal mucociliary clearance of a
99mxc-macroaggregated albumin suspension that had been instilled
intratracheally in rats 1 wk before exposure to diesel exhaust for 1,
6, or 12 wk. They also examined the morphology of the lung and trachea
with scanning electron microscopy. In the group of animals that had
been exposed at high dosages (cumulative particle exposure of 151, 822,
or 1,310 mg/m^-h, respectively, after 1, 6, or 12 wk of exposure),
clearance of the suspension was increased after 1 wk; by 12 wk, it was
below normal. In this group, a tendency toward reduced numbers of
ciliated cells was noted. Furthermore, a dose-related increase in
pulmonary macrophages was apparent. Many of these cells contained
diesel particles as inclusions. No changes were seen in the morphology
of the alveoli or airways. In the groups of animals that were exposed
during the same times at lower dosages (30.6, 203, or 317 mg/m^-h) , a
reduction in clearance was the more prevalent response. However, in
the pulmonary function part of this study, in no group of exposed mice
or rats was any significant alteration in pulmonary function observed.
A similar lack of effect on pulmonary function after diesel-exhaust
exposure of rats was reported by Pepelko " and by Gross. ^ In the
study of Pepelko e£ £l. , rats were exposed for 20 h/d, 7 d/wk, for 28 d
to a 1:4 raw or irradiated exhaust from a six-cylinder Nissan diesel
engine. Gross exposed rats to diesel-exhaust particles at 1,500
Ug/nr* for 20 h/d, 5 d/wk, for up to 267 d, but suggested that a
longer chronic exposure of the rats to the particles might result in
lung disease.
Because guinea pigs are generally more susceptible to pulmonary
lesions, they were similarly exposed to diesel-exhaust particles5 for
periods varying from 2 wk to 3 mo at 250-6,000 yg/m . As reported
for other species, the number of pulmonary alveolar macrophages
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increased, and they tended to accumulate at the bronchoalveolar
junctions. Occasional localization of the particles in alveolar Type I
epithelial cells and sporadic increases in Type II cells were
observed. However, all these morphologic changes would be classified
as minimal with regard to pulmonary toxicity.
A different tack was taken by Mauderly et_£l_.110 and by Campbell
e_t aj... to demonstrate an alteration in pulmonary function after
exposure of rats or mice to diesel-exhaust particles. In the former
study, exposed rats were inoculated with -^^P-labeled Pseudomonas
aeruginosa at the oropharynx, and the killing and clearance of these
organisms were ascertained 48 h later; no significant difference in
either measure was observed. In the latter study, mice that had been
exposed to light-duty diesel exhaust (up to 8 h/d, 7 d/wk, for 46 wk)
were treated with aerosols of Streptococcus pyogenes or Salmonella
typhimurium; mice that had been exposed to exhaust showed slightly
increased toxic responses to streptococci. These results were
confirmed in later studies by Campbell e_t al. showing greater
mortality of infected mice exposed to diesel than to gasoline
(catalyst-treated) engine exhausts.
In brief, minimal changes are observed in pulmonary function and
morphology after exposure to diesel-exhaust particles. Although many
morphologic studies have been conducted in animals that have received
some individual PAHs intratracheally or otherwise, there is little
information on resulting alterations in pulmonary function. The
morphologic changes that are generally classified as metaplastic are
discussed later in this chapter.
NERVOUS SYSTEM
Evaluating the effects of any potential toxin on the development
and function of the nervous system experimentally is very difficult.
Laurie and colleagues''''" set about to determine the effects of
chronic diesel-exhaust exposure of neonatal rats on spontaneous
locomotor activity and on performance in a bar-pressing task. The
neonatal rats were exposed to the exhaust at 6 mg/nr for 8-20 h/d for
17-42 d, starting on day 1 or 2 of life. Performance was assessed
during weeks 5-16. The activity was depressed both during exposure and
in the group tested after exposure, compared with a control group;
i.e., they required more extensive training. Because published reports
had indicated that the gaseous components lacked any such effect, the
authors concluded that the particles or their PAH components were the
responsible factors. Laurie and Boyes*' measured the somatosensory
and visual evoked potentials in control rats and rats that had been
exposed to diesel exhaust during neonatal life. Although only small
abnormalities were noted in the visual evoked potential, significantly
longer latencies for all the peaks of somatosensory evoked potential
were seen in the exposed rats. Because the latter potentials are in
the central nervous system, the authors suggested that diesel-exhaust
exposure may lead to failure to develop a normally functioning nervous
system. These types of studies have not been conducted with animals
exposed to individual PAHs, or to mixtures thereof, so it is not known
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whether these organic components are responsible for the nervous system
1a Q i nnfl.
lesions
IMMUNE SYSTEM
The effects of chronic diesel-exhaust exposure of rats on the
immune system were assessed by Mauderly et_ a_l. They placed rats
in chambers and exposed them to diesel-exhaust particles for various
periods under dynamic conditions and gave them sheep red blood cells
intratracheally. The numbers of lymphoid cells that produced IgM
antibody to sheep red cells were determined in lymph nodes and spleen 7
d after inoculation with the sheep red cells. Only minimal or no
effects on the induction of immunity were observed.
Those studies are of particular interest, in view of the long-
known damaging effect of some PAHs on lymphoid tissue. In 1937, Haddow
and co-workers^' reported the systemic toxic effects of PAHs, calling
attention particularly to damage to lymphoid tissue. Acute exposure of
mice to 3-methylcholanthrene reportedly resulted in damage to the
thymus that was followed by thyraoma formation, a marked reduction in
the weights of the spleen and the mesenteric lymph nodes, and
degeneration of bone marrow cells. Newborns appeared particularly
sensitive, suffering a wasting disease that culminated in death. This
toxic effect has also been noted after administration of
7,12-dimethylbenzanthracene (7,12-DMBA) to rats.23'128 Repeated
administrations of dibenz[ahjanthracene, benz[a]anthracene, or anthra-
cene to mice resulted in an increase in stem cells in lymph glands, a
decrease in mature lymphoid cells, and a decrease in spleen weight
(only for dibenz[ahjanthracene). In rats, findings were similar
after treatment with dibenz[ahjanthracene; the effects with anthracene
were much less dramatic.
The total immune response of an organism is an expression of the
sum of humoral and cell-mediated effects. Humoral effects derive from
the activity of B lymphocytes, which on maturation to plasma cells
elaborate immunoglobulins; cell-mediated immunity is expressed by T
cells. The effects of chronic administration of benzo[a]pyrene (daily
subcutaneous injection for 14 d, for a total of 50-400 mg/kg of body
weight) on the humoral immune response were summarized by Dean et_
a_l.3° There was a marked decrease in this response. (The
noncarcinogen benzo[e]pyrene (BeP) was without effect.) A variety of
T-cell responses have been tested for sensitivity to benzo[a]pyrene
(BaP) administration.12'38'39'104 The effects of chronic BaP
administration, to a total of 400 mg/kg of body weight, on T-cell
function were much less marked than those on B-cell activity. Little
effect on the incidence of the B6 tumor in inoculated mice or on the
growth of the B16 melanoma after intravenous challenge was observed.
Furthermore, the resistance of mice to Listeria monocytogenes was
unaltered by administration of BaP, although the expulsion of the
parasite Trichinella spiralis was reduced. BaP administration resulted
4-4
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in myelotoxicity, as determined by in vitro clonal bone-marrow assays.
But delayed hypersensitivity reactions in the host were unaffected. In
summary, the effects of BaP administration on a variety of T-cell
functions were not very significant.
It has long been known that carcinogenic PAHs are imraunosuppres-
sive; indeed, this aspect of their action was believed responsible, at
least in part, for their ability to cause neoplasia. After treatment
of mice with 3-methylcholanthrene (3-MC), dibenz[ahjanthracene, or BaP,
a prolonged depression of the immune response to sheep red cells was
noted; the noncarcinogens BeP and anthracene were ineffective in this
regard.107'163 The effects of the PAHs have been reviewed by
Baldwin, who reported a good correlation between degree of iramuno-
suppression and carcinogenicity.
Although the previously cited work implied a link between the two
activities, Dale and Hedges37 and Stutman16^ definitively
dissociated immunosuppression from carcinogenicity. Using guinea pigs,
Dale and Hedges concluded that the effects of the PAHs were due to
generalized toxicity and were not likely to persist long enough to lead
to neoplasia. Stutraan produced tumors in mice with very low doses of
3-MC—doses that did not influence the immune status of the animals.
To conclude, some PAHs at high doses can alter the immune status of
animals when administered to the point of general toxicity, whereas
exhaust and emission have not been shown to do so.
SKIN
The major changes occurring in skin after application of emission
or PAHs are associated with neoplasia and are discussed later in this
chapter.
KIDNEY
The toxicity of diesel fuel to kidney and other tissues has been
described in only one report: a sailor cleaned his hair with diesel
fuel and was later hospitalized for renal failure. This acute
intoxication also resulted in damage to the liver, the gastrointestinal
tract, and the lungs. The information presented does not allow further
definition of the toxic components responsible for the pathologic
condition.
GLANDS, REPRODUCTION, AND TERATOLOGY
Although individual PAHs have pathologic effects on some glandular
tissue, little toxicity has been reported after administration of
various kinds of emission. The oral administration of 7,12-DMBA to
4-5
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female mice caused the destruction of small oocytes and reduction in
the number of growing and large oocytes. 3 This compound also caused
specific destruction of the adrenal cortex in the rat.13 3-MC
administration resulted in destruction of the primordial oocytes in the
mouse. 1-09 3-MC or BaP given intraperitoneally produced abnormally
shaped sperm indicative of damage to the primary spermatocytes and
spermatogonia. ^
With regard to reproduction and teratology, only few PAHs have been
tested. The feeding of BaP to female rats resulted in no abnormalities
in their ovarian cycle, ovulation, fertilization, or implantation, and
few resorptions were observed in treated pregnant rats.^°^ Similar
findings have been reported for the mouse. ^'
SHORT-TERM MODEL SYSTEMS FOR DETECTING EFFECTS
Whole-animal experiments for assessing toxicity are often expensive
and time-consuming. Therefore, alternative approaches have been
developed. A variety of short-term biologic model systems are avail-
able for assessing the effects of exhaust, its particulate components,
and pyrene analogues. These systems are characterized by the use of
multiple end points to measure genotoxicity, the use of both bacterial
and mammalian cell lines, the use of end points that can be evaluated
in relatively short periods (i.e., 1 d to 6 wk), and the incorporation
of an exogenous source of metabolic activation for generating the
active PAH metabolites. Each end point in concert with a particular
cell system has its own unique strengths and weaknesses. Recognizing
this fact, the regulatory agencies have required a battery of
short-term tests, to provide a more complete picture of the potential
biologic activity of a test chemical. The categories of available
short-term tests are presented in Table 4-1, with a partial list of
some of the particular tests given in Table 4-2 (see Hollstein et_
a_U74 for details).
Examples of the use of these tests in a short-term battery are
presented in Tables 4-3 and 4-4. Table 4-3 demonstrates the guidelines
that the Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA)
recommended; tests recommended by the Organisation for Economic
Cooperation and Development (OECD) are presented in Table 4-4. In
general, these batteries include the evaluation of three or more end
points from the following list: toxic effects, mutagenesis, DNA damage
and repair, chromosomal alteration, and neoplastic transformation.
TOXIC EFFECTS
Toxicity is usually manifested by such end points as cell death,
increase in generation time, decrease in respiration, decrease in rate
of raacromolecular synthesis, and release of particular cell-bound
proteins. Many of these end points have been used in bacteria,
protozoa, ^ algae,^ ^ invertebrates,^'^ fish,^4^ and mammalian
4-6
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cells.88 Cell death is also used concoraitantly with virtually every
assay system, to determine the numbers of cells at risk.
MUTAGENESIS
A mutation is any heritable change in the phenotype of an organism
or cell resulting from a change in its DNA. A mutation need not be
reflected in a change in function. The phenotypic expression of such a
change can be detected in a variety of cellular proteins. Examples of
genetic markers that use mutagenesis as the end point are given in
Table 4-5.
The bacterial systems are best exemplified by the Salmonella
typhimurium test strains developed by Ames jat al•.•* This system
measures the reversion rate to histidine prototrophy in five test
strains that carry specific frameshift and base-pair substitutions at
the his locus and a series of mutations at the other loci to make the
bacteria more sensitive to chemically induced mutation. The deep rough
mutation (rfa~), ultraviolet-light sensitivity (uvr B~), nitrate
reductase deficiency (chl~). biotin deficiency (bio"), and
introduction of R factor plasmids are examples of alterations of these
test strains to make them more sensitive to chemically induced
mutagenesis. Because reversion to his prototrophy is being measured, a
battery of strains (three to five) must be tested, to ensure detection
of point mutation, frameshift mutation, and intragenic deletion. To
circumvent this problem, there have been attempts to standardize a
forward-mutation assay with S. typhimurium. Forward mutations at both
^ T T fc^ T ~TO
the arabinose-resistance (arar) *•**>> L/Q an(j 8-azaguanine-resistance
(8-Azr)156 genes have been described. These assays have the
advantages of detecting virtually all mutagenic events, detecting
mutagenesis at more than one genetic locus (probably at least three),
and requiring the use of only one test strain.
Mutagenesis testing in mammalian cells has used cell types that
range from the rapidly growing, easily handled cell lines, such as
CHO*55 and V-79,26'27 to the more difficult testing of in
vivo-derived human lymphocytes. Advantages of the CHO and V-79
cells include high plating and cloning efficiencies, pseudodiploidy,
and the ability to monitor mutagenesis at a variety of genetic loci. A
disadvantage is that these cells have little or no capacity to
metabolize xenobiotics, especially pyrenes. Recent results5^ suggest
that hamster-derived cell lines, such as CHO, have limited capacity to
remove 06-alkylated guanine; thus, they may be deficient in DNA repair.
The merits and limitations of the three most widely used loci for
testing with mammlian cells are presented in Table 4-6. Such end
points as resistance to purine analogues, to 5-bromodeoxyuridine
(5-BUdR), and to ouabain collectively can detect most of the potential
genotoxic effects of PAHs.77 These end points are now being used
simultaneously to limit the possibility of false-negative conclusions.
In assays for purine-analogue resistance, mutants lacking the enzyme of
the purine salvage pathway, hypoxanthine-guanine phosphoribosyl
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transferase (HGPRT), are identified by their resistance to toxic
analogues, such as 8-azaguanine or 6-thioguanine. In assays for
ouabain resistance, mutants are detected by their ability to grow in
the presence of the glycoside ouabain. The basis of the latter muta-
tion is an alteration in the receptor for the membranal sodium-
potassium ATPase system. In the assay for 5-BUdR resistance, an
alteration of the enzyme thymidine kinase is responsible for the mutant
phenotype. The altered enzyme is unable to "activate" 5-BUdR by
catalyzing its conversion to a deoxyribonucleotide; the latter is
required for cell death.
Such cells as C3H10T1/2 and BALB/3T3 have also been used in
mutagenesis studies; these cells have easily detectable hydrocarbon-
metabolizing activity.^°>^ However, they are hypotetraploid, may
not detect some recessive mutations, and may not detect some mutations
that are expressed codominantly. ^ These cells express a high degree
of contact inhibition and low saturation density and thus can be used
in bioassays of neoplastic transformation. Recent studies have
suggested that such cells can be used to detect simultaneously the
mutagenic and transforming capacities of test chemicals. -* Primary
cell strains and in vivo-derived cells have been used in mutagenesis
assays; although they have high PAH-raetabolizing capacities and are
diploid, the difficulty in growing, handling, and evaluating data from
these mixtures of cells is an important disadvantage.
DNA DAMAGE AND REPAIR
Assays for DNA damage and repair have also used both bacterial and
mammalian cells. Primary DNA damage in mammalian cells has been
measured by such end points as selective toxicity in strains of cells
deficient in DNA repair, increase in rate of DNA elution under
1 f\ S
alkaline conditions, LD:3 formation of specific pyrene-DNA
adducts,^3 increase in rate of unscheduled DNA synthesis, ^-°^
increase in incorporation of specific dyes,^'- and increase in
incidence of sister chromatid exchange.180 QJJ^ repair is a specific
response to DNA damage. The covalent interaction of chemicals with DNA
provokes an enzymatic repair of the damaged regions of DNA.^*^
Repair synthesis can be measured in a variety of ways, but
incorporation of radioactive precursors into DNA is the
simplest. ',60 ^ jj^ damage-repair system that shows promise in
detecting chemically induced DNA alteration uses the rat
hepatocyte.1'* This assay has the advantages of using nondividing
cells (normal semiconservative DNA replication is suppressed) and using
freshly cultured cells that have high endogenous capacity for
carcinogen metabolism or activation. It has recently been shown to be
effective in detecting the ability of a variety of chemical carcinogens
(including many different PAHs) to damage DNA.^36
An increase in sister chromatid exchange (SCE) may be one of the
best measures of DNA damage in humans. This end point, which involves
incorporation of 5-BUdR into DNA during two cycles of replication and
making the two chromacids stain differently, so that exchanges of
4-8
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material are scorable, seems to develop as the consequence of
presumably long-lived DNA lesions in the S phase of the cell cycle.
The exact mechanism of SCE formation is not understood, although it
is well known that the frequency of SCE is increased by exposure of
cells to known mutagens in vivo or in vitro.^»127,161 jn fact( a
linear correlation between mutations induced at specific loci and SCE
has been demonstrated in CHO cells. This assay has been used to
monitor the exposure of humans to potentially harmful
chemicala51>94,118 and even to cigarette smoke.^^
The in vivo techniques for detection of SCE can be applied in two
basic ways. One method involves 5-BUdR incorporation into bone-marrow
cell DNA by inoculation of solutions or implantations of 5-BUdR pellets
directly into animals^ and exposure of the animals to the chemical
under study; this method has been used to detect in vivo DNA damage
via such substances as cyclophosphamide," styrene, ^ benzene,^-^
urethane,^^ and cigarette smoke." The second method involves the
incorporation of 5-BUdR into lymphocyte cultures during mitogen-induced
activation in vitro; this has been used in the human studies mentioned
above. Good baseline data on the incidence and variation of SCE in
humans now exist.^° SCE has also been shown to persist for several
days or even months after'chemical exposure and thus can serve as an
index of acute or chronic exposure to chemicals.94,118,161
Comparison of rates of formation of SCE and specific DNA adducts
suggests that, for several types of mutagens, induction of SCE does not
necessarily result from a single specific DNA lesion.
CHROMOSOMAL ABERRATION
Assays for chromosomal aberration are also used to monitor for the
mutagenic activity of test chemicals. These assays detect major
rearrangements in the chromosomal or chromatid structure and include
such end points as chromosomal or chromatid breaks, chromatid trans-
location, dicentric chromosomes, ring chromosomes, balanced transloca-
tion, and inversion. "»^° Another test for acutely altered chromo-
somes is the micronucleus test, in which chromosomal damage leads to
fragmentation of chromosomes or malfunction of the spindle apparatus,
so that whole chromosomes lag behind the rest and, accordingly, form
micronuclei.^i These techniques can be used with tissues derived
either in vitro or in vivo much like those used for analysis of SCE.
Generally, agents that induce point mutation also induce chromosomal
aberration. In humans, mitogen-activated lymphocytes can be used to
monitor for the effects of exposure to physical and chemical agents.
Exposure to radiation, to such chemicals as alcohol and vinyl chloride,
and to cigarette smoke causes increases in chromosomal aberra-
ion
tion.lzu Cytogenic end points of aberration are useful, but one
should remember that often chemicals induce very few aberrations at
concentrations that permit the end point of gene mutation to be readily
observed.^^ In recent comparisons of three cytogenetic tests—
4-9
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induction of chromosomal aberration, induction of micronuclei, and
induction of SCE—the third proved to be the most sensitive in testing
with several PAHs.8
NEOPLASTIC TRANSFORMATION
Neoplastic transformation has been assayed by a variety of in vitro
systems, and it is not possible to review all the pertinent literature
here. The reader is directed to the recent reviews of Casto and
Carver,21 Heidelberger,64 and Mishra e_t a_l.115
Specific cells that have been used for assay of in vitro chemically
induced neoplastic transformation include normal rodent (diploid) cell
strains,42>129 established aneuploid rodent cell lines,36'49'64'84
cell lines derived from human tumors, 0»138 and cell lines initiated
from apparently healthy human tissue. 0>"» 113,154 Xable 4-7 com-
pares properties of some mammalian-transformation systems. These cell
lines share the following properties to some degree: They exhibit
density-dependent inhibition of cell division and reach a defined
saturation density, do not form colonies on soft agar or agarose, and
do not give rise to tumors when inoculated into immunosuppressed
syngeneic hosts. After transformation by chemicals, they lose the
density-dependent inhibition of cell division and form piled-up,
criss-crossed foci; they grow on soft agar or agarose, and they form
tumors when inoculated into host animals. In addition, many trans-
formed cells exhibit increased fibrinolytic activity,1*4 altered
morphology in the scanning electron microscope,1^6 specific chromo-
somal arrangement, ^-O*134 and specific DNA sequences that can be
transfected into normal cells, resulting in formation of the trans-
formed phenotype. 33> I-*2 Although each of these cell systems has been
successfully used to ascertain the biologic activity of chemical
agents, none appears to be capable of universally detecting all classes
of chemical carcinogens, low concentrations of all such agents, and
relatively weak biologic activity of some chemicals.
MUTAGENESIS
As just discussed, a number of model systems are available for
assessing the mutagenic activity of emission, individual PAHs, and
their mixtures. These are in two categories: bacterial systems and
mammalian cell-culture systems. The activity of emission and its PAH
constituents is discussed below relative to both kinds of model.
BACTERIAL MUTAGENESIS
Particulate matter from city air has been tested for mutagenic
activity with the Salmonella/microsome system.131»16'•169 In all
cases, a positive response was obtained. Furthermore, many of the
samples exhibited direct-acting mutagenic activity, i.e., the addition
of activating enzymes present in a liver S-9 fraction was not required
4-10
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for mutagenic activity.30'132'167'168'171 Wang e£ al.177
collected air samples from a residential area at an intersection of two
heavily trafficked crossroads in the Buffalo, New York, area.
Extraction of the particulate fraction with acetone resulted in a
preparation highly mutagenic in Salmonella strains TA 98, TA 100, and
TA 1537. These investigators also obtained a positive direct mutagenic
response with automobile-exhaust samples from a spark-ignition
internal-combustion engine (with leaded gas as the fuel). The
mutagenic ingredients appeared to originate in motor oil during the
combustion process and were not due to lead. Similar results have been
obtained by Pitts e_t_ a_l.132 with atmospheric particulate extracts
from the Los Angeles basin, by Teranishi &t_ a_l.,168 by Tokiwa et_
al.171 with extracts from several Japanese cities, and by Talcott and
Wei167 and Commoner e_£ al.30 with extracts from other American
cities. Unfortunately, the quantitation of some of these studies may
be open to question because of filter artifacts. The disposition of
the filter apparatus in relation to sunlight, temperature, etc., is
important, because these factors may facilitate chemical reactions
involving PAHs and may result in artifactual formation of mutagens.
This aspect is discussed in Chapter 3.
Soot makes up 2-15% of the mass of fine particles that are present
in urban atmospheres. ^" A number of studies have been conducted to
establish its mutagenic potential. Kadin e_t jl.83 have experimen-
tally generated soot from ingredients with varied sulfur composition
—i.e., from pyridine, decalin, and o-xylene or from thiophene,
decalin, and £-xylene—and have compared its mutagenicity with that of
soot obtained from burned kerosene. Dichlororaethane extracts of all
the soots were mutagenic in a bacterial assay in which a forward muta-
tion of 8-azaguanine resistance was measured. The soots generated from
the sulfur-containing and nitrogen-containing ingredients, as well as
soots from kerosene or furnace black, exhibited 10-17% of the rautagenic
activity of authentic BaP (on a weight basis).
Emission from spark-ignition combustion and diesel engines has been
tested for rautagenic activity in the Salmonella system.27,79,83,101
It is known that particulate emission from light-duty diesel engines is
considerably greater than that from light-duty catalyst-equipped
spark-ignition engines—i.e.. 0.2-1.0 vs. 0.006-0.02 g/mi.147 Table
4-8 presents data of Claxton^8 relative to comparative mutagenic
activity of emission of diesel and spark-ignition engines, of
cigarette-smoke condensate, of coke-oven emission, of roofing-tar
emission, and of BaP (positive control). The results are reported in
terms of revertants/100 Mg of soluble dichloromethane organic
compounds; the soluble organic components represent approximately 25%
of the total mass of the particles. As is evident from the table,
cigarette-smoke condensate, roofing tar, and BaP required metabolic
activation by an S-9 fraction, whereas diesel-engine exhaust was
directly mutagenic. The other kinds of emission were both directly and
indirectly mutagenic. The diesel exhaust exhibited a wide range of
rautagenic activity, although the high value is probably peculiar to the
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particular engine that generated the emission. The activity of BaP is
far greater than that of emission.
Naraan and Clark^' have determined the quantity of particles
emitted and the mutagenic activity of extracts of the exhausts of
several spark-ignition engines that burned gasoline, a 90X/10X ethanol
blend, or commercial gasohol. The results are presented in Table 4-9.
Although the number of revertants per mile differed for each of the
four-cylinder engines, the addition of ethanol clearly reduced the
direct mutagenic capacity.
Several investigators have determined the rautagenic activity of
respirable coal fly ash ,'5,69 which does have mutagenic activity in
the Salmonella/microsome assay. Virtually all the emissions yield a
mutagenic response in this test system.
Extracts from the various kinds of emission contain a large number
of PAHs, among which is BaP-1^8 >89»176 Extracts of diesel particles
have been separated on Sephadex LH-20 into six fractions;^! the
contribution of each to the total mass of the diesel extract obtained
from a low-sulfur and high-sulfur fuel is shown in Table 4-10.
Furthermore, each of these exhausts was obtained before or after
passage through an oxidative catalyst. Fraction 1 contained most of
the mass of the extracts from both fuel exhausts. However, fractions 3
and 4 contained most of the mutagenic activity. Fraction 3 from the
low-sulfur exhaust contained the bulk of the PAHs, including
phenanthrene, methylphenanthrenes, fluoranthene, pyrene, methylpyrenes,
benzo[ghi]fluoranthene, benzanthracene (BA), chrysene (or
benzo[c]phenanthrene), methyl-BAa, and BeP (or perylene). * With the
high-sulfur fuel, one found, in addition, the methylbenzothiophenes.
It is of interest that the low-sulfur fuel gave an exhaust whose
mutagenic activity was increased after passage through a catalyst. The
reverse was true for the high-sulfur fuel. Furthermore, fraction 4
from the high-sulfur fuel, before oxidative catalysis, proved the most
mutagenic.
The major identified components of emission have been tested for
mutagenic activity with the Salmonella forward-mutation assay of Thilly
and co-workers.83,101 jn tnis aaaay% mutants that are resistant to
the purine analogue 8-azaguanine are scored. Of the components present
in kerosene-soot extract, cyclopenta[cd]pyrene proved the most
mutagenic; it was also present in the highest concentration (see Table
4-11). Cyclopenta[cd]pyrene is a known component of all
soots ,53,174,175 of cigarette smoke,^' of automobile exhaust,
and of coal fly ash.25 The sum of the mutagenicities of the
identified individual PAHs was slightly greater than that of the
kerosene-soot extract itself. The total mutagenic activity of the
kerosene-soot extract could almost be reproduced by that of the
cyclopenta[cd]pyrene.
The investigators compared the mutagenic efficacy of additional
PAHs with and without an S-9 preparation, using induced cells from the
liver; the results are in Table 4-12. Methylation of several of the
4-12
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inactive PAHs, such as anthracene and phenanthrene, resulted in the
acquisition of mutagenicity. Preliminary evidence has led the Thilly
group to suspect the presence of alkyl-substituted anthracene and
phenanthrene in diesel-soot fractions that were mutagenic in
"• bacteria. ^ In this series, the most active compound was perylene,
which was followed by cyclopenta[cd]pyrene. The mutagenicity of
cyclopenta[cd]pyrene in the Salraonella/raicrosome assay has been
confirmed by Eisenstadt and Gold; ^ metabolic activation by the S-9
fraction was required before this mutagenic property was elicited.
Nitrated PAHs
As mentioned previously, emission from either diesel or
spark-ignition engines exhibits considerable direct-acting mutagenic
activity in the Salmonella/microsome assay, whereas cigarette-smoke
condensates, roofing-tar extracts, and BaP do not. This has led
several investigators to study engine exhaust for the presence of
direct-acting PAH derivatives that might have been produced by gaseous
exhaust components—e.g., nitrogen oxides—or by atmospheric oxidative
reactions involving ozone. Various nitropyrenes and other analogues
have been assayed for mutagenic activity in the bacterial
system.52,96,111,112,124,130 These substances exhibit potent
activity in the Salmonella mutagenesis assay. Indeed, it has been
estimated by Gorse (personal communication) that the concentration of
nitropyrene alone in diesel particulate extracts could account for
13-24% of the total direct mutagenic activity with TA 98. Tokiwa e_£
gl.1^0 have assayed the mutagenicity of the nitrophenanthrenes,
1-nitropyrene, 3-nitrofluoranthene, and 6-nitrochrysene. Each of the
parent PAHs was inactive as a direct mutagen, but 6-nitrochrysene was
slightly active, nitrophenanthrene was active, and 1-nitropyrene was
most active against TA 98 and TA 100. S-9 was not required for this
demonstration of mutagenic activity. Pitts ejt a^.132 reported the
direct mutagenic activity of 1-, 3-, and 6-nitrobenzo[a]pyrene in the
Salmonella/microsome assay. Perylene, another exhaust constituent that
is converted to 3-nitroperylene, demonstrated mutagenesis ."^ in a
similar fashion, nitrated derivatives of anthracene, fluoranthene,
benzfa]anthracene, benzo[k]fluoranthene, and benzo[ghi]perylene—all of
which are present in diesel exhaust—exhibited potent mutagenic
activity in the Salmonella assay.
The nitropyrenes have been reported as contaminants of xerographic
copiers and toners, which may therefore contribute to the problem of
mutagenicity.^"^' "^ Rosenkranz et^^.^-* have demonstrated the
presence of such a mutagenic activity with various Salmonella test
strains; they have traced this property to nitropyrenes that were
present as impurities in carbon black. In addition to mononitrated
components, they were able to identify the 1,3-, 1,6-, and
1,8-dinitropyrenes, 1,3,6-trinitropyrene, and 1,3,6,8-tetranitropyrene
as contaminants. All these derivatives demonstrated direct mutagenic
activity (see Table 4-13) with both nitroreductase-positive and
-negative variants of Salmonella. The mutagenic property of
4-13
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1-nitropyrene and 1,3-dinitropyrene depended heavily on the endogenous
bacterial nitroreductase activity. Insertion of two nitro groups in
the pyrene moiety increased mutagenic activity by a factor of approxi-
mately 100, although information is insufficient to extrapolate to
other PAHs. The most potent of these derivatives was 1,8-dinitro-
pyrene. It was striking that, of the three dinitro derivatives, two
acted independently of endogenous nitroreductase; equal numbers of
revertants per nanomole are observed in both TA 98 and TA 98 NR. A
similar situation occurred with the trinitropyrenes and tetranitro-
pyrene. The mutagenic activity of 1,8-dinitropyrene is the highest
ever recorded in the literature. ^ The presence of the 1,6- and
1,8-dinitropyrenes as predominant mutagenic components in diesel-
particle extract has been confirmed by Pederson and Siak,^^ who
estimated that 15-20% of the total mutagenic activity of the extract
may be contributed by these dinitropyrenes (in addition to as much as
24% contributed by 1-nitropyrene).
Sulfur-Containing. PAHs
The presence of sulfur-containing heterocyclic PAHs has been
reported in various combustion products, particularly from high-sulfur
petroleum products (see Chapter 1). In many heterocyclic structures,
one aromatic ring has been replaced by thiophene.®" It is anti-
cipated that the increase in the use of coal, particularly with high
sulfur content, will result in substantial environmental pollution with
these ingredients. Thus, it is imperative to have a better understand-
ing of the biologic effects of these sulfur-containing heterocyclic
PAHs.
The mutagenicity of several sulfur-containing PAHs has been
determined in the Salraonella/microsome assay by Karcher e£ al. ,°' and
the results are presented in Table 4-14. Of the isomers listed,
benzo[2,3]phenanthro[4,5-bcd]thiophene was the most potent; its ring
configuration corresponds to that of BaP, although it is more mutagenic
than the latter.
ANIMAL-CELL MUTAGENESIS
A number of animal-cell model systems have been used to ascertain
the mutagenic effects of combustion-engine emission, as well as other
exhaust. These have been reviewed in previous monographs on
PAHs.174>175 Many of the tests depend on the selection of variants
on the basis of resistance to 8-azaguanine, 6-thioguanine, ouabain, or
deoxythymidine analogues.
Comparative data on the development of 6-thioguanine resistance in
Chinese hamster ovary (CHO) cells have been reported by Casto e£
a1., who used extracts of diesel-exhaust particles and coke-oven
emission (see Table 4-15). All extracts yielded the same number of
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mutant cells, which was comparable with that of the positive control,
methyl methanesulfonate, a known direct-acting methylating agent.
Curren ejt al_.^ have tested the production of ouabain-resistant
BALB/c 3T3 cells when the latter were exposed to a variety of agents
(see Table 4-16). The spark-ignition-engine extract was considerably
more mutagenic in this assay than the diesel extracts. A roofing-tar
pot sample and coke-oven emission also exhibited greater mutagenic
efficacy. The presence of an activating system did not significantly
affect rautagenicity. The extract from the gasoline-engine exhaust
appeared more mutagenic than the extracts from various diesel engines,
and coke-oven pot samples were even more active. Mutability at several
different genetic loci by PAHs has been determined by Huberman and
Sachs™ (Table 4-17). Good mutagenic activity with respect to the
HGPRT locus was manifested by dibenz[ac]anthracene,
dibenz[ah]anthracene, 7-methylbenzanthracene, BaP, 7,12-DMBA, and
3-MC. The last four compounds named were also mutagenic with respect
to ouabain resistance. At both loci, 7,12-DMBA was most active.
As indicated previously, diesel exhaust demonstrates considerable
direct mutagenic activity in the Salmonella/microsome assay. The
nitro-PAHs have been considered as likely candidates for this
activity. Thilly and colleagues have been unable to demonstrate
any direct mutagenic activity with human lymphoblasts as the target
cells, although, in the presence of an activating system, a consider-
able amount of 6-thioguanine resistance and trifluorothymidine
resistance resulted after addition of diesel extract to the culture
media. These experiments suggest that the nitrated PAHs, if present in
the diesel extracts, are rapidly inactivated by the lyraphoblasts or
require for activation a nitroreductase (or other enzyme) that is
absent from these cells. Indeed, application of the term "direct-
acting rautagen" to the nitrated PAHs is not entirely correct. It is
postulated that these analogues undergo a reduction, catalyzed by a
nitroreductase, to an amino derivative that may be further transformed
into reactive hydroxylamino PAHs (see Chapter 3). The latter would
easily form electrophilic substances that could interact with DNA in
causing a mutation. What is needed is additional experimentation on
the mechanism of action of the nitrated PAHs in both bacterial and
mammalian-cell systems.
Sister chromatid exchange has been used to assess genotoxic
activity of various kinds of emission. Unfortunately, SCE appears to
be more predictive of point mutation than of frameshift mutation,1-'
whereas most of the PAHs produce the latter damage. The experiments of
Mitchell et al.,116 which used CHO cells, indicated that all the
emission extracts were inferior to BaP in inducing SCE. Of the emis-
sions, coke-oven extracts proved the most active, and the heavy-duty
Caterpillar diesel-engine exhaust was the least potent. Intermediate in
activity were cigarette-smoke condensate, roofing-tar emission, Mustang
gasoline-engine emission, and other diesel-engine emission. None of
these required metabolic activation for SCE activity.
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The induction of SCE has been performed in vivo with Chinese
hamsters that were given various PAHs intraperitoneally. After
two injections, the bone marrow was aspirated, and the SCEs per
metaphase cell were determined. Although the positive control, BaP,
did produce SCE, there was little correlation between the quantitative
aspects and the carcinogenic potential of the PAHs. No comparable
experiments were performed with the various kinds of emission. The
experiments of Schbnwald ejc a^.^" also showed a lack of correlation
between carcinogenicity and SCE. These investigators determined SCE
induced by BaP with human lymphocytes obtained from normal persons and
lung-cancer patients; no difference was observed. Guerrero etal.^Sa
intratracheally exposed Syrian hamsters to 200 ng of BaP over a 10-wk
period, examined in vitro cultures of lung tissue for sister chromatid
exchange (SCE), and concluded from the results that BaP was
metabolically activated by lung cells in vivo. In other studies,
diesel exhaust particles (DEP) in doses of 0-20 mg per hamster were
administered over a 24-h period; although the study was limited in
scope, the results demonstrated that DEP can induce genotoxic damage.
CARCINOGENESIS
SKIN
Kotin and colleagues"'"*^ first reported the presence of carcino-
genic substances in the exhaust of gasoline and diesel engines. Benzene
extracts of particles from these sources produced both papillomas and
carcinomas when applied to the skin of mice. These studies were
extended by Wynder and Hoffmann, " who compared the carcinogenicity
of cigarette tar with that of organic extracts of gasoline-engine
exhaust particles. The latter, obtained from a 1958 gasoline engine
without a catalytic converter, proved twice as active (on a weight
basis) as cigarette tar. Many studies have since been conducted with
skin as the target tissue; only a few are described here.
Automobile-exhaust condensate has been partitioned into a number of
fractions by Pott e_t a_l., ^ with the PAHs predominantly found in
fraction IV, the nitromethane phase. Each of these fractions was
tested^ for ability to produce papillomas and carcinomas in life-
long mouse skin-painting experiments in which combined initiator and
promoter activity was measured. BaP, the positive control, at
1.92-7.68 ug/treatment caused tumor formation in 15-60% of the mice.
The exhaust condensate at 0.53-4.7 mg/treatment, equivalent to BaP at
0.15-1.35 vig/treatment, produced tumors in 1-72X of the mice, and the
tumors arose after a shorter latent period. The major tumor-producing
activity was noted in fraction IV, which contained the PAHs. In this
fraction, however, BaP is responsible for only 91 of the carcino-
genicity of automobile-exhaust condensate (AEC). Agents other than
BaP, acting either alone or synergistically with AEC, are responsible
for the major carcinogenicity of AEC and probably of diesel exhaust.
The tumor-producing effects of AEC in the carcinogen mouse model
have been contrasted with those of 15 PAHs that occur as major
4-16
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components of AEG. These components and their relative
concentrations in a simulated AEC mixture are shown in Table 4-18.
AEC, diesel-exhaust condensate (DEC), BaP (the positive control), and
the mixture of PAHs were tested for their comparative potency (see
Table 4-19). The data indicate the greater potency of AEC than of
DEC. If the relative potency of AEC were accepted as 1, the
corresponding values for DEC, BaP, and the PAH mixture would be 0.02,
187, and 68, respectively. The proportions of the carcinogenic potency
of AEC and DEC attributable to the selected PAHs can be calculated.
BaP would account for only 9.62 of this potency in AEC, and the
selected PAHs, only 412. In DEC, the contribution of BaP is
approximately 16%. These results indicate that compounds other than
the selected PAHs contribute to the carcinogenic potency of AEC or DEC.
Slaga and associates'** used a mouse that had been bred for
quickness of response in the initiation-promotion skin-carcinogenesis
model—the SENCA& mouse—to study comparative biologic potency of
various kinds of emission and PAHs (see Table 4-20). The exhausts were
relatively ineffective, in comparison with purified BaP, in causing
papilloma formation. Indeed, 10 aig each of emission from roofing tar,
coke ovens, and the Nissan diesel engine was equivalent in response to
50, 60, and 80 g of BaP, respectively. In no case did 10 mg of
emission extract contain that much BaP- The activity of anthracene,
pyrene, dibenz[ah]anthracene, dibenz[ac]anthracene, benz[a]anthracene,
2-hydroxybenzo[a]pyrene, and BaP as complete carcinogens and as tumor
initiators was compared in this mouse strain. " Their relative
potencies were 0, 0, 20, 0, 5, 30, and 30, respectively, compared with
7,12-DMBA, set at a potency of 100. Seaman1 and colleagues extended
these studies by determining whether groups of nonactive PAHs would
interact with the carcinogens in a synergistic or inhibitory
manner.1*' The proportions of the various compounds were chosen on
the basis of their relative concentrations in automobile exhaust. The
groups of carcinogens and noncarcinogens are shown in Table 4-21, and
the percent tumor formation after lifetime application is shown in
Table 4-22. Mixtures of the four carcinogens were more effective than
a comparable dose of BaP alone. Of greater importance, no evidence of
synergism or inhibition could be found when mixtures of carcinogens and
noncarcinogens were applied.
The application of multiple PAHs to mouse skin has often resulted
in data that were confusing, with regard to carcinogenesis. Thus, in
opposition to the above discussion, Steiner^^ reported that the
combination of two weak carcinogens, benz[aJanthracene and chrysene,
resulted in a synergistic-tumorigenic response; benz[a]anthracene and
dibenz[ah]anthracene yielded fewer tumors than expected; and
dibenz[ah]anthracene and 3-MC yielded the sum of individual tumorigenic
potentials. Falk and co-workers reported much lower tumor
production after the simultaneous administration of BaP and several
noncarcinogenic hydrocarbons. Van Duuren and Goldschmidt noted
that repeated application of the weak carcinogen BeP and the
noncarcinogen pyrene to mouse skin with BaP resulted in a
coca re ino genie effect. DiGiovanni e£ £l.« found that mouse-skin
carcinogenesis induced by 7,12-DMBA was inhibited when BeP, pyrene, or
fluoranthene was applied 5 min before the initiator. The apparent
paradox was explained by the later studies of DiGiovanni and
4-17
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Slaga.40 They used either 7,12-DMBA, BaP, or 3-MC as an initiator
and tetradecanoyl phorbol acetate (TPA) as the promoter. BeP or
dibenz [ac]anthracene was applied 5 min before the initiator in all
cases. With 7,12-DMBA as the initiator, BeP and dibenz [acjanthracene
each reduced tumor igenes is by more than 80%. However, with BaP as
initiator, dibenz [ac] anthracene exerted no effect and BeP stimulated
tumor formation by 30%. If dibenz [ac]anthracene was applied 12, 24. or
36 h before BaP, a reduction in tumorigenesis was observed. With 3-MCT
as initiator, dibenz[ac] anthracene inhibited tumor formation, whereas
BeP was without effect. BeP apparently exerts its effect on the
7 , 12-DMBA-initiated system by profoundly inhibiting the ring
hydroxylation of this initiator and reducing the covalent binding to
DNA. Thus, the order of application of the multiple noncarcinogenic
with carcinogenic PAHs can have serious effects on carcinogenesis .
Finally, with regard to mouse-skin tumorigenesis, cyclopenta[cd]-
pyrene , a major component of soot that can transform mouse fibroblasts
oncogenically; 122 was tested for tumor-initiating activity on mouse
skin by Wood e_t a^. Although tumorigenic, cyclopenta[cd]pyrene
was weaker than BeP.
TISSUES OTHER THAN SKIN
PAHs and exhaust condensates have been administered to experimental
animals in ways other than topically. The subcutaneous injection of
AEG and fractions thereof into mice produced sarcomatous lesions; ^5
administration of 20-60 mg yielded tumors in up to 8% of mice, and
administration of 10 or 90 ug of BaP yielded tumors in 17% or 75% of
the animals, respectively. Simultaneous administration of 20 mg of AEC
with 90 pg of BaP yielded lower tumorigenesis. The most active
fraction from AEC was the nitromethane phase, which contained the
various PAHs.
Sellakumar and Shubik*-^ studied benz[a]anthracene, benzofb]-
fluoranthene, dibenz[ah]anthracene, dibenzo[ai]pyrene, and pyrene.
They mixed the PAHs with a hematite dust (at 1:1), suspended the
mixture in 0.9% saline, and instilled it intratracheally at weekly
intervals into Syrian golden hamsters. Most of the PAHs were not
carcinogenic in this limited series, but dibenzo[ai]pyrene produced a
high incidence of carcinomas. With multiple doses that totaled 8 mg of
this substance, 47% of the hamsters had respiratory tract tumors
(squamous cell carcinomas); with 12 mg, 89% of the animals were
affected. This degree of carcinogenicity is greater than that of BaP.
Reznik-Schuller and Mohr^" have compared the carcinogenicity of
AEC with that of several major PAH constituents in the Syrian golden
hamster intratracheal model. The hamsters were given AEC at 2.5 or 5
mg/animal every 2 wk intratracheally, corresponding to a total
administration of 42.5-75 or 75-150 mg of AEC. The total was
equivalent to 11.56-25.5 or 25.5-51 ug of BaP. In all animals,
multiple pulmonary adenomas were observed. This strikingly high
incidence of neoplasia could not be explained by the BaP content of the
AEC. It is of interest, however, that no carcinomas were observed.
4-18
-------
As indicated earlier and as is discussed more fully in Chapter 6,
ingestion of PAHs , whose presence may be attributed to vehicular
exhaust, appears to be a major route of entry in animal systems. Yet,
the literature pertinent to this form of administration of exhaust
particles, their major PAHs, and mixtures thereof is very limited.
Neal and Rigdon*-2'- » *^° have examined the effects of oral administra-
tion of BaP on tumor formation in mice. No gastric tumors developed in
any of the 289 mice that were fed a control ration; the incidence of
tumors in the BaP-fed mice depended on concentration in the food and on
the number of days of feeding. **• These investigators^^ also
established that the incidences of pulmonary adenomas, gastric tumors,
and leukemia in BaP-fed mice were genetically determined. No relation-
ship, however; was observed between the relative incidences of these
two types of neoplasms within a given mouse. Studies of these types
would be useful, with regard to other PAHs and their mixtures. The
interpretation of these studies is colored by the failure to house the
mice in metabolic chambers, which would eliminate the contribution of
coprophagy.
Another series of studies took advantage of the susceptibility of
the A strain mouse to pulmonary adenoma formation, particularly after
the intravenous administration of selected PAHs. ^ Shimkin and
- were able to calculate the amount of each agent that had to
be injected for the induction of one pulmonary adenoma in this strain
of mouse. The compounds tested were 3-MC, dibenz [ah] anthracene,
7H-dibenzo [cglcarbazoyl, BaP, dibenz [aj]aceanthrylene, and dibenz[ah]-
acridine. The respective values were 0.9, 1.0, 6.0, 9.5, 14, and 18
utnol/kg of body weight for one adenoma. Benz [ajanthracene was essen-
tially inactive. The objection to the use of the A strain mouse for
these types of studies rests on its extraordinary sensitivity to
pulmonary adenoma formation. In fact, if the A strain mouse is allowed
to survive long enough, almost all the untreated animals will develop
these tumors — they are already "initiated."
ALKYLATED PAHa, MUTAGENESIS, AND CARCINOGENESIS
Because of the presence of alkylated PAHs in cigarette smoke and
various coal-derived liquids and tars,^6»62'71'7^ their biologic
effects are of paramount interest. Perhaps the most thoroughly studied
of the alkylated PAHs are the methylbenz[a]anthracenes,
methylchrysenes, methylanthracenes, and methylphenanthrenes.
Some of the earliest studies, in which the effects of a methyl
group on the carcinogenicity of benz[a]anthracene (BA) were
investigated, were conducted by Dunning and Curtiss and by Huggins's
laboratory. 2 These investigators monitored sarcoma incidence in
rats to which the various PAHs had been administered subcutaneously.
Their results, which were in remarkable agreement, indicated that the
insertion of a methyl group at position 6 or 7 of BA increased
tumorigenicity to the extent that 70-100Z of the rats were affected.
4-19
-------
However, 8- or 12-methyl-BA resulted in tumor formation in only 50-69?
of the rats, and 1-, 2-, 3-, 4-, 5-, 9-, 10-, or 11-methyl-BA proved
noncarcinogenic.
The nature of the alkyl group was an important consideration:
substitution of an ethyl group at position 7 or 12 of BA greatly
diminished tumor incidence, compared with that of the methyl
congeners. ^ Pataki and Huggins have also studied the
structure-activity relationship in the BA series when two methyl groups
were inserted. A marked increase in tumorigenicity—shown by sarcoma
formation—was observed with 6,7-dimethylbenz[ajanthracene (DMBA) and
6,8-, 6,12-, 7,8-, 7,12-, and 8,12-DMBA. But, 1,12-, 3,9-, and 9,10-
DMBA were essentially nontumorigenic. Of the trimethylated BA
derivatives, 6,7,8-, 6,7,12-, and 7,8,12-trimethylbenz[ajanthracenes
were all very tumorigenic.
The mutagenicity and tumor-initiating activity of methylated
fluorenes, phenanthrenes, anthracenes, and benzofluorenes
were studied by LaVoie ^t al_.'' > ^"" Only the 9-methylf luorene was
more mutagenic in the Salmonella/microsome assay than the parent
compounds; the 1-, 2-, 3-, and 4-methylfluorenes were as poor mutagens
as fluorene itself. In the dimethyl series, 1,9-dimethylfluorene was a
potent mutagen in Salmonella TA 100, and the 2,3- and
9,9-dimethylfluorenes were relatively ineffective. Benzofa]fluorene,
benzofb]fluorene, and benzo[c]fluorene were poor mutagens in the
organism, but the 11-methyl derivatives of the first two and the
7-methyl derivative of the latter were more effective. Of this series
of methyl derivatives, ll-methylbenzo[b]fluorene was the best mutagen.
In the phenanthrene series,'' only the 1- and 9-methyl analogues
exhibited greater mutagenicity than phenanthrene itself. Equal
mutagenic activity was manifested by phenanthrene, the 2-, 3-, and
4-raethyl analogues, and the 3,6- and 2,7-diraethyl analogues. The poor
mutagenic activity of anthracene was not altered by substitution of a
methyl group in position 1, 2, or 9.
Tumor-initiating activity of several of these alkylated PAHs was
determined with the mouse-skin two-stage carcinogenesis model. In
a series of fluorene, 9-methylfluorene, 1,9-dimethylfluorene,
benzo[a]fluorene, benzo[b]fluorene, benzo[c]fluorene,
11-methylbenzo[a]fluorene, ll-methylbenzo[b]fluorene, and
7-raethylbenzo[c]fluorene, only 11-methylbenzo[b]fluorene resulted in a
marked increase in tumorigenicity. All other compounds exhibited
rather weak initiator activity.
The methylchrysenes are known respiratory pollutants that occur in
substantial amounts in cigarette smoke—approximately 18
ng/cigarette.62 Although chrysene itself is generally inactive,
several of the methylated species are carcinogenic. In early studies,
Gough and Shoppee" and Dunlap and Warren^ showed that the 1-, 4-,
-------
and 6-methylchrysenes demonstrated only weak tutnorigenicity. The
1,11-diraethvl derivative, however, was moderately active as a skin
carcinogen,** although less so than 3-MC.
Hecht and colleagues studied a series of methylated chrysenes
as both complete carcinogens and initiators. As a complete carcinogen,
5-methylchrysene was far superior to chrysene and the other
monomethylated derivatives; it was almost equivalent in carcinogenic
potency to BaP. 2-Methylchrysene exhibited about 50% of the carcino-
genicity of the 5-methyl analogue. As an initiator, 5-raethylchrysene
was also the most potent of the methylated derivatives, yielding tumors
in 50% of the mice by 14 wk. Next in potency was 3-raethylchrysene.
The 1-, 4-, and 6-methylchrysenes were all much less effective as tumor
initiators. These investigators considered whether 5-methyl-
chrysene, rather than BaP, would be a major contributor to the carcino-
genicity of tobacco smoke; but, in view of its small concentration in
tobacco smoke, compared with that of BaP (0.6 ng/cigarette vs. 30
ng/cigarette), it is unlikely that this is so.
A series of methylated BaPs were tested for tumor-initiating
activity with the mouse-skin carcinogenesis model. "•" Several of the
methylated derivatives exhibited greater initiating activity than the
parent compound, namely, the 1-, 3-, and 11-methyl analogues. Several
were completely ineffective in this regard: the 7-, 8-, 9-, and
10-raethyl analogues. The 4-methyl derivative was about equal to BaP in
initiating potency.
From these examples, it is apparent that some methylated PAHs are
strong carcinogens and therefore should be reckoned with as environ-
mental contaminants.
TOBACCO-SMOKE CARCINOGENESIS
Although the topic has been discussed extensively, several of the
potent carcinogenic PAHs that are present in tobacco smoke should be
mentioned here.
Approximately half of tobacco smoke consists of particulate consti-
tuents in which over 2,000 compounds are represented. The carcino-
genicity of cigarette smoke was demonstrated through skin application
to the backs of mice and the ears of rabbits' and has been confirmed
repeatedly in a number of laboratories. Unfortunately, inhalation
experiments have not led to as clear-cut a conclusion. When Syrian
hamsters were exposed to diluted smoke (smoke-to-air ratio, 1:7) for 10
min twice a day for 18 mo, precancerous lesions were observed in 30%,
esophageal tumors in 5%, and laryngeal carcinomas in 10% of the
animals; no bronchial or tracheal cancer was seen.
Cigarette-smoke condensates have been partitioned into a number of
fractions, of which the most carcinogenic is the "neutral fraction,"
representing 57% of the mass of the condensate. ^ Although the
4-21
-------
weakly acidic fraction (approximately 22 of the condensate mass) con-
tained little carcinogenicity itself, 80% of the tumorigenic property
of the total condensate could be reproduced in conjunction with the
neutral fraction. The neutral fraction was further fractionated by
silica-gel chromatography and partitioning between n-hexane and nitro-
methane into a preparation that contained 0.6% of the total mass, but
much of the carcinogenicity. The active nitromethane preparation was
further fractionated into two components, each of which contained
PAHs. The relative tumor-initiating activity and concentration of
several of these ingredients are shown in Table 4-23. BaP, dibenztah]-
anthracene, benzo[b]fluoranthene, benzo[j]fluoranthene, and dibenz[a]-
acridine—all present in substantial amounts in cigarette-smoke conden-
sate—are potent carcinogens.
Cocarcinogenicity was demonstrated by applying cigarette-smoke
"tar" to the backs of mice in combination with a mixture of 17 major
PAHs found in smoke. The concentration of the 17 PAHs was such as not
to be tumorigenic. However, the combination of tar and PAHs resulted
in tumor formation in 55% of the mice at 13 wk, whereas tar alone
yielded tumors in only 18% of the mice. It should be mentioned that
various neutral fractions obtained from the cigarette-smoke condensate
significantly increased the tumorigenicity of BaP applied topically to
mice.
SUMMARY OF ANIMAL-CELL MUTAGENESIS AND CARCINOGENESIS DATA
Although extracts of automobile emission demonstrate mutagenicity
in both bacterial and animal-cell systems, their carcinogenicity is not
great. Furthermore, attempts to reproduce their pharmacologic activity
by assembling mixtures of the major PAH constituents have not been
successful. The activity of the major PAHs as carcinogens or mutagens
is depicted in summary fashion in Table 4-24. The most potent
derivatives in eliciting a mutagenic response in the Salmonella/
microsome assay are the nitropyrenes. These have not been tested for
carcinogenicity. However, a number of PAHs, such as BaP and some
benzofluoranthenes (Table 4-24), are potent carcinogens. There is
moderate agreement between the mutagenicity and carcinogenicity of the
individual PAHs, although some exceptions are apparent, e.g.,
fluoranthene. A better predictor of carcinogenicity would consist of a
battery of four tests—assays for mutation, chromosomal aberration,
primary DNA damage, and morphologic transformation.
4-22
-------
TABLE 4-1
Categories of Short-Tera Tests3
No. Methods
Test Category Identified
Tests in bacteria and phage 13
Tests in eukaryotic microorganisms 19
Mammalian-cell mutagenesis tests 21
In vitro transformation tests 18
Tests of DNA repair and other effects 14
In vivo tests in mammals 14
Tests in insects 4
Mammalian cytogenetics tests 13
aData from Hollstein j£ jd.^* Many assays detect the same genetic
event, but are considered separate systems because of other
differences, such as in target organism or cell line. Decisions to
regard methods sufficiently distinct to be considered separately are
arbitrary.
4-23
-------
Bacteria tests:
TABLE 4-2
Representative Short-Term Screening Systems
Cytogenetics tests:
Salraonella/microsome test
Poly A test (E. coli)
Yeast tests:
Mitotic recombination or gene
conversion (Saccharomyces
cerevisiae)
Mammalian mutagenesis tests:
Mouse lymphoma TK +•/-
CHO/HGPRT
Chinese hamster V79
BALB 3T3 OuaR
Insect test:
Drosophila sex-linked reces-
sive-lethal test
In vivo tests in mammals:
Sperm-abnormality test
Dominant-lethal test
Mouse specific-locus spot test
Sister chromatid exchange
in vivo
Sister chromatid exchange
in vitro
Chromosomal aberrations in v
Chromosomal aberrations in
vitro
Micronucleus test
Tests of DNA effects and other
effects:
Unscheduled DNA synthesis
(UDS) in human fibroblasts
UDS in hepatocyte primary-
culture/DNA-repair tests
UDS in other target cells
In vitro transformation tests:
Baby hamster kidney cells
BALB 3T3 and other cells
Hamster embryo cells
Enhancement of viral trans-
formation
4-24
-------
TABLE 4-3
Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA)
Short-Term Battery
For detecting gene mutations, three of the following:
Bacterial mutagenesis assay, with and without activation
Eukaryotic microorganisms, with and without activation
Insects (sex-linked recessive-lethal test)
Mammalian somatic cells in culture, with and without metabolic
activation
Mouse specific-locus test
For detecting chromosomal aberrations, three of the following:
In vivo cytogenetics tests in mammals
Insect tests for heritable chromosomal effects
Dominant lethal effects in rodents
Heritable-translocation tests in rodents
For detecting primary DNA damage, two of the following:
DNA repair in bacteria, with and without activation
Unscheduled DNA synthesis (repair test) in mammalian cells,
with and without activation
Mitotic recombination and/or gene conversion in yeast cells,
with and without activation
Sister chroraatid exchange (SCE) in mammalian cells,
with and without activation
4-25
-------
TABLE 4-4
Organisation for Economic Co-operation and Development (OECD)
Test Guidelines
For detection of gene mutations:
Salmonella/microsome assay with and without exogenous mammalian
(S-9) enzyme activity
Mammalian-cell point-mutation assay with and without exogenous
metabolic activity
For detection of chromosomal damage:
In vivo cytogenetics assay in rodents
In vitro cytogenetics assays with and without exogenous metabolic
activity:
Sister chromatid exchange
Chromosomal aberrations
4-26
-------
TABLE 4-5
Genetic Markers Developed in Cultured Mammalian Cells
Resistance to cytotoxic chemicals—e.g., 8-azaguanine
(8-AGR), 6-thioguanine (6-TGR), 5-bromo-2'-deoxyuridine
(BUdR*). ct-amanitin, aminopterin, ouabain
cytosine arabinoaide, diphtheria toxin
Glutamine or asparagine independence
Auxotrophy (e.g., adenine or proline dependence)
Temperature sensitivity (TS)
4-27
-------
TABLE 4-6
Markers for Evaluating Mutagenesis in Cultured Mammalian Cells
Merits Limitations
Purine-analogue resistance
Specificity:
Low spontaneous background
No lethal mutants, because non-
essential pathway involved
Detects both base-pair and frame-
shift alterations, with latter
more efficient
Dominance: X-linked
5-Bromodeoxyuridinea resistance
Specificity:
No lethal mutants, because non-
essential pathway involved
Detects both base-pair and frame-
shift alterations, with latter
more efficient
May detect chromosomal altera-
tions
Expression time: short
Ouabain resistance
Specificity:
ground
low spontaneous back-
Dominance: independence of ploidy
and genotype
Artifacts: minimal
Croas-feeding occurs; need for
refeeding; use of special
selection medium
Influenced by ploidy; must be
heterozygote, because auto-
somal trait; need for preselec-
tion of population before use
Selective responsiveness (reacts
only to ouabain mutagens—
base substitution); limited
spectrum and frequency of
mutants; no simple back-
selection
aTrifluorothymidine also used.
4-28
-------
TABLE 4-7
Comparison of Properties of Some Mammalian Transformation Systems
System
Fischer rat
embryo
(F1706)
Advantages
Hamster in
vitro colony
assay
BALB/3T3
clone A31 or
C3H IOT 1/2
clone B
1. Shown to correlate with
in vivo test results in
double-blind study
2. Easy to discriminate
between normal and trans-
formed morphology
3. Transformed cells do not
need to be cloned before
inoculation into the syn-
geneic host
1. High levels of mixed-
function oxidase activity
2. Diploid chromosome
complement
3. Rapidity and reproduci-
bility of test
4. Characterized independ-
ently in several
laboratories
5. Can be used with meta-
bolic activation
6. Quantitative results
possible
1. Fairly rapid
2. Easy to discriminate
between normal and trans-
formed colonies
3. Well characterized by
independent laboratories
4. Can be made quantitative
5. High correlation between
phenotypic morphology and
tumorigenesis
Possible Disadvantages
1. Not a cloned population
2. Only useful at certain
passage levels
3. Low passage cells must be
preinfected with a type
"C" RNA virus
4. Aneuploid
5. Difficult to quantitate
transformation
1. Low cloning efficiency
2. Heterogeneous population
3. Discrimination between
normal and transformed
morphologies somewhat
subjective
4. Need for pretested,
specific lota of fetal
calf serum
5. Variation in sensitivity
between different embryo
pools
1. Period of usefulness in
terms of sensitivity to
focal transformation by
chemical carcinogens is
unknown
2. Aneuploid
3. Need for pretested,
specific lots of serum
4. Intermediate level mixed-
function oxidase activity
4-29
-------
Table 4-7 (continued)
System
Advantages
Human cells
1. May reflect more
accurately human in vivo
conditions
2. Diploid chromosome
complement (some systems)
Possible Disadvantages
1. Not well characterized
2. Some systems require
genetically aberrant
target cells
3. Experience necessary t<
recognize transformed
phenotypes
4. Usually a very long
latency period
5. Low levels mixed-functi
oxidase activity
4-30
-------
TABLE 4-8
Mutagenic Activity of Various Particulate Emissions and of
Cigarette-Smoke Condensates, Compared with Benzo[a]pyrenea
Mutagenic Activity, revertants/
100 ug of organic material
Source of Emission
Spark-ignition engine
Diesel engine"
Cigarette-smoke condensates
Coke ovens
Roofing tar
Benzo[a]pyrene
Without S-9
138
66
1,225
0
164
0
0
With S-9
342
Unaltered
Unaltered
98
252
99
15,202
aData from Claxton.^S wno used TA98 strains of Salmonella typhimurium
with and without the S-9 activating systems.
''Different values were obtained for various diesel engines; the lowest
and highest are given here.
4-31
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TABLE 4-9
Influence of Alcohol on Direct Mutagenicity of Particulate
Extracts of Spark-Ignition-Engine Exhaust3
Vehicle and Fuel
Ford Escort:
Gasoline
Ethanol blendb
Gasohol
Oldsmobile Cutlass:
Gasoline
Ethanol blend*5
Gasohol
Chevrolet Citation:
Gasoline
Ethanol blendb
Gasohol
Mercury Monarch:
Gasoline
Ethanol blend6
Gasohol
TA 100 Revertants/
g of Extract
10
9
4
10
5
13
17
14
10
16
12
20
Particulate
Emission,
mg/mi
1.5
1.1
1.2
1.7
0.6
0.6
1.9
0.9
1.0
7.1
2.8
1.1
Revertants/
mile
15,000
9,900
4,800
17,000
3,000
7,800
32,300
12,600
10,000
114,000
34,000
44,000
aData from Naman and Clark.
b90X gasoline-IOZ ethanol.
4-32
-------
TABLE 4-10
Mutagenic Activity of Sephadex LH-20 Fractions of Extracts of Diesel Exhaust
Obtained from Combustion of Low-Sulfur and High-Sulfur Fuel3
Mutagenic Activity (Revertants/ug)
Low-Sulfur Fuel
Fraction
% Mass
4^
^ Total extract
LO
Fraction
Fraction
Fraction
Fraction
Fraction
Fraction
1
2
3
4
5
6
70
21
3
1
1
4
With Without
Catalyst Catalyst
1.0
0
0
6
8
3
0
0.5
0
0
8
3
0.3
0
High-Sulfur Fuel
% Mass
59
31
3
1
1
5
With
Catalyst
0.1
0
0
0.6
5.0
0.6
0
Without
Catalyst
0
0
0
0
23
0
0
.5
.4
.8
aData from Hanson e_t^ £l_. ^ Diesel exhaust was collected after combustion of 0.25%-sulfur
(low-sulfur) or 0.772-sulfur (high-sulfur) fuel. Catalytic converter used was of monolithic
oxidation type. Diesel particles were collected in filters and extracted with dichloro-
methane—i.e., total extract. Extract was then fractionated by Sephadex LH-20 chromatography.
-------
TABLE 4-11
Bacterial Mutagenic Activity of PAHs in Kerosene-Soot Extracta
Mutation Contribution, induced
mutant fraction x 10^
Compound
Cyclopenta[cd]pyrene
Pyrene
Benzo[ghi]perylene
and anthanthrene
Coronene
Phenanthrene and
anthracene
Perylene
Benzo[a]pyrene and
benzo[e]pyrene
Uncharacterized
Total CH2C12
aData from Kadin et al.aj
Weight %
in Extract
15
8
8
5
2
2
1
18.3
100
Cone. Extract
= 20 ug/mlb
30
0
2.6
0
0
1.4
0.6
Cone. Extract
= 100
165C
1.7
3.4
105
0
34
3.4
20
150
°Kadin et_ al. determined the amounts of the individual PAHs in the kero-
sene soot and, knowing the mutant fractions that these amounts would
induce from a dose-response curve, were able to estimate mutagenic con-
tributions that compounds would elicit. The induced mutant fraction "
[(no. colonies exhibiting azaguanine resistance in presence of mutagen)/
(no. azaguanine-resistant colonies in absence of mutagen)](dilution
factor).
cBecause of nonlinearity of dose-response relationship, compounds may
contribute differently to mutagenic response, depending on amount of
soot extract and therefore on amount of individual PAH.
4-34
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TABLE 4-12
Mutagenic Efficacy of PAHs in Relation to Benzofa]pyrene3
Concentration, Relative Mutagenic
Potencyb
0.15
0.08
0.50
0.30
0.07
0.05
1.51
0.14
0.20
0
1.00
1.00
0.11
6.00
0.08
0.77
0.08
0
0
0
0
^Relative to benzo[a]pyrene, set at 1.00; rate-limiting factor is
concentration that produced too much cell death.
4-35
Compound
2 -Me thy 1 anthracene
9-Methyl anthracene
1-Methylphenanthrene
2-Methylphenanthrene
Pyrene
1-Methvl pyrene
Cyclopentafcdl pyrene
Benz [a 1 anthracene
Chrysene
1 ,2~Benzodibenzo[hd ] thiophene
Fluoranthene
Benzo [a] pyrene
Benzo[e] pyrene
Perylene
Anthanthrene
Dibenz [ac] anthracene
Dibenz [ ah ] anthracene
Coronene
or Anthracene
or Phenanthrene
or Dibenzo[bd] thiophene
ug/ml
15.4
14.4
15.4
7.7
28.3
17.3
13.9
14.8
10.3
117.0
1.0
1.3
22.7
2.8
12.1
3.6
20.9
51.0
40.0
53.4
55.2
aData from Kadin et al.8J
-------
TABLE 4-13
Mutagenicity of Nitrated Pyrenes in Salmonella typhimurium
Compound
1-Nitropyrene
1 ,3-Dinitropyrene
1 , 6-Dinitropyrene
1 ,8-Dinitropyrene
1,3, 6-Trinitropyrene
1,3,6, 8-Tetranitropyrene
TA 98 and
TA 98b
484
28,600
36,350
75,500
31,400
7,700
TA 98 NRa
TA 98 NRb
35
4,900
37,850
75,500
28,220
5,200
TA 98/TA 98 NR
14
5.8
1.0
1.0
1.1
1.5
aData from Mennelstein £t_ al.
^Strains TA 98 and TA 98 NR are nitroreductase-positive and -negative
respectively.
4-36
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TABLE 4-14
Mutagenicity of Sulfur-Containing Heterocycles3
Material
Control
Benzo[alpyrene
Dibenzo[bd]thiophene
Phenanthro[4,5-bcd]-
thiophene
Benzo[b]naphtho-
[2,1-dlthiophene
Benzo[2,3]phenanthro-
[4,5-bcdlthiophene
Triphenylene[4 , 5-bcd]-
thiophene
Dinaphtho[2,l-b;l' ,
2'-d]thiophene
Pose, yg
0.5
2.0
20.0
0.5
1.0
10.0
0.5
10
10
No. Revertants
(TA 98)b
36
78
277
27
73
85
44
122
42
36
Occurrence;
Tobacco smoke
Coal tar;
carbon black
Coke-plant
effluent
aData from Karcher ^t al.87
''In presence of activating S-9.
4-37
-------
TABLE 4-15
Mutagenesis in CHO Cells3
Concentration Yielding
Comparable Mutation
Addition Frequency, \ig/m\
Positive control—methyl methanesulfonate 175
Diesel-exhaust particulate extracts0 100-275
Spark-ignition engine extract 200
Coke-oven emission 175
aResults extrapolated from data of Casto e_t^ al_.^ CHO cells were
treated with test agent at various doses for 16-24 h. Cells were
collected, and 10^ cells were inoculated into dishes. Mutant cells
were selected for resistance to 6-thioguanine.
bConcentration of test agent yielding mutation frequency of 5 x 10^.
cHxtracts obtained from two different engines.
4-38
-------
TABLE 4-16
Mutation Frequency of Test Agents with Bal/Bc 3T3 Cells3
Mutation Frequency^
Source of Emission Extract"
Solvent control
Positive control, MNNG
(1 ug/ml)
Diesel engine
Spark-ignition engine
Roofing-tar pot sample
Coke oven
Without Activation
0.18
35.5
0.18-1.06d
4.49
3.14
8.17
With Activation
0.26
—
0.20-1
3.97
1.73
__
.81
Benzofa]pyrene — 14.2
(12.5 ug/ml)
aData from Curren et al.
34
^Particles were extracted with dichloromethane. Extract was used
at seven concentrations in mutagenesis assay in absence of metabolic
activation. Dose ranges included: diesel extract, 10-300 ug/ml;
roofing tar, 10-300 ug/ml; spark-ignition engine, 2.5-500 ug/ml; and
coke oven, 10-1,000 yg/ml.
°Number of ouabain-resistant colonies per million viable exposed cells.
^Diesel exhausts from one heavy-duty and two light-duty engines are
included; former yielded lower value.
4-39
-------
TABLE 4-17
Induction of Ouahain- and 8-Azaguanine-Resistant Mutants by PAHsa
Mutants/106 Survivors
Treatment15
Solvent
Pyrene
Phenanthrene
Chrysene
Benz [a] anthracene
Dibenzfac] anthracene
Dibenz [ah] anthracene
7-Methylbenz[a]-
Cloning
Efficiency, %
92
94
79
85
92
95
79
61
Ouabain-
Resis tant
1
1
1
2
2
3
4
24
8-Azaguanine-
Resistant
6
5
8
9
9
22
17
75
anthracene
Benzo[a]pyrene 27 45
(0.3 pg/ral)
7,12-Diraethyl- 50 22
benzanthracene
(0.01 ug/ml)
3-Methylcholanthrene 41 38
(0.3 ug/ml)
128
41
152
aData from Huberman and Sachs.^®
''All compounds added at 1 yg/ml unless otherwise stated.
4-40
-------
TABLE 4-18
Weight Proportion of Various PAHs in a Simulated "AEC" Mixture3
Component Weight, ug
Benzofe]phenanthrene 0.08
Cyelopentenopyrene 1.85
Benz[a]anthracene 0.09
Chrysene 0.21
Benzofb]fluoranthene 0.17
Benzo[k]fluoranthene 0.06
Benzofj]fluoranthene 0.09
Benzo[a]pyrene 0.30
1,12-Methylenebenzo[e]pyrene 0.14
10,ll-Methylenebenzo[a]pyrene 0.05
DibenzofajJanthracene 0.10
IndenoU, 2, 3-od] pyrene 0.21
Dibenz[ah]anthracene 0.02
aData from Misfeld.114
4-41
-------
TABLE 4-19
Carcinogenic Activity of AEC, DEC, and PAHs on Mouse Skin3
Treatment, p g % Tumors Latency Period, wk
Solvent
Benzofa]
AEC:b
DEC:b
Mixture
control
Ipyrene: 3.85
7.69
15.4
290
880
2,630
4,300
8,600
17,150
of PAHs:c
3.5
10.5
0
32
60
89
10
44
83
0
2
12
1
38
.8
.9
.1
.3
.3
.3
.6
.7
.3
.7
—
74
61
44
72
72
52
0
102
76
91
73
aData from Misfeld.114
bObtained with leaded fuel.
cSee composition in Table 4-18.
4-42
-------
TABLE 4-20
Carcinogenic Potency of Various Emissions and PAHsa
Potency,
Substance" pap illomas/mouse-trig
Benzo[a]pyrene 46
Roofing-tar emission 0.2
Coke-oven emission 0.3
Caterpillar diesel exhaust 0
Oldsmobile diesel exhaust 0.1
Nissan diesel exhaust 0.3
Mustang gasoline-engine exhaust 0.1
Cigarette-smoke condensate 0
aData from Slaga ejt al_
158
^Material was applied to mice once as initiator. TPA (2 yg), twice
a week, was used as promoter. Amount of emission condensate that
yielded linear response of tumors vs. dose was used.
4-43
-------
TABLE 4-21
Mixtures of PAHs and Their Proportions3
Carcinogens
Benzo[a]pyrene
Dibenz[ah]anthracene
Benz[c]anthracene
Benzo[b]fluoranthene
Total
Noncarcinogens
Phenanthrene
Anthracene
Fluoranthene
Pyrene
Chrysene
Benzo[e]pyrene
Benzo[gh i]pyrene
Total
Carcinogens + Noncarcinogens
Carcinogens
Noncarcinogens
Total
Amounts , n g
4
27.3
8.5
10.8
13.8
1.2
0.6
3.1
65.3
1
1.0
0.7
1.4
0.9
4.0
5
81.0
25.5
32.4
41.4
3.6
1.8
9.3
195.0
8
4.0
65.3
69.3
2
1.7
1.2
2.4
1.5
6.8
6
243.0
76.5
97.2
124.2
10.8
5.4
27.9
585.0
9
6.8
110.5
117.3
3
3.0
2.1
4.2
2.7
12.0
7
729.0
229.5
291.6
372.6
32.4
16.2
83.7
1,755.0
10
12.0
195.0
207.0
aData from Schmahl et
4-44
-------
TABLE 4-22
Carcinogenicity of PAHs in Combination3
Treatment, ug % Papillomas + Carcinomas
Solvent 0
Benzo[a]pyrene:
Carcinogens:
Noncarcinogens:
Carcinogens + noncarcinogens:
aData from Schmahl et al.
149
1.0
1.7
3.0
4.0
6.8
12.0
65
195
585
1,755
69
117.3
207.0
14
28
56
36
68
71
1
0
1
17
50
60
70
4-45
-------
TABLE 4-23
Tumor-Initiating Activity of Cigarette-Smoke Ingredients3
Relative Tumorigenic Concentration
Compound _ Activity _ ng/cigarette
Benzo[a]pyrene +++ 10-50
5-Methylchrysene +++ 0.6
Dibenz[ah]anthracene +•»• 40
Benzofb] fluoranthene ++ 30
Benzo[ j] f luoranthene ++ 60
Dibenz[a]acridine ++ 3-10
Indeno[ 1 ,2 ,3-cd] pyrene + 4
Benz[a]anthracene + 40-70
aData from Hoffmann e£ al.72
^Mouse-skin carcinogenesis.
4-46
-------
TABLE 4-24
Summary of Carcinogenicity and Mutagenicity of PAHs in
Various Emissions
Compound
Anthracene
2- or 9-Methylanthracene
1,2-, 1,3-, 1,4-, or 2,3-
Diraethylanthracene
9,10-Dimethylanthracene
Phenanthrene
1- or 2-Methylphenanthrene
Fluoranthene
2- or 3-Methylfluoranthene
Pyrene
1- or 2-Methylpyrene
1-Nitropyrene
1,3-Dinitropyrene
1,6-Dinitropyrene
1,8-Dinitropyrene
1,3,6-Trinitropyrene
1,3,6 ,8-Tetranitropyrene
Cyclopenta[cd]pyrene
Benz[a]anthracene (BA)
1-, 3-, 4-, 5-, or 11-Methyl-BA
2-Methyl-BA
6-, 7-, 8-, 9-, or 12-Methyl-BA
1,7-, 1,12-, 2,9-, 2,10-, 3,9-,
3,10-, 4,2-, 4,12-, 5,12-,
or 8,11-Dimethyl-BA
4,5-, 6,7-, 6,8-, 6,12-, 7,8-,
7,11-, 7,12-, 8,9-, or 8,12-
Dimethyl-BA
9,10- or 9,ll-Dimethyl-BA
Fluorene
9-Methylfluorine
Acridine
Anthanthrene
Chrysene
1-Methylchrysene
2-, 3-, 4-, or 6-Methylchrysene
5-Methylchrysene
Benzo[b]fluoranthene
Benzofj]fluoranthene
Benzo[ghi]perylene
Carcinogenic
Activity3
0
0
0
0/ +
0
0
0
+
0
0
0/ +
0
0
0
0
0/+
•f
Relative In Vitro
Mutagenic Activity"
Animal Bacteria
0 0
+ 0
-t-n-
•n-
•t-f
Kd +d
4-47
-------
TABLE 4-24 (contd)
Compound
Carcinogenic
Activity3
In Vitro
Mutagenic Activityb
Ajiimal Bacteria
Benzo[k] fluoranthene
Benzo[a]pyrene (B[a]P)
2-, 3-, 4-, 6-, 11-, or 12-
Methyl-B[a]P
5-Methyl-BfalP
8-Methyl-B[a)P
1,2-, 1,3-, 1,4-, 1,6-, 2,3-,
3,6-, 3,12-, or 4 ,5-Diraethyl-
B[a]P
Benzo[e] pyrene
Perylene
3-Methylcholanthrene
IndenoU ,2 ,3-cd] pyrene
Dibenzt ah] anthracene (DBA)
2-, 3-, or 6-Methyl-DBA
7-Methyl-DBA
Coronene
Benz[alaeridine
Dibenzo[bd] thiophene
Dibenzf ac]anthracene
0/ +
0
++
+
•f
+
0
•n-
•»•
0/+
0
•»•
aO, no tumors ; +, tumors in up to 33% of animals; ++, tumors in over 33Z
of the animals.
hBenzo[ a] pyrene rautagenicity set at ++.
c7-Methyl-BA.
d7,12-Dimethyl-BA.
4-48
-------
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Cyclopenta(£d)pyrene and acepyrene. Environ. Sci. Technol. 9:
143-145, 1975.
176. Waller, R. E. The benzpyrene content of town air. Brit. J Canc§
6:8-21, 1952.
177- Wang, Y. Y., S. M. Rappaport, R. F. Sawyer, R. E. Talcott, and
E. T. Wei. Direct-acting mutagens in automobile exhaust. Cancer
Lett. 5:39-47, 1978.
178. Whong, W-Z., J. Stewart, and T-M. Ong. Use of the improved
arabinose-resistant assay system of Salmonella typhimurium for
mutagenesis testing. Environ. Mutagen. 3:95-99, 1981.
179. Williams, G. M. The detection of chemical mutagens/carcinogens
by DNA repair and mutagenesis in liver cultures, pp. 61-79. In A.
Hollaender and F. DeSerres, Eds. Chemical Mutagens: Principles an>
Methods for Their Detection. Vol. 6. New York: Plenum Press, 191
180. Wolff, S. Sister chromatid exchange. Ann. Rev. Genet. 11:
183-201, 1977.
181. Wood, A. W., W. Levin, R. L. Chang, M-T. Huang, D. E. Ryan,
P. E. Thomas, R. E. Lehr, S. Kumar, M. Koreeda, H. Akagi,
Y. Ittah, P. Dansette, H. Yagi, D. M. Jerina, and A. H. Conney.
Mutagenicity and tumor-initiating activity of cyclopenta(c,d)-
pyrene and structurally related compounds. Cancer Res. 40:
642-649, 1980.
182. Wynder, E. L. , and D. Hoffmann. A study of air pollution carci.no-
genesis. III. Carcinogenic activity of gasoline engine exhaust
condensaste. Cancer 15:103-108, 1962.
183. Wyrobek, A. J., and W. R. Bruce. Chemical induction of sperm
abnormalities in mice. Proc. Natl. Acad. Sci. USA 72:4425-4429, 19
184. Yasuhira, K. Damage to the thymus and other lymphoid tissues
from 3-methylcholanthrene, and subsequent thymoma production, in
mice. Cancer Res. 244558-569, 1964.
4-62
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5
EFFECTIVE BIOLOGIC DOSE
In the class of polycyolic aromatic hydrocarbons (PAHs), there are
several chemicals that are environmental pollutants; some are carcino-
genic in experimental animals, and some are suggested to be carcino-
genic in humans. ° In the body, they are enzymatically converted to
reactive forms that bind extensively and covalently to cellular macro-
molecules. °3,80,151,182 .jftg covalent binding of reactive metabolites
of PAHs to DNA is considered to be an essential first step in PAH
induction of neoplasia.63'80'8^.128,151,182 The damaged DNA cannot
be fixed and results in a mutation within the cell unless enzymatic
repair occurs first. There are many pharaacokinetic and enzymatic
processes involved before the formation of reactive metabolites of
PAHs, which may ultimately form adducts with DNA. Thus, the
concentration to which a person is exposed is probably not a good
measure of the biologic dose that causes neoplasia^i^l* 149 or other
PAH-induced toxicoses (see Chapter 4). This chapter develops the theme
that some degree of PAH metabolite-DNA adduct formation in the target
tissue can be used as a measure of effective biologic dose. The
effective biologic dose of a substance is a reflection of its
absorption, distribution, metabolism (activation or detoxification),
and excretion. In the case of an alkylating substance, such as a'PAH,
that dose can be measured directly on the basis of the amount of
alkylated DNA, itself a reflection of adduct formation. If the
accumulation of adducts in DNA is greater than the capacity of the
tissue to repair such lesions accurately and greater than the capacity
of the tissue to replicate its DNA, then the presence of adducts will
be indicative of the effective biologic dose. The chapter begins with
a brief discussion of the phannacokinetics of PAHs. That is followed
by a discussion of the metabolism of selected PAHs. The in vivo
formation and disappearance of PAH metabolite-DNA adducts are next
reviewed in detail. Finally, there is a discussion of the possibility
of using PAH metabolite-DNA adduct content as a measure of effective
biologic dose for in vitro mutagenesis, initiation of carcinogenesis,
and inhibition of replication and transcription.
PHA&MACOKINETICS
Many pharmacokinetic and enzymatic processes are involved before a
PAH reaches a target cell and is metabolized to reactive metabolites
that interact with DNA and other cellular macromolecules (see Figure
5-1). The oxidative metabolism of PAHa is usually by cytochrome P-450,
and the formation of excretable glutathiorie, glucuronide, and sulfate
conjugates results in a very complex metabolic profile. Thus, pharma-
cokinetic information that would enable one to construct mathematical
5-1
-------
models of the tissue distribution, metabolism, covalent binding to
cellular macromolecules, and excretion of PAHs and metabolites as func-
tions of exposure dose are nonexistent. However, sufficient studies
have been done to allow some generalization regarding absorption, tis-
sue distribution, and elimination of PAHs (see Santodonato et al., ^Q
pp. 6-1 through 6-27). Most of these studies have only followed radio-
activity in various tissues, urine, and feces after administration of
radiolabeled PAHs.
PAHs are readily absorbed after administration by various routes
and are then rapidly removed from the blood and distributed into a
variety of body tissues. Kotin e^ al_. ^ examined the radioactivity
derived from C-labeled benzo[aJpyrene (BaP) in various tissues of
rats and mice after intravenous, subcutaneous, and intratracheal
administration. The blood concentrations resulting from intravenous
injection were hardly detectable after 10 min. Radioactivity was found
in stomach, intestine, liver, kidney, lung, spleen, testis, myocardium,
urine, and feces. The pattern of distribution was independent of route
of administration, except that particularly high lung concentrations
followed intratracheal administration. These workers did not examine
fat or mammary gland. Other investigators have shown that nonmetabo-
lized BaP,19 3-methylcholanthrene (3-MC),19>4° and dimethylbenz[a]-
anthracene (DMBA)1''58 accumulate and persist more in fat and mammary
tissue than in other tissues. Some PAHs induce neoplasia in the
mammary glands of rats.
Rees e£ £!.• examined the mechanisms by which BaP and other
PAHs are absorbed from the gut. Accumulation of BaP in everted sacs of
small intestine increased exponentially with incubation-medium
concentration. The transport of BaP from the sac tissue to the inside
medium was found to be proportional to the concentration in the sac
tissue. Thus, if the capacity of other tissues to absorb BaP from
extracellular fluid (and blood) is proportional to the concentration of
BaP in the fluid, then accumulation in the tissues should also be
proportional to intragastric concentration. For example, this
relationship was observed in adipose and mammary tissue .18 h after oral
administration of BaP. Rees e_t_ al. postulated a mechanism of physical
adsorption onto the intestinal mucosal surface and then passive
diffusion into and through the intestinal wall. The proportional
nature of the accumulation in the tissue can be accounted for by two
phases of adsorption, one unilayer and the other multilayer. Even if
tissue accumulation of PAHs is proportionally related to exposure dose,
these results should not be overinterpreted. The situation is dynamic;
the accumulation is transient, in that PAHs are rapidly metabolized and
removed from the body. Rees et_ al. observed that BaP disappeared very
rapidly from the thoracic duct lymph. Moreover, PAH metabolite-DNA
adduct content in various tissues is not linearly related to exposure
dose (as discussed later).
A relevant route of environmental exposure to PAHs is deposition in
the lung of particles with PAHs on their surfaces. In general, the
degree of retention of PAHs in the lung is a function of the size and
5-2
-------
composition of the particles carrying them. Several investigators have
shown that BaP retention by the lung is higher when it is adsorbed on
particulate carbon, ' dust, ferric oxide, aluminum
oxide, and talc^- than when it is not; carbon-particle size
affects BaP retention, but the size of particulate ferric oxide or
aluminum oxide does not. However, some recent studies have sug-
gested that particulate adsorption of PAHs does not alter retention
time in the lung or their distribution to other tissues. Adsorption on
ferric oxide did not increase the retention time of BaP in hamster lung
after intratracheal instillation.5'' Pylev e_t al. ^5 examined the
clearance of intratracheally instilled BaP from the hamster lung; the
disposition and clearance from liver, kidney, and blood; and excretion
into feces and urine. BaP was instilled alone or adsorbed on asbestos
or carbon black. Although these studies were limited in scope, it was
found that the disposition of BaP from lung to other tissues, the rate
of tissue clearance of BaP, and the pattern of BaP excretion were not
altered by the introduction of BaP into the hamster either in free form
or bound to particles. Obviously, more studies on rates of clearance
from the lung and the later fate of particle-adsorbed PAHs are needed
to clarify the effects of particle size and composition. However, it
can be concluded that distribution to other tissues occurs after
pulmonary exposure to particles on which PAHs are adsorbed.
Elimination of PAHs in animals occurs mainly by excretion of con-
jugated metabolites into the feces.4'23'109»16l»162 There is some
excretion of metabolites into the urine—approximately 10% in the study
by Kotin et_ a^. • Excretion into bile can be very rapid. For
example, 6 h after intravenous injection of [ H]BaP. 60-70% of the
tritium appeared in bile or conjugated metabolites. ^ PAH clearance
from an animal probably is not limited by metabolic rates or biliary
clearance of metabolites, but rather is affected by the persistence of
nonmetabolized compound in various tissues (such as fat, skin, and
mammary gland) or perhaps by adsorption on particles.
The pharmacokinetics of a PAH will be influenced by prior treatment
with chemicals capable of inducing enzyme systems that metabolize it.
Schlede ot_ al_.^>^^ have shown that pretreatment of rats with
unlabeled BaP markedly increased the plasma-disappearance rate of a
tritiated dose of the same compound given intravenously; the effect was
especially marked during the first 5 min after the intravenous
administration of the radiolabeled material, and increased clearance
lasted for 6 h. This effect of pretreatment with the compound was
paralleled by a lower concentration of [ H]BaP in brain, liver, and
fatty tissues; similar but more variable results were observed in lung
tissue. These influences of BaP pretreatraent on a later intravenous
dose of [^H]BaP were also observed when the radiolabeled compound was
administered orally. 3-MC and DMBA pretreatment of animals produced
comparable effects on the metabolic disposition and tissue content of
radiolabeled BaP. Pyrene and anthracene pretreatment had little or no
such effect on the in vivo disposition of this compound, nor did
phenobarbital. In other studies, the biliary excretion of [^
5-3
-------
was shown to be increased by pretreatment with the unlabeled compound-
however, ao increase in excretion of the C-labeled metabolites of
BaP into bile was observed after pretreatment with this compound.
These findings suggest that conversion of BaP to its metabolites maybe
the rate-limiting step in its biliary excretion.
METABOLISM OF PAHs
An organism's processing of xenobiotic chemicals is determined by
their physical and chemical characteristics. Figure 5-2 summarizes the
possible events leading to carcinogenesis in a cell exposed to a
xenobiotic toxic chemical. After uptake, the cell may simply excrete
the chemical unchanged, as is the case with some metals and apparently
inert materials, such as asbestos. A toxicant may contain functional
attachment groups, such as hydroxyl or ketone, that can be conjugated
to deactivating moieties like glutathione or glucuronic acid by
cytoplasmic transferase. If the toxicant is a PAH or other relatively
stable molecule, it will be attacked by the microsomal raonooxygenases
and form an electrophilic intermediate, which can later be conjugated
to a deactivating moiety, detoxified, and excreted.
Once an activated electrophile is formed, it can readily attack
nucleophilic sites other than the detoxifying substrates, such as
nucleic acids and proteins. The formation of adducts between
electrophile metabolites of PAHs and DNA is probably a necessary first
step in the initiation of carcinogenesis by PAHs. The in vivo
formation' of PAH metabolite-DNA adducts is discussed later in this
chapter.
These biochemical changes to biologically active intermediates
depend on the balance between enzyme systems: those enzymes generating
and those detoxifying the intermediates. One of the major enzymes
involved in activation is aryl hydrocarbon hydroxylase (AHH). It is
found in virtually all eukaryotes (and some prokaryotes), has a wide
range of specificities for substrate activity, uses a variety of
iron-containing cytosolic pigments as the active sites for chemical
oxidation (e.g., cytochrome P-450), and is substrate-induc-
ible.6J«134 Many PAHs are capable of inducing one or more forms of
cytochrome P-450. There is some evidence that induction is regulated
by one gene or a relatively small number of genes in animal-model
systems ' and perhaps even in humans (see Chapter 7). The basis
for genetic regulation appears to reside in a balance of inducers and
receptors that are activated by PAH metabolites; after binding, trans-
location to the nucleus, expression of induction-specific RNA, and
protein synthesis, the generation of specific cytochrome P-450 is
observed. In the murine-model systems, genetically controlled AHH
activity is correlated with cancer formation caused by PAHs, such as
BaP,110 3-MC,110 dibenz[a]anthracene,1U and DMBA.110
5-4
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Examples of enzymes that can detoxify these metabolic intermediates
are UDP-glucuronosyltransferase, glutathione-S-epoxide transferase,
aryl sulfatase, and epoxide hydrase. These enzymes catalyze the
conjugation of the primary oxidative species formed as a result of AHH
activity to forms that are sufficiently polar to be excreted from cells
and from the body. Some of the conjugating enzymes are also under a
form of genetic control, *^ but their role in PAH carcinogenesis is
not completely defined. Epoxide hydrase is one of the enzymes that had
been thought to function in a manner that results in the detoxification
of PAHs; however, it is now established that, for a variety of PAHs,
epoxide hydrase can catalyze the formation of dihydrodiol derivatives
of PAHs and that these diols may serve as substrates for monooxygenase
activity again--resulting in the formation of diol-epoxides.63 The
diol-epoxides constitute at least one of the ultimate mucagenic and
carcinogenic forms of PAHs.
Over the last decade, BaP has been the most exhaustively studied
PAH carcinogen and has been the prototype compound in developing the
mechanism of action of the cellular monooxygenase and cytoplasmic
transferases necessary to activate and detoxify PAH carcinogens. A
recent exhaustive summary of BaP metabolism dealt with its activation,
carcinogenesis, and role in the regulation of mixed-function oxidases
and related enzymes.^3 A composite of metabolic products of BaP is
shown in Figure 5-3. BaP has been studied in a large number of in vivo
and in vitro systems, as well as in cell-free preparations using
homogenates, microsomal fractions, and purified enzymes. BaP may form
epoxides at several sites around its ring system, and three epoxides
(4,5-,7,8-, and 9,10-) have been identified. Research over the last
half-decade has implicated the 7,8-diol (bay region*)^ as the
primary precursor for the second round of activation by mixed-function
oxidases, both cytoplasmic and nuclear,80 that form the highly
electrophilic 7,8-diol-9,10-epoxide (Figure 5-4), which opens to form a
triol carbonium intermediate. This reactive molecule has been shown to
be the major species that binds to nucleic acids via the C-10 position
of BaP and to exocyclic amino groups of guanine.
Metabolism of many PAHs other than BaP has also been shown to
proceed via diol-epoxides, such as benzf ajanthracene , l'*» *-°°
chrysene,116'189 dibenz[ah]anthracene,l'° 5-methylchrysene.81
7-methylbenzanthracene (7-MBA),35-127 DMBA,16,46,91,132,179 and
3-MC. 1"5,179 ^jje ease Of formation of carbonium ions by these
diol-epoxides parallels the observed biologic activity of the parent
chemicals. " Metabolic profiles on some PAHs other than BaP are
available, and salient features of their metabolism are presented below.
*The bay region is a molecular region between adjacent fused aromatic
rings (see reference 115).
5-5
-------
BENZO[e]PYREME
B*nzo[e]pyrene is a marginally carcinogenic structural isomer of
the strong environmental carcinogen BaP. It contains two bay regions
and, by theoretical calculations, should approximate BaP in carcino-
genic activity.93 Metabolic studies have determined that the prob-
able reason for its lack of carcinogenicity is that its major metabo-
lism is distal to the bay region, so that the molecule does not favor
formation of diol-epoxide intermediates. Its metabolism has been
studied in hamster embryonic cells and in cell-free preparations froa
rat liver. Its major metabolite is 4,5-dihydro-4,5-dihydroxybenzo-
[ejpyrene. Large-scale experiments with microsooes positively identi-
fied 9,10-dihydro-9,10-dihydroxybenzo[e]pyrene, but it constituted leu
than 12 of the total metabolites.
4,5-Dthydro-O-
dihydromy-
»,10-blhydro*y-
5-6
-------
PYRENE
Early studies on pyrene metabolism were in rats and showed
increased urinary excretion of sulfuric acid esters and glucuronic acid
conjugates.^3 Later, l-hydroxypyrene and 1,6-dihydroxypyrene were
identified.78 More definitive studies of pyrene metabolism were
performed in rabbits and rats by analysis of urinary metabolites after
intraperitoneal injection. ^2 NO direct structural analysis was
performed, but the results of a number of chromatographic and spectral
analyses were compared with synthetic standards. 1-Hydroxypyrene, 1,6-
and 1,8-dihydroxypyrene, 4,5-dihydro-4 ,5-dihydroxypyrene, and N^acetyl-
£-(4,5-dihydro-4-hydroxy-5-pyrene)-L-cysteine were identified." The
latter compound was also isolated from bile in rats. More recent
studies with gas-liquid chromatography and mass spectroraeCry have con-
firmed the presence of 1-phenolic and 1-dihydroxydihydro derivatives
from rat-liver microsomal incubation and shown a marked increase in
mutagenesis in the Salmonella I 90 and I 100 strains.79
HO H
4,5-Olhydro-
4.S-dihydroxypyr«n«
OH
N — •c*cylcyic«ln<
HO'
1,6-Dlhyd r oxypyrene
l,8-Dlhydroxypyr«n«
BENZ(a 1ANTHRACENE
Benz[a]anthracene is a marginally carcinogenic PAH that has both
bay-region and K-region areas. The original metabolic studies with
benz[a]anthracene were done with thin-layer
chromatography.21,26-28,75,169,171 (see reference 115 for explana-
tion of the K-region.)
5-7
-------
HO H
1.2-Dihydro-l,J-
dlhydroiy-
3,*-Dthydro-J,4-
dlhydroxy-
H
OH
8,»-DihyATO-*
HO
S.6-Dihydro-5,6.-
-------
sulfate and glucuronic conjugates at the 3, 4, 8, and 9 positions,
presumably as products of phenols; and 10,ll-dihydro-10,ll-dihydroxy-
benz[a]anthracene were also reported. The 3,4-dihydro-3,4-dihydroxy
derivative of benz[a]anthracene was later confirmed with high-pressure
liquid chromatography as a metabolite formed by rat-liver micro-
somes.* ' This 3,4-dihydrodiol adjacent to a bay region leads to the
idea of benz[ajanchracene bay-region activation, including the possi-
bility of an isolated double bond in the 1,2 position after formation
of a diol-epoxide. That 3,4-dihydro-3,4-dihyroxybenz[a]anthracene is a
minor product quantitatively, as opposed to the less active 5,6-diol,
may explain the weak carcinogenicity of benz[a]anthracene.
CHRYSENE (1,2-BENZOPHENANTHRENE)
This molecule is composed of two linearly annellated rings formed
by pyrocondensation of carbonaceous material and is therefore present
in coal tar in substantial quantities.^ Metabolism of chrysene has
been studied in rodents and in cell-free and organ-culture systems.
Incubation with rat-liver microsomes produced a series of hydroxylated
Chry««n«
OH
5-9
-------
metabolites, as seen by high-pressure liquid-chromatographic (HPLC)
separation. Several of these metabolites have been identified with the
use of synthetic standards.122 Three dihydrodiols have been char-
acterized: the 1,2-, 3,4-, and 5,6-dihydrodiols. This metabolic
profile has been concerned with the use of rat or mouse skin-organ
culture. ^ The dihydrodiol metabolites are presumably formed
through reactive epoxide intermediates by the P-450 mixed-function
oxidases. However, no phenolic or quinoid structures have been
identified from the remaining peaks in the HPLC separation."9
5-METHYLCHRYSENE
Of all the methylchrysenes studied, only 5-methylchrysene shows any
substantial carcinogenicity. Metabolism of this compound has been
studied in the 9,000-g supernatant from rat liver.81-83 Liver homo-
genates used for this work were prepared from Aroclor-treated male
F-344 rats, and HPLC of metabolites showed nine peaks, of which seven
had been identified (according to their relative abundance) as
5-hydroxy-5-methylchrysene, 5-methylchrysene 1,2-diol, 7-hydroxy-5-
methylchrysene, 5-methylchrysene 9,10-diol, 9-hydroxy-5-methylchrysene,
l-hydroxy-5-methylchrysene, and 5-methylchrysene 7,8-diol. Two minor
metabolites have not been identified. The bay-region theory would
predict that 5-methylchrysene 1,2-diol and 5-methylchrysene 7,8-diol
are primary candidates for active carcinogenic intermediates. However,
experiments with liver homogenates indicated that formation of 5-methyl-
chrysene 1,2-diol is favored over that of 5-methylchrysene 7,8-diol.
No other biologic system has been used to study metabolism of 5-methyl-
chrysene, so it is not possible to make any pertinent comparisons with
other tissues or between intact-cell activation and detoxification.
5-10
-------
H.
7,8-Dihydro-
7,8-dihydroxy-
5-Hydroxy-
CH3
9,10-Dlhydro-9,10-dihydroxy-
FLUORENONE
There appear to have been no direct studies on the metabolism of
fluorenone. but fl-3-fluorenyl acetamide (3-FAA) yielded two metabolic
products.102 Because the parent compound is carcinogenic, it appears
that the derivatives are detoxification products. The authors sug-
gested that metabolism of 3-FAA consists of two sequential reactions:
the initial formation of 9-hydroxy-3-FAA as an intermediate to the for-
mation of 3-acetamido-9-fluorene hydroperoxide, which is then dehydrog-
genated to form 9-oxo-3-FAA. Exposure of rainbow trout to a number of
hydrocarbons showed no bioaccumulation of fluorenone; the compound is
most likely metabolized to excretible products."' However, no at-
tempt was made to analyze any metabolic products. There is no litera-
ture on the isolation and identification of fluorenone metabolites.
5-11
-------
9-Fluorenone
H H
N-3-fluorenyl »c«t*mide (3 FAA)
3-Ac« t»mld«-J-f luorint-
hydroptroxld*
METHYLFLUORENE
There is no literature on the metabolism of methylfluorene, but
there haa been a major study on methylfluorene-2-acetic acid (MFA)
(Cycloprofen, Squibb Institute for Medical Research). This compound is
an anti-inflammatory agent whose metabolism has been studied in rats;
its metabolites have been isolated from urine and identified. Its
major metabolite is substituted at the 7 position on the aromatic ring,
so its metabolism may be similar to that of methylfluorene. This
congener was given both orally and intraperitoneally. Analysis of the
metabolites by thin-layer chromatography yielded six peaks, of which
four have been identified. The major metabolite, consisting of 472 of
the material, was 7-hydroxy-MFA, with approximately lOt each of
9-hydroxy-MFA and 7,9-dihydroxy-MFA.42.i!2
5-12
-------
H«thylfluor«n«
Mtthylfluona*-
2-*c«tlcacld
CYCLOPENTA[cd1PYRENE
Cyclopenta[cd]pyrene has been identified as • component of carbon
black.66i»9 It§ Bcc^boliim has been studied in rat-liver micro-
•omea.^^ The major metabolite isolated by HPLC has been identified
as trans-3,4-dihydroxy-3,4-dihydrocyclopenta[cd]pyrene. Several com-
ponents not yet well characterized consisted presumably of phenolic
derivatives, as well as metabolitea that appear to have saturation of
the ethylene bridge.
5-13
-------
Cyclop«nt«[cd]pyr«n«
Tr«ni-3,4-dihydroxy-
3,4-dlhydrocyclop«nti(cd|pjr
DIBENZOTHIOPHENE AND BENZOTHIOPHENE
Dibenzothiophene and benzothiophene are biodegradable in both
eutrophic and oligotrophic pond waters.™ Their major metabolite is
l,2-dihydro-l,2-dihydroxydibenrothiophene, with later ring degradation
to benzothiophenedione. Benzothiophene and dibenzothiophene form a
thioketone, a dihydrodiol (cia and trans isomers and a diketone).
There have been no studies dealing with further metabolism.
Enzymes other than microsomal monooxygenases may also be involved
in the metabolic activation of PAHs. Eling ££ £l_." and Marnett^5
have shown that numerous xenobiotics, including the dihydrodiol
metabolites of PAHs, can be cooxygenated during the oxygenation of
arachidonic acid by prostaglandin synthetase. In the case of PAHs,
when the dihydrodiols are generated, this novel pathway could lead to
an alternative pathway for the formation of diol-epoxides. These
studies have been done on in vitro model systems. The relevance of
their pathway in vivo ia unknown.
There is a suggestion in the literature that nitropyrenea are
metabolized by bacteria, presumably via a nitroreductase, to produce a
high mutagenicity; however, there has been no isolation or character-
ization of metabolites. Because mammalian cells have much-reduced
nitroreductase activity, this rationale has been used to explain the
lack of activity in mammalian cells.^^
5-14
-------
H OH
B*nxothloph*n«dlonc
Bcnsothiophanc
5-15
-------
IN VIVO FORMATION AND DISAPPEARANCE OF PAH METABOLITE-OKA ADDUCTS
HISTORICAL PERSPECTIVE
Brookes and co-workers observed chat ^H-labeled PAHs applied to
the backs of mice or incubated with mouse-embryo celLs resulted
in covalent binding of radioactivity to DMA, RNA, and protein. Grover
and Sims and Gelboin showed that PAHs require metabolic activa-
tion by mixed-function oxidases if they are to react covalently with
cellular macromolecules. The interactions between PAH metabolites and
nucleic acids have since received considerable
attention.63'80'151'182 Identification of the reactive PAH metab-
olites that form adducts with DNA has been emphasized, because forma-
tion of these adducts is believed to be essential for tumor initiation,
although interaction with RNA and protein may also be important.
Initial attempts to identify the reactive metabolites that bound to
DNA focused on the arene oxide intermediates proposed by Boyland,~5
and especially on the K-region arene oxides, because, according to
quantum mechanics, carcinogenic PAHs are distinguished by an electron-
rich K region.10'153 They induce malignant transformation of cells
and are active in mutagenicity tests. ' In addition, they are
metabolites of PAHs70'72~74'l04'l63'17° and will bind to DNA in
1 8
vitro. However, it became obvious that K-region epoxide-DNA
adducts were not the adducts formed in vivo between BaP metabolites and
18 17
DNA. A similar conclusion was reached in studies with 7-MBA.1'
Borgen et a_l_- found that, in a microsomal activating system,
the 7,8-diol BaP metabolite bound to DNA to a much greater extent than
any other known diol or phenol. Sims et_ al_. provided evidence
that the BaP raetabolite-DNA adduct formed by BaP metabolism by Syrian
hamster-embryo cells in culture was chromatographically identical with
an adduct formed by metabolism of BaP 7,8-diol and proposed that the
reactive metabolite was a diol-epoxide (DE). Studies in various in
vivo model systems, such as cell cultures and organ
1^1R?Q/17QQT17?
explants, ' ' > ' ' ' and in in vivo skin, lung, Liver, and
forestomach of mice have shown that the major BaP metabolite-DNA adduct
observed after exposure of these tissues to BaP is the (+)-BaP DE
I-deoxyguanosine adduct. (+)-BaP DE I is apparently the major enzy-
matic metabolite of (-)-trans BaP 7,8-diol. The adduce results
from the interaction of (+)-BaP DE I with two amino groups of guanine.
The cis isomer (-)-BaP DE II is also formed enzymaticai.lv from
(-)-trans-BaP 7,8-diol.1" The (-)-BaP DE II-deoxyguanosLne adduct
is formed to the same extent as the (+)-BaP DE I-deoxyguanosine adduct
in lung and liver of rabbits, as opposed to the results in -nice (C.
Bixler and M. W. Anderson, unpublished data). Structures of the BaP DE
isomers are shown in Figure 5-5. Although the predominant binding of
BaP DE is to the 2-amino gf°UP ^f guanine. these diol-epoxides can also
bind to the N7 of guanine,140 adenine,98• " '13°•175 and
cytidine175 and to phosphate residues.60'108
5-16
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Q C 107
Evidence is accumulating that other PAHs~e.g., 7-MBA ->J'l^'
benzanthracene,174'188 chrysene 116'189 5-methylchrysene 81
dibenzanthracene,190 3-MC,1°5.179 and OMBA1*.^! ,1.32,179__are
similarly converted to highly reactive diol-epoxides, which then
interact with DNA in vivo. All these diol-epoxides have a structural
similarity. Jerina and Daly9 pointed out that the epoxide ring is
in a bay region and suggested the term "bay-region diol-epoxides'* for
these highly reactive metabolites of PAHs.
The bay-region diol-epoxides are
mutagenic,89'95.115.117.1",136,174,185,186 have trangforining
activity in mammalian cells, »^° are carcinogenic in newborn
miceJl,100,101,117 and Chinese hamster living cells.l^7'187,and are
initiators in cells of mouse skin. 31'-55 >9b> 100> 101>115 >117 »173 The
mutagenicity and carcinogenicity, combined with the observation that
bay-region diol-epoxide-DNA adducts are the major adducts formed in
vivo i-n target tissue, have led to the hypothesis that bay-region
diol-epoxide adducts are the ultimate carcinogens generated by
metabolism of most PAHs.96'115 However, it should be pointed out
that many nonraetabolized PAHs that are termed carcinogenic either lack
a bay-region benzene ring or contain nonreactive substitutes in this
molecular region. In addition, PAH metabolite-DNA adduct formation
in vivo has been examined only for BaP, DMBA, and 3-MC, and these
studies have concentrated on target tissues in mice (see Table 5-1).
PAH metabolites other than bay-region diol-epoxides can also bind
to DNA. The K-region epoxide of BaP binds DNA covalently.168>17°
Incubation of BaP with microsomes in the presence of exogenous DNA
results in a variety of BaP metabolite-DNA adducts.12•14B In
particular, adducts are formed from further metabolism of 9-hydroxy-
BaP, possibly the 4,5-epoxy-9-hydroxy-BaP metabolite.12'106 The
major DNA adduct observed after exposure of hepatocytes in culture to
BaP resulted from the further metabolism of 9-hydroxy-BaP.9' A
BaP-phenol-oxide-DNA adduct was the major adduct observed in rat lung
and liver after intravenous administration of BaP. 3 Various
structural modifications of PAH diol-epoxide metabolites do not inhibit
binding to DNA.80'86'87
Dose-response relationships for formation of PAH metabolite-DNA
adducts in target tissue would be helpful in the low-dose extrapolation
problem for PAH carcinogenesis.'.61 Many pharmacokinetic processes
determine the extent of formation of PAH metabolite-DNA adducts in an
organ after exposure of an animal to a PAH (see Figure 5-1). Although
most of these processes have not been completely characterized, some
generalizations regarding the extent of adduct formation in vivo can be
made from recent reports (see Table 5-1). Previous reviews of covalent
f PAHs to DNA have not analyzed in vivo adduct formation in
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CHARACTERIZATION OF PAH METABOLITE-DNA ABDUCTS
Table 5-1 lists the studies concerned with the in vivo formation of
PAH metabolite-DNA adducts. Most of them used mice and BaP.
HPLC analysis of BaP-deoxyribonucleoside adducts formed in lung,
liver, and forestomach of A/HeJ mice after oral administration of BaP
(3 mg/mouse) is shown in Figure 5-6. • 1°4 HPLC analysis is shown for
DNA samples isolated by the hydroxylapatite and precipitation pro-
cedures. Radioactivity eluted in the water wash (water fraction, WF)
and in early portions of the water:methanol gradient that varied from
40% to 70%. This uncharacterized early-eluting radioactivity was much
higher in DNA samples isolated by precipitation than by the hydroxyl-
apatite method (see Figure 5-6), although it was still substantial,
especially in liver and forestomach, in samples isolated by the
hydroxylapatite procedure. Three distinct peaks—I, II, and III in
Figure 5-6—were observed in the gradient portion of the chroma-
tography. The specific activity (picomoles per milligram of DNA) asso-
ciated with these peaks is independent of the procedure used to isolate
DNA. Peaks II and III have been identified as (+)-BaP DE I-deoxy-
guanosine and BaP DE II-deoxyguanosine adducts, respectively. Peak
I is probably generated from 9-hydroxy-BaP, although it could be a BaP
DE I-deoxycytosine adduct. Small, late-eluting peaks were con-
sistently observed, especially in lung samples (Figure 5-6). They
could be BaP DE I-deoxyadenosine or BaP 4,5-oxide adducts. Similar
BaP metabolite-DNA adduct profiles were observed in lung, liver, and
forestomach from ICR/Ha and C57BL/6J mice after oral administration of
BaP. ''^ Eastman e_t^ £!_• examined the in vivo binding of BaP to
DNA in lung, liver, and kidney of Aroclor 1254-treated A/J mice after
intravenous administration of BaP. The only identified adducts
observed by Sephadex LH20 chromatography were BaP DE-DNA adducts.
Early-eluting radioactivity was present in the chromatograph.
Eastman and Bresnick^O and Eastman e_t a_l.51 used Sephadex
LH20 chromatography to analyze the 3-MC metabolite-DNA adduct profile
in lung and liver of several mouse strains after intravenous admin-
istration of 3-MC (Figure 5-7). Two major 3-MC-deoxyribonucleoside
adduct peaks were observed in lung and liver of each mouse strain
examined. HPLC analysis^ demonstrated seven 3-MC metabolite-DNA
adduct peaks in lung and liver of C57BL/6J mice, with the two major
adduct peaks corresponding to those observed by Sephadex LH20 chroma-
tography.^'^ Early-eluting peaks (Figure 5-7) were also present in
the chromatographs of these studies with 3-MC.
Binding of 3-MC, BaP, and DMBA to DNA has been examined in
skin of several mouse strains (Table 5-1). In each study with BaP,
the major adduct observed was BaP DE-deoxyguanosine. Sephadex LH20
chromatography revealed only one adduct peak. The HPLC adduct profile
in skin was virtually identical with that in Figure 5-6 for lung,
5-18
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forestomach, and liver.U>14>107>149 The adduct profile for 3-MC
in mouse skin was the same in each strain examined, but was slightly
different from that for lung and liver (Figure 5-7), in that three
3-MC metabolite-deoxyribonucleoside peaks were observed in the
Sephadex LH20 chromatograph of akin, whereas only two peaks were
observed for lung and liver. * ^> Sephadex LH20 chromatography
of DMBA-deoxyribonucleoside adducts in skin was similar in each mouse
strain studied; three peaks were observed. ^ Early-eluting peaks
were also observed in the chromatographs of these investigations of PAH
metabolite-DNA adduct formation in mouse skin.
The formation of DNA adducts of the carcinogen 15,16-dihydro-
ll-methylcyclopenta[a]phenanthrene-17-one (11-methvlketone) was
examined in liver, lung, and skin of TO raice.^'2' HPLC analysis
revealed eight 11-methylketone metabolite-DNA adduct peaks in each of
the tissues. The major adduct was generated from the interaction
between the anti-3,4-dihydro-3,4-trans-dihydroxy-l,2-dihydro-l,2-
epoxide (diol-egoxide) metabolite of 11-methylketone and deoxy-
guanosine.l» ' There is no major qualitative difference in the
adduct pattern among the three tissues. The adduct profile for each of
the three tissues was not substantially altered by the route of
administration (intramuscular, topical, and'intraperitoneal).
In the studies with mice, the PAH metabolite-DNA binding profiles
are very similar in all tissues of all strains examined. For BaP, the
predominant characterized adduct is the BaP DE I-deoxyguanosine
adduct. When HPLC analysis was used, a BaP DE II-DNA adduct was also
observed, as well as an adduct probably generated from 9-hydroxy-BaP.
There is also evidence of BaP DE-deoxyadenosine adducts, although in
relatively small amounts. The DNA adduct profiles for 3-MC are the
same in lung and liver of each mouse strain examined and only slightly
different in skin. The pattern of DMBA metabolite-DNA adducts in skin
is the same for all strains examined. The HPLC profiles for 11-methyl-
ketone metabolite-DNA adducrts are very similar in lung, liver, and
skin, with the major adduct being a diol-epoxide metabolite-
deoxyguanosine adduct.
The in vivo formation of BaP metabolite-DNA adducts has recently
been examined in male Sprague-Dawley rats and male New Zealand rabbits
(Table 5-1). In rats, BaP was administered intravenously at 1.0 and
10.0 ymol/kg. Several chromatographically distinct nucleoside-bound
adducts were observed in lung, whereas only one adduct was apparent in
the liver. The predominant BaP-nucleoside adduct formed in vivo
in rat lung and liver was chromatographically identical with adducts
formed on further metabolism of BaP phenols, possibly because of the
interaction of 9-hydroxy-BaP 4,5-oxide with DNA.6»12'23 The BaP DE
adducts were not detected in rat liver, and only a relatively small
amount was observed in rat lung. The BaP DE adducts in rat lung
accounted for only 1.4% of total DNA binding and 3.32 of the adducts
generated by BaP phenol(s). Thus, the in vivo BaP metabolite-DNA
adduct profiles obtained in lung and liver of Sprague-Dawley rats are
5-19
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distinctly different from those observed in various mouse strains.
This is the only known case in which the BaP DE adduct is not the
predominant BaP metabolite-DNA adduct formed in vivo.
In an examination of BaP metabolite-DNA adduct profiles in lung and
liver of male New Zealand rabbits, BaP was administered either orally
or intraperitoneally at 50 mg/kg. The DNA adduct profiles were
identical with those in mice (Figure 5-6), with one notable difference
(Bixler and Anderson, unpublished data): In rabbits, there was
approximately 75% as much BaP DE II-deoxyguanosine adduct as BaP DE I
adduct, whereas in mice, there was only 10Z as much BaP DE II adduct as
BaP DE I adduct. This is the only known case in vivo in which the BaP
DE II-deoxyguanosine adduct approaches the BaP DE I adduct in amount.
It should be emphasized that, in these investigations of PAH
raetabolite-DNA adduct formation, large amounts of the DNA-associated
radioactivity chromatographed not as nucleoside-bound adducts, but
rather as uncharacterized fast-eluting peaks (Figures 5-6 and 5-7).
Although Adriaenssens £££!.• showed that isolation of DNA by a
hydroxylapatite procedure, instead of by the precipitation method,
significantly reduces the amounts of" these early-eluting peaks, the
peaks still account for a large proportion (especially in liver and
forestomach) of the total radioactivity eluted in chromatography. Some
workers have ignored these early-eluting peaks by a pre-elution step
with Sephadex LH20 chromatography (e.g., Figures 5-6 and 5-7). These
peaks are also observed in in vitro studies and in in vivo model
systems (see Boroujerdi ^t £l. ). The radioactivity appears to
reflect some tissue-specific reactions, such as those exhibited by the
different patterns in lung and liver. "•-'I Eastman and Bresnick^
showed, by using borate-eluted Sephadex LH20 and DEAE-Sephadex
chromatography, that early-eluting radioactivity contains numerous
constituents. Studies that used [14C]BaP and [3H]BaP5'141>l49
and the results of Eastman and Bresnick^" and Eastman &t_ al.
suggested that only a small amount of this radioactivity is due to
tritium exchange, whereas experiments of other investigators^1^'
suggested the opposite. The results of Eastman and Bresnick^"
suggested that only a small amount of the radioactivity in the early
peaks is related to oligonucleotides. Phosphotriesters might
contribute to the early-eluting radioactivity. ' In any case,
because a considerable amount of radioactivity appears in the early-
eluting peaks, their identification deserves further consideration.
COMPARISON OF EXTENT OF PAH METABOLITE-DNA ADDUCT FORMATION BETWEEN
TISSUES AND BETWEEN SPECIES
Specific activities (SAs), in picomoles per milligram of DNA, of
PAH metabolite-DNA adducts have been determined in several tissues
after administration of PAHs. Table 5-2 gives the SAs of BaP DE
adducts in lung and liver of various mouse strains and New Zealand
rabbits. The amounts of BaP DE adducts are very similar in lung and
5-20
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Q
liver for each study reported in Table 5-2. Anderson et al.
examined the BaP DE adducts in lung, liver, and forestomach of A/HeJ
mice for oral administration of BaP at 2-1,350 ymol/kg. The SAs of the
BaP DE adducts in lung and liver were similar over the entire dose
range and ranged over 3 orders of magnitude in this study. The SA of
BaP DE adducts in forestomach of mice is very similar to that in lung
^ n 90
and liver after oral administration of BaP. ' ' The- similarity of
the BaP DE adduct amounts in lung, liver; and forestoraach is rather
surprising, inasmuch as the disposition and rate of metabolism of BaP
in these tissues are probably very dissimilar. The higher rate of BaP
metabolism in liver is reflected in the greater total DNA binding and
protein binding in liver, compared with lung and forestomach.
The SAs of the BaP phenol-oxide adduct were also very similar in
lung and liver of the Sprague-Dawley rat for each intravenous dose of
BaP (Table 5-3). As mentioned previously, this was the predominant
adduct observed in lung and liver of this species. As with mice,
the total DNA binding was significantly higher in liver.
Eastman and Bresnick^ examined 3-MC metabolite-DNA adduct
formation (Figure 5-7) in lung and liver of several mouse strains at
several points after intravenous injection of 3-MC (12.6 rag/mouse).
Mixed-function oxidases were induced with Aroclor 1254 24 h before
injection of 3-MC. The amounts of adducts were significantly higher in
lung than in liver in each mouse strain and at each time (Table 5-4).
Total DNA-associated radioactivity in liver is not significantly
different from that observed in lung. Thus, the relative binding of
3-MC to DNA of lung and liver of Aroclor 1254-treated mice is
distinctly different from that of BaP in untreated mice. The ratio of
3-MC-DNA binding in liver to that in lung is smaller than the ratio for
BaP for both nucleoside-bound adducts and total DNA-associated
radioactivity (Tables 5-3 and 5-4). At present, BaP and 3-MC are the
only PAHs for which the amounts of PAH metabolite-DNA adducts can be
compared between lung and liver.
It should be emphasized that the SAs for the in vivo studies
reported in Table 5-1 are calculated on the basis of the total DNA in
the organ. These values for the BaP DE adducts (Table 5-2) do not
differentiate between lung and liver and therefore do not appear to
offer any explanation for susceptibility of the lung and resistance of
the liver to BaP-induced neoplasia in, for example, A/HeJ and A/J
mice. However, it is likely that the amounts of adducts formed in
different cell types vary considerably. This possibility has the
greatest implication for organs, such as the lung, that contain a
multitude of cell types. Although little is known about the
localization of carcinogen-DNA adducts in lung, cytochrome
p-450-dependent monooxygenase enzymes appear to be much more localized
in lung than in liver.*^» 1", ^°° The nonciliated bronchiolar
epithelial (Clara) cells of rabbit lung have been identified as having
high concentrations of these enzymes—a finding that correlates with
the observed pulmonary toxicity of 4-iporaeanol, which is thought to
5-21
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result from high covalent binding of a reactive metabolite to proteins
in the Clara cells of a number of species. The Clara cells amount
to only 1% of the pulmonary cells. The high concentration of
cytochrome P-450 in Clara cells may also be important in
nitrosamine-induced pulmonary carcinogenesis. Because the average
SAs of BaP DE-DNA adducts are similar in lung and liver, the above
considerations suggest that adduct contents might be much higher in
some pulmonary cell types than in hepatocytes. Also, the removal rates
of the BaP DE-DNA adducts might vary considerably in different cell
types in the lung. Examination of adducts in individual cell types
might allow differentiation of tissues with respect to susceptibility
and resistance to PAH-induced neoplasia.
Results of several studies allow comparisons to be made between
amounts of PAH metabolite-DNA adducts in the same tissue from different
species (strains). BaP DE adducts and 3-MC metabolite-nucleoside
adducts (Figure 5-7) can be compared in lung and liver of different
strains of mice under the same dosage regimen (Tables 5-2 and 5-4).
The BaP DE adduct amounts are very similar in lung of A/HeJ and
C57BL/6J mice and 10 times smaller in ICR/Ha mice (Table 5-2). A/HeJ
mice exhibit high susceptibility, ICR/Ha moderate susceptibility, and
C57BL/6J high resistance to BaP-induced pulmonary neoplasia.
Obviously, the SAs of the BaP DE adducts do not differentiate between
species in susceptibility to BaP-induced neoplasia. Also, the
disappearance rates of BaP DE adducts are similar in C57BL/6J and A/HeJ
mice. In contrast, Eastman and Bresnick claimed that their
results regarding 3-MC metabolite-nucleoside adducts formed and their
disappearance rates do differentiate between 3-MC-induced pulmonary
neoplasia in the various mice strains (Table 5-4). It is hard to
understand why adduct contents or their disappearance rates based on
total organ DNA would differentiate between mouse strains on the basis
of pulmonary susceptibility to one PAH and not another. Again,
examination of SAs in individual pulmonary cell types might unravel
this dilemma.
Phillips et_ £1^.150 examined the covalent binding of DMBA, 3-MC,
and BaP to DNA in the skin of mice of various strains. Neither the
amounts of PAH metabolite-DNA adducts nor their disappearance rates
showed a correlation with the reported susceptibilities of the strains
to PAH-induced skin carcinogenesis. Thus, the results of this study
based on average cellular SAs of the organ are in agreement with the
BaP, but not the 3-MC, study of pulmonary adducts in the different
mouse strains. '^» Ashurst and Cohen confirmed the results of
Phillips et±I'15° with HPLC analysis. Baer-Dubowska and
Alexandrov examined the binding of BaP to skin of rats and mice
under conditions known to initiate tumorigenesis in the skin of mice.
The patterns of BaP metabolite-DNA adduct profiles were identical in
the two species and very similar to those in Figure 5-6. The amounts
of the BaP DE-deoxyguanosine adducts and total DNA-associated radio-
activity were 3 times higher in mouse skin. The adduct difference
probably does not differentiate between the mouse susceptibility and
rat resistance to BaP-initiated skin tumorigenesis.
5-22
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DOSE-RESPONSE RELATIONSHIPS OF PAH METABOLITE-DNA ADDUCTS
Although there have been a considerable number of in vivo studies
on PAH metabolite-DNA adduct formation (Table 5-1). there are only
three reported dose-response studies of amounts of PAH raetabolite-DNA
adducts .
Phillips e£ al.150 treated C57BL mice topically with DMBA at
0.025-1.0 umol/raouse. DMBA metabolite-DNA adducts in skin were
determined by Sephadex LH20 chroraatography. Three DMBA metabolite-
deoxynucleoside adduct peaks were present at each dose, and the SAs
plotted in Figure 5-8 are the sums of these three peaks.
Pereira e_t a_l. ' examined the formation of epidermal BaP-DNA
adducts in ICR/Ha mice at topically applied doses of 0.01-300
mg/mouse. The SAs in Figure 5-9 are essentially linear with dose
throughout the dose range; the log-log plots have slopes of
approximately 1. Peak I in Figure 5-9 represents early-eluting peaks,
and peak III in Figure 5-9 represents the BaP DE-deoxyguanosine adducts.
o
Anderson et al. investigated the dose dependence of BaP
metabolite-DNA adducts in the lung, liver, and forestomach of A/HeJ
mice (Figure 5-10). BaP was administered orally at 0.048-29.7
uraol/raouse, and an-imals were sacrificed 48 h later. The SAs plotted in
Figure 5-10 are for the BaP DE-deoxyguanosine adducts. Similar
dose-response curves are obtained for the early-eluting peaks (WF and
IP in Figure 5-6) and the BaP-phenol-oxide adduct (Peak I in Figure
5-6). The curves in Figure 5-10 are either linear with a slope greater
than 1 and concave downward or linear with a slope of 1 and concave
upward (Figures 5-8 and 5-9, respectively). This means that the
percentage of the dose that becomes bound to DNA as BaP DE adducts
decreases as the dose decreases in the tissues of the A/HeJ mice
(Figure 5-10), whereas the percentage of the dose that becomes bound to
epidermal DNA is constant or actually increases as dose decreases
(Figures 5-5 and 5-6, respectively). However, the values of the BaP DE
adducts in lung and forestomach of A/HeJ mice at the lowest dose
examined are only approximately 50% lower than those predicted by
simple proportion from the adduct values at the highest dose (Figure
5-10). Thus, the results of these dose-response studies do not reveal
the existence of any threshold dose below which binding of PAH
metabolites to DNA does not occur.
EFFECT OF AHH INDUCERS ON PAH METABOLITE-DNA ADDUCT FORMATION
The effect of AHH inducers on the in vivo binding of BaP to DNA has
been examined in several tissues of various mouse strains. »'
In a study by Wilson et al., adducts were determined under
conditions known to resuTt in inhibition of BaP-induced pulmonary
5-23
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neoplasia by g-NF. Female A/HeJ mice were fed either a g-NF diet (0.3J
of diet) or a control diet for 16 d and then given BaP (6 mg/mouse).
The animals were returned to a control diet for 2 wk and then placed on
either a g-NF diet or a control diet for another 16 d. This regimen
results in a 94% reduction in BaP-induced pulmonary adenomas. ^ In
the adduct study, [3H]BaP (6 mg/mouse) was administered to one group
(A) of animals, both control and g-NF-treated, on the sixteenth day of
the first NF feeding. ^ Another group (B) of mice, control and
g-NF-treated, were given unlabeled BaP (6 mg/mouse) on the sixteenth
day of the first g-NF feeding and [^H]BaP (6 mg/raouse) on the six-
teenth day of the second g-NF feeding. Animals in both groups were re-
turned to control diets after administration of [^H]BaP and
sacrificed 48 h later. Table 5-5 gives the percentage decrease in BaP
DE-DNA adduct formation in the lung and liver of g-NF-treated mice.
The decrease in the amount of BaP DE-DNA adducts in the lung, 86-93%,
appears to correlate with the inhibition of pulmonary adenoma
formation (94%).181>184 No BaP DE-DNA adducts were detected in the
liver of g-NF-treated mice. Wilson et_ j_l. also examined the
effects of two other AHH inducer's, TCDD and Aroclor 1254, on BaP-DNA
adduct formation in lung and liver of A/HeJ mice. These inducers,
like 6-NF, markedly decreased the formation of the BaP DE adducts in
lung and liver (Table 5-5). It is interesting that TCDD and Aroclor
1254 had little effect on or actually increased total DNA-associated
radioactivity in the early-eluting peaks. In a similar study, loannou
ej: £l.9° investigated the effect of 6-NF treatment on BaP DE adduct
formation in ICR/Ha mice. g-NF treatment also resulted in a
significant reduction, 80-90%, in BaP DE adducts in lung, liver, and
forestomach of this strain.
Cohen et_ a_l.^ examined the effect of TCDD treatment on BaP
DE-DNA adduct formation and BaP-induced tumor initiation in the skin of
SENCAR and CD-I mice. BaP DE adduct formation in skin was completely
inhibited in both strains of mice, and BaP-initiated papilloma forma-
tion was inhibited by 93%. Total DNA-associated radioactivity and
covalent binding of BaP to skin protein were increased (Table 5-6).
The increase in total DNA-associated radioactivity was due to the
increase in the amounts of the early-eluting peaks. Again,
inhibition of BaP-induced tumor formation correlates with inhibition of
BaP DE adduct formation in the target tissue, but not with total DNA-
associated radioactivity on covalent binding of BaP to protein. In a
continuation of their studies on the anticarcinogenic effects of TCDD,
DiGiovanni e£ *i« showed that the time course for induction of
epidermal AHH and UDPGT correlated with the time course for the
inhibitory effects of TCDD on tumor initiation with DMBA, BaP, and 3-MC.
Thus, AHH inducers inhibit in vivo BaP DE-DNA adduct formation in
every tissue of every mouse strain examined. The effects of AHH
inducers on BaP DE-DNA adduct formation in vivo contrast markedly with
their effects in vitro, i.e., in lung and liver microsomes, isolated
peripheral lung and liver tissue slices, and hepatocytes. Treatment of
animals with AHH inducers stimulates the formation of BaP DE adducts in
5-24
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vitro (see Wilson £t..il' for references). The effect of AHH
inducers on total DNA-associated radioactivity and on covalent binding
of BaP to protein is similar in vivo and in vitro. The reason for the
disparity between the in vivo and in vitro results for BaP DE adduct
formation is unclear — in particular, the relationship between induction
of AHH and decreased in vivo BaP DE adduct formation.
Induction of AHH has been postulated as a mechanism for the anti-
c action of a wide variety of
, 90,154,180,181, 184 The resultg of wilgon et
al. 4 and Cohen £t £l.36 discussed above suggest that AHH
inducers, such as g-NF and TCDD, inhibit BaP-induced neoplasia by
reducing the amount of BaP DE-DNA adducts formed in the target tissue.
The data of Pyerin e± a_l. also suggest that prior AHH induction is
a protective mechanism against DMBA-induced tumor initiation in mouse
skin. The concept of a protective effect of prior treatment with AHH
inducers against PAH-induced tumor initiation might appear to be
incompatible with the suggestion that AHH inducibility and PAH-induced
tumor initiation may be causally related.64' 103 • 133 > l35 • 143' L47> 17°
Relationships between AHH inducibility and PAH-induced tumor
susceptibility have been examined in a great variety of tissues and
animals, including man. It was proposed that humans with intermediate
and high AHH inducibility carry a higher risk for lung and laryngeal
cancer than persons with low inducibility. ^3, 1/8 However, these
findings could not be confirmed. In various strains of mice, AHH
inducibility correlates well with susceptibility to PAH-induced
carcinogenesis for some PAHs but not for others.^4' l33 ' *•"» 147
However, none of these studies allows any assessment of whether AHH
induction in the target tissue preceded PAH-induced tumor initiation.
Even in cases in which positive trends exist between AHH inducibility
and susceptibility to PAH-induced tumor initiation, prior treatment
with AHH inducers will probably protect against tumor initiation by the
PAH. Whether the prior induction of AHH in the target tissue is part
of the protective mechanism is unclear.
IN VIVO DISAPPEARANCE OF PAH METABOLITE-DNA ADDUCTS
Mutations and malignant transformations of cells by chemicals may
be a consequence of DNA synthesis on parent-strand templates containing
unexcised chemically induced lesions. Thus, the ability of a cell to
repair the damaged DNA by an error-free pathway could constitute a
critical protective mechanism against mutagenesis and carcino-
genesis.^'12^ However, Feldman e_£ a_l.55>5° and Shinohara and
Cerutti reported the persistence of BaP DE-DNA adducts in human
alveolar tumor cell A549 and secondary mouse embryo fibroblast,
respectively. Dipple and Roberts also noticed the persistence of
7-bromomethylbenz[a]anthracene metabolite-DNA adducts during
replication in cell cultures. Cerutti e_t £l..33'34 observed that
significant fractions of BaP metabolite-DNA adducts were still present
in the DNA when the alkaline-elution profile of the parent DNA had
5-25
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returned to normal. The persistent lesions could result from the loss
of excision repair during prolonged incubation or from the adducts1
becoming part of a portion of DNA that cannot be repaired by the
excision pathways.
There have been only a few studies on the in vivo disappearance of
PAH metabolite-DNA adducts. Before these results are discussed, it
should be emphasized that the interpretation of in vivo rates of
disappearance of carcinogen-DN^ adducts has several limitations. When
the SAs of the adducts are based on total DNA content in an organ, the
disappearance rates are average cellular values. The disappearance
rate in the target cells could be masked by the rates in nontargec
cells, especially if the target cells are only a small fraction of the
total cells in the organ. In addition, in vivo rates of disappearance
of specific activities of adducts cannot be unequivocally equated with
enzymatic excision rates, because cell turnover will also result in a
decrease in the SA of the adduct.
Eastman and Bresnick-^ examined the disappearance of 3-MC
raetabolite-DNA adducts in lung and liver of several mouse strains
(Table 5-4). In each strain examined, the adduct decreased
significantly faster in liver than in lung. The adducts are more
persistent in the lungs of A/J and C3H/HeJ mice—strains susceptible to
3-MC-induced pulmonary adenomas—than in the lungs of highly resistant
strains. Whether these differences in disappearance rates and initial
amounts of 3-MC metabolite-DNA adducts are causally related to
differences in tissue and species susceptibility to 3-MC-induced
tumorigenesis is unclear.
Anderson and Wilson' examined the disappearence of BaP
metabolite-DNA adducts in lung and liver of A/HeJ and C57BL/6J mice
(Table 5-7). The disappearance rates in A/HeJ were examined at oral
doses of 0.011 and 6.0 mg/mouse. The amount of BaP DE-DNA adducts in
lung and liver decreased monoexponentially with time at each BaP dose.
Although the initial BaP DE-DNA adduct amount at the higher dose was
more than 1,000 times larger than that at the lower dose in lung
(liver), the half-life of BaP DE adducts in lung (liver) was similar at
the two doses (Table 5-7); thus, there is no apparent threshold dose
for removal of BaP DE adducts in lung and liver of A/HeJ mice. BaP
DE-DNA adducts had a biphasic decay in the lung and liver of C57BL/6J
mice (BaP at 6 mg/mouse). Also, the BaP DE adducts in the lung and
liver of C57BL/6J mice could be approaching a constant, nondecaying
value. Examination of adduct amounts at longer times after the initial
dose will be required to determine whether persistent adducts exist.
The half-life of the adducts in the terminal phase was similar to that
observed in the A/HeJ mice (Table 5-7). In contrast with the results
of Eastman and Bresnick5 with 3-MC, these data do not appear to
offer any explanation for the tissue and species differences in
susceptibility to BaP-induced neoplasia.
5-26
-------
The in vivo disappearance of 11-raethylketone metabolite-DNA adducts
was examined in lung, liver, and skin of TO mice for 14 d after initial
treatment. Rate of DNA turnover was also examined over this same
time span. Rates of disappearance of the major adducts in skin and
lung could not be measured above the normal rate of DNA turnover,
whereas, in the liver, adducts were removed rapidly (half-life, about
2.5 d) relative to the DNA turnover rate. Thus, enzymatic repair of
adducts would be occurring in liver, whereas adducts might be
persistent in some cell types of skin and lung.^ Again, whether
these apparently persistent adducts of skin and lung are causally
related to the initiation of neoplasia in these tissues by 11-methyl-
ketone awaits further investigation.
The disappearance of PAH metabolite-DNA adducts has also been
examined in skin of mice. Phillips e£ £l.^^ studied the
disappearance of DMBA metabolite-DNA adducts in skin of several mouse
strains. The half-life of the adducts was 1-2 d in each strain
examined. These results do not explain the strain difference in
susceptibility to DMBA-induced neoplasia in skin of mice. Similar
conclusions were obtained with BaP and 3-MC, although adduct amounts
were examined at only two points. ^ Rayman and Dipple"° examined
the formation and disappearance of 7-bromomethylbenz[a]anthracene and
7-bromoraethyl-12-methylbenz[a]anthracene metabolite-DNA adducts in skin
of Swiss S mice. The adduct amounts of the less carcinogenic
7-broraomethylbenz[a]anthracene were higher and required a longer time
to reach maximums—4 d vs. 1 d—after a single topical application of
1 uraol. For both chemicals, the adducts decayed rapidly from the
maximum, the half-lives being less than 24 h. The data do not differ-
entiate between the carcinogenic potency of these two PAHs in mouse
skin. Pelkonen et_ £l.. examined the disappearance of BaP
metabolite-DNA adducts in skin and subcutaneous tissue of C3H and
C57BL/6 mice. No strain difference was observed in the disappearance
in either strain of the BaP 4,5-oxide or the BaP DE adducts. Thus,
rates of disappearance of the adducts do not differentiate between the
C57BL/6 resistance and the C3H susceptibility to BaP-initiated
subcutaneous fibrosarcomas.
In summary, no generalization regarding these data on in vivo
disappearance of PAH metabolite-DNA adducts can be made now. In some
investigations, extrapolation of the adduct-time curve suggests
complete removal of the adducts, whereas adducts appear to persist in
other studies. Similar results were obtained in'studies of adduct
removal in cell cultures. And adduct removal rates differ
significantly between tissues susceptible and resistant to PAH-induced
neoplasia in some cases and not in others. It is possible that the in
vivo results on adduct disappearance rates based on total organ DNA
content are misleading, in that PAH metabolite-DNA adducts might be
persistent in some cells that constitute only a small fraction of the
total cell population.
5-27
-------
PAH METABOLITE-DMA ADDUCT AMOUNTS AS A MEASURE OF EFFECTIVE
BIOLOGIC DOSE OF SOME PAH TOXICITY
It has been suggested that the amount of carcinogen-DNA binding is
a measure of the effective dose of a carcinogen. This proposal is
consistent with the somatic-mutation theory of tumor initiation by
chemical carcinogens. Although the extent of carcinogen-DNA binding
has been successfully used to rank a series of carcinogens, such as
*J A *5 7 Q Q
PAHs and alkylating agents, for carcinogenic potency, '• there
have been few, if any, serious attempts to use the amount of
carcinogen-DNA adducts formed in the target tissue as a predictive tool
for low-dose extrapolation in carcinogenesis. Anderson e_t^ a_l- and
Gehring and Blau *• suggested a general scheme for the incorporation
of pharmacokinetics in low-dose risk estimation for chemical
carcinogenesis. This section discusses the feasibility of using some
measure of chemical-induced DNA damage as an effective biologic dose of
a carcinogen.
IN VITRO MUTAGENESIS
Several recent studies have examined the extent of PAH-induced DNA
damage in cells undergoing mutation and have attempted to relate the
PAH-induced mutation frequencies to the amounts of PAH metabolite-DNA
adducts. Wigley e_t a_l. and Newbold e_t al_. examined the
mutagenicity of benz[a]anthracene, 3-MC, DMBA, BaP, and 7-MBA in a
cell-mediated mutagenesis system, using BHK 21 cells to metabolize Che
PAH and Chinese hamster (V-79) cells as targets for mutation. The
frequencies of PAH-induced mutation were not significantly different at
equivalent amounts of PAH metabolite-DNA adducts. Yang et al.^3
examined the induced-mutation frequency in normal fibroblasts and
XP12BE cells as a function of the number of BaP DE-DNA adducts in the
cells when they were released from confluence and plated for the
expression of 6-thioguanine resistance. The mutation frequency is
linearly related to the BaP DE I-DNA adduct amount when the cells were
released from confluence. Fahl e£ a_l. examined the induction
of Hist* reverse-mutation frequencies by BaP DE I, BaP DE II, and
9-hydroxybenzo[a]pyrene in Salmonella typhimurium cells (TA 98 and TA
100) as a function of the number of bacterial DNA bases modified by the
electrophilic BaP metabolites. The induced-mutation frequencies were
linearly related to the DNA adduct amounts in each case (Figure 5-11).
Newbold et. &±-^^ investigated the mutagenicity of BaP DE I and
BaP DE II in V-79 cells as a function of BaP DE-DNA adduct amount in
the V-79 cells. At sublethal doses of the diol-epoxides, the
induced-mutation frequencies were linearly related to the DNA adduct
amounts in the cells for both BaP DE I and BaP DE II. The relation of
mutation frequency to DNA adduct amount, however, becomes exponential
at toxic amounts of BaP DE-DNA adducts. Newbold and Brookes'^ also
5-28
-------
observed a linear relationship between induced-mutation frequency and
the concentrations of BaP DE I in the medium as long as that
concentration was sublethal.
Mutation induction by PAHs, as well as by other mutagens and UV
radiation, appears to result from the cells' attempt to deal with the
unexcised chemical-induced DNA lesions at the time of replication.
Moreover, induced-mutation frequencies are linearly related, at least
at sublethal doses of the mutagen, to the amounts of mutagen-DNA
adducts present when the cells are undergoing replication. This linear
relationship makes DNA adduct amounts a good measure of the effective
biologic dose of mutagens in in vitro assay systems. Comparisons of
the mutagenic potency of PAHs should be based on the slope of the curve
relating induced-mutation frequency to PAH raetabolite-DNA adduct
amounts.
CARCINOGENESIS
In addition to the results obtained in in vitro mutagenic assay
systems, the in vivo results discussed below suggest that PAH
metabolite-DNA adduct amounts in a known target tissue are a good
measure of the effective biologic dose of PAHs for initiation of
neoplasia.
Several investigations have shown a positive correlation
between the carcinogenicity of a series of PAHs of widely differing
carcinogenic gotencies and their extent of reaction with
DNA.30'32'68'88'150 This correlation was not observed with the
binding of the reactive metabolite of the PAHs to protein and
30 flfi
RNA. >° Similarly, the binding of S-propiolactone and similar
alkylating agents to DNA, but not to RNA and protein, corresponds to
their tumor-initiating potency.37 Total DNA binding (picomoles of
radioactivity associated with DNA per milligram of DNA) was used to
rank the carcinogens for carcinogenic potency. Recent studies have
suggested that amounts of specific carcinogen-DNA adducts should be
used as a measure of effective dose, instead of total DNA binding. For
example, for the nitrosaraines and nitrosamides, correlation between
carcinogenicity and nucleic acid alkylation has been observed only with
06 alkylation of guanine, and not with N7 alkylation of guanine, even
though the latter alkylation is approximately 10 times greater than the
former. 21,124,144 Current evidence suggests that the bay-region
diol-epoxides , such as BP DE, are the ultimate carcinogenic forms for
most PAHs. The use of specific carcinogen-DNA adduct amounts, instead
of total DNA binding, should improve the ability to rank a series of
carcinogens.
Studies with inhibitors of carcinogenesis have shown that tumor
response changes quantitatively with PAH metabolite-DNA adduct
•36,90,184
amount. •,, Swann et ^l.. showed that changes in the
incidence of dimethylnitrosamine-induced kidney tumors produced by
5-29
-------
changes in the diet and by treatment with BaP correspond to the changes
that these treatments produce in the alkylation of the target-tissue
DNA by dimethylnitrosamine.
Janss and Ben^2 found a correlation between the amount of DMBA
bound to DNA and the incidence of mammary tumors in rats of different
ages.
Several studies have shown that, although the time for appearance
of the first tumor was not necessarily dose-dependent, the average tine
between the administration of a single dose of a carcinogen and the
individual appearances of the several tumors was definitely
dose-dependent. This increase in the average time for the development
of tumors as the dose is lowered led Swann et _§!,• to suggest that
the time needed to achieve full malignancy Ts" a function of the amount
of the initial preneoplastic lesion, i.e., the formation of
carcinogen-DNA adducts.
Most in vivo studies of carcinogen-DNA adduct formation have used
total DNA content in an organ to calculate the specific activities of
the adducts. However, it is likely that the amounts of adducts formed
.vary considerably in different cell types. This possibility has the
greatest implication for organs (such as the lung) that contain a
multitude of cell types and in which the cytochrome P-450-dependent
monooxygenase enzymes are localized to a few cell types.^'1»5,166
Obviously, the specific activities of carcinogen-DNA adducts in
individual cell populations would be a better measure of the effective
dose of a carcinogen than the values based on the total DNA content of
the organ.
In addition to the extent of carcinogen-induced DNA damage, the
capacity of cells to repair such damage and the degree of cell
replication are critical in the initiation of carcinogenesis. As with
the measurements of SAs of carcinogen-DNA adducts, most in vivo studies
of repair of carcinogen-induced DNA damage have been based on total
cell populations and thus represent average cellular values. Lewis and
Swenberg11 did study the differential repair of 06-methylguanine in
the DNA of rat hepatocytes and nonparenchymal cells (NPCs) after
administration of 1,2-dimethylhydrazine. The NPCs are the target cells
in 1,2-dimethylhydrazine-induced liver neoplasia. Although the initial
alkylation was similar in both cell types, the NPCs repaired
06-methylguanine more slowly than the hepatocytes. This led to a much
greater accumulation of the promutagenic lesion in the target cells
(NPCs). The rate of cell division was also much higher in the target
cells.119 Thus, in this model system, carcinogen-DNA repair rates
and cell division rates were definitely correlated with target-cell
susceptibility. Such studies of PAHs are needed to identify the target
cells for PAH-induced neoplasia in such organs as the lung.
Carcinogen-DNA adduct amounts, their rates of removal, and cell
turnover rates may not be able to explain the difference in organ and
species susceptibility to chemical-induced neoplasia. Promotional
5-30
-------
aspects of carcinogenesis might be required to explain these
differences. Determination of whether the initiation
characteristics—adduct amounts, DNA-repair rates, and cell turnover
rates—can explain the species, strain, and organ differences in
susceptibility to PAH-induced neoplasia will require more detailed
examination of individual cell types in the organs, such as the studies
by Lewis and Swenberg.118•119
Even if the initiation characteristics cannot explain organ,
species, and strain differences, this does not detract from the use of
PAH metabolite-DNA adduct amounts as a measure of the effective
biologic dose of a carcinogen in a known target tissue. For low-dose
extrapolation of carcinogenic data, for the ranking of a series of
similar carcinogens, and for determining the effect on neoplasia of
pretreatments that alter the metabolism of a carcinogen, the results
discussed above strongly suggest that specific PAH metabolite-DNA
adduct amounts are a good measure of effective biologic dose. Adduct
amounts in individual cell types of a target organ would probably be an
even better measure of effective biologic dose. However, they are not
now practical, because the separation of cell types is not generally
feasible. We should attempt to incorporate DNA-repair and cell
turnover rates into the effective biologic dose of a carcinogen. The
use of these initiation characteristics as a measure of effective
biologic dose has practical value, because they can usually be studied
at doses much lower than those used in bioassay studies.
TRANSCRIPTION AND REPLICATION
Several studies have shown that the functions of DNA during
transcription and replication are inhibited by the presence of
carcinogen-DNA adducts on the DNA template (Grunberger and
Weinstein' ). Mizusawa and Kakefuda131 concluded that the BaP DE
I-DNA adduct inhibits chain elongation with little effect on initiation
1 an
of DNA replication. These authors, Yamaura et al. ,^'~ and Hsu et
a_l/ suggested that elongation of the deoxypolynucleotide chain was
terminated at each BaP DE I binding site of the template. Leffler e_t
al. ^ demonstrated a progressive inhibition of transcription with
increasing amounts of BaP DE I adducts on the template. However, in
contrast with the above-mentioned results on replication, Pulkrabek e£
a_l. ^ concluded that, with some degree of frequency. RNA polymerase
can bypass BaP DE I adducts in a template to permit continued chain
elongation. They also showed that BaP DE I-modified plasmid DNA could
not transfect a receptive Escherichia coli strain to antibiotic
resistance. The modification of DNA by metabolites of aflatoxin and
2-acetamidofluorene inhibits replication and transcription on the
modified template in a similar manner.
Most of the detailed studies of the effects of carcinogen-modified
DNA on transcription and replication have been performed in in vitro
model systems. However, because the DNA adducts formed in vivo after
5-31
-------
exposure to BaP are, at least in most cases, the same as those which
were used in the in vitro studies, these basic functions of DNA might
also be affected after in vivo exposure to BaP. Dose-response studies
of formation of PAH metabolite-DNA adducts revealed that adducts will
exist after environmental exposure to PAHs. Moreover, recent studies
by Anderson e_t al. (unpublished data) have shown that BaP
raetabolite-DNA adducts are formed in many organs, both susceptible and
nonsusceptible to BaP-induced neoplasia. BaP DE I-DNA adducts were
observed in lung, liver, forestomach, brain, kidney, and colon of mice
after oral exposure to BaP. Even if the environmental exposures to
PAHs are too small to induce neoplasia, the formation of DNA adducts
after such exposures could produce aberrations in the transcription
(replication) of genetic information in many organs and perhaps lead to
subtle toxic effects. Obviously, the amounts of DNA adducts
constitute the appropriate measure of effective biologic dose for these
considerations.
5-32
-------
TABLE 5-1
In Vivo Formation of PAH Metabolite-DNA Adducts
Species8
PAH
Doseb
Route
Organ
Reference
A/HeJ
A/HeJ
A/HeJ
C57BL/6J
ICR/Ha
A/J
A/J
C57BL/6J
A/J, C3H/HeJ, DBA/2 J,
C57BL/6J
C57BL, DBA/2, Swiss
Y1 C57BL
£ C57BL/6J, DBA/2, Swiss
Ha/ICR
Swiss
SENCAR, CD-I
C57B1
C57BL/6J
C57BL
C57BL
C57BL
C57BL
TO
TO
TO
Wistar rat
Sprague-Dawley rat
New Zealand rabbit
BP
BP
BP
BP
BP
BP
3-MC
3-MC
3-MC
BP, DMBA, 3-MC
DMBA
BP
BP
BP
BP
BP
BP
BP
BP-diolsc
3-OH-BP
3-MC, DMBA
11-me thy Ike tone
11-me thy Ike tone
11-methylketone
BP
BP
BP
23.8
11.9
0.048-29.7
23.8
23.8
0.05
0.03
0.06
0.03
0.1, 1.0
0.025-1.0
0.2
0.004-1.2
0.25
0.1
1.0
0.4
0.2, 1.0
0.2
0.2
1.0
12.9
1.3
1.7
2.9
1.0, 10. 0
198
P.O.
P.O.
P.O.
P.O.
P.O.
I.V.
I.V.
I.V.
I.V.
Topical
Topical
Topical
Topical
Topical
Topical
Topical
Topical
Topical
Topical
Topical
Topical
I.M.
I. P.
Topical
Topical
I.V.
I.V., I. P.
Lung,
Lung,
Lung,
Lung,
Lung,
Lung,
Lung,
Lung,
Lung,
Skin
Skin
Skin
Skin
Skin
Skin
Skin
Skin
Skin
Skin
Skin
Skin
Lung,
Lung,
Skin
Skin
Lung,
Lung,
liver
liver, forestoraach
liver, forestomach
liver
liver, forestomach
liver, kidney
liver
liver
liver
liver, skin
liver
liver
liver
6,184
3
8
9
90
51
51
49
50
150
150
11
149
14
36
41
107
71
71
71
179
2
2
1,2,38
14
23
d
aMouse strain unless otherwise designated.
"Micromoles per animal for mice. Micromoles per kilogram for other species. Mouse weight 18-25 g in these
studies.
c7,8-Dihydro-7,8-dihydroxy-, 9,10-dihydro-9,10-dihydroxy-, and 4,5-dihydro-4,5-dihydroxy-.
and Anderson, unpublished data.
-------
TABLE 5-2
Formation In Vivo of BP DE-DNA Adducts
SfieEiw.*- ^-^-^-^..
A/HeJ
A/HeJ
A/HeJ
A/HeJ
ICR/Ha
ICR/Ha
ICR/Ha
C57BL/6J
o, C57BL/6J
^ A/J
*• A/J
A/J
New Zealand rabbit
New Zealand rabbit
New Zealand rabbit
New Zealand rabbit
Tiwue^ _.
Lung
Liver
Lung
Liver
Lung
Liver
Fores tomach
Lung
Liver
Lung
Liver
Kidney
Lung
Liver
Lung
Liver
Do8e,b
mg/mg
240
240
2.4
2.4
240
240
240
240
240
0.5
0.5
0.5
50
50
50
50
Specific Activity of
BP DE-DNA Adducts, c
Route
P.O.
P.O.
P.O.
P.O.
P.O.
P.O.
P.O.
P.O.
P.O.
I.V.
I.V.
I.V.
P.O.
P.O.
I. P.
I. P.
pmol/mg of DNA
6.97
6.02
1.68
0.98
0.42
0.44
0.43
5.80
3.50
0.14, 0.09
0.10, 0.08
0.03, 0.03
0.58
0.31
0.23
0.29
Reference
9
9
9
9
90
90
90
9
9
51
51
51
d
d
d
d
aMouae strain unless otherwise designated.
''Mice assumed to weigh 25 g. In all P.O. studies in mice, two equal doses of
were given 2 h apart. Values represent total dose.
cValues represent sums of BP DE I and BP DE II adducts. Adducts determined 48 h after
[^HjBP dose, except for studies in A/J mice, in which first value is 4 h after I.V.
dose and second is 24 h after dose. Average of at least two determinations.
-------
TABLE 5-3
In Vivo Formation of BP Metabolite-DNA Adduces
in Lung and Liver of Rats
Specific Activity,b fmol/mg of DMA
Tissue
Lung
Liver
Lung
Liver
Dose,8
yimol/kg
1.0
1
10
10
.0
.0
.0
Total
DNA Binding
133 +
730 +
680 •*•
2,500 >
11
220
165
610
BP-phenol-oxide-
DNA Adduct
55.5
48.2
178
168
+ 6.0
± 5'5
± 41
± 25
aI.V. injection.
bDetermined 1 h after I.V. injection of [3H]BP. Data from
Boroujerdi e_t *_!• Mean +_ S.E. for three experiments.
5-35
-------
TABLE 5-4
Binding of 3-MC to DNA in Lung and Liver of Mice
Strain
A/J
C3H/HeJ
DBA/2J
C57BL/6J
Tissue
Lung
Liver
Lung
Liver
Lung
Liver
Lung
Liver
In Vivo
Exposure Time8
4 h
7 d
28 d
4 h
7 d
28 d
4 h
7 d
28 d
4 h
7 d
28 d
4
7
28
4 h
7 d
28 d
4 h
7 d
28 d
4 h
7 d
28 d
Specific Activity,b fmol/mg of pi
Nucleoside-
Total0 Bound Addui
113
74
106
135
96
96
76
43
43
45
20
29
68
35
24
64
50
28
173
103
83
139
135
107
44.9
30.0
15.9
5.5
5.5
0
17.6
8.3
6.1
4.4
0.6
0
16.3
6.3
1.1
4.8
1.3
0
16.1
2.7
1.7
9.7
0
0
aTime of sacrifice after I.V. injection of [^H]3-MC (12.6 rag/mouse).
bData from Eastman and Bresnick.50
cMean of 6-10 mice (3-5 determinations). Fairly large individual
variation was observed with S.D. of +_ 252 of mean.
Peaks V and VI in Figure 5-4. Values obtained by multiplying X of
radioactivity eluting from Sephadex LH20 as peaks V and VI by total
DNA-associated radioactivity.
5-36
-------
TABLE 5-5
Effect of AHH Inducters on In Vivo Formation of BP DE-DNA Adducts
and on Total DNA-Associated Radioactivity in Mice3
Specific Activity0
(pmol/mg of DNA) of Total
BP DE-DNA Adducts DNA-Associated
Radioactivity,
Z of control
43
67
46
48
73
182
123
101
42
55
40
aUntreated or treated mice were killed 48 h after oral dose of [3H]BP
(6 mg/mouse). DNA isolated from tissue was enzymatically digested, and
deoxyribonucleosides were chromatographed on HPLC. Specific activity
of BP DE-DNA adducts calculated from HPLC chromatogram.
bMice were fed &WF (3 mg/g of diet) for 2 wk before [3H]BP administra-
tion. See text for discussion of Group A and B animals. Animals
treated with TCDD (80 nmol/kg) 4 d before [3H]BP administration. Aroclor
1254-induced mice received inducer (500 mg/kg) 48 h before administration
of [3H]BP.
Strain
A/HeJ
A/HeJ
A/HeJ
A/HeJ
A/HeJ
A/HeJ
A/HeJ
A/HeJ
ICR/Ha
ICR/Ha
ICR/Ha
Treatment1*
6-NF (Group A)
B-NF (Group A)
6-NF (Group B)
8-NF (Broup B)
TCDD
TCDD
Aroclor 1254
Aroclor 1254
6-NF
6-NF
6-NF
Tissue
Lung
Liver
Lung
Liver
Lung
Liver
Lung
Liver
Lung
Liver
Forestomach
j.ii xLeaueu Lii.i^e,
Z of control
15
0
8
0
5
0
9
0
16
7
16
cData for A/HeJ mice from Wilson £t £l. Data for ICR/Ha mice from
loannou et al. Zero means adduct not detected in treated animals.
5-37
-------
TABLE 5-6
Effects of Pretreatment with TCDD on In Vivo Covalent Binding
of BP to Mouse Epidermal DNA, RNA, and Protein
Hydrocarbon Bound to Macromolecules
(praol/mg) in Treated Animals,0
Strain
SENCAR
SENCAR
CD-I
BP Dose,a
ing/mouse
25
25
50
Tirae,b h
3
24
24
% of control
DNA
250
278
318
RNA
217
133
100
Protein
1 •
255
260
242
aMice received single topical application of [3H]BP. Animals received
single topical application of TCDD (1 mg in 0.2 ml of acetone) or acetone
72 h before carcinogen.
bTime of sacrifice after application of BP.
cData from Cohen et al.36
5-38
-------
TABLE 5-7
In Vivo Disappearance Rates of BP DE-DNA Adducts in Lung
and Liver of Mice
Strain
A/HeJ
A/HeJ
A/HeJ
A/HeJ
C57BL/6J
C57BL/6J
Tissue
Lung
Lung
Liver
Liver
Lung
Liver
P.O. Dose,
nig/mouse
6.0
0.012
6.0
0.012
6.0
6.0
Half-life for Dis-
appearance of BP DE-
DNA Adducts,3 days
17
19
9
16
19
14
aAnimals sacrificed at intervals of 10 h to 28 d after oral dose of
[3H]BP. Specific activity of BP DE-DNA adducts decreased mono-
exponentially in lung and liver of A/HeJ mice. There was biphasic
decay of adducts in lung and liver of C57BL/6J mice. Half-life of
terminal phase given here. Values are sums of BP DEI- and BP DEII-
DNA adducts. Data from Anderson and Wilson.
5-39
-------
Exposure to
Carcinogen
w
w
Accumulation
in body through
portal of entry
k.
F
Transport to
target tissue
- - *
/
Entry into
target cell
1
4>
O
f
Interaction
with critical
receptor
Transport to
critical reception
(DMA, RNA, protein)
Transport to
site of metabolism
Conversion to
appropriate metabolite
-------
Eiposun
Uptake
Biotront/orma Fion
Chemical binding
Repair
Genetic lesions
Corrinogenesis r 4
r2
FIGURE 5-2. Sequence of possible eventi from exposure to
carcinogenesis.
5-41
-------
9. l6-epox 9.10-diol
BENZO(o)PYRENE
XI
6-OH-Me
i-PHENOXY
RADICAL Jl
6-OH
7,8-epox
7-OH
1—
CONJUGATES
BOUND MACROMOLECULES
DNA
RNA
PROTEIN
FIGURE 5-3. Composite of metabolic products of benzola]pyrene.
-------
FIGURE 5-4. Sequences in formation of highly
electrophilic 7,8-diol-9,10-epoxide.
5-43
-------
FIGURE 5-5. Structures of BPDE; both isouers represent
enantioneric mixtures.
5-44
-------
GF
GF
Ul
I
lOOOr
8OO
Lung
6O
8O 100
20
4O 6O 6O
Fraction no
KX>
20 40 60 8O KX)
FIGURE 5-6. Comparison of HPLC analysis of nucleoslde-bound BP metabolites obtained from DNA Isolated by pre-
cipitation ( ) and hydroxyapatlte (^ ) procedures. When curves are Identical, only solid curve Is
Illustrated. Mice were killed 48 h after p.o. dose of 3H[BPJ (3 mg/mouse). DNA Isolated from lung, fore-
stomach, and liver was enzymatIcally digested, and deoxyrlbonucleosldes were chromatographed on Cio-sllica
column with linear aqueous-methanol gradient. WF, water fraction; PF, pregradlent fraction; GF, gradient
fraction. Peaks I, II, and III observed in gradient fraction are discussed In text. Specific activities of
various peaks are given In Table 5-2. Reprinted with permission from Adrlaenssens et al.;3 copyright
Academic Press.
-------
CLUTION VOLUME M)
FIGURE 5-7. Sephadex LH-20 column chromatography of
deoxyribonucleosides from DMA of mice treated with
[3H]-3MC. Chromatography was performed on 0.9 x 30-cm
column eluted vith 140-ml gradient of 30-1001 methanol in
water. This ia composite graph to indicate possible p»«ki
of radioactivity obtainable. Peaka I to IV, early
peaks obtainable from liver DMA. Peaka V and VI, nucleotide-
bound adducti that appear maximally in A/J lung DMA. Aj^Q'
unmodified deoxyribonucleoaide; arrov at bottom, position of
elution of marker, 4-(£-nitrobenzyl)pyridine. Reprinted with
permission from Eastman and Bresnick.*"
5-46
-------
IE
3
us is tn
\ Mol DMSA / Mouse
u
FIGURE 5-8. Binding of 7,12-dimethylbenz[a]anthracene
Co DNA of skin of C57BL mice 19 h after topical
application. Binding determined from elution profiles
of DNA hydrolysates prepared from treated skin. Reprinted
emission from Phillips et al.^0
with permission
5-47
-------
FIGURE 5-9. Dose dependence of BaP adduct formation in epidermal
DNA. Groups of mice, 15-20 each, were treated topically with
1, 2, 5, 10, 25, and 100 pg of [^JBP and sacrificed 24 h later.
Enzymatically digested epidermal DNA was chromatographed on
Sephadex LH-30 or Waters C^g-yBondapak column. A, dose dependene
of peak I formation; B, dose dependence of peak III formation.
Reprinted with permission from Pereira et «1.
5-48
-------
LUNG
LIVER
FORESTOMACH
Ul
10.0
i.o
0.10
o.ooi
1
10
100
1000 1
10
100
1003 1
10
100
1000
FIGURE 5-10. Dose-response relation for BP DE-DNA adducts in lung, liver and forestomach nf A/H i
was administered orally to .nice at 2-1 351 pmol/kg. Mice were sacrif iced'^h faTer "oNA sot t d
tissue was enzymatically digested, and deoxyribonucleosides were chroraatographed on HPLC SoeciH? • •
of BP DE adducts were determined as described in Adriaenssens et al.3 Specific activities
-------
= 1000 •
750 •
500 •
25O •
c
o
Z
C 100 200 JOC
Ne BP Molteultt/Solmoittilo C*nomt
FIGURE 5-11. Linear relation between BP DE I, BP DE II, and
9-hydroxybenzo[a]pyrene-induced mutation frequencies at
histidine locus in Salmonella typhinturium strains TA 100 and
TA 98 and BP metabolite-DMA adducts in bacterial cells. Data
from Fahl et al.
54
5-50
-------
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186. Wood, A. W., R. L. Chang, W. Levin, R. E. Lehr, M. Schaefer-
Ridder, J. M. Karle, D. M. Jerina, and A. H. Conney.
Mutagenicity and cytotoxicity of benz[a]anthracene diol
epoxides and tetrahydroepoxides: Exceptional activity of the
bay region 1,2-epoxides. Proc. Natl. Acad. Sci. USA 74:2746-
2750, 1977.
187. Wood, A. W., W. Levin, R. L. Chang, R. E. Lehr, M. Schaefer-
Ridder, J. M. Karle, D. M. Jerina, and A. H. Conney. Tumori-
genicity of five dihydrodiols of benz(a)anthracene on mouse
skin: Exceptional activity of benz(a)anthracene 3,4-dihydro-
diol. Proc. Natl. Acad. Sci. USA 74:3176-3179, 1977-
188. Wood, A. W., W. Levin, A. Y. H. Lu, D. Ryan, S. B. West, R. E.
Lehr, M. Schaefer-Ridder, D. M. Jerina, and A. H. Conney.
Mutagenicity of metabolically activated benzo[a]anthracene 3,4-
dihydrodiol: Evidence for bay region activation of carcinogenic
polycyclic hydrocarbons. Biochem. Biophys. Res. Commun. 72:
680-686, 1976.
189. Wood, A. W., W. Levin, D. Ryan, P. E. Thomas, H. Yagi, H. D.
Mah, D. R. Thakker, D. M. Jerina, and A. H. Conney. High
mutagenicity of metabolically activated chrysene 1,2-dihydro-
diol: Evidence for bay region activation of chrysene. Biochem.
Biophys. Res. Commun. 78:847-854, 1977.
190. Wood, A. W., W. Levin, P. E. Thomas, D. Ryan, J. M. Karle, H.
Yagi, D. M. Jerina, and A. H. Conney. Metabolic activation of
dibenzo(a,h)anthracene and its dihydrodiols to bacterial
mutagens. Cancer Res. 38:1967-1973, 1978.
191. Yamagawa, K., and K. Ichikawa. Experimentelle Studie u'ber die
PaChogenese der Epithelialgeschwulste. Mitt. a.d. med. Fakult.
d. k. Univ. zu Tokyo 15:295-344, 1915-1916. (Ind. Medicus
14:294, 1916)
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192. Yamaura, I., B. H. Rosenberg, and L. F. Cavalieri. The aajor
adduces of cis and trans benzo[a]pyrene diol epoxides cause
chain termination during DNA synthesis in vitro. Chen. BioL.
Interact. 37:171-180, 1981.
193. Yang, L. L., V. M. Maher, and J. J. McConnick. Error-free
excision of the cytotoxic and mutagenic N^-deoxyguanosine
DNA adduct formed in human fibroblasts by 00-76, 8cl-dihydroxy-
9i, 10a-epoxy-7,8 ,19,10-tetrahydrobenzo[a]pyrene. Proc. Natl.
Acad. Sci. USA 77:5933-5937, 1980.
194. Yang, S. K., P. P. Roller, and H. V. Gelboin. Benzo[a]pyrene
metabolism: Mechanism in the formation of epoxides, phenols,
dihydrodiols, and the 7,8-diol-9,10-epoxides, pp. 285-301. In
P. W. Jones and R. I. Freudenthal, Eds. Carcinogenesis—A
Comprehensive Survey. Vol. 3. Polynuclear Aromatic Hydro-
carbons: Second International Symposium on Analysis, Chemistry,
and Biology. New York: Raven Press, 1978.
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POLYCYCLIC AROMATIC HYDROCARBONS IN FOOD AND WATER
AND THEIR METABOLISM BY HUMAN TISSUES
This chapter deals with the relation of PAHs to human metabolism.
Specifically, its purposes are to collate a large volume of literature
dealing with the capacities of a number of human tissues to interact
with and biotransform selected PAHs; to define, where possible, the
effects of these compounds on human tissues; and to examine the
principal sources of human exposure to PAHs through food and water.
PAH METABOLISM BY HUMAN TISSUES
The abilities of various human tissues to metabolize PAHs have been
extensively studied, with emphasis on the chemical biotransformations
that are catalyzed by tissues that can be readily sampled (such as
blood cells, skin, and placenta) or that can be biopsied or cultured
(such as fibroblasts, liver, and intestinal and tracheohronchial
epithelium). The chemical biotransformations of selected PAHs that
such tissues carry out are in general qualitatively similar to those
demonstrated in animal tissues, although there are considerable species
and organ differences in catalytic activities of relevant enzymes.
These differences may be great enough to preclude comparative
generalizations; and for the most part the relation between in vitro
and in vivo enzymatic activities is unclear. Moreover, it is apparent
from the findings reviewed in this chapter that the enzymatic capacity
to biotransform PAHs to ultimate carcinogens in various tissues is not
necessarily correlated with the demonstrated ability of PAHs to produce
cancers in those tissues.
SKIN
That benzo[a)pyrene hydroxylase can be induced in cultured human
skin was first demonstrated in 1972. Foreskins from children who
were circumcised 2-4 d after birth were shown to contain an enzyme that
hydroxylates the carcinogen benzo[alpyrene (BaP), and induction of the
enzyme (by a factor of 2-5) was demonstrated when the foreskins were
cultured for 16 h in the presence of 10 yM henz[a]anthracene. Among a
group of 13 skin samples studied, control enzymatic activities extended
over a threefold range and were not correlated with race, age of
mother, or medications given to mother or infant. The enzyme had an
absolute requirement for NADPH and molecular oxygen and was completely
inhibited by CO; these findings suggested the involvement of a species
of cytochrome P-450 in the hydroxylation reaction. The presence of
this heme protein in low concentrations in cutaneous tissue was later
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demonstrated by Bickers e_t a_1.22 Coal-tar products, which are widely
used in dermatologic practice in conjunction with exposure to
ultraviolet light (§.g., the Goeckerman regimen for psoriasis62,63^
were also shown to induce aryl hydrocarbon hydroxylase (AHH) signifi-
cantly in patients with dermatologic disease where the coal tar was
applied, but not in skin distant from the site of application. *• In
concurrent studies in neonatal rats, although (as in humans)
distant skin sites were unaffected by the coal-tar application, AHH
activity in the livers of treated animals increased to more than 20
times the control values. Among five identifiable constituents of coal
tar studied for their AHH-inducing properties in human skin, BaP
was the most potent; pyrene and anthracene also caused significant
induction of the enzyme. In isolated cultured human hair follicles,
Vermorken e_t a_l. , using radioactive BaP as substrate, not only
demonstrated the presence of the hydroxylase, but identified the
formation of the 3-OH, 7,8-dihydro-7,8-dihydroxy, and 9,10-dihydro-
9,10-dihydroxy metabolites of this PAH.
BaP clearly has cytotoxic effects on cultured human skin
fibroblasts, although relatively high concentrations are required for
cytotoxicity. Milo e_t a_l.12^ studied the influence of the three
carcinogenic PAHs, 7,12-ditnethylbenzanthracene (7,12-DMBA), 3-methyl-
cholanthrene (3-MC), and BaP on mixed-function oxidase activity,
cell proliferation kinetics, and DMA damage in cultured fibroblasts.
They found that only BaP, at 10 ug/ml or higher, affected all the
cellular metabolic characteristics examined. 7,12-DMBA at 6 ug/ml or
higher induced the mixed-function oxidase system and stimulated DNA
synthesis; 3-MC at concentrations as high as 15 ug/ml produced no
significant cellular alterations. Similarly. 5-fluoro-7,12-DMBA,
anthracene, and phenanthrene had no effects on these human cells. The
authors concluded that BaP alone could initiate all the biochemical
events probably necessary to trigger transformation of human cells in
vitro.
PAH-induced cytotoxicity to cultured human fibroblasts has also
172 8
been demonstrated by Strniste and Brake and Aust et al. In the
former study, normal fibroblasts and xeroderma pigmentosum (XP) cells
were used, and BaP was "activated" by light radiation (near
ultraviolet), rather than enzymes. Photoactivation (at 300-400 nm) of
BaP produced at least three identifiable quinones (1,6-, 3,6-, and
6,12- isomers), as well as more hydrophilic products, depending on the
duration of light exposure. Formation of these products was
oxygen-dependent. The irradiation products led to several types of DNA
*"Mixed-function oxidase" refers to the NADPH-dependent enzyme complex
containing cytochrome P-450s in the membranes of the endoplasmic
reticulum, which catalyzes the oxidation of numerous structurally
diverse molecules, including drugs, steroid hormones, and carcinogens.
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damage, with covalently bound hydrocarbons constituting the major
lesion under all conditions studied. XP cells were more sensitive to
damage than normal cells (by a factor of 1.7-2), and sensitivity
increased by a factor of 10 when long-wavelength ultraviolet light was
used. 7,12-DMBA, 3-MC, and BaP were also examined; the order of
phototoxicity was 7,12-DMBA> BaP > benzo[ejpyrene > 3-MC.
In the study of Aust ji_t ji_l.,° a human epithelial cell-mediated
cytotoxicity and rcutagenicity assay system for BaP was developed with
human fibroblasts as the target cells. Lethally x-irradiated human
kidney-carcinoma epithelial cells were cocultivated with human XP skin
fibroblasts (XP12BE) lacking excision-repair capability for BaP-DNA
adducts. Under defined conditions, the frequency of mutation to
6-thioguanine resistance and PAH binding to DNA were shown to be
concentration-dependent. Two principal BaP-DNA adduct peaks could be
identified--a major peak consistent with an adduct standard synthesized
from th.e anti-isomeric 7,8-dihydrodiol-9,10-epoxide of the hydrocarbon
and a minor peak consistent with the syn-isomeric form of this
metabolite. The results are consistent with those in other reports on
BaP adducts formed in human explant tissue from lung, "*• colon, 0
esophagus,°° and bronchus,?5 and they represent an advance in the
development of sensitive assay systems for detecting biologic responses
to human epithelial-cell activation of BaP.
Direct neoplastic transformation of human fibroblasts by
carcinogens has also been demonstrated. In the study of Kakunaga, *
normal human adult fibroblasts exposed to the carcinogen 4-nitro-
quinoline 1-oxide underwent malignant transformation in a process
requiring numerous cell divisions. When injected into athymic (nude)
mice, the transformed cells produced solid tumors at the site of
inoculation. Because it could not be metabolically activated by the
target cells used, 3-MC was unable to effect transformation; the use of
other PAHs and induced microsomes with high concentrations of
cytochrome P-450 to activate 3-MC was not examined. Normal human
foreskin cell populations were neoplastically transformed in studies by
19 L. 1 9 ?
Milo and DiPaolo ^' with a number of non-PAH carcinogens; and
treatment with a tumor promoter alone (phorbol ester) has been
shown*" to induce neoplastic transformation in fibroblasts from
humans genetically predisposed to cancer (familial polyposis of the
colon). Thus, it can be inferred that cells already in an "initiated
state" as a result of a genetic defect represent a novel fibroblast
system that may provide a means for exploring separately the roles of
initiators and promoters in carcinogenesis. Painter^5 U3ed HeLa
cells to develop a rapid screening test to detect agents that damage
human DNA. The test measures thymidine uptake into the cells at
various times (principally 1-2 h) after treatment with a presumptive
carcinogen or mutagen. In this test system, BaP was inert unless
metabolically activated by incubation with rat-liver microsomes.
Brookes and Duncan2 compared the effects of PAHs on primary human
embryo cells and HeLa cells. Fibroblasts from skin, lung, muscle, and
gut were cultured and treated with 3H-labeled BaP and 7,12-DMBA.
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Both hydrocarbons were metabolized in the cultures, 7,12-DMBA more
slowly than BaP; among the cell types studied, lung fibroblasts
metabolized the compounds more efficiently than others and retained
this capacity well in subculture. The binding of BaP and 7,12-DMBA to
DNA, RNA, and protein of these primary lung-derived fibroblasts was
also studied (metabolism of each hydrocarbon exceeded 752 during the
48-96 h of treatment) and was found to be significantly greater for BaP
than for 7,12-DMBA. The data in this study also established
parallelisms, at least for BaP, between hydrocarbon binding to cellular
macromolecules in fibroblast cultures derived from mouse embryos and
those derived from human lung cells. Such parallelism is of more than
casual interest, in view of the susceptibility of the mouse to
hydrocarbon carcinogenesis and the known correlation between
hydrocarbon-DNA binding and cancer-producing activity in mouse skin.
The effects of pyrene and BaP in the human diploid fibroblast
culture WI-38 were studied by Weinstein e_t a_l.l°" Neither caused
significant damage (compared with controls), as assessed by mitotic
index or chromosomal breaks after 1-h pulse exposures. However,
metabolic activation of BaP with mj-crosomes resulted in a dramatic
decrease in mitotic index and a significant increase in breakage.
Microsomal incubation did not alter the inertness of pyrene in this
test system. Freeman ejt a_K have made interesting observations on
comparative aspects of hydrocarbon metabolism in skin epithelial and
fibroblast cultures. A comparison of the ability of epithelial-cell
colonies and of fibroblast colonies from the same 13 subjects to
metabolize BaP to a water-soluble form demonstrated clearly the
markedly greater metabolizing capacity of epithelial cells. There was
a 20-fold difference in this capacity of epithelial cells; within
individual subjects, the ability of epithelial cells to metabolize the
PAH exceeded that of fibroblasts by as much as a factor of 40. There
appeared to be a major effect of culture age (6-55 d) on the ability of
epithelial cells to metabolize BaP.
A direct toxicity of BaP to normal human epithelial-cell cultures
has been described by Dietz and Flaxman. " This toxicity was
reflected in a dramatic reduction in epidermal-cell outgrowth, a
decrease in mitotic indexes, a loss of the well-ordered cell
relationships, and the early appearance of giant cells ranging in
diameter from 100 to 200 pm. In a clinical study that clearly could
not be carried out today, Cottini and Mazzone^ (i.n 1939) applied a
It solution of BaP daily (up to 120 d) to the skin of 26 normal
subjects and patients with various dermatologic disorders and examined
the gross and histologic consequences. The sequential epidermal
changes, of which gross pigmentation and verrucae were the most
frequent, and histologic alterations (which regressed within several
months when treatment was terminated) led the authors to conclude Chat
"benzopyrene, if applied to human skin for protracted periods, would be
carcinogenic as it is in animals."
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The metabolism of benz[a]anthracene, 7,12-DMBA, and BaP by human
mammary epithelial-cell aggregates in culture has been investigated by
Grover et £^. with nonneoplastic tissue obtained from eight
patients undergoing reduction mammoplasty. All three PAHs were
metabolized to water-soluble and organic-solvent-soluble products; the
latter included K-region and non-K-region dihydrodiols. The major
dihydrodiols detected as metabolites of the parent PAHs were the
8,9-dihydrodiols of BaP and 7,12-DMBA and the 9,10-dihydrodiol of BaP.
The hydrocarbons bound to the proteins and DNA of the epithelial cells,
but there were wide differences between different PAHs in extent of
binding between tissue preparations from different patients. Some of
the PAH-deoxyribonucleoside adducts formed from 7,12-DMBA and BaP
appeared to have been produced through reactions of bay-region
diol-epoxides with DNA, but little reaction with DNA was detected in
tissue preparations treated with BaP.
Unscheduled DNA synthesis induced by DNA-damaging chemicals has
been measured in nonreplicating human fibroblasts by autoradiographic
methods that are not readily applicable to organotypic epithelial-cell
cultures. To evaluate the range of chemical sensitivity and DNA-repair
responses of human skin epithelial cells, Lake e_t a_1.^5 developed a
semiquantitative in vitro method for measuring unscheduled DNA
synthesis in normal foreskin epithelial cells. On serial subculture of
organotypic primary skin cultures, the unscheduled DNA synthesis
response elicited by 3-MC decreased in parallel with the ability of
cells to metabolize PAHs to water-soluble metabolites. The working
hypothesis was that procarcinogens that are efficiently activated by
human skin-specific metabolism will be detected with unscheduled DNA
synthesis as an end point.
LIVER AND INTESTINE
Obana e_t aJL.^0 analyzed quantitatively and qualitatively the PAH
content of samples of human liver and fatty tissue. Six samples of
liver and 10 samples of fat were obtained at autopsy from 10 persons
who died of unstated causes (although the tissues were reported to be
"free from cancer"). Smoking habits, occupations, etc., were not
described. The tissue samples analyzed were quite large (40-120 g),
and the PAHs were determined without complex pretreatment. Table 6-1
shows the analytic results for liver, and Table 6-2 the comparable data
for fat tissue. Note that PAH concentrations are expressed as parts
per trillion (ppt), not parts per billion (ppb), and are in general
extremely low. Nevertheless, the data indicate that, on the average,
the PAH concentration in liver was one-third that in fat. Pyrene had
the highest concentration, followed by anthracene. Although the number
of samples was small, no sex or age differences were evident. The
known carcinogens benz[a]anthracene and dibenz[ah]anthracene were not
detected in either tissue. However, BaP was detected in small amounts
(20 ppt) in both liver and fat. This finding should be compared with
that of Tomingas £t £^.,178 who detected BaP at 1-15,000 ppb in
6-5
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human bronchial-carcinoma tissue (24 samples). Obana et al.130
called attention to the fact that the PAHs in the human tissues they
examined were different, in both concentration and composition, from
the PAHs that had been identified in marine samples. For example,
pyrene was found in oysters at 7-52 ppb and in Wakame seaweed at 12-41
ppb; the comparable figures for BaP were 0.3-2.6 ppb and 0.6-9 ppb,
respectively. Moreover, although pyrene was the most abundant PAH in
all cases, the next most abundant in the human tissues was anthracene,
whereas in the marine samples the next most abundant were
benzole]pyrene and benzofb]fluoranthene. These qualitative and
quantitative distinctions, especially the marked concentration
differences between nontumorous and tumorous tissues1-'8 and
between a common food source in the area and the human tissue samples,
need to be recalled in evaluating the importance of the food content of
PAHs, as well as the role of malignant pathology, when trying to
determine the significance of the body or tissue burden of these
hydrocarbons.
The liver contains the highest concentration of cytochrome P-45U in
the body. The activity of the pathway by which heme, the prosthetic
group of this heme protein, is synthesized can be greatly induced by a
host of foreign chemicals and can approach the rates of heme synthesis
in erythroid cells; and the enzymatic capacity of hepatic cells to
carry out the biotransformat ions that characterize the great variety of
PAH metabolites formed in vitro, and probably in vivo, has been well
defined. Only selected PAH transformations catalyzed by liver cells
are reviewed here, with some emphasis on the relationships of PAH- and
drug-metabolizing capabilities and on recent data indicating that
carcinogen metabolism may be increased by direct actions on relevant
membrane-bound enzymes, as well as by the conventionally assumed
process of increased dg novo synthesis of enzyme protein, i.e.,
indue tion.
Dybing et a_l> have examined the in vitro metabolism and
metabolic activation of several carcinogenic PAHs in subcellular
fractions from seven human livers. The patients all suffered from
total cerebral infarction and were serving as potential kidney donors
(maintained temporarily by life-support systems) at the Huddinge
University Hospital in Sweden. At the appropriate time, liver
extirpation was performed; within 20 min after the procedure, perfusion
had been completed and the tissue frozen in liquid nitrogen. This
study cnay mimic the enzymatic properties of human liver cells in the
living subject as closely as experimentally possible, other than by
direct biopsy or surgery in a living patient. Because of the unique
source of the tissue studied, some of the data merit recording here.
Microsomal cytochrome P-450 content (seven livers) was 0.16-0.60
nmol/mg of protein, with a mean ^ S.D. of 0.36 ^ 0.15, and AHH activity
averaged 175 + 138 pmol/mg of protein per minute; one sample had a
value of 483 pmol/mg. These activities are approximately the same as
those in liver raicrosoraes of untreated mice and rats. AHH activities
expressed per nanomole of cytochrome P-450 varied by a factor of 2.8
among the seven liver samples.
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Conney and co-workers27.28'40.41. 73'81 .82,94, 156 - "7,174
conducted a series of studies of direct liver-cell metabolism of
carcinogens and compared such metabolism with that of drugs. They
established that carcinogen metabolism may be increased not only
through enzyme induction, but through enzyme activation as well. They
compared the oxidative metabolism of BaP with that of antipyrine,
hexobarbital, coumarin, zoxazolamine, and 7-ethoxycoumarin in 32 adult
liver samples obtained at autopsy some 8-20 h after death. When enzyme
activity for one substrate was plotted against enzyme activity for a
second substrate for each of the 32 liver samples, significant
correlations were found. For example, for BaP paired against anti-
pyrine, hexobarbital, zoxazolamine, and coumarin, the correlation
coefficients were 0.85, 0.72, 0.69, and 0.57, respectively. Some
drug-drug metabolizing activities also showed a high correlation (e.g.,
antipyrine and hexabarbital, r » 0.79; antipyrine and coumarin, r =
0.72), whereas in other instances, metabolizing capacities did not have
a high correlation, e.g., carcinogen vs. drug (BaP and
7-ethoxycoumarin, r » 0.35) and drug vs. drug (e.g., hexobarbital and
7-ethoxycoumarin, r - 0.37). The findings raise the possibility that
an in vivo drug-metabolism assay (e.g., a plasma disappearance-rate
study of a suitable test drug) might predict some carcinogen-
metabolizing capabilities of a person and suggest the presence in
humans of multiple monooxygenase systems for the substrates studied, as
well as their heterogeneous distribution in the population. Individual
differences in the rates of metabolism of BaP (7-fold) in these and
other liver samples studied^ were considerable, although, they did
not approach the known species differences^" in rates of metabolism
of drugs.
The effects of PAH administration on the metabolic disposition of a
specific carcinogen, such as BaP, have not been studied in humans, but
Schlede e£ a_l« ' recently examined the metabolic disposition of
radiolabeled BaP in rats, and the results probably can be extrapolated
to humans. Pretreatment of rats with unlabeled BaP greatly increased
the plasma disappearance rate of a tritiated dose of the same compound
given intravenously; the effect was especially marked during the first
5 min after the intravenous dose of the radiolabeled material, and the
increased rate lasted for at least 6 h. This effect of pretreatment
with the compound was paralleled by a lower concentration of PH]BaP
in brain, liver, and fatty tissues; similar but more varied results
were observed in lung tissue. These influences of BaP pretreatment on
a later intravenous dose of the •'H-labeled chemical were also
observed when the radiolabeled PAH was administered orally. 3-MC and
7,12-DMBA pretreatment of animals produced comparable effects on the
metabolic disposition and tissue contents of radiolabeled BaP. Pyrene,
anthracene, and phenobarbital had little or no such effect on the in
vivo disposition of this compound. In other studies, the biliary
excretion of [^^C]flaP was shown to be increased by pretreatment with
the unlabeled compound; however, no increase in excretion into bile of
the ^C-labeled metabolites of BaP was observed after prior
administration of this PAH. These findings suggest that conversion of
6-7
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BaP to its metabolites may be the rate-limiting step in the biliary
excretion of the compound. Phenobarbital had no effects on the pharma-
cokinetics (plasma disappearance) of [*• C]BaP and its metabolites ;
this drug might stimulate the conjugation of hydroxylated BaP
derivatives before their excretion into bile. The relevance of these
findings to humans is related not only to the (probably) qualitatively
similar pharaacokinetics of such a chemical as BaP — particularly its
extensive excretion via bile — but also to the potentially important use
of selected drugs, singly or multiply, to increase disposal of PAHs and
their metabolites from the body by stimulating conjugation and biliary
excretion and by increasing otherwise innocuous metabolic
bio trans format ions .
Prough e_t a_l_. have also studied BaP metabolism in human liver,
kidney, and lung, and they characterized the metabolites formed by HPLC
techniques. Tissue samples were obtained within 1-5 h after death, and
assays were completed within the succeeding 2-4 h. In the analysis of
metabolites formed from BaP, quinones, three classes of dihydrodiols ,
and two classes of phenols were categorized. Table 6-3 summarizes the
rates of formation of these metabolites by tumor, liver, kidney, and
lung microsomes and presents comparable data in rodents. There was a
very large variation in the human metabolism of BaP, compared with that
demonstrated in studies carried out concurrently in rodents. That
might reflect, as the authors noted, either the controlled environment
of the animals studied or a genetic variation in humans. In the human
liver, activities for metabolite formation were substantially lower
than those in rat microsomal fractions, and there were significant
differences in the BaP-metabolite profiles. A greater proportion of
benzene-ring metabolites was formed by human lung microsomes than by
human liver and kidney microsomes (or rodent lung microsomes). The
relative increase in the 9, 10-dihydrodiol product, as well as some
increase in the 7 ,8-dihydrodiol metabolite, accounted for the larger
portion of this difference among lung, liver, and kidney microsomes.
There is an apparent biologic inconsistency in these findings: although
the human lung is the principal site of PAH tumorigenesis , as the
authors observed, the 9 , 10-dihydrodiol product is suggested to have
little biologic activity on further metabolism, whereas other tissues,
such as the liver, formed large concentrations of the 7 ,8-dihydrodiol,
a "proximate carcinogen." Nevertheless, the findings are important,
providing not only data, on comparative rates of formation of metabo-
lites constituting the "HPLC profiles" in man and rodents, but also
intertissue metabolic profiles of BaP biotrans format ion for three major
organ systems in man. Thus, they extended the earlier findings of
Selkirk e_£ a_1.^0 on liver cells and lymphocytes.
A major advance in defining the role of the liver in PAH metabolism
and the factors that regulate liver monooxygenase activity has been the
demonstration that hepatic microsomal oxidation of specific substrates
can be directly increased in vitro, in addition to their property of
being induced in vivo, by various chemical treatments . ^' • ^° ''^'
Conney and colleagues have shown that 7 ,8-benzof lavone added to
homogenates of human liver samples can increase the rate of BaP
6-8
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hydroxylation by a factor of up to 11. Benzoflavone also increased
some drug hydroxylations substantially, although benzoflavone at
concentrations lower by a factor of about 100 paradoxically inhibited
these reactions. Marked individuality for activating and inhibiting
effects of benzoflavone were noted, and no significant effects on the
oxidation of some drug substrates (e.g., coumarin and hexobarbital)
were observed.
The enhancing effect of 7,8-benzoflavone on BaP oxidation was shown
to extend to other flavonoids, such as flavone, nobiletin, and
tangeritin. Related compounds—such as apigenin, chrysin, fisetin,
flavonone, galargin, hesperitin, kaempferal, raorin, myricitin,
naringerin, and quercetin—inhibited BaP oxidation. The stimulatory
effect of 7,8-benzoflavone on BaP 9,10-dihydrodiol oxidation to
bay-region epoxides was also studied and shown to have significant
species-specific characteristics. With untreated hamster microsomes,
more than 60% of the total metabolites of the hydrocarbon were
bay-region diol-epoxides, whereas human liver formed less than 5% of
such metabolites. Addition of 7,8-benzoflavone to the microsomal
incubations dramatically stimulated the formation of these metabolites
in human (and rabbit) liver microsomes. The stimulatory effects of
flavonoida on hydrocarbon oxidation have recently been shown to occur
in vivo as well.^O so the biologic importance of this phenomenon for
the intact host is likely to be considerable. The mechanism of
flavonoid activation of BaP hydroxylation has recently been explored in
detail by Huang e_t a_l- jhe flavone stimulates the NADPH reduction
of cytochrome P-450, but not that of cytochrome c, by NADPH-cytochrorae
c reductase; this finding supports the idea that the catalytic sites
for these substrates of the reductase are different. Other evidence
that these catalytic sites are different has also recently been
provided by the studies of Yoshinaga et al.197-200
Metabolic transformation of PAHs and their binding to cellular
macromolecules in cultured human gut tissues have been described.
Harris &t_ £1.^8 examined the metabolic fate of BaP and several other
compounds in cultured esophogeal explants from eight patients, six of
whom had esophageal carcinomas. Metabolism of the -'H-labeled PAH to
water-soluble metabolites varied among the eight patients over the
range of 1-68Z of total metabolism; the variation found within a single
case, however, in relation to different anatomic segments of the
esophagus (proximal, middle, and distal) was quite narrow—2%. In
spite of the 68-fold variation in metabolism among subjects, the
patterns of conjugates formed from metabolites in general were
qualitatively similar: sulfate esters, 21-55%; glucuronide conjugates,
7-37Z; and glutathione conjugates, 24-66X. Most of the radioactivity
of the organic-solvent-soluble metabolites of BaP cochromatographed
with authentic metabolites of this compound, including its proximate
carcinogenic, (-)-trans-7,8-dihydro-7,8-dihydroxy, derivative. Despite
the predominant occurrence of esophageal cancer in the distal segment,
the patterns of metabolites formed in all segments of the organ were
similar. Binding of ^H-labeled PAHs to DNA and protein was detected
in all eight cases, with binding to protein greater than to DNA in each
6-9
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instance. Binding among the eight cases varied 99-fold, and at least
three hydrocarbon-DNA adducts, including the specific guanine adduce,
were recognized. The adducts appeared identical with those previously
found in human colon and bronchus; and the patterns of both metabolism
and adduct formation with BaP were analogous to those found in
experimental animals susceptible to the carcinogenic action of PAHs.
Aucrup e_t a_l.,9 in an earlier and less detailed but essentially
similar study, had reported comparable data based on human colon
explants of tumor-free tissue. The binding of labeled BaP to cellular
protein was several times higher than that to DNA; however, hydrocarbon
binding to DNA correlated with tissue AHH activity (r » 0.87), whereas
no such correlation existed for protein binding. DNA binding (BaP,
picomoles bound per 10 rag) among seven tissue samples varied over a
25-fold range; variation within subjects was small. In an extension of
this work, Autrup e_t a_l.*• studied the comparative metabolism of BaP
(and aflatoxin Bl) and hydrocarbon-DNA adduct formation in cultured
human and rat colon explants. Adduct formation (in 103 cases) varied
over a 125-fold range in the tumor samples and in the same subject over
a 3- to 10-fold range in different segments of the organ. A number of
hydrocarbon-DNA adducts were identified, and both qualitative and
quantitative rat-human differences were demonstrated. Although overall
BaP metabolism was similar in rat and human colon tissue, the ratio of
organic-solvent-soluble to water-soluble metabolites was higher in the
human; sulfate esters predominated in rat colon, whereas equivalent
quantities of sulfate esters and glutathione conjugates were formed in
the human tissue; and hydrocarbon-DNA binding waa distinctly greater in
human colon, although, as noted, there was marked variability in adduce
formation within a given subject.
The comparative hydration of styrene 7,8-oxide, octene 1,2-oxide,
naphthalene 1,2-oxide, phenanthrene 9,10-oxide, benz[ajanthracene
5,6-oxide, 3-MC 11,12-oxide, dibenz[ah]anthracene 5,6-oxide, and BaP
4,5-, 7,8-, 9,10-, and 11,12-oxides to their dihydrodiols was
investigated in microsomes from nine human liver samples obtained at
autopsy. 0 The substrate specificity of the epoxide hydratase in
human liver microsomes was very similar to that of the epoxide
hydratase in rat liver microsomes. Phenanthrene 9,10-oxide was the
best substrate for the human and rat epoxide hydratases, and
dibenz[ah]anthracene 5,6-oxide and BaP 11,12-oxide were the poorest
substrates. Plotting epoxide hydratase activity obtained with one
substrate against epoxide hydratase activity for another substrate for
each of the nine human livers revealed excellent correlations for all
combinations of the 11 substrates studied (r » 0.87-0.99). The data
suggest the presence in human liver of a single epoxide hydratase with
broad substrate specificity, although the results do not exclude the
possible presence in human liver of several epoxide hydratases that are
•inder similar regulatory control.
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CIRCULATORY SYSTEM
Juchau et al.^i?? summarized a body of literature bearing on the
hypothesis that PAHs may play an important role in the pathogenesis of
arteriosclerotic lesions. The validity of this hypothesis apart, these
investigators clearly demonstrated that aortic tissues from a number of
species, including man, have detectable, albeit low, monooxygenase
activities using BaP and 7,12-DMBA as substrates. Enzyme activities
were comparable with those characterizing mouse skin. Cytochrome P-450
could be detected in primate aortas, and epoxide hydratase activity for
BaP 4,5-oxide was identified in homogenates of the arterial walls of
chickens and rabbits. The characteristics of the aortic monooxvgenase
for BaP resembled those of the enzyme system found in other tissues.
It could be markedly induced, for example, by 3-MC, polychlorinated
biphenyls (PCBs), and 5,6-benzoflavone; and, surprisingly, aortic
homogenates produced higher than expected quantities (by as much as a
factor of 28) of alkali-extractable metabolites when hematin was added
to the reaction mixtures. Interestingly, hematin has been shown in
other studies to degrade, in vitro, components of the monooxygenase
system. The primary BaP metabolites formed in rabbit aortic
homogenates were the 3-OH and 9-OH derivatives, phenolic compounds
known to be cytotoxic. The authors cited unpublished data to show that
the aortic metabolites of BaP form covalent bonds with such macro-
molecules as calf-thymus DNA. Treatment of chickens with the inducer
3-MC markedly increased the amount of the PAH-DNA adducts, whereas
addition of 7,8-benzoflavone in vitro inhibited binding. Aortic
enzymes also have been shown to catalyze the formation of rautagenic
metabolites from 7,12-DMBA. Thus, both cytotoxic and mutagenic
metabolites of PAHs can be generated in vascular tissues. The possible
relation of the formation of these compounds to the initial vascular
injury that may presage the local development of an atherosclerotic
plaque is of considerable interest.
The interaction of benz[a]anthracene and BaP with crystalline human
serum albumin in solution has been studied fluorimetrically by Ma ££
a_l. 1TO Equilibrium studies indicated that both PAHs bind to the pro-
tein to the same extent. Evidence of energy transfer from the trypto-
phan residue of the protein (increase in the weak B region—395-420
nm—fluorescence of the PAHs) permitted an assessment of the mean dis-
tance between the tryptophan and the bound ligand, thus identifying two
different binding sites in the same general area. The authors sug-
gested that structural differences among hydrocarbons, which may
greatly affect their orientations on the protein molecule, influence
mainly the selection of the binding site, rather than the binding
equilibrium.
In vivo BaP associates very little with serum albumin in the
presence of lipoproteins. The kinetics of BaP transfer between human
plasma lipoproteins have been examined by Smith and Doody163 with
high-density lipoproteins (HDL), low-density lipoproteins (LDL). and
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very-low- isitv lipoproteins (VLDL) prepared from fresh unfrozen human
plasma by altracentrifugal flotation. BaP-lipoprotein interactions
were analyzed fluorimetrically, and kinetic measurements were
determined by stopped-flow techniques. The half-times of BaP transfer
between HDLs, between LDLs, and between VLDLs were 40, 180, and 390 ras,
respectively. The transfer of these PAHs among lipoproteins of the
same density class was about one-twentieth that of pyrene under the
same conditions. The rate of BaP transfer between lipoproteins also
decreased with increasing size of lipoprotein; at equilibrium in vitro,
VLDLs contain about 10 times more of the BaP than LDLs, and LDLs
contain 20-50 times more than HDLs. The distribution between plasma
and erythrccvtes is different for 7,12-DMBA, BaP, benzanthracene, and
anthracene rhe mass of the PAH being associated with red cells (50,
70, 93, an: IOOT, respectively). Plasma lipid concentrations and the
dynamics or lipid and lipoprotein metabolism clearly may have an impact
on PAH distribution in blood and into specific tissues. For example,
transfer of BaP is quite rapid, compared with the half-time for either
hydrolysis of chylomicron triglyceride (about 2-5 min in humans) or
clearance of the most abundant lipoproteins from the circulation (3-5 d
in humans). The data of Smith and Doody concerning the role of
plasma lipoprotins in the transport of PAHs corroborated and extended
the findings of other investigators who examined the interaction of
PAHs and plasma proteins.13'3''32'57'162
The specific process of BaP uptake from human LDLs into cultured
human cells was examined by Remsen and Shireman. The cell lines
used were WI-38, a human embryonic lung-fibroblast line, and CM 1915, a
skin-fibroblast line derived from a patient with homozygous familial
hypercholesterolemia; the former cells are LDL-receptor-positive, and
the latter LDL-receptor-negative. Thus, in these studies, it was
possible to explore the role of LDL receptors in the cellular uptake of
PAHs that enter the bloodstream transported by chylomicrons and plasma
lipoproteins. The results indicated that cellular uptake of the
tritiated PAH by hoth cell lines from delipidated or serum-free medium
varied linearly with concentration, whereas incorporation of PAH bound
to LDLs was much less and, at higher lipoprotein concentrations, varied
nonlinearly. The presence of the PAH in the LDL preparation did not
affect the binding of I-labeled lipoprotein to receptor-positive
cells. The study provided several findings of special importance
relative to the biologic impact of PAHs—or at lease BaP as a model
compound—on tissues in vivo. Clearly, although LDLs carry substantial
amounts of PAH, the presence of LDL receptors on cells is not necessary
for tissue uptake. The fact that PAH bound to LDL was incorporated intc
cells more slowly than PAH in a delipidated serum or serum-free medium
raises questions about the biologic significance of experimental models
in which increased incorporation of BaP from particles into lipid
vesicles has been demonstrated. The data from these experiments also
indicate that cells that may be directly exposed to a PAH (i.e.,
tracheobronchial, intestinal, and cutaneous cells) before the compound
reaches the bloodstream may accumulate PAH in much higher
concentrations than cells exposed to the PAH bound to lipoproteins,
inasmuch as the latter significantly slowed as well as limited the
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cellular uptake of BaP. Finally, the report indicated that BaP
previously incorporated into WI-38 cells could be substantially removed
(by 55-792) in a 120-min post treatment study period by 107. delipidated
serum or LDL-containinsz medium. This finding implies a potential for
considerable PAH redistribution and a requirement for a not
insignificant period for progression of the hydrocarbon from the plasma
membrane to the endoplasmic reticulum, where metabolism takes place.
The ability of human monocytes to oxidize BaP and the induction of
this enzyme activity by benzanthracene have been demonstrated by
opuera 1 i nvp!^ i oaf-nra . ' ^ > > *• T.akp and rr> 1 1 pa mips'"
investigators. 0>L'' 70 'L J Lake and colleagues-
re-examined this problem with the goal of developing a practical assay
for measuring whole-cell metabolism of BaP under highly standardized
conditions, eliminating—among other problems—the need for a large
volume of blood (50 ml) in the fluorometric assay developed earlier for
AHH activitity in this cell type. By measuring whole-cell generation
of water-soluble BaP metabolites over a 3-d culture period, using
H-labeled substrate and closely controlling other character-
istics, they provided a useful alternative cell system to that using
mitogen-stimulated lymphocytes for characterizing BaP oxidation
activity in humans.
Because of the advantage gained by the much greater inducibility of
AHH activity (up to 40-fold) in cultured monocytes, compared with
mitogen-stimulated lymphocytes (about 5-fold), the raonocyte system was
used by Okuda ££ a_l« to study the contribution of genetic factors
to the control of individual variation in AHH inducibility. Ten sets
of monozygotic tissues were assayed two to four times and 17 sets of
dizygotic tissues one to three times for basal and induced monocyte AHH
activity. The results indicated that 55-70? of the individual
variation in AHH inducibility of raonocytes was genetically determined.
Variation in AHH inducibility within subjects in repeat assays was wide
and approached the magnitude of the variation between subjects. Thus,
a single AHH assay is an imprecise biochemical characterization of a
subject. Alternatively expressed, the method then available (late
1977) made it impractical to characterize a population with genetically
distinct differences in AHH inducibility. The large intrasubject
variation in AHH inducibility of monocytes also indicated that, in
addition to the clear genetic influences on this process, unknown
environmental or technical factors expressed themselves in the test
procedure.
An abundant literature exists related to the monooxygenase activity
of lymphocytes; the inducibility of this activity by mitogens, which
have the property of stimulating lymphocyte transformation, during
which a number of metabolic activities are concurrently greatlv
increased; and the use of mitogen-stimulated lymphocytes to study the
genetic control of AHH in man and its relation to the occurrence of
some human cancers—notably those of the lung. Kouri and colleagues
have reviewed key aspects of this subject, ' McLemore e_t
al. H9-122 have also provided a detailed analysis of the genetics of
AHH and its purported relation to human cancer. Only a brief summary
of these findings can be included here.
6-13
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The identification of AHH activity in lymphocytes in
and its increase during lymphocyte blastogenesis led quickly to
clinical studies, the earliest being that of Kellermann e_t a_l.,"7 in
which this induced enzyme activity was measured in cultured lymphocytes
of normal controls, non-lung-tumor controls, and lung-cancer patients.
In a preceding study in the same year, this group had examined the
genetic variation in AHH activity in lymphocytes of 353 normal subjects
and had categorized the population into three groups—low, inter-
mediate, and high responders with respect to AHH inducibilicy; the
population frequencies were about 50%, 407, and 10%, respectively. The
conclusion was reached that the enzyme activity was controlled by two
alleles at a single gene locus and that the high and low responders
were homozygous and the intermediate group heterozygous for those
alleles. In the initial lung-cancer study, there was a virtual
absence of cases in the low-inducibility population, and all but two
cases were in the intermediate- and high-inducibility categories. All
the lung-cancer cases were in heavy smokers; of the 50 subjects, 48 had
an average consumption of two packs of cigarettes per day. When the
two control groups (normal subjects and a non-lung-cancer tumor group)
and the lung-cancer group were compared for risk of lung cancer, those
with intermediate and high inducibility (48 of the 50 lung-cancer
cases) had risks for lung cancer 16 and 36 times, respectively, the
risk in the low-inducibility group. This study prompted considerable
controversy over the next few years, during which the findings of
Kellermann and associates were cast in doubt.
A strong correlation (r =* 0.923) was also found by Kellermann e_t
al. between the plasma elimination rate of antipyrine and the rate
of BaP metabolism in human lymphocytes from a "carefully selected
homogeneous" population, compared with the much lower correlation (r =
0.425) found in a "heterogeneous" population. The authors interpreted
their findings as supporting the existence of common oxidative systems
or common genetic control of the systems for antipyrine and BaP
oxidation. Atlas e_t_ £_!• confirmed that plasma antipyrine half-life
is correlated to some extent with AHH inducibility (r » 0.84), although
no intrasubject correlations were found between AHH inducibility and
the oxidation of other drug substrates, such as phenylbutazone and
bishydroxycoumarin. Most importantly, this group, while affirming a
significant heritable determinant of AHH inducibility in human
lymphocytes, failed to confirm the monogenic model and trimodal distri-
bution of AHH inducibility in the general population, proposed by
Kellermann et_ a_l. ; rather, the population distributions for AHH
inducibility (and for plasma antipyrine half-life) were consistent with
polygenic control of both traits in man. In other studies in which the
relation of AHH inducibility to the occurrence of lung cancer was
re-examined by Paigen et^ al., ' low AHH activity was found in
half the tumor patients studied, in contrast with the earlier findings
of Kellermann et_ al_. , and no characteristic alterations in this
enzyme activity were found in the progeny of these patients. A con-
siderable number of technical problems related to the lymphocyte-AHH
assay may confound the results obtained in studies of this enzyme
6-14
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activity and its relation to human cancer, as noted by Kouri e_t
a_l.**- However, recent methodologic advances made by this group,
particularly the use of cryopreserved lymphocytes and close control of
a number of assay variables, have added an important degree of
precision to the assay.
Chrysene, one of several PAH derivatives (benzanthracene is
another), has been shown by Snodgrass et a_K ^4 Co induce AHH
activity in cultured human lymphocytes~Tfrom normal subjects) with BaP
as substrate. The individual variation in the monooxygenase activity
observed with other inducers was also seen with chrysene.
The comparative metabolism of BaP in human lymphocytes and human
liver microsomes has been studied by Selkirk e£ ajU , 16° who examined
the nature of the metabolites formed by each cellular system. The
patterns of metabolites formed in both cell systems had characteristics
quite similar to each other, with some exceptions — for example, among
the derivatives formed in a 30-min incubation, all three dihydrodiols
produced by liver were absent in the lymphocyte incubation mixture. In
a 24-h incubation of lymphocytes, however, all three dihydrodiols
formed by liver microsomes were also formed by the blood cells, and new
metabolite peaks were observed, presumably reflecting more extensive
biotransformation of already formed metabolites in the reaction
mixture. The authors concluded that, although the ratios of some
metabolites may differ and although lymphocytes form several more
derivatives than does liver, many identical metabolites are produced in
these two human cell types.
Schbnwald e_t a_1. studied the effect of BaP on sister chromatid
exchange in mitogen-stimulated lymphocytes of 11 normal subjects and 18
patients with lung cancer. Patients and controls differed neither with
respect to the spontaneous rate of sister chromatid exchange nor in
their responses to the hydrocarbon, although it did double the number
of exchanges in both population groups.
Barfknecht e_t £1^5 9tudied the ability of dichloromethane
extracts of automobile diesel soot at high concentrations (100 mg/m3)
to induce trif luorothymidine-resistant mutants in human lymphocytes
incubated in the presence of rat-liver postmitochondrial supernatant.
A significant induction of such mutants was observed. Anthracene,
phenanthrene, and their alkylated derivatives accounted for one-fourth
of the observed biologic activity. Among eight related compounds,
there was general agreement between responses in lymphoblasts and in
bacterial test systems. Phenanthrene was an exception, in that it was
positive in the human-lymphoblast test system, but negative in bacteria
at a concentration 60 times higher. The data in this report indicate
that methyl substitution at some sites of anthracene and phenanthrene
greatly increases their mutagenicity in both S_. typhimurium and human
lymphoblasts. A similar effect for chrysene has been observed.
Methylationa at the 1 and 3 positions of phenanthrene and the 2 and 9
positions of anthracene result in PAHs that are particularly mutagenic
6-15
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in the human and bacterial test systems used. Methylations at other
positions had the capability of eliminating the mutagenic activity of
the PAH derivative. No correlation between the results of the
mutagenesis studies with the soot-derived PAHs and the reported
capacity of the compounds studied to elicit neoplastic or carcinogenic
responses in test animals could be made.
REPRODUCTION
The title of this section refers collectively to studies related to
the ability of some genital tissues (including the placenta) to
metabolize or otherwise respond biochemically to pAHs. There is an
abundant and detailed literature on transplacental and peri-
natal^2 carcinogenesis. These and related topics in reproduction
were reviewed in a 1981 special issue of the Journal of Environmental
Pathology and Toxicology^- and are not summarized here. It is
perhaps appropriate, however, to refer to the report by Sir Percival
Pott in 1775, in which there was first described an increased
incidence of scrotal cancer in chimney sweeps exposed to soot, and to
note that almost 150 yr elapsed before Yamagiwa and Ichikawa1
demonstrated that the repetitive application of crude coal tar to the
rabbit ear produced skin cancer and that the identification of specific
carcinogenic coal-tar constituents, such as BaP, required the passage
of additional decades.20'^ ' Over this period, the question of why
only scrotal cancers, and not other genital cancers or even other
cancers in general, were found in excess in chimney sweeps appears to
have remained unanswered.
Grover e_t a_1.^6 investigated the metabolism—including the
specific identification of biotransformation products—of three
H-labeled PAHs by nonneoplastic human mammary epithelial-cell
aggregates maintained in culture. The lobuloalveolar units from which
these aggregates are derived are thought to be the site of origin of
many human mammary carcinomas; two of the PAHs studied, 7,12-DMBA and
BaP, are known to be relatively potent mammary carcinogens in rats,
whereas benz[a]anthracene is not a mammary carcinogen in rats. Tissues
from eight patients were studied. The extent of metabolism of the PAHs
is summarized in Table 6-4. There was considerable individual
variation in PAH metabolism among the subjects studied, but the
formation of water-soluble metabolites by the tissue samples accounted,
in each instance, for a major portion of the total of each PAH
metabolized. The extent of binding of each PAH to cellular DNA and
proteins also varied considerably. Interestingly, the extent to which
H-labeled metabolites of benz[a]anthracene—a noncarcinogen for
mammary tissue in the rat—were bound seemed, from'the limited data
obtained, to be consistently lower than the binding displayed by the
other two PAHs. The results of chromatographic characterization of
PAH-DNA adducts formed suggested that, with BaP, the hydrocarbon was
activated by the cultured cells through the formation of antj^-BaP
7,8-diol-9,10-oxide, a bay-region diol-epoxide that appears to be
responsible for most of the nucleic acid adducts formed in several
6-16
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other biologic sysceras. The situation was less clear with 7,12-DMBA,
although a portion of the adducts formed with this PAH
cochroraatographed with adducts present in a DNA hydrolysate that had
been treated with anti-7, 12-DMBA 3,4-diol-l , 2-oxide — a derivative that
is also classified as a bay-region diol-epoxide. The authors
interpreted their data with caution, considering all the factors known
to bear on the development of mammary cancer; but the possibility of
partial causal relationships among the PAHs , their metabolic
transformations, and tumor stimulation is implicit in this work.
Stampfer and colleagues did similar studies with BaP and
cultured mammary epithelial cells and fibroblasts. They showed that
the breast epithelial cells were 50-100 times more sensitive (growth
inhibition) to BaP than the fibroblasts; that the epithelial cells
formed adducts as early as 6 h after addition of the PAH to the
cultures; and that the adducts between the 7R anti stereoisomer of BaP
diol-epoxide and deoxyguanosine predominated at all times and, with two
minor adducts that were consistently present, persisted in the
epithelial cells for at least 72 h in a BaP-free medium. No adducts
were detected in fibroblasts until 96 h after exposure to the PAH, at
which time the type and extent of adduct formation were similar to
those observed with epithelial cells. As with the report of Grover e_t
a_l.,°° caution concerning the direct relation of these findings to
the role of PAHs in mammary carcinogenesis is necessary. On this
matter, Stampfer and co-workers *-°° stated, however, that "chemical
carcinogens, particularly BaP, should not be minimized as possible
factors in the initiation of breast cancer."
Mass e_t £_!• studied 26 specimens of normal human endometrium
to determine the patterns of metabolism of [ H]BaP in short-term
explant cultures. Three of the tissue samples were from postmenopausal
women; of the remaining 23, it was possible to approximate the stage of
the menstrual cycle at which the tissue was removed during surgery.
Eight of the latter subjects were smokers. In summary, it was clear
that normal human endometrium could enzymatically convert BaP to a wide
variety of oxygenated derivatives that cochromatographed with
dihydrodiols , quinones, and monohydroxy products of the PAH; sulfation
was also identified. HPLC analysis of metabolites revealed marked
individual variation in metabolite formation among the subjects
studied; smoking did not account for this difference, but some evidence
of hormonal influences on the patterns of PAH metabolism was adduced.
In a study by Dorman e_t a_l.,^ BaP binding to DNA in human
endometrial tissue was studied in samples obtained from 41 subjects
and, again, a striking (70-fold) range in the observed specific
activities of carcinogen binding to DNA was identified (see Figure
6-1). Tissues obtained late in the proliferative phase or early in the
secretory phase of the menstrual cycle had the highest mean specific
activity of PAH-DNA binding (Table 6-5). Binding was significantly
reduced when tissue specimens from low-estrogen periods of the
menstrual cycle were studied. The reason for this apparent association
between estrogen content (actually, the estimated phase of the cycle)
6-17
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and PAH-DNA binding is obscure, but clearly merits further study. Such
study would have to deal with the important confounding factor of the
broad range of individual variation in binding, which may mask
systematic but small changes that can occur during a menstrual cycle
but which cannot now be detected.
1 7 ft
Namkung and Juchau1^0 studied the oxidative biotransformat ion of
BaP in preparations of human placental microsomes with HPLC. The
investigations revealed that the use of substrate concentrations high
enough to ensure zero-order reaction kinetics markedly inhibited the
formation of dihydrodiols in the reaction mixtures. The relative
quantities of dihydrodiols generated increased with decreasing
substrate concentrations between 200 and 2.7 uM. Addition of manganese
or ferric ions to reaction mixtures altered the ratios of generated
phenols to dihydrodiols. Identical results were obtained with ^C-
and -^-labeled BaP as substrate. The data suggested that
considerable amounts of 7,8-dihydroxy-7,8-dihydro-BaP, a proximate
rautagen-carcinogen, may be generated in vivo by placental tissues of
women who smoke.
The formation of PAH metabolite-nucleoside adducts when human tumor
placental microsomes were incubated with HH]BaP and salmon sperm DNA
has been studied by Pelkonen and Saarni.^^ There were significant
differences between the PAH metabolite patterns and the nucleoside-
metabolite complexes formed, compared with rat liver, for example.
Specifically, in the human placenta microsomes, the absence of the
nucleoside complex of 9-hydroxy-4,5-oxide implied the inability of this
tissue to form 4,5-oxides of BaP. Indirect evidence of epoxide
hydratase activity in placental tissue was obtained. The extent of
PAH-DNA bine rig in mis tissue correlated significantly with both
7,8-diol mec^oolite rorraation and fluororaetrically determined AHH
activity. The question of whether the 7,8-diol-9,10-epoxide of BaP is
formed by the human placenta in vivo could not be answered unequi-
vocally, but the authors' inferential conclusion is that it is probably
formed in the human host. The interplay of possible genetic influences
and clearly established regulatory influences of environmental factors
on human placental AHH has been incisively discussed by the same
group.13®
Cigarette-smoking has been shown by Conney and
associates^' 189,190 co ^e one ojr Cne raogt pocent and consistent
inducers of human placental AHH activity yet identified. In the
initial report of the group,^°* the enzymatic hydroxylation of BaP
could 'not be detected in nonsmokers in homogenatea of placentas frozen
immediately after birth and studied within 48 h. In contrast, the
enzyme activity was present in all 11 placentas from women who smoked
during gestation, although enzyme activity in this small group did not
correlate with the number of cigarettes smoked. BaP administration to
pregnant rats also was shown to induce AHH activity in the placenta.
The effect was related to PAH dose. This study constituted the first
demonstration that compounds in cigarette smoke could induce a
carcinogen-metabolizing enzyme in human tissues. These studies were
6-18
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extendedl90 to related enzymatic reactions in human placentas and to
other types of pyrenes as probes for AHH-inducing activity in rat
placenta (see Table 6-6). Extremely active inducers included chrysene,
1,2-benzanthracene, pyrene, 3 ,4-benzofluorene, and a number of related
compounds.190 The wide variability in the induction of AHH activity
in human placentas is exemplified by the data in Table 6-7—a range in
activity of the enzyme in smokers approaching 1,000-fold (a nearly
2,000-fold range if smokers are compared with nonsmokers). The basis
for this extreme range of responses to a chemical exposure (15-20
cigarettes/d for each subject) is not known. However, data presented
by Harris et al.'0a suggests that pulmonary alveolar macrophages can
metabolize BaP to proximate and ultimate mutagens released into extra-
cellular space.
LUNG
The respiratory tract comprises an extremely disparate and complex
set of tissues containing some 40 different cell types.^6 ^s
Devereux e_t a_l.^' have noted, whereas pulmonary cytochrome P-450 and
the metabolism of xenobiotics have been studied with various
preparations of lung tissue (microsomes, isolated perfused lung, cells
obtained by pulmonary gavage, direct instillation of xenobiotics in
various portions of the respiratory tract, etc.), little is known about
the localization of the cytochrome P-450 monooxygenase components in
the pulmonary system. This section deals exclusively with the
metabolic properties of human respiratory tissues with respect to PAH
metabolism, but the lack of information just cited needs to be kept in
mind. There are facets of the investigation of Devereux e_t a_l_.^' in
rabbits that probably bear significantly on problems of human pulmonary
tissue biotransformations that depend on cytochrome P-450; these
aspects include the observation that the alveolar macrophage that
accumulates PAH has little or no measurable cytochrome P-450 or
monooxygenase activity5**, 71,148 an(j that there is selective cellular
distribution of cytochrome P-450 species.
The ability of human bronchial epithelial cells to bind and
presumably to activate such PAHa as 7,12-DHBA, 3-MC, BaP, and
dibenzfahjanthracene was described by Harris and colleagues in
1974.™ Four tissue samples were studied (one control and three lung
cancer) in explant cultures, and radiolabeled PAHs were used;
radioactivity from all four compounds tested was found in both
cytoplasm and nuclei and in all tissue samples studied (see Table
6-8). The number of tissues examined precluded comparisons between
normal and tumorous lung PAH metabolism, and no studies of PAH-DNA
adduct identification were carried out, although, as noted, radio-
activity from the labeled PAHs was found tightly bound to DNA isolated
by CsCl gradient. A more detailed study by this group195 used
tissues obtained from an additional four patients, three of whom had
pulmonary malignancy.
Explants of human bronchi also metabolized BaP and released
derivatives that are mitogenic in the Chinese hamster V-79 cell
line.'2 The 7,8-diol of BaP was approximately 5 times more potent as
a promutagen than the parent PAH; binding of the diol to DNA was 5-20
6-19
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times greater than that found with BaP. When 13 samples of bronchial
cells were studied with cloned Chinese hamster V-79-4A cells, a
positive correlation between DNA-PAH binding (in the cultured bronchial
cells) and induction of Or (ouabain-resistant) mutants was found, but
no correlation between this mutation frequency and AHH activity was
identified. This may be attributable, as the authors noted, to the
difficulty in correlating AHH activity with the consequences of the
multistep pathway of metabolic activation for BaP. The individual
variation in mutation frequency was 9-fold, and the variation in
binding of PAH to DNA 5-fold. This important investigation pointed the
way toward study of the metabolic activation of chemical carcinogens
into promutagens and mutagens directly in differentiated epithelial
cells derived from human tissues; and the human tissue-mediated rautagen
assay opened the possibility of testing the hypothesis that people
differ in mutagenic and oncogenic susceptibility to environmental
chemicals, depending on individual capacity to activate and deactivate
chemical procarcinogens. Autrup e_t aj.^ compared the metabolism of
BaP by cultured tracheobronchial tissues from humans and four other
species (mice, hamsters, rats, and cows). They provided evidence that
the metabolism of BaP is qualitatively similar in tracheobronchial
tissues from humans and from animal species in which PAHs have been
shown experimentally to be carcinogenic.
A similar study limited to a comparison of human lung microsomal
fractions and'rat microsomes was carried out by Prough et_ a_l- The
results indicated that human microsomes form a higher percentage of
dihydrodiol products from BaP than do rat microsomes. The wide
variation of PAH metabolite profiles formed by the 15 samples of human
lung studied may be due in part to differences in clinical diagnosis
when the samples were obtained. Bronchial tissues cultured in a
chemically defined medium were exposed to radiolabeled BaP or its
metabolites, and their binding to DNA was measured. Radiolabeled
metabolites were prepared by incubating the parent PAH with rat liver
microsomes and then purifying and identifying with silica gel and
HPLC. The binding data showed that (-)-trans-7,8-diol bound to
bronchial mucosal DNA to a considerably greater degree (5- to 23-fold)
than did BaP; binding was also much greater (25- to 80-fold) than with
the (-)-trans-9,10-diol. The trans-7,8-diol constituted 3-6% of the
total identified metabolites when human bronchi were exposed to BaP.
Diol-epoxidea were formed from (-)-trans-7,8-diol in two of the
bronchial explants, and strong evidence was provided that the major
tumor bronchial mucosal DNA-binding BaP metabolite is in fact derived
from (-)-trans-7,8-diol.*95 The specific adducts formed between DNA
and the metabolic intermediates of BaP were not isolated, but the
author concluded that the predominant bound metabolite is a single
enantiomer of diol-epoxide I derived as indicated above.
In an extension of their earlier work, Harris and colleagues"'
examined the metabolism of BaP and 7,12-DMBA in explants of human
bronchus and made a metabolic comparison with human pancreatic duct
explants. As in the prior study, both normal and malignant human
bronchi (37 subjects) metabolized BaP actively and in generally similar
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fashion, except for a higher percentage of organic-solvent-extractable
metabolites formed by bronchi from noncancer patients. In addition,
prior exposure of the bronchial explants to benz[a]anthracene altered
the qualitative features of the metabolite profile of BaP, as analyzed
by HPLC. Benzfa]anthracene specifically increased the binding of BaP
to cellular DNA and the activity of AHH. Among a group of 28 of the
patients' tissues studied, 7,12-DMBA was bound to DNA more often (26 of
28) than BaP. In che comparison with pancreatic duct explants,
7,12-DMBA-DNA binding was consistently lower in the latter tissue than
in the bronchial explants.
Cohen e_t a_l. showed, with cultured human bronchial epithelium,
that BaP was converted promptly to metabolites that cochromatographed
with 9,10-dihydro-9,10-dihydroxy-BaP and 7,8-dihydro-7,8-dihydroxy-
BaP. Similar results were obtained with human lung cultures, except
that a major metabolite, benzofajpyrene-3-yl hydrogen sulfate, was
identified. The biologic activity of this sulfate ester of 3-hydroxy-
BaP is of interest, because, owing to its physicochemical properties,
it could be extremely persistent in man.
Covalent adducts between DNA and BaP in treated cultured explants
of peripheral human lung tissue and in the continuous human alveolar
tumor cell line were identified by Shinohara and Cerutti.^°^ From
the chromatographic analysis of digests of DNA extracted from these
tissues, it was concluded that both the lung specimens and the human
alveolar tumor (A549) cells metabolized BaP to diastereomeric
7,8-dihydroxy-9,10-epoxytetrahydro-BaP intermediates that mostly
reacted with the exocyclic amino groups of deoxyguanosine to form
N2-(10-f78,8ci,9a- and 96-trihydroxy-7,8,9,10-tetrahydro-
benzo[a]pyrene]yl)deoxyguanosine (dGua-BaP I and II). Although
comparable amounts of dGua-BaP I and II were formed in A549 cells,
dGua-BaP I was the predominant adduct in the DNA of lung specimens from
six different donors.
The wide range of metabolic capacities for PAHs exhibited by other
human tissues studied also extends to. lung tissue, as shown by Cohen e_t
al. They observed a 44-fold variation in the ability of short-term
organ cultures of peripheral lung tissues from human cancer patients to
metabolize BaP to organic-solvent-soluble derivatives. The total
amounts metabolized ranged from 1% to 96.1T> in a 24-h culture period.
The authors concluded that, although caution must be exercised in
measuring metabolic activities of human tissues derived from diseased
patients, the use of short-term organ explant cultures mimics the in
vivo metabolic disposition of PAH better than the use of lymphocyte AHH
activity would. A solution to the practical problem of obtaining lung
tissue from large populations to study the validity of this conclusion
is not apparent.
Kahng et al. concluded from a study of 11 immediately autopsied
subjects that bronchial tissue exposed to benz[a]anthracene produced
induction responses of AHH that correlated with induced AHH activity in
raonocytes from the same subjects. A reconfiraation of the wide range
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of individual di: -jrences in AHH activity of surgically obtained
specimens or normal lung tissue (86 subjects) came from a detailed
study by Sabadie e£ £l. Briefly, AHH activity was lower Chan
normal in tumorous lung sections in 73 of the 86 patients; and in 21
tumor tissue samples, no AHH activity was detected at all. Individual
variation (excluding the 22 subjects) in lung-tumor AHH activity wa ?
20-fold, which approximated the variation observed in other studies
including those in which PAH-DNA binding and pulmonary tract tisaue.v
were studied. BaP metabolite formation was analyzed, and che resales
generally conformed with the data of other investigators.
Interestingly, BaP (but not pyrene) induces AHH and prolyl
hydroxylase activity in neonatal rat lungs in organ culture.'^
Because prolyl hydroxylase is an indicator of collagen synthesis and
increased activity of this enzyme in lung reflects increased collagen
formation, the authors, Hussain &t_ aK , hypothesized that the earliast
events in BaP-induced lung injury may include alterations in collagen
metabolism. In a study of the effect of tobacco-smoke compounds on the
plasma membrane of cultured human lung fibroblasts, Thelestam a_t_
a_U^5 examineci 464 compounds, of which nearly one-fourth gave risa
to severe membrane damage. PAHs proved inactive in this test system;
the PAHs tested included anthracene, benz (a]anthracene , chrysene,
pyrene, BaP, perylene, f luoranthene , and coronene. The significance of
these findings is not entirely clear, but, inasmuch as very large
concentrations of the compounds were used (25 mM) . the failure of all
PAHs tested to cause substantial release of the radiolabeled nucleocida
material from the cells suggests that PAH entry into cells of organs in
which their carcinogenic potential is expressed does not require as an
initial event plasma membrane damage by the active chemical species.
Lung damage by ozone-^' and nitrates ^"^ showed contradictory
effects: in the former case, adaptation may become apparent, and, in
the latter, susceptibility to infection may increase. In the case of
asbestos-produced damage, as well as damage produced by other
particles — such as iron oxide, silica, and carbon black — cellular
uptake and availability of BaP increase. '"'* Asbestos, of the
several particles tested, was particularly effective in increasing
microsomal uptake of the PAH, although clearly adsorption of the PAH on
the particles — rather than simple mixture of the two — is required for
the increase in cellular uptake to become evident. The relevance of
these findings to the phenomenon of particle-PAH cocarcinogenesis is
clear. * BaP elution from typical soot from pollution sources, as
well as from soot in lungs (11 cases), has been carefully studied by
Falk £t a_1.56 strikingly, this PAH could not be recovered from soot
in human lungs without malignancy (Table 6-9). whereas the
noncarcinogen pyrene could be identified (in much lower concentration*
than expected). Adequate controls appeared to ensure that tha
disappearance of the carcinogenic PAH was a biologic phenomenon taking
place in vivo; the authors concluded that elution must have occurred in
the host through an undefined mechanism. In another study,
involving 21 bronchial carcinomas, a search was made for 12 PAHs in the
tissues with chromatographic and fluorescence techniques. Only four of
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the 12 PAHs sought were found: BaP, fluoranthene, perylene, and
benzo[b]fluoranthene (Table 6-10). BaP was found in all tumors;
fluoranthene and banzofb]fluoranthene were sometimes present, as was
perylene. Coronene, dibenz[ah]anthracene, pyrene, benz[a]anthracene,
chrysene, benzotghi]perylene, benzo(k)fluoranthene, and benzofejpyrane
were, if present, below the limits of detection.
HUMAN EXPOSURES TO PAHs: A BRIEF SUMMARY
The studies reviewed in the preceding sections were related
primarily to the metabolic interactions of PAHs and human tissues and
focused principally on the oxidative reactions known to convert ntanv of
these compounds to potent mutagens and carcinogens. This section
reviews a number of reports dealing with possible detrimental health
effects of specific workplace exposure to PAHs and representative
reports dealing with PAH contamination of the aquatic environment and
of foods. The literature on atmospheric exposure to PAHs is dealt with
elsewhere, except for exposures that are discrete and intense, as in
some working environments. In the light of this review, one cannot
avoid the conclusion that the greatest present source of human PAH
exposure is through the gastrointestinal tract; nor can one disagree
with the statement in the 1970 Royal College of Physicians report'-3
that, to the extent that PAHs are involved in the genesis of pulmonary
malignancies, "by far the most important matter affecting all . . .
aspects of mortality from lung cancer is smoking." The equally
emphatic conclusion of Pike and Henderson that "the epidemiologic
evidence implicating cigarette smoking as the major cause of lung
cancer is overwhelming" puts the clinical studies reviewed here related
to the potential pulmonary hazards of atmospheric PAHs in proper
perspective.
WORKPLACE EXPOSURE
Schenker in 1980^" reviewed the question of whether diesel
exhaust is an occupational carcinogen and summarized a number of the
principal studies (Table 6-11) on the question of cancer incidence in
populations of workers exposed to diesel exhaust. Data on environ-
mental and occupational BaP and total suspended particles in various
urban and rural sites and specific occupations were also provided
(Table 6-12). These epidemiologic data emphasize the conclusion that
"the carcinogenicity of workplace exposure to diesel engine exhaust is
suggested ... but the existing data are sparse and contradictory."
Table 6-11 shows only concentrations of BaP, and the values are in
units of micrograns per 1,000 cubic meters. Because the air breathed
by a normal adult approximates 15-20 m /d, the highest PAH concentra-
tion shown indicates a potential exposure dose of about 700 -g/d in a.
work setting (coal and pitch-coking plant) known to have one of the
aiost intense PAH exposures. This figure exceeds by orders of magnitude
the exposure produced by the heaviest smoking, and such an occupational
locale would thus be expected to elicit detrimental and clearly
detectable health effects in man. The same consideration applies to
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the data on workers in gasworks retort houses and roof tarrers. But
beyond these specific occupational sites, the respiratory intake of
BaP — even if, for occupational purposes, a person had to remain for 24
h/d in Blackwall Tunnel, London (Table 6-11) — would approximate that
from about a pack of old-style cigarettes per day. The improbability
of such occupational exposures emphasizes the difficulty of measuring
the health hazards of atmospheric PAH sources in the general sense
(i.e., in the 28 rural and 24 urban sites depicted in Table 6-11).
A number of occupational-epidetniologic studies have emphasized the
difficulties of reaching firm conclusions with respect to the direct
(or measurable) health risks of PAHs in work environments, whether the
suspected hydrocarbon comes from diesel or other automotive exhausts or
from chemicals, such as petroleum sources, that are intrinsic in the
occupation itself. Battigelli e_t a_l. studied 210 locomotive
repairmen (average age, 50 yr; average work period, 10 yr) considered
to be regularly exposed to diesel exhaust and 154 "control" railroad
workers. The studies were carried out in two railroad shops in
Pittsburgh, Pa. The clinical data were scanty, and it was not possible
to differentiate the exposed from the nonexposed worker population on
the basis of pulmonary-function tests. However, smoking clearly
impaired the pulmonary functional performance of workers. A somewhat
comparable environmental study carried out by El Batawi and Noweir"
in two diesel-bus garages in Egypt raised the possibility of clinically
detrimental, synergistic effects of smoke and acrolein gas, which is
known to be present in exhaust of diesel engines. Ventilatory-function
changes over a workshift in coal miners exposed to diesel emission were
studied by Reger ej: a_l. ; the only positive finding in this study
of 800 men was that smokers suffered consistently greater pulmonary-
function decrements over a workshift than nonsmokers. In a retro-
spective study of mortality statistics, Kaplan could identify no
higher than normal rates of death from bronchopulmonary carcinoma in
workers exposed to fumes from diesel engines among the medical records
of 6,500 deceased railroad workers, including 818 deaths from malignant
diseases .
Lloyd e_t a_1. reported that the mortality from respiratory
cancer for men employed in a coke plant was twice the rate generally
observed among steelworkers ; the whole difference was accounted for by
a threefold excess for nonwhite workers. A more detailed analysis1-08
showed the following: The excess of respiratory cancer previously
reported for coke-plant workers was limited to men employed at the coke
ovens, the relative mortality for this disease being 2.5 times that
predicted. The greatest part of the excess was accounted for by an
almost fivefold risk of lung cancer in men working on the tops of the
coke ovens. A 10-fold risk of lung cancer was observed for men
employed 5 yr or more at full-time topside jobs; 15 lung-cancer deaths
were observed among the 132 men in the topside group, compared with 1.5
expected. The apparent differential in respiratory-cancer rates for
white and nonwhite coke-plant workers reported in an earlier paper was
accounted for by differing distributions by work area and the unusually
high lung-cancer risk for topside workers; lung-cancer mortalities for
6-24
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white and nonwhite coke-plant workers employed at work stations other
than topside were comparable. A deficit of deaths from heart disease,
previously reported for similar occupational groups, was also seen for
coke-oven workers. Coke-plant workers employed only in nonoven areas
may be at excess risk of digestive cancer.
A review of the literature on cancer mortality of men employed in
the coal-tar industries showed that all these occupations evidence
excess cancer at one or more sites. The lung-cancer excess in coke-
oven workers also was observed in other groups engaged in coal carbon-
ization, and it appeared that the lung-cancer response was positively
correlated with the temperature of carbonization.
Among coke-oven workers studied by Mazumdar e_t_ a_l., excessive
deaths from respiratory malignancy were reported. As in the study of
Lloyd e_£ a_l . , there was a tendency for the death rates of nonwhite
workers to be higher than those of white workers. Measured concentra-
tions of coal-tar pitch volatiles in the environment of men who worked
at the top of coke ovens were 2-3 times higher than in that of men
employed at the side of the ovens. High BaP emission, among others,
has been measured in the gaseous discharge--including the coal-tar
pitch volatiles — of coke ovens in the steel industry, a rough estimate
being that 1.8 g of this chemical is emitted per ton of coke
produced. ^ As in the Lloyd e£ a_K study,10' the overall
cancer-death risk for coke-oven workers was distinctly higher than that
for normal persons in the age group over 55 yr, and the age-adjusted
death rates for lung cancer showed a strong relationship between extent
of exposure to coal-tar pitch volatiles and lung-cancer mortality. The
lowest-exposure group11^ nad death rates similar to those of nonoven
workers, but all higher-exposure groups had age-adjusted rates that
ranged from 3 to 10 times those of the comparison group with increasing
exposure. The data in this study also confirmed the long latency in
cancer formation, even under the conditions of high exposure to
carcinogens characterizing coke-oven workers; the time between first
exposure to coal-tar pitch volatiles and death from lung cancer varied
from 10 to 40 yr, with an average of 25 yr.
Toxicologic experience with workers in the developing shale-oil
industry is incomplete, although historical evidence indicates that
potential health hazards related to malignancy may exist in the
processes involved in oil extraction. Some data on the content of
BaP and pyrene analogues from shale materials, as reported by Weaver
and Gibson, are useful to record here (Tables 6-13 and 6-14).
Because the industry is still in its developmental stage in this
country, the overall health impact that may be attributed to exposure
to these PAHs — as well as to other contaminants, such as arsenic,
beryllium, cadmium, lead, mercury, and nickel1" — is difficult to
estimate.
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EXPOSURE TO PAHs VIA THE GASTROINTESTINAL TRACT
The exposure of humans to PAHs may be considered to be almost
exclusively via the respiratory and gastrointestinal tracts. Some
occupational groups (e.g., the chimney sweeps studied by Pott) may have
an intense local cutaneous exposure to PAHs, but the significance of
percutaneous absorption of these compounds for the general population
is not known. Such substances as the polychlorinated biphenyls"^ and
constituents of coal tar^"- can pass through the skin and induce liver
oxidative enzymes in animals, so it is possible for some (undoubtedly
small) degree of PAH accumulation to occur in humans systemically via
skin exposure.
Several major reviews of the importance of water and food as
vehicles of human exposure to PAHs have been published in the last 5
yr. These include a special issue of the Journal of Environmental
1 ^A ~^^™—^—^^^"^^~"~—~"~~™^~—~—^~^~^—^—
Pathology and Toxicology devoted to the health aspects of PAHs and
several monographs focusing on PAHs in drinking-water sources and on
PAHs in the marine environment. *•'• ^9 >
PAHs in Water
It can be stated at the outset that human exposure to PAHs through
the ingestion of water is quantitatively insignificant, compared with
exposure through food—the contribution of drinking water is estimated
to be only about 0.1% of the total PAH derived via the gastrointestinal
tract in humans. This estimate, carrying with it an implicit
assumption of relative biologic safety (at least compared with foods as
a source of PAHs). is probably valid except perhaps for some
surface-water sources, which, because of location (e.g., downstream
from shale-oil effluent or coke-byproduct discharge sites—see Table
5-12 of Santodonato e_t a_l. ^4), may be heavily contaminated by such
PAHs as BaP. Groundwater concentrations of this prototype PAH
determined in multiple German and American sources are extremely low
(see Table 5-11 of Santodonato st_ a_l_. ), ranging from a fraction of
a nanograra per liter to several nanograms per liter. The average
"total" PAH content is, of course, greater, but still in the same
range. In contrast, low- to medium-concentration contaminated surface
waters may contain PAHs 5-20 times higher, and this pollution may be
increased by several orders of magnitude in sewage water or in surface
waters adjacent to industrial sites. Treatment of surface water to
obtain drinking water can nevertheless remove the bulk (95% or more) of
the PAHs, particularly with activated-carbon filtration. This reflects
the fact that much of the PAH in water subject to pollution is quickly
adsorbed on suspended solids or is found in sedimented particulate
matter. The majority of PAH entering surface water is concentrated
locally; although PAH can probably be considered ubiquitous in water,
the amounts involved are substantially lower than those found in air or
129
on land. Neff has pointed out that, if all PAHs found in the
aqueous environment were distributed evenly throughout the oceans and
fresh-water bodies, they would be undetectable and inconsequential.
6-26
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As noted, the PAH content of drinking water is, with an occasional
exception, low, as expressed as BaP and total PAH (Table 6-15).^4
Among the general class of PAHs, the compounds that have been detected
by high-resolution gas chromatography after extraction from
tapwater^^ are listed in Table 6-16 with their concentrations. Such
contamination at a typical, most proximate (tapwater) drinking-water
source represents only trace contamination, compared with the PAH
content of original fresh-water sources, marine and estuarine waters,
fresh-water and marine sediments, and some alcoholic beverages.^7^
The occurrence of PAHs in saltwater sources has for several reasons
more potential biologic importance than the occurrence of these
compounds in drinking water. The oceans provide a very large surface
area for deposition of airborne PAHs via rain and dry fallout. Runoff
of PAHs from the land surface also contributes substantially to
marine-water content, as do direct effluents from sewage and industry.
Carcinogenic PAHs occur in crude and, particularly, refined oils,3
and oil spills may contribute in a major way to marine pollution with
these compounds, especially on a local scale. The oceans constitute an
ecosystem in which varied animal and plant life can participate in the
metabolic processes involved in the uptake, storage, concentration,
biotransfortnation, and discharge of PAHs. Thus, the consumption of
fish and shellfish of predominantly saltwater, compared with
fresh-water, origin (88Z vs. 12? of the seafood in the diet) gives
special importance to the PAH contamination of the aquatic environment
that these food species inhabit.
PAHs are universally, although unevenly, distributed throughout the
marine (saltwater) environment. They are derived principally from
atmospheric fallout, terrestrial runoff, and spills of petroleum pro-
ducts. The contribution, if any, of marine organisms to PAH pollution
by
-------
depending on the depth and turbidity of water and other factors; hut
persistence of PAHs is much greater in water than in air, because the
particulate matter on which these compounds are mostly adsorbed
provides a storage pool from which they may be slowly returned to the
water by leaching or through biologic processes involving marine
organisms.
The characteristics of marine pollution by PAHs are such as to
suggest the occurrence of multiple varieties of discrete ecosystems
with relatively high concentrations of these compounds in sediments and
local plant and animal species—all existing in a vastly larger aquatic
environment characterized by a smaller degree of PAH contamination. In
the local marine areas of high PAH pollution—principally river basins
and estuarine and coastal waters—the degree of PAH contamination and
the PAH composition in water, sediments, and nonraigratory marine life
are determined by the nature of the point sources of contamination. In
the organisms found in these areas, the PAH composition depends on
metabolic processes related to the selective bioconcentration,
biotransformation, and accumulation of the PAHs or metabolites or their
discharge into the aquatic environment.
The fate of PAHs in marine ecosystems has been studied by Lee e_t_
a_K , 01 who used as a model Prudhoe crude oil enriched with a number
of PAHs dispersed into a controlled ecosystem (polyethylene enclosure 2
m wide and 15 m deep) suspended in Sadnich Inlet, Canada. The oil was
estimated to contribute PAHs at concentrations ranging from BaP at
100 ug to naphthalene at 300 x 10 ug per 100 g. Multiple water and
sediment sampling, microbial-degradation studies, analysis of bio-
accumulation by oysters, and analysis of adsorption to sediments with
[ C]PAH were carried out. The results demonstrated a rapid, marked
decrease in PAHs from water (half-life, 3-4 d) and a variable recovery,
depending on the PAH, in the sediment. For the low-molecular-weight
PAH naphthalene, this recovery was only 7% after 1 wk; for BaP, it was
39%. Oysters rapidly took up all PAHs, but released naphthalene to
such an extent that it was not detectable in the organisms after 23 d.
In contrast, benz[a]anthracene and BaP were released much more slowly,
with estimated half-lives (assuming exponential discharge) of 9 and 18
d, respectively. Thus, the higher-weight PAHs persisted much longer in
the organisms than the lower-weight PAHs. Other degradation studies
involving mussels collected from oil-contaminated waters also have
shown the persistence of the higher-molecular-weight PAHs.51-'53
Evaporative loss of lower-weight PAHs, such as naphthalene, in the
upper waters was expected, whereas this would be limited for higher-
weight PAHs. Microbial degradation of naphthalene and anthracene was
measurably increased in oil-contaminated water, compared with control
water (4 h vs. 48 h, respectively, for appreciable degradation)—a
finding consistent with those of other studies showing higher numbers
of oil-degrading microorganisms in polluted than in control or
unpolluted waters. ' Photochemical degradation of PAHs was
inferred; for BaP, this was considered to account for an amount that
could approximate about 50% of the compound, inasmuch as no microbial
degradation of the compound was demonstrated and 40% was recovered in
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Che bottom sediment. The study permitted several conclusions that
probably have general relevance. The half-lives of PAHs in marine
waters are short (a few days); for lower-weight PAHs, microbial
degradation and evaporative loss may be primary removal processes; for
higher-weight PAHs, such as BaP, sedimentation and photodegradation are
the most important removal means; and, by inference, for higher-weight
PAHs after sedimentation, biologic degradation and interaction between
plant and animal life in the sediment are important factors in removal.
These-processes (biologic degradation and interactions) have been
extensively studied with a wide variety of aquatic species. It is clear
that, as with terrestrial fauna, the capacity of marine animal species
to effect the metabolic transformation of PAHs can be considered to be
universally distributed. Reviews of the results and other aspects of
such studies have been published elsewhere,3>^i37|38,129,181 ancj on^y
representative reports are summarized here. PAHs in the marine
environment can be metabolized by aquatic bacteria and fungi;^29 for
some species of bacteria, a monocyclic aromatic hydrocarbon, such as
benzene, can serve as a sole carbon source. PAHs, such as BaP and
benz[a]anthracene, can also be oxidatively metabolized to hydroxylated
derivatives comparable with those produced in the livers of
vertebrates. PAHs can be degraded to CC>2 to a considerable degree
(13-68%*-29) by aquatic microorganisms. PAH metabolism by fungi also
occurs; these organisms contain cytochrome P-450 and can carry out the
initial oxidative metabolism of PAHs in a manner resembling that
catalyzed in vertebrate liver. Marine fungi isolated from oil-polluted
water or oil slicks have a substantial ability to assimilate petroleum
hydrocarbons, and this hydrocarbon-degrading capacity can permit use of
a PAH as a growth substrate.
Fish and crustaceans (and some worms) can oxidize PAHs—as measured
by AHH activity—and cytochrome P-450 has been identified in a number
of these species. Most oxidative metabolism in these aquatic animals
is in the liver, as it is in mammals. Induction of cytochrome P-450
(not always correlated with an associated P-450-dependent increase in
chemical oxidation) in fish has been produced by benzfalanthracene,
chrysene, BaP, and other organic substances-^ |61| {29,140,169 to wni.ch.
fish may be exposed in their natural environments or under experi-
mental conditions. The products of the oxidative metabolism of PAHs in
fish resemble those produced in mammalian liver and include diols,
epoxides, phenols, quinones, and all principal types of conjugates
formed from PAH metabolites in mammalian liver.
Seasonal changes in P-450-dependent oxidation have been reported in
fish, * and alterations in this enzymatic activity have been related
to ambient temperature, food status, and exposure to inducing chemicals
in their natural habitat.^8,49 Apart from carrying out biotrans-
formation, the capacity of marine species to accumulate and discharge
PAHs from the surrounding waters is important in relation to the
pattern of distribution of these compounds in the marine environment
and to the use of marine species as food, in view of their contribution
to the exposure of humans to PAHs via the gastrointestinal tract.
6-29
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Marine animals readily accumulate PAHs from Che surrounding waters
and can discharge both the untransformed PAHs and their metabolic
products into the aqueous environment. The rates of release of
accumulated PAHs may vary substantially from species to species (aad
compound to compound), and half-lives can range from hours to many
days. The substantial concentration gradients of PAHs that may occur
between an organism and its aqueous environment can have importance for
man in relation to marine species that are eaten by man or by edible
species . ^' ^2 > '•"•^ Whether these concentration gradients involve an
active uptake mechanism is not known; but they do not depend solely on
solubility, inasmuch as polar metabolites of a PAH can be retained
longer than the more lipophilic parent compound. ^ This may be due
to the electrophilic nature of these metabolites and their consequent
binding to tissue macroraolecules.^'
Oysters have been shown to concentrate hydrocarbons from diesel-oil-
contaminated waters to concentrations over 300 ug of total hydrocarbons
per gram of wet weight over a 7-wk period. ' These hydrocarbons
were rich in aromatics, compared with the contaminating oil. In clean
seawater, the hydrocarbon concentrations decreased dramatically (by 90Z
in 4 wk). Other marine species show the same biologic characteristics,
although uptake and release of accumulated hydrocarbons vary. The con-
centration factor (i.e., tissue vs. water concentration) may reach
1,000-fold in marine animals that cannot escape a contaminated
environment. The potential importance for humans of this capacity for
bioaccumulation in edible marine species is evident. PAHs can, as
expected, accumulate rapidly in fish from contaminated sediments, as
McCain e_t a_l. ^ have shown, although this process is less efficient
than uptake from water.
The biologic impact of contaminating PAHs on marine species has
been thoroughly reviewed recent ly^' • -*° >^° ' *•*' ^9 and is not
summarized here. Toxic effects of these and related pollutants have
been described across the spectrum of marine life, from bacteria and
fungi to plants and animals; and they range from the "tainting" of
commercial species^ >^ to the development of cancer and cancer-like
growths in aquatic animals. °>°3
PAHs in Food
The exposure of humans to PAHs from dietary constituents greatly
exceeds that from any other sources except specific hazardous occupa-
tional settings. PAHs are ubiquitous contaminants of foods and—
depending on the extent of atmospheric and soil pollution in crop areas
and on methods of processing, preservation, and preparation—can become
highly concentrated in selected foodstuffs. At least 100 types of PAHs
have been identified in foods. ^"- Some of these have been shown to
have well-defined carcinogenic properties in experimental animals.
Epidemiologic studies have suggested an association between the consump-
tion of high-PAH foods and gastrointestinal malignancies in selected
6-30
-------
populations, *®^> 165,177 ^ut £t ^g difficult Co extend this associa-
tion to the general population or to define the biologic risk of PAHs
in foods in more direct terras. Nevertheless, the quantitative dimen-
sions of PAH exposure via the diet and the established carcinogenic
potential of some of the compounds frequently identified in foods
suggest that the health risks from this source of exposure, although
still incompletely defined, may be important for various groups.
Edible marine species may contain variable amounts of PAHs derived
principally from polluted terrestrial runoff waters, from marine sedi-
ments, and from petroleum-contaminated aquatic environments. As noted
above, such environments are largely in-shore (e.g., estuaries and
river basins), with pollution diminishing rapidly in the open seas.
Bioaccumulation of PAHs in the marine food chain may be substantial in
some fauna, and, of course, national predilections for such modes of
seafood preparation as smoking ^> 64,67 > H^» ^6 can j_ncrease to high
values'the content of PAHs in such foods. The potential for bioraagni-
fication of PAHs in aquatic food chains is clear, but the extent to
which this process results in contamination of seafood ingested by
humans is not known (the subject has been reviewed by Neff^9). For
some crustaceans and fish, PAH uptake through the food chain can be
more efficient than uptake from the surrounding waters ,^> ^2, ^3 and
storage of such compounds in these species can be substantial. PAHs
thus stored may or may not undergo extensive biotransfonnation. The
processes of storage, uptake, metabolism, accumulation, and excretion
have generally large interspecies variation; but crustaceans appear
relatively efficient in their uptake of PAHs from food and other
sources. Table 6-17 shows an analysis of PAHs in oysters
collected in a moderately polluted harbor area by Cahnmann and
Kuratsune. 0 The comparative BaP and benzanthracene contents of a
variety of foodstuffs are shown in Table 6-18. (Also, see Table 6-21
for similar information on benzo[e]pyrene, chrysene, and
dibenz[ahjanthracene.) The extent of and striking variation in PAH
contamination of marine species are evident in the data (Table 6-19) of
1 7 f*
Mix and Schaffer, z who examined BaP concentrations in mussels
(Mytilus edulis) in Yaquina Bay, Ore., at multiple sites over a 2-yr
period. The variations have a time component, geographic determinants,
seasonal and environmental elements, and unknown biologic influences
that make generalizations from such data extremely difficult and
perhaps impossible. The BaP concentrations in mussels reported by this
study exemplify, however, the extent to which marine species have che
potential for representing a considerable exposure source of PAHs in
the human diet.
A variety of foodstuffs of terrestrial origin have been analyzed
for PAH contamination, and many PAHs have been identified. They
include the polycyclic compounds listed in Table 6-20, some of which
have known carcinogenicity.l26 The known carcinogens 7,12-DMBA,
cholanthrene, and dibenzo[ai]pyrene have also been identified in curing
smoke. The relative concentrations of five carcinogenic PAHs in a sam-
pling of foodstuffs are shown in Table 6-21. * It is clear from
these data that amounts of some of these foodstuffs that are well
6-31
-------
within the amounts ingestible within a I-d period constitute a PAH
exposure via the gastrointestinal tract that can greatly exceed the
pulmonary exposure of a very heavy smoker to PAHs.
Large amounts of PAHs can be found in soils and can enter food
crops from this source. Table 6-22^ shows results of a. sampling of
soils in the Northeast for BaP. The concentration of PAHs in soil can
vary over an enormous range; for the prototype compound BaP, Baum^
has summarized World Health Organization data showing a range (in
micrograms per kilogram of soil) extending from around 100
(nonindustrial sites) through 1,000 (towns and vicinity) and 2,000
(soil near traffic) to 200,000 (soil near an oil refinery) or even over
600,000 (soil directly contaminated by coal-tar pitch). The higher
figures reflect particle deposition, local atmospheric fallout, and
direct waste discharge; the origin of the PAHs in forest samples (whose
soil concentrations range up to 1,300 ug/kg ) is not certain, but
must include a large contribution from natural combustion.
PAHs in food crops are probably derived in part from polluted
soils, although the relative contribution of this source, compared with
irrigation water or atmospheric pollution, is not established. PAHs in
soils can be translocated to plants, probably through root adsorption,
but the extent to which this occurs does not seem to correlate with the
PAH content of soi1.89'^9'^ Uptake of PAHs may also vary with
plant species. The aboveground parts of edible plants can, of course,
also concentrate PAHs through surface absorption from deposited dusts
containing these compounds. Through this process, the aboveground
parts of food crops can accumulate a gradient of PAH contamination
exceeding that in root parts by a factor as high as 10, * and the
bulk of this contamination in such edible crops as leafy vegetables
(e.g., lettuce, spinach, and kale) and tomatoes cannot readily be
removed by washing. ^^ PAH contamination of irrigation wastes also
contributes to an unknown extent to the contamination of edible
plants. In the processing of foods, packing materials and additives
are other sources of potential PAH contamination.
By far the largest sources of PAH contamination of foods are curing
and preserving processes and cooking, especially of meats. Apart from
shellfish, the "intrinsic" content of PAHs in most foods is low; for
example, uncooked pork and beef may contain only 0.1 pg/kg. This con-
centration can increase substantially as a result of any cooking pro-
cess (see Table 6-21) and especially as a result of smoking, curing, or
broiling under a direct flame in which food drippings can be pyrolyzed.
PAH contamination of foods associated with smoke-curing results in
part from the resinous condcnsates of liquid smoke flavors and from
food combustion products.64-106>l50-I77 • I-91 The type of smoke
generation and other characteristics of the smoking process can
influence the amounts and types of PAHs produced—e.g., the temperature
of combustion, the air supply, the length of smoke ducts, and the
density and temperature of the smoke-cure. ' Domestic smoking
clearly produces more PAH than the commercial process,^ probably
6-32
-------
because Che procedure is less controlled and, as a result, entails
heavier and more prolonged smoke exposure. A general survey by
the Food and Drug Administration and the U.S. Department of Agriculture
of PAH content of smoked foods prepared commercially was reported by
Malonoski ejt al. ^^
The broiling of meats over an open flame in which fat drippings can
be pyrolyzed probably contributes more to diet-derived PAH exposure
than any other method of food processing or preparation. Potent
mutagens can be produced from amino acids and proteins in foods by
high-temperature cooking.36>^,127,167,173,196 Tn£g mode Oj cooking
also increases the carcinogenic PAHs in meats to very high
concentrations. ^°
The concentrations of 15 PAHs found in the outer surfaces of
charcoal-broiled steaks by Lijinsky and Shubik^6 are recorded
in Table 6-23. These concentrations are not unusually high for
broiled or smoked meat (as the data in Table 6-21 indicate), nor
for dietary constituents that are known to have a high PAH content,
such as yeast. oilSi some leafy vegetables and fruits, roasted coffee,
and teas .24,64,65,201 pAHg forraed t»y pyrolysis can be derived (at
least with pure substrates) from carbohydrates, fatty acids, and amino
acids, and the extent of their production depends on temperature.
The data of Masuda e_t a_l.^ (Table 6-24) show the amounts of 19 PAHs
formed from combustion of six potential substrates at 500 or 700°C.
Combustion took place in a nitrogen atmosphere; at 300°C, no PAHs
were formed from any of the starting materials, but at the highest
temperature studied, large amounts were produced from each. Clearly,
substantial quantities of PAHs can be formed from these substrates
under the pyrolytic conditions used, and, although ordinary pyrolysis
takes place in air, the substrates tested are common constituents of
foods and common broiling temperatures are within the range of those
used in this study.
The conditions of broiling heavily influence the amounts of PAHs
produced. Fatty meat produces more PAH after broiling than lean meat,
and it has been suggested^" that pyrolysis of fats dripping onto
red-hot coals is the most likely source of PAHs. PAH production in
broiled meat clearly depends, in addition, on the closeness of the
meat to the heat source, on whether meat drippings reach the heat
source (i.e., heating from the top, rather than the bottom), and on
whether cooking is quick at high temperatures or slow at low
temperatures.104-107 Toxins other than PAHs are also produced by
high-temperature cooking; these include the mutagenic-carcinogenic
amino acid pyrolysis products described by Japanese and American
workers and the N-nitroso compounds formed in cured-meat products,
especially bacon and ham.64 It should be noted that these non-PAH
substances can be produced at temperatures distinctly lower than those
used in conventional broiling and that a large fraction of them may be
volatile; thus, redeposition of these airborne substances and their
inhalation during cooking are additional toxin exposures that can be
related to the diet.146
6-33
-------
An approximate "balance sheet" of Che estimated PAH exposure of
humans from air, water, and food is shown in Table 6-25. ^4 Despite
a degree of inexactness in these figures—especially for foods — they
provide a reasonable perspective of the sources of PAHs that might have
an impact on man. It should be evident from these estimates that food
constitutes the predominant source of PAHs for humans; even if che
contribution from smoking were included, the diet would still be the
dominant source.
The health impact of the PAHs in the human diet is not known,
although, as noted above, an association between the intake of these
substances in smoked foods and the occurrence of gastrointestinal
malignancies in select populations has been inferred. The remarkably
large amounts of PAHs that are ingested, compared with those to which
the pulmonary system is exposed (even in heavy smokers), makes it clear
that there must be tissue-specific factors related to the disposition
of or metabolic responses to PAHs that protect the gut from the
deleterious impact that might be anticipated from such exposure. The
possibility of detrimental effects of diet-derived PAHs on the gastro-
intestinal system will not be so amenable to quantitation as has bean
the case with respect to smoking and the development of lung pathology.
An approach to defining the human metabolic impact of diet con-
stituents in general and of charcoal-broiled meats in particular has
been taken in the clinical-nutrition studies recently summarized by
Anderson et_ al. Several dietary factors were shown to influence
potently the oxidative metabolism of various drugs used as model sub-
strates for cytochrome P-450- and cytochrorae P-448-mediated chemical
transformations. It has been shown that isocaloric substitution of
dietary protein for carbohydrate substantially shortens the plasma
half-times of such drugs as antipyrine and theophy11ine; i.e., a pro-
tein-enriched diet increases the oxidative metabolism of these com-
pounds. Opposite changes were observed during periods of high-
carbohydrate feeding. Substitution of protein for fat in the diet
(a nonisocaloric change) also stimulates the oxidative metabolism of
these drug substrates; however, neither high-unsaturated-fat nor high-
saturated-fat diets produce alterations in drug oxidation distinct from
those produced by high-carbohydrate diets alone. Thus, with respect to
influences on microsomal mixed-function oxidases, carbohydrate and fac
in the diet appear to be interchangeable.
Feeding rats charcoal-broiled beef is known to increase intestinal
metabolism of phenacetin. Increased oxidative metabolism of this
drug, as well as of antipyrine and theophylline, was also observed in
the test subjects after short-term feeding (2 portions/d for 4 d) of
normal portions of charcoal-broiled beef at mealtimes.™'°^> ^ The
effect of broiling (in control diets, the beef was protected from the
cooking fire with aluminum foil) was striking; during the test-diet
period, there was a pronounced decrease in the mean plasma concentra-
tion of phenacetin and a comparable decrease in the area under the
curve for plasma phenacetin concentration plotted against time. The
6-34
-------
ratios of the mean concentrations of metabolite and unchanged
phenacetin at each point studied increased markedly during the
charcoal-broiled-beef test period, compared with control periods. The
findings suggested that charcoal-broiled beef greatly stimulated the
metabolism of this model drug substrate in the gastrointestinal trace
or during its first pass through the liver. Smaller, but still
substantial, increases in antipyrine and theophylline metabolism during
the ingestion of the charcoal-broiled-beef test diet were also observed
These systematic and pronounced effects of specific dietary
manipulations on the metabolism of model drug substrates by the
cytochrome P-450-dependent mixed-function oxidase system provide a
valuable means for defining the metabolic responses of both normal and
ill subjects to the ingestion of various foodstuffs or foods prepared
in various ways. The physiologic import of such clinical studies can
be greatly extended by the judicious selection of suitable chemical
substrates for the metabolic systems under investigation. The extent
to which individuality in man characterizes specific chemical
biotrans format ions can also be explored by these metabolic techniques.
Finally, it may be possible through such clinical studies—in which
each subject serves as his own control—to.identify patterns of
biologic responses to specific foods or food components that might
otherwise be obscured by the genetic and environmental diversity of
large population groups.
6-35
-------
TABLE 6-1
PAHs in Human Liver0
Concentration (wet basis), ppt
PAH
Anthracene
Pyrene
Benz[a]anthracene
Benzo[e]pyrene
Benzo[b]fluoranthene
Benzo[k]fluoranthene
Benzo(a]pyrene
Benzo[ghi]perylene
Dibenz[ah]anthracene
1
(F.54)
200
450 .
ND
ND
88
15
13
59
ND
2
(F.17)
240
460
ND
ND
81
23
21
48
ND
3
(F.65)
170
340
ND
ND
87
10
19
36
ND
4
(M.65)
180
470
ND
ND
68
17
22
45
ND
5
(M.51)
140
310
ND
ND
53
8
10
21
ND
6
(M.41)
110
270
ND
ND
33
6
11
17
ND
aReprinted with permission from Obana e_t_ a_l_. " Numbers in
parentheses are sex and age of subject. ND » not detectable.
6-36
-------
TABLE 6-2
PAHa in Human Fata
Concentration (wet baaia), ppt
PAH
Anthracene
Pyrene
Benz [a] anthracene
Benzo [ e ] py rene
o> Benzo (b) f luoranthene
i
OJ
Benzo[k] f luoranthene
Benzol a) py rene
Benzo ( ghi ] pery lene
Dibenzf ah] anthracene
1
(F.54)
575
780
ND
71
260
28
31
110
ND
2
(F.27)
440
920
ND
110
190
40
25
62
ND
3
(F.66)
260
890
ND
64
240
35
24
61
ND
4
(M.65)
190
650
ND
57
77
17
18
54
ND
5
(M.51)
420
1,500
ND
140
250
43
59
69
ND
6
(M.41)
140
590
ND
83
120
27
18
42
ND
7
(F.35)
ND
49
ND
49
56
ND
ND
13
ND
8
(M.52)
25
2,000
ND
30
95
11
12
23
ND
9
(M.35)
140
1,300
ND
41
110
11
16
19
ND
10
(M.66)
390
2,700
ND
150
160
42
19
32
ND
aReprinted with permission from Obana et^ a_l.l-*" Numbers in parentheses are sex and age of sub-
ject. ND - not detected.
-------
TABLE 6-3
Rates of Formation of Benzo[a)pyrene Metabolites by Hamster, Rat, and
Human Microsomes3
Spec ies and
Organ
Rat 1 i ver
(n = 3)
Hamster liver
(n = 3)
Human liver
(n = 9)
Rat lung
(n = 3)
Hamster lung
(n - 3)
Human lung
(n = 19)
Human kidney
(n = 10)
An imal
Pret reat-
ment
None
PB
3-MC
None
3-MC
None
None
Pfl
3-MC
None
3-MC
None
None
Distribut ion
Total
bol ic
3,350
5,420
8,010
4,380
3,320
516
340
23
28
112
21
59
11
4
16
10
Meta-
Rate
.0
.0
.0
.0
.0
.0
.0
.2
.9
.8
.9
.0
.5
.8
.5
.2
of Metabol i tes
,c %
Dihydrod iol s
1
7
9
11
2
3
5
2
4
3
15
4
9
9
5
4
3
.8
.2
.6
.5
.0
.8
.4
.7
.8
.2
.1
.0
.3
.1
.7
.3
2
2. 1
9.4
5.5
15.5
20.5
7.1
1.5
7.8
6.2
8.0
6.9
5.9
6.4
3.6
5.5
3.2
3
3.9
3.9
8.0
2.7
2.7
5.5
1.9
6.5
6.6
9.7
8.7
9.3
12.5
7.0
5.0
2.5
Qu
26
27
20
29
33
54
6
30
29
14
43
36
33
13
57
11
inones
.8
.6
.6
.7
. I
.7
.0
.6
.4
.5
.8
.8
.9
.0
.3
.6
Phenols
1
7.8
2.8
13.1
5.0
4.2
6.9
3.1
15.9
16.6
14.7
9.6
5.1
11.0
7.2
8.0
,. 1-4
2
51.8
47.1
41.2
44.6
36.5
20.0
4.6
34.5
37.4
37.9
26.9
33.9
26.9
9.7
17.9
6.1
aReprinted with permission from Prough e_t^ al.
''Means, expressed as picomoles of product formed per minute per milligram of microsoraal protein.
cCrouped into three classes of dihydrodiols and two classes of phenols, because the radioactive
peaks cochroraatographed with the authentic standards of the dihydrodiols (dihydrodiol L, 9, 10-dihydro-
diol; dihydrodiol 2, 4,5-dihydrodio1-BaP; and dihydrodiol 3, 7,8-dihydrodio1) and phenols (phenol 1,
9-hvdroKv-BaP. and Dhenol 2. 3-hvdroxv-BaP).
-------
TABLE 6-4
Metabolism of Benz[a]anthracene, 7 ,12-DimethyIbenz[a]anthracene, and
Benzofa]pyrene by Human Mammary Tissue3
Hydrocarbon
Benz I a ) anthracene
7, 12-Dime thy 1-
benz [a] anthracene
Benzo(aJ pyrene
Patient
6
7
8
1
2
3
3
4
5
Protein, mg
19.7
22.5
22.2
64.3
33.3
20.5
16.0
18.8
78.4
Water-
soluble
Metab-
ol i tes ,
nmol/mg
prote in
24.0
3.3
22.2
4.6
9.9
20.3
12.0
8.4
1.6
Ether-
soluble
Metab-
olites ,
nmol/mg
protein
2.2
4.5
1.7
1.6
1.7
1.0
1.0
0.1
Protein-
bound
Metab-
olites ,
pmo 1 / mg
protein
72
129
30
260
47
371
28
47
90
Hydrocarbon-
DNA Adducts,
pmo 1 /nig
protein
<0.1
<0.l
<0.1
0.5
2.7
1.8
0.4
0.7
0.7
Total Metabolism
of Administered
Hydrocarbon, %
30.8
8.0
27.9
26.6
21.1
28.7
13.1
11.1
8.9
aReprinted with permission from Grover et aj_. " Metabolism and activation measured with epithelial-cell
aggregates in culture prepared from nonneoplastic mammary tissue.
-------
TABLE 6-5
Binding of PH]Benzo[a]pyrene Co DMA in Human Endomecrial Tissue Taken
Throughouc Menstrual Cycle and Before and Afcer Menopause3
[3H]B[a]P Binding to DNA,
dpm/ug DMA
Hormonal Status
Early and tnidprol iterative
Late proliferative and early secretory
Midsecretory and late secretory
Premenopausal
Pos tmenopausal
Mean *. S.E.
15.0 * 3.69
24.5 + 6.12
6.7 +_ 2.12
16.8 *• 2.70
4.7 * 1.67
No. Cases
11
16
10
37
3
aReprinted with permission from Donnan e_t a_1.
6-40
-------
TABLE 6-6
Effect of PAHs in Cigarette Smoke on Benzopyrene Hydroxylase
Activity in Rat Placenta4
PAH
Control
1,2-Benzanthracene
1,2,5,6-Dibenzanthracene
3,4-Benzopyrene
Chrysene
3,4-Benzofluorene
Anthracene
Pyrene
Fluoranthene
Perylene
Phenanthrene
8-Hydroxybenzopyrene formed,
ng/g-h
218 +_ 81
4,034 + 519
3,577 + 494
3,543 + 114
3,267 ^ 117
1,939 * 98
1,377 + 316
1,232 + 306
1,123 + 129
805 ^ 159
721 + 155
aReprinted with permission from Welch et^ al
190
'•'Rats pregnant for 18 d were given PAH orally at 40 mg/kg.
Placenta was assayed for benzopyrene hydroxylase activity 21 h
after the dose. Each value represents the mean _+ S.E. from
three rats.
6-41
-------
TABLE 6-7
Variability in Induction of Benzopyrene Hydroxylase
Activity in Human Placenta3
Hydroxybenzopyrene
Formed by Placenta,
Subject
L.B.
G.A.
P.C.
C.G.
A.T.
J.K.
L.C.
C.J.
E.R.
D.B.
D.A.
H.J.
M.N.
ng/g-h
240
260
547
643
1,269
1,317
1,860
4,289
4,390
5,267
15,181
16,524
17,100
aReprinted with permission from Conney et_ £l. All
subjects in this study were Caucasian, and all smoked 15-20
cigarettes daily during pregnancy. Variability in benzopyrene
hydroxylase activity was not related to medication taken during
or before delivery.
6-42
-------
TABLE 6-8
Specific Activities of Binding of Tritium-Labeled PAHs
to Human Bronchial DNAa
PAH
7, 12-Dimethylbenzanthracene
Benzo[ a]pyrene
3-Methylcholanthrene
Dibenz[ ah] anthracene
No.
Cases
3
4
2
3
Specific Activity
dpm/y g of DNA pmol/mg of DNA
170 jf
224 _*
38 ±
15 *
22
77
9
3
53 i
40 _*
34 jf
28 *_
7
14
8
6
aReprinted with permission from Harris ^£ a_l.'^ Nature, Vol. 252,
pp. 68-69, copyright 1974 Macmillan Journals Limited.
''Mean + S.E. Amount of DNA and dpm determined from peak DNA fraction of
CsCl gradients.
6-43
-------
TABLE 6-9
Disappearance of PAHs from Soot in Human Lungs3
Lung
No.
2
3
I 4
•C-
5
6
7
8
9
11
Pt. Age,
yr
90
71
70
74
71
81
62
70
65
Lung
Weight,
g (wet wt.)
950
1,335
1,570
1,570
710
1,260
1,120
910
1,420
Soot,
720
360
660
360
390
190
540
830
5,010
Ash,
mg
480
200
770
170
1,070
150
510
310
3,460
Pyrene
Found ,
0.9
1.9
1.9
3.6
1.3
2.3
4.4
5.1
1.3
Expected ,
Ug
6-27
3-13
6-24
3-13
3-14
2-7
5-20
7-31
45-185
Benzopyrene and
Benzopery lene _
Found, Expected,
34-172
17-86
31-158
17-86
18-93
9-45
25-130
Trace 38-200
233-1,200
aReprinted with permission from Falk et aj_.
-------
TABLE 6-10
PAH* in Human Lung Sample* from Surgical Operations*
&•
Ul
Sample
Ho.
1
2
3
4
5
6
7
a
9
10
11
12
13
14
IS
16
17
IB
19
20
21
22
23
24
25
26
27
28
29
30
Tiaaue
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Bronchial
Adjoining
Adjoining
Adjoining
Adjoining
Adjoining
Adjoining
Adjoining
Adjoining
Adjoining
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoam
carcinoau
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
carcinoma
tiaaue
tiaaue
tiaaue
tiaaue
ciaiue
tiaaue
tiaaue
tiaaue
tiaaue
Weight of
Sample, g
20.0
25.0
35.0
20.0
23.0
a.o
18.9
26.5
27.0
43.4
66.4
39.8
31.2
39.7
67.8
41.0
148.1
112.4
116.7
114.0
87.4
53.0
97.0
26.0
62.0
24.0
42.0
32.0
51.0
54.0
Benzojalpyrene
Total.
300,000
3,300
34,000
1,640
5.760
690
80
220
80
93
570
97
91
<2
<2
290
640
28
190
100
96
<2
<2
<2
<2
<2
<2
<2
<2
<2
ng
.0
.0
.0
.3
.0
.7
.7
.2
.2
.0
.0
.7
.0
.0
.6
.2
.2
.2
.2
.2
.2
.2
.2
.2
ng/g
15,000
130
900
80
250
80
4.0
8.3
3.0
2.2
8.7
2.4
2.9
—
—
7.1
4.3
0.3
1.6
0.9
1.1
—
—
—
—
—
--
—
—
Fluoranthene Benzo(b) f luoranthene Pervlene
Total, ng ng/g Total
<8.8
<8.8
<8.8
<8.8
400
<8.8
63.6
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
650
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8.8
<8
<8
<8
<8
17 <8
<8
3.4 <8
<8
-------
TABLE 6-11
Environmental and Occupational Benzo[a]pyrene and Total
Suspended Particulate Concentrations9
Location
28 rural sites, U.S.
94 urban sites, U.S.,
1958-1959
All rural sites, U.S.
291 urban sites, U.S.,
1960-1965
Merton bus garage,
June 1956
Blackwall Tunnel, London,
summer, 1958
Sumner Tunnel, Boston,
summer, 1961
Diesel locomotive cab
Diesel roundhouse
Coal and pitch coking plants
Gasworks retort houses
Roof tarring, outdoor
Benzofa)pyrene,
ug/1.000 m3
0.01-1.9, annual avgs.
0.11-61.0, annual avga.
Total Suspended Particles,b
ug/m3
26.5 (inside), 7-h mean
14.5 (outside), 7-h mean
350, daytime mean
690, 24-h mean
300-35,000, short-term sample
3,000, long sample means
90-14,000, short-term samples
28, arith. avg., biweekly samples
iith. avg., biweekly samples
V 1,254 maximum)
210-1,440 (inside)
140-240 (outside)
930-2,350
588 (outlet)
104 (inlet)
380 (100-800)
1,990 (70-8,700)
aReprinted with permission from Schenker. * •>->
-------
TABLE 6-12
Epidemiologic Studies of Cancer in Occupations Exposed to Diesel Exhaust3
Population Studied Findings Comment
cr>
l
Male London transportation
workers, 45-65 yr old,
1950-1955
Baltimore and Ohio Railroad
workers, 1953-1958
"Two large railroad companies,"
1939-1950
Finnish railroad workers, 30-52
yr old, 1955-1973
Massachusetts tumor registry,
1965-1972
Central state teamsters,
May-July 1976
96 cases lung cancer; no excess
lung cancer attributed to diesel-
exhaust exposure
154 lung-cancer deaths; SMR
slightly lower than national
rates; no differences in rates
for exposed and nonexposed workers
133 cases lung cancer; about 3:1
ratio observed:expected in
operating workers, compared with
nonoperating workers
47 tumors in engineers—signifi-
cantly greater number than in
trainmen or railroad clerks
91 cases oat-cell cancer; excess
in transportation-equipment
operatives
34 respiratory-tract cancer deaths;
increased SMR all age groups;
significant for age 50-59
Inadequate duration of expo-
sure and latent period at
time of study; diesel-exhaust
exposure estimated
Inadequate duration of diesel-
exhaust exposure and latent
period at time of study; no con-
sideration of transfers, retire-
ment, or duration of exposure
Inadequate duration of diesel-
exhaust exposure at time of
study; rates not calculated
Small number of cases; not
analyzed by tumor type
Not specific for diesel-exhaust
exposure
Small number of cases; short
period of observation
aReprinted with permission from Schenker et^ al.
-------
TABLE 6-13
Benzo[aJpyrene Content of Shale Materials,3 Mg/kg (ppb)
Retort Oil
i Raw Shale
oo Shale
RS-101
RS-102
RS-103
1st
2.5
2.0
1.7
2nd
1.9
1.6
1.6
Spent Shale
Shale
SS-201
SS-202
SS-203
1st
ND
<0.4
19.8
23.8
2nd
NO
<0.2
19.6
23.3
Oil
RO-1
RO-2
RO-3
Parent
1st
1,900
1,800
2,300
Methyl -Substituted
2nd
1,700
1,800
2,300
1st
5,000
21,000
11,000
2nd
3,300
22,000
12,000
aReprinted with permission from Weaver and Gibson. "'
-------
TABLE 6-14
PAHs in Shale Retort Oils,3 ppb
Component
Pyrene
Fluoranthene
Benz[ a] anthracene
Chrysene
Triphenylene
Benzo[a]pyrene
Benzo [e] pyrene
Perylene
Anthanthrene
Benzo[ghi] perylene
Coronene
Parent
17,500
5,650
1,200
2,850
540
4,250
1,950
325
275
1,900
_ _
Methyl-Substituted
50,500
8,050
12,000
23,500
5,700
8,350
2,650
1,015
455
8,650
__
aReprinted with permission from Weaver and Gibson. '•"'
6-49
-------
TABLE 6-15
PAHs in Drinking Water3
Concentration, ng/L
Source
Mixed tap water at
Mainz, Germany
Water at:b
Syracuse, N.Y.
Buffalo, N.Y.
New York, N.Y.
Lake George, N.Y.
Endicott, N.Y'.
Hammonds port , N.Y
Pittsburgh, Pa.
Philadelphia, Pa.
Huntington, W.Va.
Wheeling, W.Va.
New Orleans, La.
Appleton, Wis.
Champaign, III.
Fairborn, Ohio
Elkhart, Ind.
BaP
—
0.3
0.2
0.5
0.3
0.2
0.3
0.4
0.3
0.5
2.1
1.6
0.4
NDC
0.1
NDC
Carcinogenic
PAH
—
0.3
0.2
3.9
1.5
1.1
1.5
1.9
2.0
2.0
11.3
1.6
2.4
1.2
0.8
0.3
Total
PAH
7.0
1.1
0.9
6.4
4.2
8.3
3.5
2.8
14.9
7.1
138.5
2.2
6.1
2.8
2.9
0.3
aReprinted with permission from Santodonato et al
bOnly the six W.H.O.-recommended PAHs were analyzed, with the exception
that BjF replaced BbF.
CND =• not detected.
6-50
-------
TABLE 6-16
PAHs in Tapwater3
Concentration,
PAH ppc
Naphthalene 2.9
2-Methylnaphthalene 1.4
1-Methylnaphthalene 1.1
•Biphenyl 0.32
Acenaphthene 0.82
Dibenzofuran 0.62"
Fluorene 0.72
Dibenzothiophene 0.21
Phenanthrene 3.1
Anthracene 0.35
2-Methylanthracene 0.06
4,5-Methylenephenanthrene 0.30
1-Methylphenanthrene 0.37
Fluoranthene 2.6
Pyrene 1.1
Benzo[a]fluorene 0.05
Benzo[b]fluorene 0.05
4-Methylpyrene 0.05
Methylpyrene 0.08
1-Methylpyrene 0.05
Benz[ajanthracene 0.49
Benzo[b]fluoranthene 0.21
Benzo[jk]fluoranthenes 0.07
Benzo[ejpyrene 0.20
Benzo[a]pyrene 0.05
aData from Olufaen.
6-51
-------
TABLE 6-17
PAHs in Extracts from Shucked Oysters3
Approximate
Concentration,
Compound yg/5 kg of oysters
Benzo[ghi]perylene 5-25
Benzo[a]pyrene 10-30
Benz(a]anthracene 50
Benzo[k]fluoranthene 40-60
Benzo[e]pyrene 100
Chrysene 100-200
Pyrene 500-800
Fluoranthen* 3,000-5,000
aReprinted with permission from Cahnmann and
Kuratsune;3^ copyright 1957 American Chemical
Society.
6-52
-------
TABLE 6-13
PAHs in Foodstuffs3
Concentration, k ug/kg (wet wt.)
Foodstuffs
Cooked meats, sausage
Cooked bacon
Charcoal-broiled meats
Smoked ham, sausage
Heavily smoked ham
Cooked fish
Smoked fish
Cereal grains
Flour and bread
Bakers' dry yeast (yeasts
grown on mineral oils
are lower)
Soybean
Refined vegetable oils,
fats
Margarine, mayonnaise
Salad
Tomatoes
Spinach
Kale (only 10% removed
by washing)
Apples
Fruits (not apples)
Dried prunes
Roasted coffee and
solubles
Malt coffee
Tea
Whiskey
Beer
Roasted peanuts
Milk
Benzojajpyrene
0.17-0.63
1.6-4.2
2.6-11.2
(50.4 recorded)
0.02-14.6
Up to 23
Benzanthracene
0.2-1.1
1.4-31
(107 recorded, Iceland)
0.4-9.6
Up to 12
0.9
0.3-60
(up to 37 in Japan)
0.2-4.1
0.1-4.1
1.8-40.4
3.1
0.4-36
0.2-6.8
2.8-12.8
0.2
7.4
12.6-48.1
0.1-0.5
2-8
0.2-1.5
0.1-4
Up to 15
3.7-21.3
0.04 ug/L
ND
ND
Up to 2.9
0.02-2.8
(up to 189
0.4-6.8
0.4-6.8
2.9-93.5
0.8-1.1
1.4-29.5
4.6-15.4
0.3
16.1
43.6-230
0.5-14.2
Up to 43
0.04-0.08 ug/L
ND
Up to 0.95
in Japan)
aReprinted with permission from U.N.
bND = not detected.
Food and Agriculture Organization.
180
6-53
-------
CT>
t-
TABLE 6-19
BenzoJa]pyrene in M. edulis from Yaqnina Bay, Oregon3
Date
6/15/76
7/22/76
9/24/76
11/16/76
12/16/76
2/03/77
4/08/77
6/29/77
8/29/77
10/13/77
12/08/77
2/03/78
4/28/78
6/24/78
Average
Benzo[
YIM
0.1
4.7
0.7
0.6
8.4C
3.8
1.7
6.3
1.2
0.8
1.2
3.1
0.7
0.8
2.0
ajpyrene
Y2M
30
67
34
40
12
33
22
15
5.1
5.0
15
27
20
29
26
Concentration, |_
Y3M
8.1
4.5
16
8.4
7.2
71C
7.9
3.5
2.8
5.3
NS
NS
NS
NS
6.5
Y4M
15
6.7
6.9
8.9
7.5
170C
12
5.4
2.2
1.9
6.4
13
5.5
17
8.5
ig/kg at Site.
Y5M
0.9
4.4
1.2
2.7
0.6
0.9
1.5
2.1
5.6
1.2
4.7
7.7
1.2
NS
2.7
Y6M
3.0
2.3
1.9
17
6.1
8.1
4.4
50C
4.4
3.2
3.1
32C
4.0
NS
7.5
Y7M
4.1
2.4
14
19
3.8
1.7
NS
3.5
5.8
4.2
36
NS
27
NS
a
Y8M
0.4
0.8
0.9
NS
NS
2.0
NS
1.5
NS
0.5
8.1C
2.3
2.7
NS
1.4
Y10M
5.2
10
6.3
3.6
2.8
3.0
3.8
NS
NS
4.2
9.4
10
NS
NS
6.0
YUM
0.5
0.6
NS
0.4
NS
0.7
0.7
2.0
0.4
0.1
NS
3.0
NS
NS
1.0
Y12M
0.4
0.7
0.8
0.4
0.3
0.5
0.5
0.4
0.0
0.2
0.0
1.2
0.1
NS
0.4
Y13M
0.4
0.3
0.8
0.9
0.1
0.2
NS
0.0
2.6
0.4
NS
NS
NS
NS
0.7
Y14M
4.3C
0.5
0.3
0.6
0.4
0.2
0.4
0.4
0.3
0.34
NS
NS
NS
NS
0.4
aReprinted with permission from Mix and Schaffer.126
^NS = not sampled or not yet analyzed.
cData not included in statistical analyses because of large variation (>4x) from the mean.
-------
TABLE 6-20
PAHs in Foods3
1 Anthracene
2 Benzanthracene*
3 Methylbenzanthracene
4 Dibenz[aj]anthracene*
5 Dibenz[ahjanthracene*
6 Dibenzfac]anthracene*
7 Dibenz[ai]anthracene*
8 Phenanthrene
9 3-Methylphenanthrene
10 2-Methylphenanthrene
11 9-Methylphenanthrene
12 2,6-Diraethylphenanthrene
13 Fluorene
14 Benzofa]fluorene
15 Benzofb]fluorene
16 Benzo[a]fluoranthene
17 Benzo[bjfluoranthene
18 Benzo[j]fluoranthene*
19 Benzofk]fluoranthene
20 Benzofghi]fluoranthene
21 Pyrene
22 4-Methylpyrene
23 £-Phenylenepyrene*
24 Benzo[a]pyrene*
25 Benzofe]pyrene*
26 Dibenzo[ah]pyrene*
27 Anthanthrene
28 Chrysene*
29 Perylene
30 Benzofghi]perylene
31 Acenaphthene
32 Acenaphthylene
33 2-Methylnaphthalene
34 Naphthalene
35 Acenaphthalene
aReprinted with permission from Mix and Schaffer.^-2^ Asterisk
indicates known carcinogenicity.
6-55
-------
TABLE 6-21
PAHs in Foodstuffs3
Concentration,
Foodstuff Compound ug/kg
aReprinted with permission from Zedeck.
201
Broiled sausage Benz[a]anthracene 0.2-1.1
Smoked sausage 0.4-9.9
Heavily smoked ham 12
Spinach 16
Crude coconut oil 98
Refined vegetable oil 1
Broiled sausage Benzo[a]pyrene 0.17-0.63
Charcoal-broiled meat 2.6-11.2
Smoked fish 2.1
Spinach 7.4
Tomatoes 0.2
Crude coconut oil 43.7
Roasted coffee 0.1-4
Tea 3.9-21.3
Cereals 0.2-4.1
Smoked ham Benzo[e]pyrene 5.2
Smoked fish 1.9
Spinach 6.9
Tomatoes 0.2
Crude coconut oil 32.7
Roasted coffee 0.3-7.2
Roasted peanuts 0.4
Broiled sausage Chrysene 0.5-2.6
Heavily smoked ham 21.2
Spinach 28
Tomatoes 0.5
Cereals 0.8-14.5
Roasted coffee 0.6-19.1
Black tea 4.6-6.3
Spinach Dibenz[ah]anthracene 0.3
Tomatoes 0.04
Cereals 0.1-0.6
6-56
-------
TABLE 6-22
Benzo[a]pyrene In Soils3
Benzo[a]pyrene
Origin and Type of Soil Concentration, ug/kg
Oak forest, West Falmouth, Mass. 40
Pine forest, West Falmouth, Mass. 40
Mixed forest, West Falmouth, Mass. 1,300
Mixed forest, eastern Conn. 240
Garden soil, West Falmouth, Mass. 90
Plowed field, eastern Conn. 900
aReprlnted with permission from M. Blumer, Science 134:
474-475, 1961;2^ copyright 1961 by the American Association
for the Advancement of Science.
6-57
-------
TABLE 6-23
PAHs in Charcoal-Broiled Steaks3
Concentration,
PAH US/kg of steak
Alkylbenzanthracene 2.4
Anthanthrene 2
Anthracene 4.5
Benz[a]anthracene 4.5
Benzo[b]chrysene 0.5
Benzo[ghi]perylene 4.5
Benzo[a]pyrene 8
Benzo[i]pyrene 6
Chrysene 1.4
Coronene 2.3
Dibenz[ah]anthracene 0.2
Fluoranthene 20
Perylene 2
Phenanthrene n
Pyrene 13
aReprinted with permission from Lijinsky and Shubik.
106
6-58
-------
cr>
m
TABLE 6-24
PAHa Produced by Pyrolyaia of Carbohydrates, Ami no Acida, and Fatty Acida
at Two Temperatures3
PAH Concentration,
PAH
Naphthalene
Acenaphchylene
Fluorene
Anthracene
Phenanthrene
Pyrene
Fluoranthene
Benco(a) f luorene
Benz 1 a ) anthracene
Chryaene
Perylene
Benzofalpyrene
Benzol elpyrene
Benzo(b) f luoranthene
Benzo[ j)f luoranthene
Benzol k) [luoranthene
Benzol ghilperylene
Anthanthrene
Coronene
Starch
700 C
5.140
100
1,550
814
1.560
965
790
306
315
159
24
179
82
62
32
48
134
49
9
500 C
—
—
32
48
41
13
15
22
6
2
7
3
—
--
—
3
2
. |i8/50 R
D-Clucoae
700 C
16.000
4.410
1.260
1.240
2.440
1.680
1.200
165
520
210
45
345
175
150
160
120
180
150
10
500 C
--
—
49
66
23
19
23
25
8
1
6
1
—
—
--
2
—
l.-Leuc ine
700 C 500 C
6.500
732
1.450
632
2,200
1.200
320
155
270
48
13
77
20
17
53
7
40
20
3
L-Clutdfflic acid
700 C 500 C
1,800
395
218
357
582
755 1.5
870 0.5
16
119
32
8
58
43
13
20
12
35
18
—
Palmitic acid
700 C 500 C
223,000
49,100
8,150
13,500
38,100
1 7 , 600 0.5
6.700 0.1
2,470
5.710
2.710
460
3.750
2.390 —
210
1.550 —
850
95
Stearic acid
700 C 500 C
19.000
40.400
11,600
8,880
50.100
18.700 0.7
6,590 0.5
3,410
8,410
4,550
675
4.440
2.630
352
2.740
935
52
•Reprinted with periaiasion from Masuda e_t at.
-------
TABLE 6-25
Estimated Daily Human Exposure to PAH from Air, Water,
and Food3
Source Benzo[a]pyrene, ug Total PAH, u g
Air 0.0095-0.0435 0.207
Water 0.0011 0.027
Food 0.16-1.6 1.6-16
aReprinted with permission from Santodonato e_t al.
154
6-60
-------
60
40
20
(0
—TTTTTT*
10 15 20 25 X) 31 40
SPECIMEN
FIGURE 6-1. Spectrum of specific activities of binding of [%] to
DNA in human endometrial tissue in vitro. Human endoraetrial tissue was
incubated for 18 hr in organ culture in medium containing 1 M[^H)BP.
For each of the 41 specimens of endometrial tissue examined, specific
activities of binding were determined in order to most clearly
illustrate the range of binding of [^Hj to DNA in endometrial tissue
from these patients. This histogram has been constructed with cases
enumerated in increasing order of specific activity. Reprinted with
permission from Dorraan e_£ al. 52
6-61
-------
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SOME FACTORS THAT AFFECT SUSCEPTIBILITY OF HUMANS TO
POLYCYCLIC AROMATIC HYDROCARBONS
The interaction of chemical pollutants, including the PAHs, and
mammalian cells can result in a variety of problems, including toxicity.
mutagenesis, carcinogenesis, and teratogenesis. This interaction of
chemicals with somatic cells probably results in such end points as
cancer, and the interaction of chemicals with germ cells probably results
in a variety of hereditary disorders. Many genetic disorders result in a
predilection to the development of cancer. The cancer burden in the male
population in the United States, although speculative, is distributed
approximately as follows: 402 from tobacco-smoking, 10-20% from all
diet-related causes, 5% from occupational exposures, 5% from single-gene
inheritance, and 35Z from other causes, which may include unknown genetic
predisposition and environmental effects.59 The birth-defects burden in
the United States is distributed approximately as follows: 5-10% from
known teratogens, such as viruses, chemicals, and radiation; 25% from
genetic anomalies; and 60-65% from unknown mixtures of genetic predisposi-
tion and environmental effects.^ Although monogenic disorders (includ-
ing dominants), X-linked recessive disorders, and chromosomal abnor-
malities account for only about 5% of the human disease burden, the impact
of heterozygous recessively inherited abnormalities similar to the mono-
genic disorders is very ill-identified, but could outweigh all other
contributions. *•"
The heterozygous recessively inherited disorders may be the major
reason why cancer incidences are not uniformly distributed.'^1*' In
fact, of the millions of people exposed to such environmental chemicals as
diethylstilbestrol, estrogen oral contraceptives, vinyl chloride, and
cigarette smoke, only a very small proportion develop or express the
cancer thought to be associated with these exposures. It is likely that
genetic variability within the human population accounts in part for the
distribution pattern. As depicted in Figure 7-1, cancer sensitivity can
be viewed as a function of inborn susceptibility. Where this inborn or
genetic susceptibility is low, cancer expression is low. Where this
susceptibility is high (e.g., in single-gene defects), cancer expression
is high. The major question is whether the combination of chemical
exposure and genetic susceptibility can change significantly the numbers
of persons who develop cancer.
PAHs are ubiquitous chemicals capable of producing a broad spectrum
of biologic responses. Some can cause cancer in a variety of tissues,
including lung, liver, kidney, colon, skin, and bladder. In humans,
epidemiologic evidence has demonstrated that the incidences of cancers of
stomach, nasal cavity and sinuses, lung, and to a lesser extent rectum,
testis, skin (e.g., melanoma), brain, liver, pancreas, and hemopoietic
7-1
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tissue (i.e., leukemia) are correlated with areas containing high con-
centrations of industrial pollutants.10-46>102•^>183 For many
persons, the amount of these agents in the environment may be the rate-
determining factor for cancer susceptibility. Thus, the primary need
would be to identify and measure the amount of exposure to the environ-
mental pollutants. The advent of a variety of in vitro and in vivo
bioassays promises the development of methods for identifying chemicals
that are potential carcinogens.
In animal-model systems, susceptibility to chemically induced cancers
is usually dose-related. However, route, duration, and frequency of
administration and such genetic factors as species, sex, and strain sll
tend to modify the relationship. In humans, mixtures containing PAHs can
certainly cause cancer, but inadequacies in the information on age and
trauma but especially on duration, frequency, and intensity of exposure
and on the size and characteristics of the exposed population make
quantitative estimation of dose-response- relationships and the concept of
thresholds difficult to interpret.
EFFECT OF GENETIC DIFFERENCES
The hypothetical stages in carcinogenesis are depicted in Figure
7-2. PAHs probably can show biologic effects at any of these stages.
Thus, answers are needed to the following questions: Which stages can
PAHs modify in humans? Are there naturally occurring variations ia the
expression of some of these steps in humans? Can a genetic basis be
identified for the regulation of these naturally occurring differences?
If so, can the differences result from the action of a single gene
system? Can a relationship be shown between the expression in the gene
locus and PAH-mediated effects?
PAH-induced effects in humans could depend on exposure, uptake, and
distribution of the chemicals; their metabolic activation and inactiva-
tion; DMA-repair capacity; "promoters"; and the extent of immunocompe-
tence. Each of these is discussed below.
UPTAKE AND DISTRIBUTION OF PAHs IN TISSUES
The distribution of PAHs in tissues or cells depends on the route of
exposure. According to the results of Rees e_t a_l_. , the distribution
of benzo[aIpyrene (BaP) in tissue other than at the site of absorption
(i.e., intestine) depends on two phases: accumulation of the BaP on the
tissue and passive diffusion through the tissue. These two phases
underlie these authors' views about the apparent exponential nature of the
accumulation of BaP as a function of dose. The exponential increase could
be very important, but it must be pointed out that humans are rarely
exposed to BaP at concentrations greater than 200 yM (i.e., 50 ug/ml)
under "normal" circumstances. Concentrations of a variety of PAHs (e.g«>
pyrene, anthracene, and BaP) in human tissues average about 1,100 parts
per trillion (ppt) in fat tissue and 380 ppt in liver. BaP can vary
7-2
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from 0.3 to 15,000 parts per billion (ppb) in bronchial-carcinoma
tissue. Most of the subjects in the study of Tomingas et^ a_1.
were cigarette-smokers, but no obvious correlation between BaP
concentration and extent of smoking was seen. The PAHs observed in
addition to BaP included fluoranthene, benzo[b]fluoranthene, and perylene.
Reasons for differences in tissue distribution are not known, but,
inasmuch as most of these chemicals are inducers and substrates for
microsomal enzymes, tissue variation in cytosol and nuclear receptors
could be important. In rodents, the induction of the raicrosomal raono-
oxygenase system by some PAHs depends on the presence of particular
cytosol receptor proteins.56'142'143 These receptor proteins are not
evenly distributed in all tissues, but are highest in thymus and lung,
lower in liver and kidney, lower yet in testes, brain, and skeletal
muscle, and not detectable in pancreas, adrenal, or prostate. Most
importantly, receptor proteins are found in high concentrations in strains
of animals or cultured cells in which PAHs induce the enzyme aryl
hydrocarbon hydroxylase (AHH) and are nondetectable in those in which AHH
is nonresponsive. 6>1 2 This correlation also extends to humans, in
whom the concentration of a BaP-binding plasma component is correlated
with the capacity of lymphocytes to be induced for AHH activity in
culture.10 A cytoplasmic receptor for BaP, which did not cross-react
with 7,12-diraethylbenz[a]anthracene, has also been reported for human
cells in culture. 31 The presence of some of these receptors is under
specific genetic control in animal models, 2>, so uptake and
distribution, at least in particular persons, could be under a form of
genetic control.
METABOLISM
PAHs are metabolized in a variety of ways, with the microsomal mono-
oxygenases (e.g., AHH) probably most important. Steady-state activities
of these enzymes vary in animals and are linked to susceptibility to some
PAH-mediated cancers. ' In humans, the data are much less clear.
Table 7-1 summarizes the studies that suggest a correlation between high
AHH inducibility (and usually high induced-AHH activity) and cancer
susceptibility, and Table 7-2 summarizes the studies that suggest the
converse.
Reasons for the contradictory results probably lie in methodologic
variations, such as the use of different cell types, different assays,
and different assay conditions. The most easily accessible and therefore
commonly used human tissue isfl£he peripheral blood lymphocytes. Nutri-
tional state,167 drug intake,89 age,35 and disease state74
influence the capacity of the lymphocytes to respond to mitogen. These
influences have not been assessed in determining their relationship to the
AHH activity observed in cultured lymphocytes. Variations in AHH activ-
ity in lymphocytes have been observed to occur seasonally in some geo-
graphic locations,128'129'154 but whether they result from in vivo or in
vitro factors is not known.
7-3
-------
A variety of in vitro conditions are known to influence AHH
activity. The initial concentration of lymphocytes affects the time
course and amount of control and induced AHH activity. The type and
lot of serum supplement infuence the control and induced AHH
activity.50'79 In fact, some lots of fetal-calf serum are capable of
causing mitogen activation of lymphocytes.' The numbers of cultured T
cells may affect the AHH activity observed. In studies using cultured
human tissue, two important aspects are the question of the variable
degree of AHH activity in different cell types and the question of
large variations between and within individuals in both AHH activity and
raicrosome-raediated BaP-DNA adduct formation. ' Blood monocytes and
pulmonary alveolar macrophages are examples of other human cell types
whose AHH activity is correlated with that in lymphocytes^^»^ or
cultured human tissue, but there are problems of accessibility with
each of these cell types.
If the cell samples are cultured and assayed on the same days,
the variation seems to be acceptable. ' Culturing lyrapho-
S 1 7 1 in 171
cytesj> Li > L-)U or monocytesL'iJ from fraternal or identical twins at the
same time has shown that AHH activity is under a degree of genetic
control, and the numbers of genes in question are probably small. Thus,
the genetic component most likely results in a uniraodal frequency
distribution -that is skewed in the populations of individuals toward thoa
7 A 1 1 1
with higher AHH activity, ' rather than the trimodal distribution
originally reported.
To circumvent many of the in vitro problems, the use of cryopre-
served tissue may be an alternative, in that lymphocytes can be cryo-
preserved before mitogen activation and still have the capacity to be
mitogen-activated and then assayed for AHH activity. The relative AHH
activities among the lymphocyte samples from different individuals are
similar, whether the assays are conducted on freshly cultured lymphocytes
or after cryopreservation. ^ Cryopreservation allows the culture and
assay at the same time of cells from different organisms collected in
diverse geographic locations and over extended periods.
The use of cryopreserved lymphocytes, control of some basic culture
variables—such as initial lymphocyte concentration (1.0 x 10 cells/ml)
and lot and type of serum supplement (e.g., human AB serum)—and assaying
AHH activity at two times to ensure detection of peak activity can yield
the data presented in Figure 7-3. Data were taken on a group of 51 per-
sons who were on hospital diets for at least 2 d before phlebotomy, who
were not on any medication, and who were eventually followed for complete
clinical diagnosis. Viability of cells was measured by assay for the
NADH-dependent cytochrome bj reductase (using cytochrome £ as a sub-
strate) activity (Cyt £). Carcinogen-metabolizing activity is presented
in terms of units of AHH per unit of Cyt c_. 9>1 5 The degree of mitogen
activation was also measured. Data analyses showed that:
• Cryopreserved lymphocytes from over 952 of the normal and cancer
patients were mitogen-activated.
7-4
-------
• Lymphocytes from lung-cancer patients were mitogen-activated as
efficiently as lymphocytes from noncancer patients (actually better;
£ - 0.001).
• The 14 highest AHH activities were found in patients with lung
cancer, with the mean in the 21 lung-cancer patients (0.89 unit AHH/unit
Cyt £) being significantly higher than that in the 30 non-lung-cancer
patients (0.47 unit AHH/unit Cyt c).
The higher AHH activities were not directly related to higher degrees
of blastogenesis and were not related to cigarette-smoking history, tumor
type, tumor location, or family history of cancer. Whether high AHH
activity is the cause or the result of lung cancer cannot yet be answered.
In animal-model systems, some PAHs cause tumors of the lymphoreticu-
lar system, and a genetic association for this activity at the Ah locus
has been suggested.29'113 Although this is only presumptive, there may
be a similar relationship in human leukemia patients who were recently
shown to express lower AHH activity (as in animal-model systems);^ in
other studies, the first-degree relatives of leukemia patients expressed
normal AHH activity.9^
The results of these studies are interesting and certainly need to be
confirmed and extended. The extended studies should be raultifaceted; that
is, they should simultaneously measure more than one enzymatic end point.
Perhaps an appropriate group of assays would include an assay for AHH, as
described in the literature; an assay for all BaP metabolites via HPLC; an
assay for particulate P-450s via imraunoassays; ' and an assay for
mRNA expression of the P-450 genes with cloned ONA fragments containing
the P-450 genes. Human tissues should be used where possible. There
is probably a degree of genetic control of AHH activity in the human
population, and this enzyme may play a role in determining susceptibility
to PAH-mediated cancer and other diseases.
DMA BINDING. DAMAGE. AND REPAIR
Many PAHs are converted by the microsomal monooxygenases to forms
that bind covalently to a variety of cellular macromolecules, including
nucleic acids (see Chapter 5 and Phillips and Sims141-'). Evidence of the
importance of DNA binding is exemplified by the observation that varia-
tions in DNA-repair capacity seem to play a major role in determining the
toxic, mutagenic, and transforming activities of many chemical carcino-
gens, including PAHs.3'100'149
In animal-model systems, the amount of PAH metabolism is determined
by the activity of the raicrosomal monooxygenases, and variations in these
enzymes result in concomitant changes in the binding of chemicals to
DNA. In cultured human tissue, hydrocarbon-DNA binding also occurs
as the result of microaomal monooxygenase-mediated metabolism,6'5 and
variations in metabolic activity are associated with concomitant varia-
tions in binding of hydrocarbons to DNA. The major DNA adduct often
results from the interaction of specific metabolites of PAHs (diol-
epoxides) and the N7 of deoxyguanosine. ' Other products are
7-5
-------
found, including interactions with the N4 deoxyadenosine, ' ' the back-
bone phosphates of DNA, and the exocyclic amino group of deoxy-
adenosine. ' The latter may be important, because its formation
from various PAH-like chemicals closely parallels their carcinogenic
potencies on mouse skin. No appreciablespecificity of binding with
respect to base sequence is apparent, ' ' but binding may be
influenced by chromatin structure, with a greater extent of binding
associated with internucleosomal regions. ^'
A potentially important anomaly is that, although in vitro metabolisn
of BaP to forms that bind to DNA parallels the AHH activity of the nucro-
soraal preparations and the genetic background of mice used to generate
these microsomal samples, the in vivo results from strains of mice
that differ widely in AHH activity suggest that there is very little
strain variation in BaP-DNA binding.3o>139
Probably more crucial to carcinogenicity is the geometry of the
binding in relation to later excision repair by endonucleases. The
binding of different residues and different chemical groups within
residues dramatically affects excisibility. These cheraical-DNA adducts
are either repaired, not repaired, or misrepaired (see Figures 7-2 and
7-4). The fate of these adducts determines whether a cell remains normal,
mutates, or di.es.
Repair capacity can be separated into two major types—excision
repair and postreplication repair. Excision repair is the in situ
removal and replacement of chemically modified DNA so that the original
DNA sequence is re-established. For a variety of reasons, excision-
repair systems usually do not remove all the modified bases; so the DNA
very often replicates, even though some unexcised damage may be present.
This replicated DNA usually has gaps in the newly synthesized strand
opposite the DNA adduct. The gaps are filled in by postreplication
repair—also termed "recombination repair.' Figure 7-3 depicts how
these two processes of repair contribute to the cells' survival of the
damaging effects of chemicals like PAHs. A combination of both methods is
involved in the repair of hydrocarbon-bound DNA. ^0
A large number of both constitutive and inducible enzymes are
involved in this DMA-repair process. The exact role of these enzymes
is not known, but it seems that rather small changes in any of the enzyme
activities can have great effects on the repair process and eventual bio-
logic expression of the DNA adducts. Moreover, it has been recently shown
in prokaryotes that the DNA adduct itself is not likely to be mutagenic,
but rather that the mutagenic event is induced by the action of the
DNA-repair enzymes themselves.^"
Natural variations in DNA-repair capacity occur in humans. These
variations are exemplified by the existence of genetic diseases that are
associated with defects in DNA repair. Table 7-3 presents a list of such
diseases, their modes of inheritance, the specific tumors associated with
them, and their proposed DNA-repair defect. These genetic diseases are
7-6
-------
associated with a high incidence of malignancy, compared with the
incidence in the general population, and often a specific malignancy is
involved (see Table 7-3). The incidences for persons who are genetically
horaozygous for xeroderma pigroentosum, ataxia telangiectasia, and Fanconi's
anemia are about 10""*, and for those who are heterozygous, about
10" - Those who are heterozygous for ataxia telangiectasia and are less
than 45 yr old have a fivefold increase in the risk of cancer, ^0 and
those heterozygous for Fanconi's anemia may account for 51 of all leukemia
deaths (approximately a fivefold increase in susceptibility).^^
Because these people are deficient in the ability to repair radiation-
induced DNA damage and chemical-induced DNA damage,^9 it has been
suggested that alteration in DNA-repair capacity may put them at greater
risk of chemically induced cancers.^-" It must be pointed out that many
of these diseases, especially ataxia telangiectasia, are also associated
with abnormalities of the immune system. Thus, genetic disease may result
in higher risks of cancer via deficiencies in DNA-repair capacity or
iramunocompetence. Among the normal population of humans, there are
probably subtle variations in DNA-repair capacity, but whether these
variations are genetically controlled or are related to cancer risk
remains to be determined.
PROMOTION AND COCARCINOGENESIS
Many studies have shown that a number of modifying factors can
increase the effect of low-dose or low-potency carcinogens that by
themselves would be insufficient to induce malignancies.^9,200 Many
PAHs are complete carcinogens; that is, they have both initiating and
promoting activities. Others—such as pyrene, benzo[e]pyrene,
fluoranthene, and benzofghi]perylene—are weak complete carcinogens and
weak cocarcinogens .*•'*•» ^"* jt £9 difficult to determine what role PAHs
might have in tumor promotion in humans, because there are no good methods
for measuring this activity in the human population. Such end points as
induction of ornithine decarboxylase activity,'--' phospholipid
synthesis,159,178 inflammation,^-* protease activity,*^ cellular
proliferation,5? decrease in differentiated states ,26,201 anej
formation of "dark cells"l^*»172 are manifestations of many promoters,
and many PAHs can induce at least some of these changes. 197,200 But no
single end point correlates with the promoting activity of all the
different chemicals that have promoting activity.
In animal systems, there seems to be a genetic basis for promota-
bility, in that different strains of mice express different suscepti-
bility to promotion during the standard two-stage carcinogenesis assay.
Such strains as CD-I and BALB/c are relatively resistant, whereas the
specifically derived SENCAR strain (i.e., sensitive to carcinogenesis) is
very sensitive to promotion of skin cancer.^'"" The molecular basis of
this difference has not been defined, but recent information suggests that
the skin itself has the sensitivity, inasmuch as skin from SENCAR mice
remains sensitive to promotion even after grafting to BALB/c mice.ZOJ
7-7
-------
No genetic variation in promotability in humans has been described.
However, the fact that pyrenes may have promoting and cocarcinogenic
activity, the possibility that such activity plays a major role in cancel
formation in humans,, and the absence of effective end points in the humai
population all suggest that much more work is necessary before the role c
PAHs in promotion can be understood.
IMMUNOCOMPETENCE
Substantial interest has centered on the role of the imramune system
in preventing the expression of malignancy by recognition and destruc-
tion of newly formed malignant cells. The concept of "immunosurveil-
lance," however, has not been well supported, and, in fact, "stimula-
tion" of malignant cells may even occur.'^
Immunodeficient persons do have a greatly increased risk of develop-
ing a malignancy of the lymphoreticular system. '•' >°0>'3,126 ^he exacc
mechanism responsible for the increase, however, is not clear.
A number of genetic disorders in humans are associated with immuno-
deficiencies. These disorders include ataxia telangiectasia, Wiskott-
Aldrich syndrome, Bloom's syndrome, common variable immunodeficiency,
selective IgA deficiency, Bruton's agammaglobulinemia, severe combined
immunodeficiency, selective IgM deficiency, and immunodeficiency with
normal or increased immunoglobulins.7^
These immunodeficient genetic disorders are usually heterogeneously
linked with a variety of other distinct underlying defects. For example,
persons with ataxia telangiectasia and Bloom's syndrome have severely
impaired DNA-repair capacities, ^9,195 an(j Choge wich severe combined
immunodeficiency also have adenosine deaminase deficiency. •* Therefore,
it is difficult to determine the reasons for the increased cancer
susceptibility of these persons. Epidemiologic evidence fails to support
the idea that immunosurveillance mechanisms are generally involved in
carcinogenesis, but does provide clues to immunologic processes that may
predispose to particular neoplasms.^°
In animal-model systems, PAHs can cause tumors of the lympho-
reticular system, and association with the Ah locus has been
9 W ii o -^
suggested. *>iij In humans, exposure to some hydrocarbons, such as
benzene, has been repeatedly associated with leukemia. Whether variations
in immunocompetence occur naturally in the normal population and whether
PAHs, as a group of environmental contaminants, pose a special risk to
persons with such variations are not known.
STAGE OF DEVELOPMENT
Some cell types undergo periods of heightened sensitivity to
chemicals during their normal growth cycles. For example, in animal-
model systems there are striking differences between germ-cell stages in
7-8
-------
the chemical induction of dominant lethals, translocations, and specific-
locus mutations.1^,160 Moreover, the fetus is at greater risk than
the mother, owing to high doses of environmental chemicals; the
permeability of the blood-brain barrier is greater, and liver-enzyme
conjugating function is poorer- The greater the lipid solubility of
a chemical, the greater its placental transfer; and the placenta is
readily permeable to chemicals with molecular weights less than 600.
Most PAHs fit into these categories, and in animal-model systems such
PAHs as BaP, 3-raethylcholanthrene, and 7,12-dimethylbenz[a]anthracene
cause oocyte and follicle destruction and embryo lethality and
resorption and have a greater incidence of malformation and even cancer
in surviving embryos.&°>1°^»171»190
In humans, gross congenital abnormalities occur in some 2Z of all
infants and are the cause of about 15% of the deaths of infants less
than a year old. Exposure to such agents as viruses, mercury, DDT, CO,
and polybrorainated biphenyls probably accounts for 5-10% of the birth
defects; genetic abnormalities cause 25%; and the causes of the
remainder are largely unknown.°^ Interactions in the intrauterine
environment between genetic predisposition and chemical and biologic
factors are probably responsible for these birth defects. Although
occupational exposure of human males^ and both parents^^ to PAHs
was not associated with increased cancer incidences in the offspring,
recent work has suggested that a combination of chemical exposures of
both parents (especially the mother) resulted in higher incidences of
brain tumors in the offspring. Maternal cigarette-smoking is
associated with decreased birthweight, increased perinatal morbidity and
mortality, and other harmful effects on the newborn. ^^ The PAHs in
cigarette smoke may account for some of its biologic activity, inasmuch
as a relationship has been shown between cigarette-smoking, induction of
AHH activity in human placental tissue, *••"" and a decrease in
placental size;^ PAHs are the major class of AHH inducers found in
cigarette smoke,*-' and thus it is important to note that BaP, which is
in cigarette smoke, can cross the placental barrier.^2
Because PAHs must be metabolized before they produce a biologic
effect, the impact of PAHs on maternal and fetal tissues can be quite
complex. Some examples of these complexities are differences in
developmental patterns of specific enzymes, the relative importance of
maternal and fetal metabolism, the role of metabolism in placental
tissue, the relative importance of hepatic and extrahepatic metabolism,
and sex differences in developmental patterns. The induced and control
forms of AHH and acetanilide 4-hydroxylase are temporally regulated both
before and after the birth of animals.^ The deactivation of
conjugating enzymes (e.g., UDP-glucuronyltransferase, sulfotrans-
ferase, and N-acetyltransferase) is also temporally regulated both
before and after birth, but this regulation can be quite different from
that of AHH.^ The relation between activation and inactivation can
be influenced by the sex of animals.70 Shum ejt al_.l71 showed that
both the fetal and maternal enzymes play an active role in determining
the ultimate fetal toxicity of BaP. Using specific crosses between
AHH-responsive and AHH-nonresponsive strains, these authors could show
7-9
-------
that when Che mother was nonresponsive the enzyme capacity of the fetal
tissue determined the toxicity of BaP, but that when the mother was
AHH-responsive there was no difference in fetal toxicity between
nonresponsive and responsive fetuses. Mice seem to have AHH activity as
early as about 7.5-8.5 d of gestation. ^ This activity slightly
increases before birth, but increases greatly in the first few days
after birth^ and then slowly decreases as the mouse ages.68 It
should be pointed out that in vivo exposure to BaP, in addition to
inducing higher AHH activity in mouse fetal tissue, can suppress humoral
immunity in animals that survive and can cause about a 10-fold increase
in the incidence of various tumors in surviving animals.^0 jt seemg
likely that, in rodents (and perhaps in humans), PAHs can be taken up
and distributed through the placenta intact or in the form of
metabolites, that the metabolites themselves can cause fetal toxicity or
the delayed effects of immune suppression or cancer, and that intact
PAHs can cause fetal enzyme induction, metabolism, and the sequelae
mentioned earlier.
MODIFYING FACTORS
A variety of environmental factors can mitigate or exacerbate the
inherent sensitivity of mammalian tissues to PAHs. These factors are
probably at least as important as some of the genetically controlled
differences discussed earlier and tend to make genetic differences less
distinct. Two factors known to modify PAH carcinogenesis, at least in
animal-model systems, are the physical state of the PAH and the
nutritional state of the exposed organism.
PHYSICAL STATE OF PAH
The sources and the formation of PAHs in the environment are dis-
cussed in Chapters 1-3. Most of them are found as mixtures and many are
found in association with particles, such as cigarette-smoke
particles,^4 fossil-fuel combustion products,^ coal flyash,^7
and asbesto*-fibers. • ^ This association can be important, because
PAHs in the presence of or adsorbed on particles are transported through
membranes more efficiently,'* are cleared from tissue more
slowly,25 and have a different tissue distribution—that determined
by the particle size, rather than by the size of the free PAH.^' The
increased uptake results in more efficient induction of AHH activity at
low PAH concentrations. 1-05 Those exposed to particles containing PAHs
are probably at greater risk of various cancers. 1-°° Uptake,
distribution, and metabolism of PAHs can be so altered by particles that
those who normally would be unaffected by the PAHs may be adversely
affected.
NUTRITIONAL STATE OF HOST
Nutritional status can substantially modify the toxicity of some
environmental pollutants. * For example, specific dietary
7-10
-------
deficiencies are known to increase the toxicity of pesticides—including
carbonate carbaryl, parathion, and captan1'0—and heavy
metals.37»98 Nutritional status can influence microsomal enzymes and
thus affect the toxicity of PAHs. Protein deficiencies can lower AHH
activity,^°l and the type of dietary protein can affect AHH acti-
vity.^ Nutrient deficiencies are observed in both children^ and
adults;189 deficiencies in iron, vitamin A, and vitamin C are the most
prevalent. Whether these deficiencies play a role in PAH-related
effects in humans is not known. Deficiencies or alterations in vitamins
(vitamins A and C) can influence the incidence of PAH-induced cancers in
animal-model systems.8,19,176 Dietary vitamin A (i.e., retinoids) may
also influence the expression of cancer in humans.^8 xhe effects of
vitamins seem to be centered on the later stages of carcinogenesis,
especially tumor progression. Chemoprevention shows promise for alter-
ing or controlling inherent sensitivity (or resistance) to carcino-
genesis, but it should be borne in mind that some vitamins, such as
retinoids, sometimes increase cancer expression and sometimes suppress
it.164
Diets high in fat and meat and low in fiber have been associated
with increased risk of cancer, especially cancer of the
colon.51'199'200 The effect of dietary fat may be related to
alterations in the concentration of colonic secondary bile acids, which
act as colon-tumor promoters.1-'0' ^ PAHs can act as cocarcinogens,
coautagens, or promoters, but whether they play these roles in humans
and whether the nutritional status of the host alters these roles are
not known.
7-11
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TABLE 7-1
Studies Suggesting Correlation between Carcinogen Metabolic
Capacity and Cancer Susceptibility in Humans
Disease
Lung cancer
Lung cancer
Laryngeal
cancer
Lung cancer
Renal and
ureteral cancer
Lung cancer
Tissue Assayed Reference Comments
Lymphocytes 72
49
Lung cancer
Lung cancer
Lung cancer
Lymphocytes
Lymphocytes 185
Bronchi 53
Lymphocytes 186
Lymphocytes, 106
PAMs
Lung cancer
Lung cancer
Lung cancer
Lymphocytes
Antlpyrlne
(half-life)
Lymphocytes,
PAMs,
lung tissue
32
1
107
Lymphocytes
Lymphocytes
Antipyrine
39
Assay only for "Inducibllity"
Radiometric assay; 11 cancer
patients monitored
No controls
BaP binding to DNA higher in
bronchi from lung-cancer
patients; large individual
differences
No controls
Dichotomy of AHH in lympho-
cytes and PAMs in lung-cancer
patients
71
Correlation of AHH depended on
patient—lung cancer vs. normal
Absolute AHH activity domlnantly
Inherited; values given relative
to "standard" panel; no AHH
values presented
Induclbllity determined by non-
induced AHH activity
Antipyrine half-life related to
cancer and smoking
Leukemia
Lymphocytes
11
Susceptibility related to low
AHH
7-12
-------
TABLE 7-2
Studies Suggesting Lack of Correlation between Carcinogen
Metabolic Capacity and Cancer Susceptibility in Humans
Disease
Lung cancer
COPD, chronic
bronchitis
Lung cancer
Laryngeal and
lung cancer
Lung cancer
Lung cancer
Lung cancer
Lung cancer
Tissue Assayed
Lymphocytes
Lymphocytes
Lymphocytes
Lymphocytes
Lymphocytes
(BaP binding)
Lung (organ)
cultures
Antipyrine
(metabolism)
Lymphocytes
Reference
127
110
67
194
66
24
188
91
Comments
Progeny vs. spouse; cancer
patients showed low AHH and
were not tested
Smoking, not cancer, associated
with high BaP metabolism; high
AHH correlated with lymphocyte
stimulation.
Measured AHH in disease-free
subjects; 40% on medication;
lymphocytes from all groups
grew well.
BaP-macroraolecule binding
measured.
HPLC analysis of BaP metabolites
in six cancer patients.
Nine hospitalized patients used.
7-13
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Diaeaae
Inheritance
TABLE 7-J
Human Diseases That May Be Associated with
Proposed Effect
DNA-Repair Deficiencies*
Neoplastic Disease
References
Xeroderma
classical
Xeroderma
pigmentoaum.
variant
Ataxia telangi~
ectaaia
Ataxia telangi-
ectaaia
Autoaomal receaaive.
groupa
Autoaomal recessive
Autoaomsl recessive.
2 complementation
groups
7
Defective in excision repair
of UV damage
Defective post repl icat ion
repair of UV damage
Gamma-repair defect; defect in
some groups in gaana- induced
base damage; proposed defect
in double-strand break re-
joining
Gamma -induced baae damage in
some cases
Multiple basal-cell
mas; malignant melanoma
Multiple basal-cell
aquamoua-cel 1 carcino-
mas; malignant melanoma
Lymphoma (60Z) , leu-
kemia (20Z), solid
tumor (20Z)
Gastric, gallbladder
lymphoma, especially
2,20,22,156
21
45,84,1J2,1JJ,
165,180,195
45.84.1J2.I JJ,
165,180,195
Bloon1a syndrome
Autosomal recessive,
especially in Aahken-
aric Jews
Fanconi'a anemia Autoaomal recesaive
Familial retino-
blast oma
Autoaomal dominant,
80-901 penetrance
Deflect in repair of UV
damage, recombinational
defect?; increased SCE in
peripheral blood lymphocytes
UV eiionucleaae deficiency?
Caana repair?; defect in base-
damage repair
increased in patient*
under 45
Leukemia
Leukemia, squamou**
cell carcinoma, muco-
cutaneou* junction*
Hadiation-induced sar-
comas; secondary meaen-
chymal tumors (mostly
osteosarcoma)
18.42.4J
117,144,161,179,
181
J4.65.196
-------
'•«• *fii»*?;
Diaeaaa ' . , , InMrUance
D-dalatloB retiea- BaajfMat. high
blaatoMl •vMkKCaBca'
Progeria 4«mteaa)al race**ive
(•utchiaaoa-
Gilford) ayvdrOM
PoMa'a ayadroaw autoaaa>al vaaaaaiva
Dyakaratoaia l-lfalMd racaaaiva
coageaite
Y* Cockayne 'a autoaoval racaeaive
t-> ayadroaa
Ul
actinia karatoaia f
Cutaneoua Deati«aa< , ethnic
•aligaaat clwaterjag
Propoted Effect
Caaata-rapair defect; propoied
•pacific locua on chroaoaoaa 13
t
Caeou-repair defect yielding
eiceaa chroaoaoaial aberra-
tion* after ioniaing radia-
tion
Light **n*itivity; exec** SCB
with p*oralen plu* light
Sun *en*itivity; fibroblait*
•an*itiv* to UV killing
(IV *en*itivity
Abnoraal re*i*t*nc* to UV
Heopl**tic Di*ea*e Reference*
•etiooblactoma; no 175.19}
•econdary tuaior*
T 33.118,1)3
Leukaaua 7.87.162
Laukeaia 16.155
T 2.23.163
Skin cancer 87
Melanoma with e»ta*ta*i* 44.96
•Data froai t««au*ieau* a*d HaichaelbauB and Little.195
-------
100
90
80
70
8 60
w
g
o
J3
50
40
30
O
SI 20
oo
g 10
0
Cancer threshold
Gradually increasing polygenic risk
requiring substantial environmental
exposure for tumor induction
enetic
resistance
30 40 50 60 70
Percent of Population
100
FIGURE 7-1. Interplay between genetically controlled
variations and environmental exposure leading to cancer
susceptibility. The population of humans is viewed as
a sigraoidal curve where the extremes are either
genetically resistant or genetically predisposed to
cancer. The shape of the curves would be expected to
change for given subpopulations that contain higher per-
centages of genetically resistant or genetically predisposed
persons. Reprinted with permission from Lynch; " copyright
Academic Press.
7-16
-------
INITIATION
PROXMATE
CAnONOGEM
BIOCHEMICAL PROMOTION
CELLULAR
PROMOTION
{OUESCENrl f NORMUTl
I COL \ lr»ENOTVPE]
r
GEte BJIRANSFi
EXPRESSUNn GENOTYPE
'-——, fa
n." w
GROWTH
FACTORS N GENE
EXPRESSUN '
fMUOEMOUB
FACTORS M
TIMOR GROWTH
v^
* FACTORS M CELL
PROLFERATKM
CtEMCAL PRECAnCMOQENB
MMDOM*** NNOMMMM
•ftAKUW
"l"^--^5
QEfCIC SUSCEPTBUIY
/ 1KMCOI WMAMMT
IfVtL OF EXPOSURE
FIGURE 7-2. Hypothetical stages In chemically Induced carclnogenesis. Characterization and chemical cell
Interactions are discussed In text. Reprinted with permission from R. E. Kourt, Genetic Differences In
Chemical Carctnogenesis.78 Copyright CRC Press, Inc., Boca Raton, Fla. ~~
-------
i
LUNG CANCER
OTHER PULMONARY |
DISEASE
25 M
INDIVIDUAL PATIENTS
FIGURE 7-3. Distribution of hydrocarbon metabolism given in terms of units of AHH per unit of NADH-
dependent cytochrome £ reductase (cyt £> activity for 51 patients. Reprinted with permission from
Kouri e_t aj^.80
-------
METABOLIC
ACTIVATION
DAMAGE TO
CELLULAR DMA
INCOMPLETE OH NO
EXCISION REPAIR
EXCISION REPAIR
I
I
T
DMA REPLICATION AND CELL DIVISION
FURTHER OAMAOC
IN NEW ONA
NORMAL, NEW AND OLD
ONA
INCOMPLETE OR NO
POSTREPLICATION REPAIR
REPAIR OF NEW ONA
POSTREPLICATION REPAIR
JL
ONA REPLICATION AND CELL DIVISION
CYTOXIC. MUTAQENIC,
CHROMOSOMAL EFFECTS
SOME NORMAL
PROGENY
I
T
NORMAL
PROGENY
FIGURE 7-4. Scheme depicting nuclear changes and their toxic
effects. Cytotoxlc, mutagenic, or carcinogenic effects are
thought to result from nonrepair or misrepair of particular
ONA damage. Reprinted with permission from Roberts;157
copyright Academic Press.
7-19
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SUMMARY
The present report attempts to make current the information relative
to the sources (both mobile and stationary), formation, atmospheric
transformations, biologic effects, and pharmacokinetics of a select group
of polycyclic aromatic hydrocarbons (PAHs) and mixtures thereof, to
identify populations hypersensitive to them and to determine the human
risks associated with exposure to them. The specific PAHs considered were
chosen on the basis of relative concentrations in various kinds of
emission or combustion products or because of some unique pharmacologic
property.
SOURCES. ATMOSPHERIC PERSISTENCE, AND TRANSFORMATIONS OF PAHs
The emphasis of this report is on PAHs emitted from mobile sources,
but these substances are ubiquitous—they are found in terrestrial and
aquatic plants, in soils and bottom sediments, and in fresh and marine
waters, as a result of emission from both mobile and stationary sources.
The total annual release of benzo[a]pyrene (BaP), as a surrogate PAH, in
the United States from all sources is estimated at 300-1,300 metric tons;
approximately 40 tons are produced from mobile sources. It is estimated
that by the year 2000 the atmospheric BaP concentration in highly
urbanized areas will be approximately 0.6 ng/m .
The concentration of a particular PAH depends on its source (among
other things), but the phenanthrenes (including methylated derivatives),
the fluorenes (including methylated derivatives), fluoranthene, pyrene,
BaP, benzo[ghilperylene, chrysene, perylene, dibenz[ac]anthracene, and
benz[alanthracene have many common sources. Emission from the combustion
of wood contains more alkylated PAHs than combustion products from other
sources. Wood stoves and fireplaces, nonregulated sources of PAHs, are
important contributors to environmental pollution, particularly in rural
areas with restricted airflow. Wood smoke contains considerable amounts
of particles and adsorbed PAHs, and it is anticipated that this source
will become even more significant with the increased use of wood as a
primary fuel.
Of the total motor-vehicle mileage accumulated in this country, the
light-duty passenger car with spark-ignition engine is the major
contributor, although the number of diesel engines is increasing. By the
mid-1990s, approximately 25% of the passenger fleet will probably be
powered by diesel engines. Rates of emission of particles from diesel
engines are about 2 orders of magnitude greater than those from
catalyst-equipped spark-ignition engines. The total PAH emission from
mobile sources in 1979 was approximately 6,500 metric tons; phenanthrene,
pyrene, fluoranthene, methylphenanthrene, cyclopentapyrene, anthracene,
benzofluorene, chrysene, benzofluoranthene, the benzopyrenes, and
8-1
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benzoperylene were major contributors. Nitropyrene and other nitro-PAHs
have also been found as emission products, but whether these very reactive
substances are artifacts of the sampling process or are present in the
initial emission has not been established. It has been estimated that the
total emission of PAHs in 2000 will be considerably lower because of
advances in collection devices on mobile sources.
There are large uncertainties concerning the persistence of the PAHs,
their chemical transformations, and their atmospheric transport and fate,
although some general principles can be derived. There is evidence of
long-range transport from the analysis of cores from sediments; the PAHs
can be transported over long distances in the atmosphere without important
degradation. The principal processes by which the PAHs are chemically
removed are photooxidation, reaction with ozone, and reaction with
nitrogen dioxide. The latter reaction may be responsible for the
generation of nitro-PAHs, some of which are potent mutagens. Of the PAHs
that have been selected for study, only BaP and pyrene have been
investigated in detail with respect to chemical transformations.
Considerably more study is needed.
BIOLOGIC EFFECTS OF SMOKE, EMISSION, AND SOME OF THEIR PAH COMPONENTS
It has been estimated that as much as 13% of all human cancer deaths
may be attributed to environmental factors, one of which is pollution
resulting from emission from mobile and stationary sources. When tested,
however, particles from diesel and spark-ignition engines and
organic-solvent extracts of these particles have not been very toxic to
animals. Only minimal effects on pulmonary function, reproductive
capacity, and glandular or hepatic function have been observed. The
chronic exposure of newborn rats to diesel-engine exhaust appears to
result in some abnormal development of the central nervous system, as
demonstrated by the slower acquisition of spontaneous locomotor activity
and bar-pressing ability; and small abnormalities have been noted in
visual evoked and somatosensory evoked potentials in exposed neonatal
rats. Whether these changes resulted from exposure to the PAH components
of diesel-engine exhaust has not been ascertained.
Although no imraunologic changes have been observed after exposure of
rats to diesel-engine exhaust, it is known that some PAHs are
immunosuppressive. In particular, high doses of 3-methylcholanthrene,
dibenz[ah]anthracene, 7,12-diraethylbenzanthracene, and BaP reportedly
depress the response of mice and rats to various immunologic challenges.
This imraunosuppresive effect, exhibited by some PAHs but not by exhaust or
emission, can be divorced from the carcinogenicity of these agents.
Extracts of particles from spark-ignition and diesel exhaust are
mutagenic to Salmonella typhimurium in forward- and backward-mutation
assays and in several animal-cell model systems. The extracts were
directly active in the bacterial assay, whereas emission from coke ovens,
roofing tar, cigarette-smoke condensate, wood combustion products, and BaP
8-2
-------
were positive only after metabolic activation—indirect rautagenesis.
After fractionation of the various extracts, the fraction that contained
the PAHs demonstrated the greatest mutagenicity in the bacterial assay. A
major PAH in soot, automobile exhaust, cigarette smoke, and coal fly ash
is cyclopenta[cd]pyrene; it proved to be highly mutagenic in the indirect
assay. Indeed, the total mutagenic activity of kerosene-soot extract
could almost be reproduced by cyclopenta[cd]pyrene alone.
The direct mutagenicity appeared in part to be caused by nitro-PAHs.
These substances have been found in automobile-exhaust particles and in
cigarette smoke, but not in wood combustion products. The nitro-PAHs were
much more mutagenic than the parent compounds, with 1,8-dinitropyrene
being the most mutagenic of all compounds that have been subjected to the
Salmonella/raicrosome assay. The mutagenicity of these nitro derivatives
has not been tested consistently in animal-cell models.
The mouse skin tumorigenesis model has been used to assay the
carcinogenicity of extracts of various particles. The condensates from
spark-ignition engine exhaust proved carcinogenic in this model; those
from diesel exhaust were less active. The exhaust preparations had both
initiation and promotion activities with this model. There are
conflicting reports as to whether the tumorigenicity of the extracts
reflected the additive activity of the major PAHs in the condensates.
When tested for tumorigenicity by inhalation and intratracheal
instillation, the condensates proved not very active. The literature is
contradictory on whether the incidence of neoplasia in animals receiving
automobile-exhaust condensate intratracheally reflected the BaP content of
the condensate. Of a series of compounds that were tested for carcino-
genicity in a mouse-adenoma model, 3-methylcholanthrene, dibenz[ah]anthra-
cene, and BaP proved most active.
The effect of alkylation, particularly methylation, on the carcino-
genicity of various PAHs has been determined with biologic models. The
fluorenes, phenanthrenes, and anthracenes are major components of smoke
and emission, so there has been considerable interest in determining the
effects of methylation of these agents on tumorigenicity. The insertion
of a methyl group at particular positions of the benz[a]anthracene ring
increased tumorigenicity considerably. 9-Methylfluorene was much more
mutagenic than the parent compound in the bacterial assay system. In the
phenanthrene series, the 1- and 9-methyl 'analogues were more rautagenic
than the parent compound. The methylchrysenes are known environmental
pollutants; although the parent compound is generally inactive as a
carcinogen, the 5-methyl derivative was as carcinogenic as BaP and was the
most potent of all the methylated derivatives when tested as an
initiator. Methylated BaPs have been tested for tumor initiation, and
some (the 1-, 3-, and 11-methyl) analogues have been found to be more
active in this regard than the parent compound. It is apparent that the
methylated PAHs, which are present in exhaust and smoke, can contribute to
carcinogenicity.
8-3
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EFFECTIVE BIOLOGIC DOSE
After administration to laboratory animals, PAHs are absorbed readily
and distributed to various tissues. Nonmetabolized material accumulates
and persists in body fat. This phenomenon may be useful for monitoring
the chronic exposure of various populations to emission and smoke that are
rich in PAHs. Particle-bound PAH is retained in the lung to various
degrees that depend on the size and composition of the particles. Once in
the lung, the particle-bound material can be desorbed and distributed to
other tissues. The clearance of a PAH from an animal-model system appears
to depend on the concentration of nonraetabolized compounds in the fat, the
metabolism of the PAH, and biliary, fecal, and urinary excretion. The
excreted metabolites of PAH are largely glucuronides, sulfates, and
hydroxylated and phenolic derivatives.
Virtually all tissues can metabolize PAHs, although liver exhibits the
greatest activity in this regard. The initial metabolism is conducted by
membrane-bound cytochrome P-450-dependent raonooxygenases that yield
epoxide derivatives. The latter may spontaneously rearrange to phenols
that serve as building blocks for later conjugation. The epoxides may
give rise to trans diol derivatives in reactions catalyzed by the
membrane-bound enzyme epoxide hydratase; these diol derivatives may be
excreted unchanged or conjugated as glucuronides. Secondary metabolism by
the cytochrome P-450-dependent monooxygenases yields very reactive diol-
epoxides that can spontaneously rearrange to electrophiles that can
interact with macromolecular nucleophiles, such as DNA. The activity of
the monooxygenases and epoxide hydratase is genetically determined and is
inducible by exposure of an organism to PAHs; the extent of induction is
also genetically determined.
PAHs may also be activated through an arachidonic acid-dependent
co-oxygenation step involving the prostaglandin synthetase complex.
Through this mechanism, the trans diol of BaP, for example, is transformed
to the diol-epoxide at the expense of prostaglandin §2-
The reactive metabolites of PAHs, such as diol-epoxides, interact
covalently with DNA to form adducts. The adducts of BaP diol-epoxide with
DNA have been examined in lung, liver, forestomach, colon, kidney, brain,
and muscle after oral administration of BaP to mice. Human tissues also
are able to catalyze adduct formation. The DNA-adduct profiles appear
specific for a particular tissue. The amount of BaP-DNA adduct formed in
a particular tissue is not correlated with the susceptibility of that
tissue to PAH-induced carcinogenesis. This is evident from consideration
of liver, a tissue that is not ordinarily a target organ for PAH-Lnduced
carcinogenesis, but one in which adducts readily form. The PAH-DNA
adducts have varied turnover rates in different tissues. The turnover
rate is related in part to the normal rate of replication of the cell and
in part to an enzymatic DNA-repair system. Different adducts are removed
from DNA at different rates.
8-4
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With regard to BaP, a linear dose-response relationship has been
observed with formation of DNA adducts as the end point. There appears to
be no threshold dose below which adduct formation will not occur. The
administration of a number of inducers of monooxygenases and of the
conjugating enzyme systems reduces the in vivo formation of adducts;
administration of antioxidants has a similar effect. It has been proposed
that the concentration of PAH-DMA adducts in a particular tissue can be
used as a measure of the "effective biologic dose" of a specific PAH. It
should be simple to determine this dose with currently available sensitive
radioiramunoassay methods. Such methods could be applied to readily
accessible lymphocytes of human populations.
HUMAN EXPOSURE TO AND METABOLISM OF PAHs
Humans are exposed to PAHs almost exclusively through the
gastrointestinal and respiratory tracts. Possibly 992 of exposure to
these substances is through the diet. The daily human exposure to PAHs
from air, water, and food has been estimated. Of approximately 1.8-16 yg
of total PAHs ingested or inhaled, 0.2 and 0.02 ug would be derived from
inhalation or ingestion in water, respectively, and the rest from food.
Of the total, approximately 10% would be BaP.
Although the PAHs are ubiquitous in foodstuffs, their content can be
surprisingly high in some foods as a result of pollution from soils,
irrigation waters, atmospheric fallout, and food-processing. The number
of PAHs ingested may be as high as 100, or even higher. Boiling or
barbecuing substantially affects the composition and quantity of PAHs in
foods.
Occupational exposure to PAHs can lead to inhalation of great
quantities. It has been estimated that a normal adult breathing 20 m
of air per day can inhale approximately 700 ug of PAHs per day in a work
setting that is rich in PAHs, e.g., coal and pitch-coking plants,
gasworks, and roof-tarring operations. It has also been estimated that
people who remain in tunnels with heavy motor traffic all day can inhale
BaP that would be equivalent to that found in a pack of "old-style"
cigarettes. In accordance with the occupational exposure, cancer
mortality among men employed in coal-tar industries reflects excess cancer
in one or more sites, particularly those involving the lungs.
The manner by which PAHs gain access to the systemic circulation is
not known. Serum lipoproteins may constitute a substantial circulatory
pool of the PAHs, which can be transferred into cells by a non-receptor-
mediated process. The pharmacokinetics of PAHs other than BaP in humans
are not well understood.
Normal and malignant human tissues have the metabolic capacity to
biotransform PAHs, especially BaP. The individual variation in this
capacity is very large in the human and appears to be genetically
determined. Although it has been proposed that aryl hydrocarbon
hydroxylase activity in lymphocytes and monocytes of lung-cancer patients
8-5
-------
is highly inducible, compared with that in "normal" patients, this
relationship has not been definitively established and deserves further
study.
There is little information to implicate diet-derived PAHs in any form
of clinical pathology, despite the relatively large amounts of these
compounds ingested. The gastrointestinal system, including the liver, may
be relatively "resistant" to the PAHs; the nature of such resistance
should be explored.
POPULATIONS OF HYPERSENSITIVE PERSONS
The exposure of cells or animals to pollutants, including PAHs, can
lead to toxicoses, mutagenesis, carcinogenesis, and teratogenesis.
Susceptibility to PAH-induced effects may be controlled at the level of
uptake into specific cells, metabolic activation or inactivation, DNA
repair^ expression of DNA damage and its progression to the phenotype of a
mutant cell, and immunocompetence of the person. Several of these steps
(perhaps all) are subject to genetic regulation, although information in
this regard is sketchy. Natural variations in capacity for human DNA
repair lead to increased susceptibility to cancer in some instances, but
the role of the PAHs in this development is not established. Genetically
controlled variations in iramunocompetence are observed in people with high
susceptibility to carcinogenesis; no definitve role of the PAHs has been
suggested. The physical state of a PAH and the nutritional or
developmental state of the host contribute substantially to the observed
biologic effect.
8-6
-------
9
RECOMMENDATIONS
MOBILE SOURCES
Several of the polycyclic aromatic hydrocarbons (PAHs) found in
emission from heavy-duty diesel vehicles and other vehicles are
potentially hazardous to human health. On the basis of what is currently
known, research should be conducted to discover practical and economical
adjustments in engine design for reducing particulate and gaseous PAH
emission. In vitro mutagenesis tests could be used to determine the types
of adjustments that influence the concentrations of PAH chemicals active
in these short-term tests. On preliminary testing, the nitro-PAHs have
been mutagenic; thus, they are an important subgroup of the PAHs
purportedly found in mobile-source emission. However, it is not clear
whether these compounds are formed in exhaust or are artifacts of
sampling; more information is needed to clarify this issue.
ATMOSPHERE
Data from core sampling of bottom sediments in rivers and bays show
long-range transport of presumably unreacted PAHs. PAH chemistry of urban
and industrial emission plumes should be systematically studied both
regionally and on a continental scale.
It is recommended that monitoring of wet and dry PAH deposition be
included in existing ambient-air quality monitoring networks. The
heterogeneous photooxidation and reactions of PAH with ozone and oxides of
nitrogen should be examined under experimental conditions with emphasis on
the nature and size distribution of carrier particles on both PAH and
reaction products; the findings should be correlated with findings on what
actually occurs in the ambient air.
A system for monitoring in large residential localities should be
encouraged, to determine the concentrations of PAHs emitted from
residential fireplaces, wood-burning stoves, and coal-fired heating
systems and the contributions from these sources relative to those from
industrial and commercial boilers and rural municipal waste-burning units.
Concurrently with the monitoring studies, research should be conducted
on design of equipment, technologies, or methods for controlling PAH
emission from residential fireplaces and wood- and coal-burning stoves.
Extracts of the condensates of smoke and other gaseous emission from wood,
coal, diesel and spark-ignition engines, and tobacco must continue to be
tested in in vitro mutagenicity systems, so that activity profiles can be
established and specific active PAHs identified. There is a need to
develop double checks on the findings of research on extracts of
condensates, to eliminate the uncertainty regarding artifacts that occur
in the sampling or extraction processes. The mutagenicity and
9-1
-------
carcinogenicity of each active PAH (especially aitro-PAHs and sulfur-
containing PAHs) should be determined in several animal-model systems to
guide the assessment of their contribution to human disease.
EXPERIMENTAL-ANIMAL STUDIES
Some data on cocarcinogenic activity of PAHs with other chemicals are
available, but this data base needs to be strengthened, and PAHs other
than benzo[a]pyrene (BaP) need to be studied further. Specifically, data
are needed to establish whether various PAHs exhibit cocarcinogenic
activity with other components of exhaust from mobile sources or emission
from other combustion sources, especially wood smoke. The potential
promoting activity of PAHs (including BaP) needs to be established. A
model for promotion other than the mouse skin tumorigenesis system is
needed. Of special interest would be a promotion system using human cells.
Extrapolation of findings from animal studies to humans is tentative
without additional biochemical and pharmacokinetic data. Sorting out the
toxic chemicals in any complex mixture (such as automobile exhaust, wood
smoke, or cigarette smoke) is always difficult. Animal models and
compound-specific testing systems are needed to ascertain the toxic
effects (if anv) of long-term (chronic) exposure of animals to diesel
exhaust and other complex kinds of emission. In this regard, it is
important to stress that the animal model systems include introduction of
the PAHs (alone, in mixtures, and bound to particles) into the diets of
animals in lifetime studies of carcinogenesis. Such dietary exposure is
based on the data that indicate that ingestion contributes heavily to the
body burden of the PAHs. As results from these studies begin to
distinguish the toxic components, biochemical and pharmacokinetic data on
experimental primates (e.g., squirrel monkeys) will be particularly useful
in confirming the findings in animal species and extrapolating to humans.
With improving characterization of the toxic components, studies should be
conducted on lung deposition, uptake, and clearance of PAHs. Studies on
the relationships of carrier-particle size, surface properties in the
submicrometer range, and absorption and adsorption of individual PAHs
should be continued and expanded with an eye to learning the source of the
greatest exposure to the toxic chemicals.
Preliminary studies in animal models should be conducted as soon as
possible to determine the relationship of PAH exposure to birth defects
and other genetic anomalies. Specifically, it would be important to know
whether chronic exposure of newborns to various types of exhaust and smoke
and to mixtures of PAHs and individual PAHs (present in high concentra-
tions in exhaust) affects development of the central nervous system.
DNA ADDUCTS, ENZYME INDUCERS, AND REPAIR
What is the relationship of the enzymes and their activity to the
metabolism of PAHs, other than BaP, and to the formation of PAH-DNA
adducts and their repair? A broader question is: What are the
consequences of the various DNA adducts known to be formed?
9-2
-------
To answer these questions, more sensitive and specific assays must be
developed for detecting PAH metabolite-DNA adducts, e.g., with monoclonal
antibodies. Such assays would be used to determine rates of PAH
metabolite-DNA adduct formation in individual cell types and in organs,
such as the lung, after in vivo experimental exposure to PAHs, especially
low-dose, long-term exposure. With appropriately designed cell-model
systems that use various cell types, the relationship of in vivo repair of
PAH metabolite-DNA adducts should be examined and an activity profile
developed for the individual known active PAHs. Animals other than mice
and rats should be used to examine PAH metabolite-DNA adduct formation and
the mechanisms by which phenolic antioxidants and inducers of aryl
hydrocarbon hydroxylase (AHH) inhibit the formation of adducts.
Can the PAH metabolite-DNA systems be quantified and further developed
for use in monitoring exposure to specific PAHs? The feasibility of using
adducts as a measure of effective biologic dose should be studied for
low-dose extrapolation of bioassay findings to dose-response curves that
show the rate of adduct formation and its relationship to PAH-induced
neoplasia in animal-model systems. The importance of the findings will
depend on a careful analysis of the background concentrations of PAH-DNA
adducts in tissues--i,e., "noise."
HUMAN STUDIES
Obviously, all health-related research findings are useful in
improving the protection of human health. Although research that uses
human beings directly poses difficult problems, there are various kinds of
human studies that avoid those problems. For instance, human tissues can
be used to study the relationship of specific biotransformations of PAHs
to findings of carcinogenicity in animals.
To determine the PAH dose absorbed from human lung tissue, there is a
need to know the chemical form and binding of PAHs on particles, particle
size, composition, clearance rates, and ultimate fate of inhaled
particle-adsorbed PAHs. These findings would be essential in studying the
relationship of formation of PAH metabolite-DNA adducts and the incidences
of adverse health effects found in animal studies.
Progress in understanding research findings could be greatly improved
if an "inventory" of PAHs identified and measured in normal and diseased
human tissues could'be developed. Perhaps samples of appropriate tissues
could be analyzed specifically for this purpose, and biologic and
historical information on the donors could be accumulated. The tissue
profiles of PAH raetabolite-DNA adducts or other indicators could be
compared with those derived from environmental sampling or air monitoring.
The findings in this report show that a high fraction of human
exposure to PAHs is attributable to dietary intake. The possible
relationship of ingested PAHs to increased incidences of gastrointestinal
(or other) malignancies should be included in epidemiologic analyses.
9-3
-------
Such analyses should attempt to isolate the portion of the prevailing
gastrointestinal malignancy rate in selected populations that is due to
food-derived exposure to PAHs. It is apparent that there is resistance in
the gastrointestinal system to the carcinogenic potential of the PAHs.
The mechanisms responsible for this resistance might involve a great
variety of body systems; no specific body function can be pinpointed.
However, some effort should be directed toward finding these mechanisms.
There can be few clinical parallels to this combination of (1) sustained
impingement of carcinogenic compounds on a system of tissues and (2) so
little evidence of realization of the potential deleterious effects of
such chemicals as the PAHs.
The following studies are suggested for the Eurther development and
evaluation of models for assessing the carcinogenicity relationships in
humans or cell cultures derived from humans.
• Consider the use of radiolabeled tracers or immunologic methods to
study the metabolism of select PAHs, such as benzofa]pyrene, in humans.
The absolute amounts of compound required for single-dose exposure would
be insignificant, compared with the heavy daily exposure commonly found in
foods, but the medical and scientific .value of the data obtained would be
very large indeed.
• Examine the metabolism, pharraacokinetics, and DNA binding of
nitro-PAHs.
• Conduct systematic studies of the patterns of tissue enzymatic
activities relevant to PAH metabolism as a function of age, sex, hormone
activities, nutritional state, or state of health (disease).
• Correlate enzymatic activities, especially those involved in PAH
activation to ultimate carcinogens, in one tissue type with the same
biochemical properties of other tissues in the same person. These data
would have the great advantage of eliminating the factor of genetic
diversity in assessing the pathophysiologic significance of such enzymatic
characteris tics .
• Determine which genetically controlled deficiencies in iramuno-
corapetence are related to specific immune dysfunctions.
• Develop better methods for determining the numbers of heterozygotes
at any given locus and use these methods specifically in populations
exposed to high concentrations of PAHs.
• On the basis of such data, monitor the development of DNA adducts
in humans with the hope of extrapolating to cancer risk.
• Reassess the role of genetically mediated differences in AHH
responsiveness in determining cancer susceptibility by using multiple
human tissues and multiple enzyme end points (assay for PAH receptors in
human tissue; assay for total and specific cytochrome P-450s by mono-
clononal antibodies; assay for AHH expression of these genes; use of
lymphoid, epidermal, and fibrobLastic cells as sources of tissues for
enzymatic assays; and use of multiple functional assays for AHH, e.g.,
fluorimetry, high-performance liquid chroraatography, and DNA binding and
repair).
• Determine whether the promotion-associated steps that occur in
mouse skin also occur in human skin. Attempt to develop assays to measure
for "promotability" among humans; i.e., ire there genetic variants among
humans for "promotability"?
9-4
-------
• Undertake occupational studies of persons exposed to high
concentrations of PAHs. These studies would record detailed information
on job histories and smoking habits of all persons studied, so that the
effects attributable to occupational PAH exposure and cigarette-smoking
could be assessed.
• Study the relationship of PAH measurements to the various defined
job categories. A studied control group (non-PAH-exposed) must be
included.
9-5
-------
APPENDIX A
LISTS OF POLYCYCLIC AROMATIC HYDROCARBONS
This appendix consists of four tables. The first is an
alphabetical list of polycyclic aromatic hydrocarbons (PAHs)
discussed in the report and close chemical relatives, with
molecular formulas and CAS numbers. The second is a list of
structural formulas (ordered according to structural complex-
ity) and ratings of carcinogenic activity; these ratings
indicate only relative activity. The third table lists nitro-
arenes that have been detected in particulate extracts of
diesel exhaust, and the fourth shows their structural formulas.
A-l
-------
TABLE A-l
Polycyclic Aromatic Hydrocarbons and Related Compounds:
Molecular Formulas and CAS Numbers
Name
Molecular Formula
Acenaphthylene
Acephenanthrylene
Acridine
Anthanthrene
Anthracene
9 , 10-Anthracenedione
9(10H)-Anthracenone
Anthraquinone
Anthrone
Benz[elacephenanthrylene
Benz[c]acrid ine
Benz[a 1 anthracene
7H-Benz[de]anthracen-7-one
Benzanthrone
Benzo[b]chrysene
Benzo[c]chrysene
Benzo[g]chrysene
Benzo[c1cinnoline
Benzo[aIdibenzothiophene
Benzo[b]fluoranthene
Benzo(ghilfluoranthene
Benzo f j1fluoranthene
Benzofk)fluoranthene
1 W-Benzo fa ] fluorene
1 Ill-Ben zo [b) fluorene
7^-Benzo [c 1 fluorene
Benzo[h]naphthofl,2-f]quinolene
Benzofb]naphthof2,l-d]thiophene
Benzofrstlpentaphene
Benzofghilpervlene
Benzofclphenanthrene
Benzofa]pyrene
Benzo[e1pyrene
Benzo[flquinoline
Benzo[hlquinoline
Benzo[b)triphenylene
Biphenylene
9H-Carbazole
Chrysene
Coronene
4H-Cyclopenta fdef]phenanthrene
Cyclopentafcd]pyrene
Dibenz[a,h]acridine
Dibenzfa,j]acridine
Oibenz[c,h]acridine
C16H10
C13H9N
see Dibenzo [def ,mno ]chrysene
C14H10
C14H8°2
C14H10°
see 9, 10-Anthracenedione
see 9( 10H)-Anthracenone
C20H12
C17H11
C18H12
C17H100
see 7H-Benz [de ] anthracen-7-one
C22H14
C22H14
C22H14
C12H8N2
see Benzofb |naphtho[2 , 1-dJ-
thiophene
see Benz [e ] acephenanthry lene
C18H10
C20H12
C20H12
^17H12
C17H12
C17H12
C21«13
C16H10
C24H14
C22H12
C18H12
C20H12
C20H12
C13H9
CUH9
C22H14
C12H8
°12H9
C18H12
C24H12
C15H10
C18H10
C21H13
C21H13
CAS No.
208-96-8
201-06-9
260-94-6
120-12-7
84-65-1
90-44-8
205-99-2
225-51-4
56-55-3
82-05-3
214-17-5
194-69-4
196-78-1
230-17-1
203-12-3
205-82-3
207-08-9
238-84-6
243-17-4
205-12-9
196-79-2
239-35-0
189-55-9
191-24-2
195-19-7
50-32-8
192-97-2
85-02-9
230-27-3
215-58-7
259-79-0
86-74-8
218-01-9
191-07-1
203-64-5
27208-37-3
226-36-8
224-42-0
224-53-3
A-2
-------
Table A-l (continued)
Name
Dibenzfa.c]anthracene
Dibenz[a,h]anthracene
Dibenz[a,j]anthracene
7H-Dibenzo[a,glcarbazole
13H-Dibenzo[a,ijcarbazole
7fl-Dibenzo[c ,g]carbazole
Dibenzo[b,def]chrvsene
Dibenzo[def,rano]chrysene
Dibenzo[def,p]chrysene
Dibenzo[b,hIphenanthrene
Dibenzo[a,e]pyrene
Dibenzo[a,h]pyrene
Dibenzo[a,i]pyrene
Dibenzo[a,1]pyrene
Dibenzothiophene
Fluoranthene
9H-Fluorene
9H-Fluoren-9-one
Indenofl,2,3-cd]pyrene
IH-Indole
Isoquinoline
Naphthacene
Naphthalene
Naphtho[l,2,3,4-def]chrysene
Naphtho[2,3-fjquinoline
Pentaphene
Perylene
IH-Phenalene
Phenanthraquinone
Phenanthrene
9,10-Phenanthrenedione
Phenanthridine
1,10-Phenanthroline
Phenanthro[4,5-bcd]thiophene
Phenazine
Phenazone
Picene
Pyrene
Quinoline
Triphenylene
9H-Xanthene
Molecular Formula
see Benzo[b] triphenylene
C22H14
C20H13
C20H13
C20H13
C24H14
C22H12
C24H14
see Pentaphene
see Naphthofl ,2,3,4-def Jchrysene
see Dibenzo[b,def Jchrysene
see Benzo[rst]pentaphene
see Dibenzofdef ,p]chrysene
C12H8S
C16H10
C13H10
C13H8°
C22H12
C8H7N
C9H?N
C18H12
C10H8
C24H14
C17Hn
C22H14
C20H12
C13H10
see 9, 10-Phenanthrenedione
C14H10
C14H802
C13H9N
C12H8N2
C14H8S
C12H8N2
see Benzo [c ]cinnoline
C22H14
CL6H10
C9H7N
C18H12
C13H100
CAS No.
53-07-3
224-41-9
207-84-1
239-64-5
194-59-2
189-64-0
191-26-4
191-30-0
132-65-0
206-44-0
86-73-7
484-25-9
193-39-5
120-72-9
119-65-3
92-24-0
91-20-3
192-65-4
224-98-6
222-93-5
198-55-0
203-80-5
85-01-8
84-11-7
229-87-8
66-71-7
30796-92-0
92-82-0
213-46-7
129-00-0
91-22-5
217-59-4
92-83-1
A-3
-------
TABLE A-2
Polycyclic Aromatic Hydrocarbons and Related Compounds:
Structural Formulas, Molecular Weights, and Carcinogenic Activity
Structural Formula
Name
iH-Indole
Molecular Carcinogenic
Weight Activity8
117.0578 0
Quincline
129.0578
8
Isoquinoline
129.0578
Naphthalene
128.0626 0
Acenaphthylene
Biphenylene
A-4
152.0626 0
152.0626 NA
-------
Table A-2 (continued)
Structural Formula
10
Name
Phenazine
(Phenazone)
Benzo[c]cinnoline
Molecular
Weight
Dibenzothiophene 184.0347
Carcinogenic
Activity
1,10-Phenanthroline 180.0687 NA
180.0687 NA
180.0687 NA
5 4
9H-Carbazole
167.0735
8
9H-Fluoren-9-one
180.0575 NA
A-5
-------
Table A-2 (continued)
Structural Formula
7 6
Name
Benzo[fIquinoline
Phenanthridine
Benzo[h)quinoline
Molecular
Weight
179.0735
179.0735
179.0735
Carcinogenic
Act ivity
8
Acridine
179.0735
9H-Fluorene
IH-Phenalene
A-6
166.0783
166.0783
NA
-------
Table A-2 (continued)
Structural Formula
Name
9H-Xanthene
Molecular
Weight
182.0732
9,10-Anthracenedione
(Anthraquinone)
208.0524
Carcinogenic
Activity
NA
9,10-Phenanthrenedione 208.0524 NA
(Phenanthraquinone)
NA
8
Phenanthro[4,5-bcd]- 208.0347 NA
thiophene
Phenanthrene
178.0783 0
5 10 4
Anthracene
A-7
178.0783 0
-------
Table A-2 (continued)
Structural Formula
Name
9[10H]-Anthracenone
(Anthrone)
4H-Cyclopenta[def]-
phenanthrene
Molecular Carcinogenic
Weight Activity
194.0732
190.0783
NA
Pyrene
Acephenanthrylene
Fluoranthene
Benzo[b]naphtho-
[2,l-d]thiophene
(Benzo[a]dibenzo-
thiophene)
202.0783
202.0783
202.0783
234.0503
NA
A-8
-------
Table A-2 (continued)
Structural Forrau1a
8
Name
Molecular Carcinogenic
Weight Activity
7H-Benz[de]anthracen- 230.0732 NA
7-one
(Benzanthrone)
Naphtho[2,3-f]quino-
line
229.0891 0/+
Benz[clacridine
229.0891
II
7H-Benzo[c]fluorene 216.0939 0
4 HH-Benzo[a]fluorene 216.0939
HH-Benzo[b]fluorene 216.0939 NA
A-9
-------
Table A-2 (continued)
Structural Formula
Name
Molecular
Weight
II
4 Benz[ajanthracene
Naphthacene
228.0939
Carcinogenic
Activity
Benzo[ghi]fluoranthene 226.0783 0
Cyclopenta[cd]pyrene 226.0783 +
228.0939 0
Benzo[c]phenanthrene 228.0939 •*•
Triphenylene
228.0939 0
A-10
-------
Table A-2 (continued)
Structural Formula
Name
Chrysene
Benzo[a]pyrene
Molecular
Weight
228.0939
252.0939
Carcinogenic
Activity
12
IJ
8
Benz[e]pyrene
Perylene
252.0939 0/
252.0939 0
Benzo[j]fluoranthene 252.0939
A-ll
-------
Table A-2 (continued)
Structural Formula
Name
Molecular
Weight
Carcinogenic
Activity
Benz[ejacephenan-
_ thrylene
(Benzo[b]fluoranthene
252.0939
Benzo[k]fluoranthene 252.0939
7H-Dibenzo[c,g]car-
bazole
267.1048
8
12
8
13
7H-Dibenzo[a,g]car-
bazole
13H-Dibenzo[a,i]car-
bazole
267.1048
267.1048
A-12
-------
Table A-2 (continued)
Structural Formula
Name
Molecular
Weight
Benzo[h]naphtho[l,2-f]- 279.1048
quinoline
Dibenzfa,j]acridine 279.1048
Dibenz[a,h]acridine 279.1048
Carcinogenic
Activity
Dibenz[c,h]acridine 279.1048 +
Benzo[ghi]perylene 276.0939 +
A-13
-------
Table A-2 (continued]
Structural Formula
10
Name
Dibenzo[def,mno]
chrysene
(Anthanthrene)
Molecular Carcinogenic
Weight Activity
276.0939 0
10
3 Indeno[l,2,3-cd]pyrene 276.0939
6
Dibenz[a,h]anthracene 278.1096
Benzo[c Jchrysene
Benzo[g]chrysene
278.1096
278.1096
A-14
-------
Table A-2 (continued)
Structural Formula
Name
4 Picene
Molecular
Weight
278.1096
Benzo[b]chrysene
278.1096
Benzo[b]triphenylene 278.1096
(Dibenz[a,c]anthracene)
Pentaphene
(Dibenzo[b,h]phen-
anthrene)
278.1096
Dibenzfa,jlanthracene 278.1096
Carcinogenic
Activity
A-15
-------
Table A-2 (continued)
Structural Formula
II
Name
Coronene
Molecular
Weight
300.0939
Benzo[rst]pentaphene 302.1096
(Dibenzo[a,ijpyrene)
2 Dibenzo[b,def]chryaene 302.1096
(Dibenzo[a,h]pyrene)
Dibenzo[def,p]chrysene 302.1096
(Dibenzo[a,1]pyrene)
Carcinogenic
Activity
0/+
Naphtho[l,2,3,4-def]- 302.1096
chrysene
(Dibenzo[a,e Jpyrene
aNA - not available.
A-16
-------
TABLE A-3
Nitroarenes Detected in Diesel-Exhaust Particulate Extracts:
Molecular Formulas and Molecular Weights
Struc-
ture
No.
Name
Mononitroarenes:
1 Nitroindene
2 Nitroacenaphthylene
3 Nitroacenaphthene
4 Nitrobiphenyl
5 Nitrofluorene
6 Nitromethylacenaphthylene
7 Nitromethylacenaphthene
8 Nitromethylbiphenyl
9 Nitroanthracene
10 Nitrophenanthrene
11 Nitromethylflourene
12 Nitromethylanthracene
13 Nitromethylphenanthrene
14 Nitrotrimethylnaphthylene
15 Nitrofluoraathene
16 Nitropyrene
17 Nitro(C2-alkyl)anthracene
18 Nitro(C2-alkyl)phenanthrene
19 Nitrobenzofluorene
20 Nitromethylfluoranthrene
21 Nitromethylpyrene
22 Nitro(C3-alkyl)anthracene
23 Nitro(C3-alkyl)phenanthrene
24 Nitrochrysene
25 Nitrobenzanthracene
26 Nitronaphthacene
27 Nitrotriphenylene
28 Nitromethylchrysene
29 Nitromethylbenzanthracene
30 Nitromethyltriphenylene
31 Nitrobenzopyrene
32 Nitroperylene
33 Nitrobenzofluoranthene
Polynitroarenes:
34 Dinitromethylnaphthylene
35 Dinitrofluorene
36 Dinitroraethylbiphenyl
37 Dinitrophenanthrene
38 Dinitropyrene
39 Trinitropyrene
40 Trinitro(C5-alkyl)fluorene
41 Dinitro(C^-alkyl)fluorene
42. Dinitro(C4~alkyl)pyrene
Molecular
Formula
C13H8N2°4
C13H10N2°4
C14H8N204
C16H8N2°4
C16H7N3°6
C18H17N3°6
C19H19N2°4
C20H16N204
Molecular
Weight
C9H7N02
C12H7N02
C12H9N02
C12H9N02
C13H9N02
C13H9N02
C13HUN02
C13HUN02
C14H9N02
C14H9N02
C14HUN02
C15HUN02
C15H11N02
C13H13N02
C16H9N02
C16H9N02
C16H13N02
C16H13N02
C17«11N°2
C17HUN02
C17H12N02
C17H15N02
C17H15N02
C18HUN02
C18H11N02
C18HUN02
C18HUN02
C19H13N02
C19H13N02
C19H13N02
C20H11N02
C20HUN02
C20HUN02
161.16
197.19
199.21
199.21
211.22
211.22
213.24
213.24
223.23
223.23
225.25
237.26
237.26
215.25
247.25
247.25
251.29
251.29
261.28
261.28
262.29
265.31
265.31
273.29
273.29
273.29
273.29
.287.32
287.32
287.32
297.31
297.31
297.31
233.20
256.22
258.23
268.23
292.25
337.25
371.35
339.37
348.36
A-17
-------
Table A-3 (continued)
Struc-
ture
No.a Name
Nitro-oxyarenes:
43 Nitronaphthaquinone
44 NLtrodihydroxynaphthalene
45 Nitronaphthalic acid
46 Nitrofluorenone
47 Nitroanthrone
48 Nitrophenanthrone
49 Nitroanthraquinone
50 Nitrohydroxymethylfluorene
51 Nitrofluoranthone
52 Nitrofluoranthenequinone
53 Nitropyrenequinone
54 Nitropyrone
55 Nitrodimethylanthracene
carboxaldehyde
56 Nitrodimethylphenanthrene
carboxaldehyde
Other nitrogen compounds:
Molecular
Formula
C10H5N04
CLOH8N04
CLOHgN04
CL3H7N03
CL4H9N03
C14H9N03
C14H?N04
C14HUN03
C16H8N03
C16H7N04
C16H9N03
C17H12N03
57
58
59
60
Benzocinnoline
Methylbenzocinnol ine
Phenylnaphthylamine
(C2-Alkyl)phenylnaphthylaraine
C12H8N2
C13H1QN2
^16^13^
C18H17N
Molecular
Weight
203.15
206.18
206.18
225.20
239.23
239.23
253.21
241.25
262.24
277.24
278.24
263.25
278.29
278.29
180.21
194.24
219.29
247.34
aStructure numbers refer to structures in Table A-4.
A-18
-------
TABLE A-4
Structures of Nitroarenes3
H H
NO.
N02
3.2
HH
NO
22.39
59,60
A-19
-------
NO.
57,58
U
NO.
'OH
"Numbers under structures refer to compounds listed in Table A-3.
A-20
-------
APPENDIX B
POLYCYCLIC AROMATIC HYDROCARBONS IN THE AMBIENT ATMOSPHERE
Ambient concen-
Compound tration, ng/m References
Unsubstituted:
Biphenyl a 1
Naphthalene 0.05-0.35 6
Anthracene 0.07-6.15 6
Phenanthrene 0.04-25 6
Benz[a]anthracene 0.5-22 6
Dibenz[ac]anthracene 0.03-4.5 6
Benzo[cJphenanthrene 0.04-1.0 6
Benzofa]fluorene 0.8 6
Benzo[b]fluorene 0.1-1.I 6
Dihydrobenzo[a,b, and c]fluorene3 0.03-0.9 1,6
Fluoranthene 0.1-41 6
Benzo[b]fluoranthene 0.1-7.4 6
Benzo[j1fluoranthene 0.2-4.4 6
Benzo[k]fluoranthene 0.14-20 6
Benzo[ghilfluoranthene 0.9-9.1 6
Pyrene 0.1-35 6
Benzo[a]pyrene 0.1-75 6
Benzo[e]pyrene 0.1-42 6
Anthanthrene (dibenzofcdjkjpyrene) 0.1-6 6
Dibenzopyrenes (4 isomers) a,b 4,6
Indeno(l,2,3-cd)pyrene 1-12.8 6
Chrysene 0.2-39 6
Perylene 0.1-5 6
Benzotghilperylene 0.2-46 6
Coronene 0.2-48 6
Picene a 1
Benzo[c]phenanthrene a 1
Benzo[b]chrysene a 1
Benzofc]tetraphene a 1
Hexahydrochryaene a 1
Dihydrobenzo[c]phenanthrene a 1
Dihydrobenz[a]anthracene a 1
Dihydrochrysene a 1
Benzacenaphthylene b 4
Binaphthyl (3 isoraers) b 4
Quarterphenyl b 4
Diphenylacenaphthalene b 4
B-l
-------
Ambient concen-
Compound _____ tration. ng/tn References
Alky1-substituted:
Methylanthracene 0.22-0.66 6
1-, 2-, 3-, and 9-Methylphenan-
threnes b 4
1-Methylpyrene 0.01-0.15 6
1-, 2-, and 4-Methylpyrenes b 4
Ethylanthracene0 a 1,4
Ethylphenanthrenec 1,4
Methylfluoranthene (5 isomers) a 1,4
Methylbenz[alanthracene a 1
Methylchrysene a 1
Methylbenzo[bk]fluoranthene a 1
Methylbenzofae]pyrene a 1
MethyIbenzopyrenes or benzo-
fluoranthenes (5 isomers) b 4
4]i-Cyclopenta [defl phenanthrene b 4
Methyl 4H-cyclopenta[def]phen-
anthrene b 4
Ethyl 4H-cyclopenta[def]phenanthrene
(5 isomers) b 4
Ethylmethyl 4PI-cyclopenta[def]-
phenanthrene b 4
Ethylmethyl anthracene or phenan-
threne b 4
Ethylpyrene or fluoranthene
(4 isomers) b 4
Ethylmethylpyrene or fluoranthene
(3 isomers) b 4
Methylbenzo[c]phenanthrene b 4
Me thyIbenzo[ghi]fluoranthene b 4
Ethylchrysene or benz[a]anthracene
(7 isomers) b 4
MethyIbinaphthyl (4 isomers) b 4
Methyldibenzanthracene b 4
N-Hetero (aza):
Acridine 0.04 6
Methylacridine 0.007 6
Benz[a]acridine 0.2 6
Benz[c]acridine 0.1-1.5 6
Dibenz[aj]acridine 0.04 6
Dibenz[ahlacridine 0.08-0.1 6
Carbazole 1.9 6
Quinoline 0.02-0.6 6
Methylquinoline 0.03 6
2,6-Dimethylquinoline 0.03 6
Dimethylquinolines 0.04-0.09 6
Ethylquinolines 0.01-0.02 6
03 Alkylquinolines 0.01 6
B-2
-------
Ambient concen-
Compound
Benzo[ f lqui.noline
Benzo[h]quinoline
L1-Indeno[I,2b]quincline
Phenanthridine
Isoquinoline
Me thylisoquinclines
Dimethylisoquinolines
Ethylisoquinolines
03 Alkylisoquinolines
Benz[f1isoquinolines
4-Azafluorene
4-Azapyrene and isomers
1-Azafluoranthene
Benzo[c]cinnoline
2-Methylindole
Benzo[a]carbazole
Benzo[c]carbazole
Phenoxazine
C/^ Alkylquinolines
Methylphenanthridines
Methylbenzoquinolines
Methylbenzoisoquinolines
Azabenzofluorenes
Methylazapyrenes
Methylazafluoranthenes
Azabenz[a Janthracene
Azachrysenes
Azabenzopyrenes
Azabenzofluoranthenes
Dibenzoquinolines
Dibenzoisoquinclines
Quinones:
9,10-Anthraquinone
Benzo[a]pyrene 6,12-quinone
Benzo[aJpyrene 1,6-quinone
Benzo[a]pyrene 3,6-quinone
Dibenzo[b.defJchrysene 7,14-quinone
Phenalen-1-one
Benzanthrone
Perinaphthanone
Carboxylic acids:
Naphthalene carboxylic acid
Phenanthrene carboxylic acid
Anthracene carboxylic acid
Pyrene carboxylic acid
tration, ng/m
0.01-0.2
0.01-0.3
0.1
0.02
0.14-0.18
0.17-0.31
0.06
0.07-0.16
0.03
0.03-0.11
0.005
0.02-13
trace-3
1.0
2.0
a
a
a
a
a
a
a
a
a
a
a
a
a
a
a
a
b
b
b
b
b
0.3-17
0.6-48
a
a
a
a
a
3
References
6
6
6
6
6
6
6
6
6
6
6
6
6
6
6
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
5
5
5
5
5
7
5,6
B-3
-------
Ambient concen-
Compound tration, ng/m References
Phenols:
flydroxyanthracene a 1
Hydroxyphenanthrene a 1
Hydroxypyrene a 1
Hydroxyfluoranthene a 1
S-Hetero:
BenzoChiazole 0.014-0.02 2
Dibenzothiophene b 4
Methyldibenzothiophenes (3 isomers) b 4
EthyldibenzoChiophene b 4
Benzo[def]dibenzothiophene b 4
Naphthobenzothiophenes (3 isotners) b 4
Methylnaphthobenzothiophenes b 4
(3 isomers)
Nitro derivatives:
1-Nitropyrene b 7
3-Nicrofluoranthene b 7
5-Nitroacenaphthene b 7
6-'Nitrobenzo [a] pyrene b 3
aConcentration reported in micrograms per gram of particulate matter or
micrograms per gram of benzene-soluble fraction, but not in nanograms per
cubic Tieter.
^"Compound identified, but no concentration reported.
cEight isomers of ethylanthracene/ethylphenanthrene identified.
isoraers of methylbenz[a 1anthracene/methylchrysene identified.
B-4
-------
REFERENCES
Cautreels, W., and K. Van Cauwenberghe. Determination of organic
compounds in airborne particulate matter by gas chromatography-mass
spectrometry. Atraos. Environ. 10:447-457, 1976.
Dong, M. W., D. C. Locke, and D. Hoffmann. Characterization of aza-
arenes in basic organic portion of suspended particulate matter.
Environ. Sci. Technol. 11:612-618, 1977.
Jager, J. Detection and characterization of nitro derivatives of
some polycyclic aromatic hydrocarbons by fluorescence quenching
after thin-layer chromatography. Application to air pollution
analysis. J. Chromatogr. 152:575-578, 1978.
Lee, M. L., M. Novotny, and K.D. Bartle. Gas chroraatography/mass
spectrometic and nuclear magnetic resonance determination of poly-
nuclear aromatic hydrocarbons in airborne particulates. Anal. Chera.
48:1566-1572, 1976.
Pierce, R. C., and M. Katz. Chromatographic isolation and spectral
analysis of polycyclic quinones. Application to air pollution
analysis. Environ. Sci. Technol. 10:45-51, 1976.
Santodonato, J., P. Howard, D. Basu, S. Lande, J. K. Selkirk, and P-
Sheehe. Health Assessment Document for Polycyclic Organic Matter.
EPA-600/9-79-008. Research Triangle Park, N.C.: U.S. Environmental
Protection Agency, Office of Health and Environmental Assessment,
Environmental Criteria and Assessment Office, 1979. [475] pp.
(preprint)
Tokiwa, H., R. Nakagawa, and Y. Ohnishi. Mutagenic assay of aromatic
nitro compounds with Salmonella typhimurium. Mutat. Res. 91:
321-325, 1981.
B-5
-------
APPENDIX C
HUMAN-CANCER RISK ASSESSMENT*
Malcolm C. Pike
Eplderaiologic studies, animal carcinogenesis experiments, and in
vitro mutagenesis and transformation assays all provide data relevant to
the assessment of the human-cancer risk from exposure to PAHs.
The data used in this assessment are in the main taken from epidemi-
ologic studies, because they refer directly to man. It is recognized
that an alternative approach would have been the extrapolation of experi-
mental animal data to humans, but the epidemiologic approach offers two
advantages: the avoidance of interspecies extrapolation and the
derivation of results from exposures not too different from that suffered
by the general population. Epidemiologic studies often suffer from
various inadequacies, such as imprecise dose measurements and poor
measurement of confounding factors, and exposure is invariably to a
complex mixture of PAHs and other chemicals. Extrapolation to other
complex mixtures therefore inevitably involves making assumptions, and
evidence from in vitro and in vivo experiments must be sought to provide a
rational basis for these assumptions.
At present, the two sources of human exposure to PAHs on which data
appear reliable are work around coke ovens and cigarette-smoking. The
major known human cancer associated with exposure to chemical mixtures
containing PAHs is undoubtedly lung cancer. Although cigarette-smoking is
of overwhelming importance as a cause of lung cancer*'^ and cigarette
smoke does contain PAHs, this appendix is concerned with cigarette-smoking
only insofar as the information derived from epidemiologic study of the
smoking population is essential in measuring the health effects that might
be expected when humans are exposed to other PAH-containing mixtures.
The quantitative relationship between cigarette-smoking and lung
cancer has been thoroughly explored in many epidemiologic studies4 and
is well understood.^»2" However, it is still far from established that
the PAH content of cigarette smoke is responsible for the development of
lung cancer. Epidemiologic data (mainly occupational) on the relationship
*Quantitative risk assessment is a developing, rather than a precise,
science. The numerical estimates in this appendix are based on a series
of assumptions. The use of different assumptions or extrapolations from
animal data could lead to very different conclusions. The calculated
risk values at ambient concentrations are not meant to be absolute
indicators of risk, but rather to indicate the region between the upper
bounds of risk and the lower bound of zero risk.
C-l
-------
between exposure to other PAH-containing mixtures and lung cancer are much
less precise. The lung-cancer risks (as well as the risks of cancer at
other sites) associated with such exposures have, in fact, always been
measured in relation to lung-cancer rates in the "nonexposed," and
cigarette-smoking has been responsible for some 90% of the lung cancers in
these "nonexposed."' To measure the risk, rather than the relative
risk, associated with these other exposures, it is essential to understand
the lung-cancer risk associated with cigarette-smoking.
DEFINITIONS
The incidence rate of a disease is the number of cases of the disease
that are diagnosed during a specified period per specified unit of
population. The mortality rate of a disease is the number of deaths
from the disease during a specified period per specified unit of popula-
tion. In epideraiologic studies, the unit of time is usually a year and
the unit of population is usually 100,000. All incidence (and mortality)
rates quoted here for man use a period of a year, but the unit of popula-
tion is 1, unless otherwise stated. If we write the incidence rate
without qualification (e.g., I), it is assumed to refer to the standard
condition of 1 yr and 1 person. The incidence rate is often affected by
many factors, particularly age, and, if the incidence rate is for some
particular subgroup, this is stated in referring to the incidence rate,
and the symbol, I, for incidence rate is qualified in some way, e.g., I(t)
for the incidence rate for a person of age t.
For cancers associated with a substantial cure rate or a long time
between diagnosis and death, the incidence and mortality rates may be very
different. For lung cancer—the major cancer discussed in this
chapter — the distinction is not so important, because some 75% of newly
diagnosed lung-cancer patients are dead within a year and some 90% within
3 yr.
The lifetime risk of a disease is the probability of being diagnosed
as having the disease by age 70 (a "lifetime") in the absence of other
causes of death. This measure has been found particularly useful in
comparing human data and experimental-animal data and forms the basis
of current methods of extrapolating animal data to man. The lifetime
risk is virtually identical with the cumulative incidence rate (to age 70)
used by the International Agency for Research on Cancer. '
CIGARETTE-SMOKING AS A SOURCE OF PAH EXPOSURE
Much of what has been learned about the quantitative relationship
between cigarette-smoking and lung cancer over the last 30 yr may be
summarized by the statement, "The excess lung-cancer incidence of a
smoker, compared with a nonsmoker, is proportional to the number of
cigarettes smoked per day and to the duration of smoking raised to the
C-2
-------
power 4. 5. "8.40 if we write the excess incidence — or single-cause
incidence — of a smoker aged t years who started smoking at age w years
and who smokes c cigarettes per day as Ic(t,w), that statement may be
expressed mathematically as
Ic(t,w) = ac(t - w)4-5, (1)
where the constant a is approximately 1.0 x 10"^ for U.K. smokers. &
For U.S. smokers, the constant a must be decreased by
. 12,13,19,33 Thg reagong for thig inciu(je the
25-50%. ,,, Thg reagong for thig inciu(je the use of different
tobaccos in the two countries and the mode of cigarette-smoking — in
particular, British smokers tend to smoke their cigarettes down to a
considerably shorter butt.7'43 Similar reasons probably explain the
existence of a range.
We may express the lung-cancer risks from cigarettes in the usual
risk-assessment terms of "lifetime risk" by using Equation 1.
"Lifetime" is taken as 70 yr, and exposure is taken as starting at birth.
If exposure is to c cigarettes per day, Equation 1 shows that the
lung-cancer rate at age t will be
Ic(t,0) = act4'5. (2)
The lifetime risk (cumulative incidence) can be shown to be
CIC(T) = 1 - exp[-ac(705-5/5.5)]. (3)
The lifetime lung-cancer risk associated with one U.K. cigarette per day
is 2,524 per 100,000, or 2.52%.
The lung-cancer risks associated with smoking depend strongly on age
at which one started to smoke, i.e., on duration of exposure (see Figure
C-l). The increase in lung-cancer incidence rate of a smoker at age 60
who started to smoke at age 20 is proportional to 40 ' ; if he had
started at age 15, the extra rate would be proportional to 45 ' .
Starting to smoke 5 yr earlier has thus increased the extra lung-cancer
rate by 70% [(45/40) ], or roughly 14% for each year. To make valid
comparisons between groups of persons exposed to different concentra-
tions of PAH-containing mixtures (e.g., different occupational groups), we
must therefore know their comparative smoking habits, not only in terms of
number of cigarettes smoked per day, but also in terms of age at starting
to smoke.
For a smoker of c cigarettes/d starting at age w and stopping at age
a, the extra lung-cancer incidence rate at age t is
I
c,s
(t,w) =» ac(s - w)4'5. (4)
Equation 4 states that the lung-cancer incidence rate associated with
cigarette-smoking remains constant at the value it had reached when
smoking stopped. »2',40 •££ a person aged 60 who has smoked 30
cigarettes/d from age 20 to 40 (30 pack-yr in total) is compared with a
03
-------
person at the same age (60) who has smoked 15 cigarettes/d from age 20 to
60 (also 30 pack-yr in total), calculations using Equation 4 show that the
latter person will have more than 11 times the lung-cancer incidence rate
of the former. Thus, to understand quantitatively the effect of exposure
to a PAH-containing mixture, one must know not only the total cumulative
exposure, but also the time during which it is accumulated.
Hoffmann e_t a_l.^' pointed out that the major carcinogenic activity
of cigarette smoke, resides in the particulate phase (the tar) and that
there is good experimental evidence that cigarettes with lower tar yields
are less tumorigenic to both hamster larynx and mouse skin. Lower-tar
cigarettes have also been shown to be less tumorigenic to man in all
epidemiologic studies that have investigated this question. Case-control
studies have found that people who smoke filter-tip cigarettes (in effect,
lower-tar cigarettes) have lower lung-cancer incidence rates than smokers
of plain cigarettes at the same frequency, '• and Hammond et al.^
found, in the American Cancer Society (ACS) cohort study, that persons
smoking low-tar cigarettes had lower risk of lung cancer than smokers of
high-tar cigarettes (matched for numbers of cigarettes smoked per day).
Table C-l shows the results of the ACS study: the lung-cancer
mortality ratios are clearly not decreased in men in proportion to tar
content, but they are nearly so in women. The latter finding suggests
that the added lung-cancer risk is close to being simply proportional to
tar content and that the failure to find a proportional reduction in men
arises from the male smokers' having switched from high-tar to low-tar
cigarettes. As Hammond e± al_- stated: "Cigarettes with reduced tar
and nicotine were not introduced until the mid 1950's. . . . Almost all
of the male cigarette smokers and the great majority of the female
cigarette smokers in our study began smoking cigarettes long before that
date. Therefore the subjects classified here as low [tar] cigarette
smokers were, with few exceptions, persons who smoked high [tar] or medium
[tar] cigarettes for many years and then switched to low [tar]
cigarettes." These results substantiate the linear dose-response
assumption of Equation 1.
EXPOSURES TO OTHER SOURCES OF PAH-CONTAINING MIXTURES
Large-scale studies of benzo[a]pyrene in the air of the United States
were conducted between 1958 and 1959 by Sawicki et_ a_l. The range of
BaP concentrations in urban air was from less than 1 to around 60 ng/ra
and the median was roughly 6 ng/m . In contrast. BaP concentrations in
nonurban air were almost always less than 1 ng/ra , with a median of 0.4
ng/ra . BaP concentrations have since decreased: by 1969, the median BaP
concentration in urban air was less than 2 ng/m . However, some
cities were still experiencing average annual BaP concentrations of nearly
10 ng/m3.
BaP is not a perfect indicator of either PAH in the air or its
carcinogenicity, and it accounts for a much smaller fraction of the
carcinogenicity of cigarettes than of air. ' It should be emphasized
C-4
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that BaP ia not a good surrogate for PAHs in mixtures from different
sources, although more information is available on its effects than those
of other PAHs. However, a person who lives where the air contains BaP at
10 ng/m and who breathes 15 m of air per day would breathe in
roughly the same amount of BaP as he would from smoking five old-style
cigarettes (as discussed by Hoffmann e_t a_l. '). It is therefore not
unreasonable to assume that this degree of pollution, which was very
common only 20 yr ago, may cause a significant amount of lung cancer.
Studying the problem directly proves difficult, because one must he
especially careful to ensure that an observed effect is not attributable
to differences in smoking habits between high- and low-pollution areas.
The lung-cancer incidence is affected not only by the number of cigarettes
smoked, but by the tar content of the cigarettes, by how far down the
cigarette is smoked, and by smokers' ages at starting to smoke and at
stopping (if ever); all these aspects of smoking habits have to be
considered. It is impossible to allow for all these factors accurately,
so extrapolating from an extreme situation, in which small smoking-habit
differences can be ignored, is likely to be the best method of estimating
general air-pollution effects. Men employed in some occupations are
exposed intermittently to BaP in air at up to 16,000 ng/m ,* ° and they
provide an opportunity to study lung-cancer effects in an extreme
si tuation.
OCCUPATIONAL EXPOSURE
Many epidemiologic studies of lung cancer have involved occupa-
tional exposure to PAH-containing mixtures. ' ~* They showed that
exposure to high concentrations of PAH-containing mixtures increases the
risk of lung cancer.
Assuming that the exposed and nonexposed workers have the same
smoking habits and that their observed lung-cancer incidence rates are
re and rn, respectively, we can express the lung-cancer burden from
the exposure eithe,r as a ratio, R = re/rn, or as a difference,
D = r - rn. For general risk-assessment purposes, we can express
these on the basis of per-unit exposure by dividing R or D by the
"exposure dose."
Both R and D are valid measures of the risk to the occupational group
as a group, but they implicitly make very different assumptions about the
risks to individual members of the group with different smoking habits.
The relative-risk index, R, implicitly assumes that the risk of lung
cancer is increased in proportion to the individual's "underlying" risk—a
nonsmoker's risk is multiplied by R, and a 2-packs/d smoker's risk is also
multiplied by R. The additional risk of the 2-packs/d smoker is thus an
order of magnitude greater than the additional risk of the nonsmoker and
double the risk of a 1-pack/d smoker. This multiplicative (sometimes
referred to as synergistic) phenomenon appears to hold for lung cancer
caused "jointly" by asbestos exposure and cigarette-smoking.23
C-5
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The additional-risk index, D, implicitly assumes that the amount of
increased risk of lung cancer is independent of other lung-cancer risk—a
nonsmoker's risk is increased by the same absolute amount as a 2-packs/d
smoker's risk.
None of the occupational studies of exposure to PAH-containing
mixtures and lung cancer was conducted in such a way as to provide data to
help in distinguishing between the possible models (i.e., multiplicative,
additive, or something intermediate). Studies comparing urban and rural
lung-cancer rates (or rates in "heavily polluted" and "lightly polluted"
areas) in persons with different smoking habits do provide relevant data,
but the studies generally have few deaths and do not clearly identify the
correct model. Data from the study of Stocks (Table C-2) and from the
1 A
study of Hitosugi (Table C-3) illustrate the point. In both studies,
the data from the smokers are in good agreement with an additive model for
the effect of "air pollution," and these data provide no evidence for a
multiplicative model. The data from nonsmokers, however, confuse the
picture. In Stocks's study, the effect of air pollution is smaller in the
nonsmokers; in Hitosugi"s study, there is no effect in nonsmokers. The
problem may simply result from basing the rates on such small numbers of
deaths in nonsmokers or from misclassifying the smoking habits of a few
persons who died of lung cancer. Because the additive model provides such
a. good fit to the data on smokers, we have assumed this model in our
discussion of lung-cancer risk from occupational exposure to PAH-contain-
ing mixtures.
To use occupational studies for risk-assessment purposes, we must
assume that, as far as lung cancer is concerned, occupational exposures
can be expressed as cigarette equivalents, i.e., that the form of
Equations 1 through 4 will hold for the excess lung cancer from such
exposure. We saw when discussing lung cancer and cigarette-smoking that
dose and duration of exposure are critical in determining lung-cancer
risk. The occupational studies must, at a minimum, provide a quanti-
tative estimate of the dose and duration of exposure to PAH-containing
mixtures. With this information and comparative information on the
smoking habits of the exposed and nonexposed workers, we can estimate the
absolute risk from such exposure. Unfortunately, only one occupational
study with high exposure to a PAH-containing mixture supplied even this
minimal information.
UNITED KINGDOM GASWORKERS
In a prospective study, Doll e_t a_l_. '^ followed a cohort of
carbonization workers in British gasworks for up to 12 yr. Carboni-
zation workers were exposed to BaP at an estimated average air concen-
tration of 3,000 ng/ra during an 8-h shift^O and experienced a 1422
increase in lung-cancer mortality, compared with their nonexpoaed
workmates (Table C-4). Although the smoking habits of only some 10% of
the cohort were ascertained by Doll and his colleagues,^ the exposed and
C-6
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nonexposed workers appear to have had very similar smoking habits, with an
average current consumption of approximately 10 cigarettes/d. It is
reasonable, therefore, to assign the excess lung cancer in the exposed
group to their working conditions, specifically to the air to which they
were exposed.
The current age of a smoker and his age at starting to smoke are both
important in determining his risk of lung cancer. Likewise, both current
age and age at starting as a carhonization worker are important in deter-
mining such a worker's lung-cancer risk. From the papers of Doll e_t
a_l., ' one may estimate the average age of the workers at the middle
year of the study to he approximately 58 yr and the average length of time
exposed to be approximately 23 vr. However, this does not necessarily
imply that their average age at starting such employment was 35 (58 - 23),
because "the men regularly change from one type of work to another."^
If the men started working at age 20, their average worktime BaP
exposure would be
3,000[23/(58 - 20)] = 1,816 ng/ra3.
To express this in constant-exposure terms, we may proceed as follows:
Total BaP-carbonization
breathed per year = (1,816)(9.6)(5)(49) ng
= 4.27 mg.
That is, 9.6 = nr of air breathed at work in a working day—8 h at 20
L/min; 5 = working days in a week; and 49 = working weeks in a year. The
total air breathed in a year is
(15.91)(7)(49) + (12.48)(7)(3) = 5,719 m3.
That is, 15.91 = [(17.28X5) + (12.48X2) ] /7 = average m3 breathed per
day during a working week: 12.48 = m breathed per day during a
nonworking day; 17.28 = m breathed per day during a working day—all
these values calculated with assumptions of 20 L/min at work for 8 h, 6
L/min asleep for 8 h, and 10 L/min otherwise. If the gasworkers' exposure
is expressed in constant-exposure terms, as though the men breathed such
air throughout the day every day, the average BaP-carbonization pollution
to which they were exposed is
4.27 mg/5,719 m3 = 747 ng/m3.
This led to a 142% increase in the rate of lung cancer over "background "
an estimate roughly 90% of which was caused by the men's smoking habits.
If we assume that the relation between duration of exposure and lung-
cancer risk is the same for gasworks exposure as it is for cigarette-
smoking and that the men started work and started to smoke regularly at
roughly the same age, we may write (in lung-cancer terms)
C-7
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10 U.K. cigarettes/d = 0.9
and BaP-carbonization at 747 ng/ra = 1.42.
Those two equations permit us to express BaP-carbonization in terras of
U.K. cigarettes as
BaP-carbonization at 47.3 ng/m3 = 1 U.K. cigarette.
Calculation of the effect of other ages at starting carbonization-
work exposure requires more elaborate computation, and the above esti-
mate appears to be the best that can be made with the limited data
available.30
Note that reasonable changes in the estimate of the proportion of the
background lung-cancer rate that was caused by cigarette-smoking have only
minor effects on this estimated equivalence. For example, if a figure of
807,, rather than 90%, is assumed, the equivalence is BaP-carbonization at
42.1 ng/ra3 = 1 U.K. cigarette.
We assumed in the above calculations that the gasworker breathed 9.6
m3 [(8)(60)(20) L/min] of air at work each working day. The "average"
adult breathes roughly half this amount at work. If we assume further
that gasworkers and the average man breathe similarly at other times, then
the average man breathes 4,543 m of air per year, or 79% (4,543/5,719)
as much aLr as a gasworker. The above equivalent of 47.3 must therefore
be divided by this figure to make the exposure applicable to "average"
man. Our best estimate is thus finally 59.5, i.e., BaP-carbonization at
59.5 ng/ra3 = 1 U.K. cigarette. We estimate from Equation 3 that the
lifetime lung-cancer risk associated with exposure to BaP-carbonization at
1 ng/m3 would be 43/100,000.
UNITED STATES COKE WORKERS
ry -I -3 n
Lloyd and his colleagues^1'J' found in cohort studies of U.S.
steelworkers that coke-oven workers experienced a substantial excess risk
of lung cancer. These workers, like the British gasworkers, are exposed
to the products of coal carbonization. Compared with nonoven workers at
the same plants, the coke-oven workers as a group had 2.8 times the
lung-cancer mortality rate; and coke-oven workers who had raore than 5 yr
of "topside" exposure had 6.9 times the lung-cancer mortality rate. No
data were given on the smoking habits of these workers or of nonexposed
workers, on length of employment, on age, or on average BaP exposure.
However, Jackson e_t al. found average BaP concentrations on the
battery roof of a coke-manufacturing plant of 6,700 ng/m . If this is
taken as the BaP exposure of the topside workers, these estimates of
lung-cancer risk are remarkably compatible with those from the study of
British carbonization workers.
The British carbonization workers had a relative risk of lung cancer
of 2.42 at a BaP exposure of 3,000 ng/m , so we may write
C-8
-------
Nonexposed British lung-cancer rate = 1.0,
Carbonization workers' rate = 2.42,
Increment per 1,000 ng/ra of
BaP-carbonization exposure = (2.42 - 1.0)/3 = 0.47.
At the time of these surveys, the age-adjusted U.S. national lung-cancer
mortality rate was just half the British rate.^2'-^7 Taking into account
this fact and the relative risk of 6.9 for the U.S. topside workers, we
may write
Nonexposed U.S. lung-cancer rate =0.5,
Topside workers' rate = 0.5 x 6.9 = 3.45,
Increment per 1,000 ng/ra^ of
BaP-carbonization exposure = (3.45 - 0.5)/6.7 = 0.44.
The experience of coke-oven workers in the U.S. steel industry is
thus in very close agreement with the British data on gasworkers in
BaP-exposure terms.
LONDON DIESEL-BUS GARAGE WORKERS
The lung-cancer incidence among diesel-bus garage workers employed by
the London Transport Authority (LTA) has been examined for the period
1950-1974. •3»Ji>^l- These men were exposed to more diesel emission than
other LTA employees, but they showed no greater risk of lung cancer than
the other employees.
No detailed information on the garage workers' duration of exposure
to diesel fumes has been published, but the concentration of smoke was
measured inside and outside selected garages. ' Waller^ concluded
that "the indications are that the overall exposure of garage workers to
benzofa]pyrene during their working lives would not differ much from those
of the general population." The BaP exposure of the U.K. gasworkers
discussed above was some 100 times background and was associated with a
142% increase in lung-cancer rates. It is therefore hardly surprising
that the very small increase over background pollution in a diesel garage
(certainly less than a twofold increase) did not produce an
epidemiologically measurable effect. Other possible biases in comparing
the LTA workers in different job categories were discussed at length by
Harris. The study must be considered noninforraative, rather than
negative; we have discussed it here because it was used as an important
data source in recent NRC reports^ ' on the impact of particulate
emission from diesel-powered light-duty vehicles.
OTHER OCCUPATIONALLY EXPOSED GROUPS
Results of other studies of groups of workers exposed to PAH-
containing mixtures were reviewed recently. ' None of these studies
provided evidence of very high exposure; most provided no measure of
C-9
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actual length or intensity of exposure to PAH-containing mixtures or
comparative cigarette-smoking habits. Their results are not useful for
purposes of quantitative risk, assessment.
GENERAL AIR-POLLUTION EXPOSURE
Studies of the effects of exposure to general air pollution have been
reviewed in numerous reports.~f)>> These reviews have found that
lung-cancer rates (as well as rates of cancer at almost all other sites)
are higher in urban (i.e., "polluted") than in rural areas (Santodonato et
a_l., Table 6-47). Interpretation of the increased rates is invariably
confounded, however, by lack of information on the possible contribution
of occupation-induced lung cancer, the possibility of greater accuracy of
death certification in urban areas, and, most critically, the lack of
detailed information on smoking history.
The confounding by occupationally induced lung cancer and more
accurate death certification in urban areas is unlikely to be the
explanation of most of the urban excess. The confounding by lack of
smoking-history information is likely to be the most important. We
have seen (Figure C-l) that the lung-cancer risk among cigarette-smokers
depends strongly on age at starting to smoke, and this holds true even
into old age. For valid comparison of lung-cancer rates between urban and
rural areas, which allows for smoking-habit differences, it is therefore
necessary to know, at a minimum, not only the current smoking habits in
the areas being compared, but also the past smoking habits in these
areas. In most countries, cigarette-smoking became popular much later in
rural than in urban areas; this itself ensures (even allowing for current
smoking habits) that lung-cancer rates will be higher in urban than in
rural areas of such countries.
The above arguments make urban-rural comparisons a very weak basis
for evaluating the effect of general air pollution on lung-cancer rates.
Moreover, most urban-rural comparisons are of no use for quanti-
tative risk-assessment purposes, because they include no estimate of PAH
concentrations in the air in the different areas.
LIVERPOOL-NORTH WALES COMPARISON
The urban-rural comparison study undertaken by Stocks covering
the years 1952-1954 in Liverpool (urban) and parts of North Wales (rural)
is perhaps unique, in that he not only measured air pollution, but also
addressed the issue of long-term smoking habits. The air pollution in the
two areas was measured in terms of average BaP concentration over a 2-yr
period starting in October 1954: the average BaP concentration in the air
was 6.7 ng/ra in the rural area and 59.2 ng/ra in the urban area.
Stocks addressed the issue of long-term smoking habits by showing that, in
men aged 50-59 at the time of the survey in 1953-1955, the urban-rural
contrast in smoking habits did not differ from that of 20 yr earlier.
C-10
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Table G-2 shows Che calculated age-standardized lung-cancer rates by
smoking category in the two areas. The difference in lung-cancer rates
between the two areas, averaged over the smoking categories, is
approximately 74. Assuming that this difference is due totally to general
air pollution, which was mainly the result of inefficient burning of coal,
we may express these rates approximately in terms of Equation 2, with t
taking the value 55, and hence in terms of equivalent U.K. cigarettes.
These calculations estimate the effect of the additional BaP air pollution
in the urban area as the equivalent of 1.09 U.K. cigarettes. Thus, we
estimate
BaP-coal-burning at 52.5 ng/ra = 1.09 U.K. cigarettes
or BaP-coal-burning at 48.2 ng/m = 1 U.K. cigarette.
Therefore, even though Stocks failed to address the issue of lifelong
smoking habits satisfactorily, his data suggest a figure for BaP-coal-
burning that is not much different from BaP-carbonization. If we use only
the data on nonsmokers in Table C-2 to estimate the effect of BaP-coal
burning, we find that
BaP-coal-burning at 128 ng/m3 = 1 U.K. cigarette.
RATES IN NONSMOKERS
'j Q
The study of Stocks has been criticized, because he obtained data
on many of the lung-cancer patients from relatives after the patients'
deaths. This would especially tend to exaggerate the lung-cancer rates in
the "nonsmokers." Doll suggested that a more accurate lung-cancer
figure for nonsraokers could be obtained by combining the data on lifelong
nonsmokers from the prospective studies of Kahn and Hammond1- in the
United States. The combined data (Table C-5) show a lung-cancer mortality
rate for nonsmokers roughly 45% of that found for nonsraokers in rural
North Wales by Stocks. This is the relevant comparison, because the
average BaP concentration in urban air in the United States-^6 in 1959
was roughly 6 ng/m —a figure very close to that of rural North Wales in
1954.
Doll showed that Equation 2 provided an excellent fit to the
combined nonsmoker data from Kahn and Hammond (see Table C-5), and
the best fit is obtained with the equivalent number of U.K. cigarettes
(smoked from birth) set at 0.14. If these lung cancers were due totally
to BaP-U.S. pollution, we could conclude
BaP-U.S. pollution at 6 ng/m3 = 0.14 U.K. cigarette
or BaP-U.S. pollution at 42 ng/m3 - 1 U.K. cigarette.
This may be considered a reasonable upper limit of the potency of
BaP-U.S. pollution in nonsmokers.
C-ll
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REGRESSION STUDIES
A multiple-regression analysis undertaken for the National Research
Council Subcommittee on Particulate Polycyclic Organic Matter^"
attempted to "explain" the annual lung-cancer death rates (per 100,000) in
1950-1969, Y, in the 48 conterminous states of the United States by the
independent variables
X^ = cigarette sales per person over 15 yr old (1963),
in dollars,
and X2 = BaP in air, in ng/ra3 (1967-1969).
A typical result obtained was
Y = 89.4 + 1.44 Xj_ + 7.05 X2
for white men aged 55-64. The observed average lung-cancer mortality rate
for such men for the 48 states was 140»6.
There are a number of major problems with this approach, which are
discussed at length in the report — in particular, the crudity of both the
cigarette-consumption data and the air-pollution figure for a whole
state. The regression equations also predict lung-cancer mortality rates
in the absence of smoking or air pollution that are much greater than the
observed lung-cancer incidence in nonsraokers. For example, Doll gave a
figure of 13.9 (compared with the above figure of 89.4) for the lung-
cancer mortality rate in this age group on the basis of the combined
results of Kahn^' and Hammond.
Other regression studies have similar problems, leaving them useless
for quantitative risk assessment.
COMPARATIVE CARCINOGENICITY OF DIFFERENT AIR-POLLUTION MIXTURES
The available epidemiologic evidence reviewed above suggests that the
carcinogenic potencies of various air-pollution mixtures (coal
carbonization, coal-burning, and general U.S. pollution) are similar when
expressed in terms of the BaP content of the mixtures (Table C-6). We
have no useful epidemiologic data on cases in which the major con-
tributor to air pollution has been mobile sources; to estimate the effects
of such air pollution, we must use the results of aniraal-
carcinogenesis studies and short-term mutagenesis assays.
This approach was used by Harris, Table C-7 shows the assay
results he considered. Tables C-8 and C-9 show the relative potencies of
the various contributors to air pollution computed from the data in Table
C-7. Coke-oven extract is taken as the standard, and the results are
expressed on a constant-weight-of-extract basis in Table C-8 and a
constant-weight-of-BaP basis in Table C-9. For example, with the SENCAR
C-12
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mouse assay, roofing-tar extract is 0.255 (0.535/2.101) times as potent as
coke-oven extract on an equal-weight basis and 0.137 [(0.255)(478/889)]
times as potent as coke-oven extract on a constant-weight-of-BaP basis.
Tables C-6 and C-9 may be used together to predict the lung
carcinogenicity of exposure to spark-ignition or diesel engine exhaust.
Table C-9 suggests that exposure to a fixed amount of BaP from a Mustang
mixture will be between 0.06 and 2.2 times as carcinogenic as such
exposure to coke-oven pollution. The different vehicles tested vary
widely in diesel-exhaust extract. The results shown in Table C-9 suggest
that exposure to a fixed amount of BaP from diesel exhaust will be between
0.1 and 89 times as carcinogenic as such exposure to coke-oven pollution.
If we consider the L5178Y+ assay as the assay of choice, the
predicted lifetime (age 70) lung-cancer risk associated with exposure to
air pollu-ted by a 1-ng/m BaP source for mobile-source emission is given
in Table C-10.
OTHER CANCER SITES
Increased rates of cancer at sites other than lung were observed in
the study of British gasworkers and in the study of U.S. coke-oven
workers.32
In the study of British gasworkers, an excess risk was noted for
cancer of the bladder (age-adjusted rate per 1,000 of 0.37 vs. 0.12
expected), for cancer of the skin and scrotum (0.10 vs. 0.00), and for
cancer at all other sites combined (2.73 vs. 2.27). Because the excess
risk of cancer of the skin and scrotum is extremely unlikely to be due to
inhalation exposure, the maximal excess rate of all cancer except lung
cancer that can be attributed to gasworks exposure is 0.71 (3.10 - 2.39).
The comparable figure for lung cancer is 2.12 (3.61 - 1.49). Lung cancer
therefore accounts for at least 75% (2.12/2.83) of the excess cancer
associated with this British gasworks pollution.
On
Similar calculations from the study of Redmond et al. for men
employed 5 yr or more in the most polluted area (topside) of the U.S. coke
ovens show that lung cancer accounted for at least 83% (17.6/21.1) of the
excess cancer associated with U.S. coke-oven air-pollution exposure.
FOOD
The estimated daily intake of BaP in food is 160-1,600 ng (see Table
6-25). No epidemiologic studies are available to permit one to estimate
the possible carcinogenic effect of such an intake of BaP, and recourse
must be made to animal experiments.
9 ft
The experiment of Neal and Rigdon,' referred to in Chapter 4,
found that BaP administered to mice in their diet produced forestomach
tumors. With the extrapolation procedure used by the National Research
Council Safe Drinking Water Committee, it can be calculated that a
C-13
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daily human intake of 47 ng of BaP would lead Co a lifetime risk of 1 in
100,000. With this estimate, we may calculate that the daily intake of
160-1,600 ng of BaP translates into an estimated lifetime cancer risk of
3.4-34 in 100,000. The estimated daily intake of PAHs ia food is 10 times
the intake of BaP (see Table 6-25), so one would estimate the total
lifetime cancer risk associated with exposure to BaP and other PAHs in
food at something less than 10 times these figures.
C-14
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TABLE C-l
Lung-Cancer Mortality Ratios for Smokers of High-, Medium-,
and Low-"Tar" Cigarettes, 1960-19723
"Tar"
Content, Mortality Ratio
mg/cigarette Males Females
High (30) 1.0 1.0
Medium (22.5) 0.95 0.80
Low (15) 0.81 0.60
aData from Hammond et
C-15
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TABLE C-2
Lung-Cancer Mortality Rates of Men in Rural (North Wales)
and Urban (Liverpool) Areas, 1952-1954, by Past Smoking Habits3
Lung-Cancer Rate"
Smoking Category Rural Urban
Nonsmokers 22 (2) 50 (3)
Cigarette-smokers:
App. 10 cigarettes/d 68 (23) 168 (71)
App. 20 cigarettes/d 147 (36) 248 (140)
App. 35 cigarettes/d 317 (33) 344 (138)
aData from Stocks (p. 80) ,38
'Per 100,000 per year, stai
parentheses are numbers of lung-cancer deaths.
Per 100,000 per year, standardized for age. Figures in
C-16
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TABLE C-3
Lung-Cancer Mortality Rates of Men, Aged 35-74, in Japan,
by Area Pollution and Smoking Habits3
Lung-Cancer Rate**
Low Intermediate
Smoking Category Pollution Pollution
Nonsmokers 11.5 (5) 3.8 (1)
Exsraokers 26.2 (11) 42.6 (7)
Cigarette-smokers :
1-14 cigarettes/d 10.6 (9) 14.2 (10)
15-24 cigarettes/d 14.7 (18) 19.1 (17)
25+ cigarettes/d 36.3 (19) 15.8 (4)
High
Pollution
4.9 (1)
61.7 (7)
23.5 (14)
27.0 (17)
46.4 (9)
aReprinted from National Research Council26 (Table 17-26); data derived
from Hitosugi.16
bPer 100,000 per year, standardized for age. Figures in parentheses are
numbers of lung-cancer deaths.
C-17
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TABLE C-4
Smoking Habits and Lung—Cancer Mortality Rates of
British Gasworkers
Lung-
Non- Ex- Continuing Smokers, Z Cancer
smokers, smokers, Cigarettes/d Mortality
Population Z Z Pipe Mixed 1-9 10-19 20+ Ratea
"Exposed" 8.3 10.2 6.7 4.4 18.1 38.5 13.9 3.61
gasworkers
Other 5.8 15.3 5.9 6.2 17.8 35.5 13.4 1.49
gasworkers
aPer 100,000 per year, standardized for age.
C-18
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TABLE C-5
Lung-Cancer Mortality in U.S. Nonsmokers3
Annual Mortality Rate,
Age, yr per 100,000
35-44 2.8
45-54 5.8
55-64 13.9
65-74 25.6
75-84 49.4
aReprinted with permission from Doll, based
on data from Kahn and Hammond.
C-19
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TABLE C-6
Estimates of Lifetime (70 yr) Lung-Cancer Risk from
Exposure to BaP Source at 1 ng/ra
Study
Population
Gasworkers
Liverpool-
North Wales:
All men
Nonsraokers
Nonsmokers
Lifetime Lung-Cancer
Risk, per 100,000
43
53
20
Reference
Doll et al.
6,10
Stocks
Doll5
38
C-20
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TABLE C-7
Estimates of Potency of Organic Extracts from Various Sources
of Air Pollution3
Source
BaPc
Coke oven
(gasworks)
Roofing tar
Caterpillar
3304 D
Oldsmobile
350 D
Volkswagen
Turbo D
Mustang II
302 V-8,
catalyst
478
889
2
2
26
103
Viral
SENCAR Trans- L5178Ye
Mice0 formation** - +
2.101 0.859 0.726 9.963
0.535 2.066 0.311 9.556
0.011 0.039 0.156 0.049
0.156 0.067 0.270 0.764
0.128 2.545 1.012
0.027 0.204 0.348 0.990
aData from Harris.
^Nanograras of BaP per milligram of extract.
cTumor initiation in SENCAR mice, papilloraas/mouse per milligram of
extract at 27 wk.
^Enhancement of SA7 viral transformation in Syrian hamster embryo cells,
transformations per 2 x 10 cells per nanogram of extract per milliliter.
eL5178Y mouse-lymphoraa mutagenesis assay (average mutant colonies/.lO
survivors per microgram of extract per milliliter) without (-) and with (+)
metabolic activation.
C-21
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TABLE C-8
Estimatas of Potency of Organic Extracts from Various Sources of Air
Pollution, Relative to Potency of Coke-Oven Extract4
Source BaP
Coke oven I
(gasworks )
Roofing tar 1.86
Caterpillar 0.00418
3304 D
OldsraobiU 0.00418
350 D
Volkswagen 0.0544
Turbo D
Mustang II 0.215
302 V-8,
catalyst
SENCAR
Mice
I
0.255
0.00524
0.0743
—
O.OU9
Viral
Trans- L5178Y
formation
I 1
:.4l 0.428
0.0454 0.215
0.0780 0.372
O.U^ 3.51
0.237 0,479
*
1.0
0.959
0.00492
0.0767
0.102
0.0994
sCoke-ov»n r*8ponsa and BaP contant (ng/mg of axtract) set at 1.0.
S«« Table C-7.
('.-22
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TABLE C-9
Estimates of Potency of Organic Extracts from Various Sources of Air
Pollution, in Terms of Fixed BaP Content and Relative to
Potency of Coke-Oven Extract2
Source
Coke oven (gasworks)
Roofing tar
Caterpillar 3304 D
Oldsmobile 350 D
Volkswagen Turbo D
Mustang 302 V-8,
SENCAR
Mice
1
0.137
1.25
17.7
—
0.0596
Viral
Trans-
formation
1
1.29
10.9
18.6
2.74
1.10
L5178Y
-
1
0.230
51.4
88.9
64.4
2.22
+
1
0.516
1.18
18.3
1.87
0.461
catalyst
aSee Tables C-7 and C-8.
C-23
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TABLE C-10
Estimated Cumulative Lung-Cancer Incidence to Age 70
Due to Lifelong Exposure to Various Pollutant
Sources at BaP of 1 ng/m-'3
Source
Coke oven
Caterpillar 3304 D
Oldsraobile 350 D
Volkswagen Turbo D
Mustang 302 V-8, catalyst
Cumulative
Incidence,
per 100,000
43 (0.043%)
51 (0.051%)
787 (0.787%)
80 (0.080%)
20 (0.020%)
aBased on Table C-6 and L5178Y+ in Table C-9.
C-24
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0.1%
3
0.1%
0.0%
2l-3t/d«y
I0-20/d«y
tofora IS IS-lf 20-24 21 or N«mr
owr
A9* (y«»r») when Started to Sack* Clg«r«tc«*
FIGURE C-l. Data on U.S. veterans. * Lung-cancer mortality at ages
55-64 among current smokers of cigarettes only, in relation to the age
when cigarette-smoking began (although this was perhaps not when regular
consumption of substantial numbers of cigarettes began). Reprinted from
Doll and Peto.9
C-25
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C-28
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APPENDIX D
PUBLIC DECISION-MAKING WITH RESPECT TO ATMOSPHERIC
PAH SOURCES AND EMISSIONS
Lawrence J. White
Among the possible justifications for public decision-making with
respect to PAH sources and emissions would be a finding that PAHs pose an
actual or potential (and nontrivial) threat to human health. This appendix
uses the cancer-risk estimates developed in Appendix C. It assumes that
benzo[a]pyrene (BaP) can be used as a proxy for PAHs and that human
exposure to BaP in the ambient air at an average concentration of 1 ng/ra
over an entire lifetime has the effect of increasing by 0.02-0.06% the risk
of dying prematurely (at or before the age of 70) because of lung cancer.
Although the appropriateness of BaP as a surrogate for PAHs in general has
been questioned, it has been so used extensively in the past, and much of
the available information refers to it as an indicator for exposure to
PAHs. The estimates of Appendix C are also based on this application. The
tocus of this appendix is on the lung-cancer consequences of human exposure
to atmospheric sources of PAHs.
The rationale for public decision-making with respect to PAH emissions
from atmospheric sources is explored first, followed by discussions of the
general problems of developing the appropriate decision-making tools,
deciding on appropriate levels of control, and choosing appropriate taeans
of implementing the decisions. The principles developed are then applied
to PAH emissions of various sources, within the constraints of the limited
amount of information that is available. These efforts should be viewed
primarily as illustrative and approximate, because the data available are
rough and approximate. Complete analysis would require a direct linking of
the damage caused by an air pollutant to the sources of its emission. For
that, the following would be needed: data on emissions of the pollutant, a
model of the pollutant's dispersion and possible transformation or decay
during dispersion, estimates of the resulting concentrations in the ambient
air, data on human exposure to those concentrations, and a model of the
exposure dose-response relationship. Reliable estimates of the costs and
consequences of control are also needed. With respect to all these
subjects, the relevant data on PAHs are scanty and approximate, and
compromises will have to be made. Some estimates may be in error by as
much as an order of magnitude. Nevertheless, the results should be
informative and point the way toward further appropriate study.
RATIONALE
PAH emissions from atmospheric sources are in a category of phenomena
that economists have labeled "negative externalities" or "negative
spillovers." The designations imply that people are taking actions (e.g.,
producing coke, driving vehicles, and burning refuse) that generate, as
byproducts or as incidental consequences, uncompensated costs imposed on
D-l
-------
other parties, outside of a market context; i.e., the PAH emissions pro-
duced incidentally by these activities ultimately have potentially un-
favorable health consequences for others. In such situations, persons who
are motivated largely by the prospect of private gain (or, in the case of
firms, private profit) are unlikely to take corrective action. Without
incentives for corrective action, too much of the activity will occur, and
too little effort will be devoted to reducing the costs imposed on others.
An externality is an indication of a market failure;3 i.e., even an
otherwise properly functioning competitive economy will not achieve an
optimal allocation of society's resources, because of the distortion
introduced by the externality. In a private-enterprise economy, the source
of the problem created by an externality can be traced to an ill-defined
property right (neither the emitters of PAHs nor those who are exposed
have a well-defined property right to the ambient air and its cleanness) or
to the difficulties of enforcing a property right. The latter difficulties
are usually due to the "public-goods" aspects of the phenomena; e.g.,
because an improvement in air quality in a locality will be enjoyed by all,
each individual has an incentive to let others make the necessary effort to
enforce emissions reductions, and this incentive for "free riding" leads to
too little (or no) action.
Externalities (especially those involving public-goods aspects) provide
a case for possible public intervention in a private-enterprise economy.
But whether, in practice, government intervention to correct an externality
increases or decreases societal welfare is an empirical question.
LEVELS OF CONTROL
Once an externality has been identified and the decision has been made
that some kind of corrective action is warranted, further decisions must be
made on the extent of corrective action (e.g., the desired degree of
reduction in PAH emissions or the amounts of PAHs that will still be
allowed to be emitted) and on the specific tools that are to be used to
implement the desired level of control. This section addresses the former
issue, leaving the latter for the next section.
The control of an externality brings societal benefits: a reduction in
the externality costs imposed on others. In the case of PAHs, reductions
in PAH emissions that translate into reductions in human exposure to PAHs
mean the avoidance of some premature deaths (frequently termed "the saving
of lives") and the avoidance of PAH-induced illness. But the achievement
of these benefits almost always involves societal costs: individuals and
firms must be induced to change their behavior with respect to emissions,
engage in less of their desired activities, and incur costs (use real
resources) to reduce emissions.
Society's resources are scarce—in essence, society does not have
limitless resources and cannot achieve all its desired objectives
simultaneously, but must choose among them—and any level of externality
0-2
-------
control involves both societal benefits and societal costs; therefore,
decisions concerning levels of control should focus on levels that best use
society's scarce resources in trying to maximize societal welfare—i.e.,
society ought to aim for levels of control that provide the greatest margin
of benefits relative to costs.
Two main analytic tools have been developed that can aid decision-
makers in choosing the appropriate levels of control: cost-effectiveness
analysis and cost-benefit analysis. Cost-effectiveness analysis is the
more limited of the two. It takes, as a given, a specific societal goal
(objective)—e.g., a reduction in emissions by X tons of a specific
pollutant or the incurring of only up to Y dollars for the reduction of
emissions from a specific source of that pollutant. The principle of
cost-effectiveness requires a search to identify the least costly way of
achieving a reduction in pollutant emissions. If all sources of the
pollutant have equal environmental consequences, then the emission source
with the lowest marginal (incremental) cost of control should be chosen.
For example, if one source has a marginal control cost of $500/ton and
another a marginal cost of $3,000/ton, the first should be chosen over the
second. The choice of the first will mean that the achievement of emission
reduction by X tons will require less resources, or the expenditure of Y
dollars will achieve a greater reduction. The formal principle is that, Ln
achieving the goal, the marginal costs of control from all sources ought to
be equated. If this principle is violated, then the cost of achieving a
given level of overall control could be reduced (or the level of overall
control achieved at given costs could be increased) by increasing the
stringency of control from the low-marginal-cost sources and decreasing the
stringency of control from the high-marginal-cost sources.
Cost-effectiveness analysis can be a useful tool for improving the
efficiency of individual programs and for comparing the effectiveness of
similar programs. But cost-effectiveness analysis cannot be used to answer
the ultimate policy questions: "Should X tons or 10X tons of
pollutant-emission reduction be the appropriate societal goal?" "Should a
cost of Y dollars or 20Y dollars be incurred to achieve emission
reduction?" But cost-benefit analysis can provide an analytic basis for
making these decisions.
There are only a few primary steps in a cost-benefit analysis. The
societal benefits and societal costs should be estimated and converted into
dollar equivalents (if they are not already in dollars). An interest rate
(discount rate) must be used to convert future benefits and costs into
present-value equivalents. The projects (or alternative versions of a
project, e.g., alternative levels of stringency of required emission
reductions) with the highest margins of benefits relative to costs should
be the ones chosen. An equivalent principle is that, in choosing among
alternative versions of a project (say, alternative levels of emission
control stringency), stringency should be adjusted until the marginal
benefits of extra stringency are just equal to the marginal costSt The
basic methods of cost-benefit analysis are, by now, standard; >
D-3
-------
controversies remain, however, as to the interest rate that should be used
for discounting, whether the income-distribution consequences of
projects should be considered explicitly in the analysis, how to
incorporate risk and uncertainty into the analysis, and how (and whether)
to place dollar values on nonmarket items and concepts.
In this last category, a frequent question that arises in the context
of cost-benefit analysis applied to projects or programs that have
mortality or morbidity consequences (e.g., many pollutant-emission control
programs) is how (and whether) to evaluate the benefits of mortality or
morbidity reduction. Claims that "a life is priceless" and that "one
cannot put a value on a life or on pain and suffering" are often heard. A
logical implication of these claims seems to be that cost-benefit analysis
is useless in such instances—that for such projects, so long as any
mortality reduction ("lives to be saved") or morbidity reduction (reduction
in "pain and suffering") can be achieved, a project or program should be
pursued (or extra stringency pursued), regardless of costs.
This approach to the benefits of reductions in mortality or morbidity
does not provide a useful guide for making societal decisions, because the
opportunities for achieving reductions in mortality and morbidity are
virtually limitless. Additional resources devoted to medical research,
medical care, accident prevention, and pollution reduction are likely to
yield reductions (albeit possibly small) in mortality and morbidity.
Society could use up its entire gross national product by devoting
ever-increasing amounts of resources to the pursuit of such reductions.
But, Ln fact, we do not. Through our societal decision-making processes,
at some point we desist. For example, in the wake of the Arab oil embargo
of 1973, the Congress enacted a law imposing a national highway speed limit
of 55 mph. The major goal of the legislation was to reduce American
gasoline consumption, but it was soon learned that the 55-mph speed limit
had the beneficial side effect of reducing highway mortality. There have
been no efforts to reduce the speed limit to, say, 45 mph, although such a
reduction would clearly reduce highway mortality even more. Similarly,
society does not build pedestrian underpasses for every busy urban
intersection and does not station ambulances near those intersections,
despite the reductions in mortality and morbidity that would be achieved.
In effect, society has decided that the extra mortality and morbidity
reductions are not worth the resources (costs) that would have to be
devoted to achieving them; lines have been drawn.
Drawing these lines has been a largely implicit process; drawing them
explicitly apparently makes many people uneasy. They are reluctant to put
a value on mortality or morbidity reductions. But a society that wishes to
achieve the best that it can from its scarce resources must understand the
uses to which those resources are put and the tradeoffs (the "opportunity
costs") involved. A society may well have multiple goals. Nevertheless,
an understanding of the tradeoffs is important in pursuing them; and the
use of explicit values for mortality and morbidity reductions is necessary
for that understanding. Furthermore, the logic of cost-effectiveness
D-4
-------
argues for the consistency of these values across projects; otherwise,
societal resources are allocated in an ineffective way, as apparently has
been the case for actual projects and programs involving mortality and
morbidity reductions.
A good case can be made, then, for using explicit values for mortality
and morbidity reductions. There are a number of candidates for
establishing the value of mortality reduction (or, alternatively, "the
value of a life"):
• The expected discounted future earnings of a person.
• The life insurance held by a person.
• The average (or some other summary measure) of the implicit
values yielded by other, recent projects or programs involving mortality
reduction.
• Compensation awarded in trials involving premature deaths.
• Estimates of the value that people, in their day-to-day
behavior, place on incurring or avoiding risks of premature death.
For the purposes of deciding on the appropriate levels of pollution
control, Bailey and Freeman1 (Chapter 4) have reviewed and criticized
these measures. The last measure (risk valuation) is most consistent with
the market valuations that are the other components of cost-effectiveness
and cost-benefit analyses. An important point here is that pollution-
reduction programs (and accident-reduction programs) do not have a knowable
effect on specific persons' lives; they do not involve before-the-fact
specific deaths. Instead, if they are effective at all, they reduce the
probabilities or risks of the premature death of exposed persons. After
the fact, this reduction in risk must mean a reduction in premature deaths;
but before the fact, the programs can be evaluated only in terms of risk.
Because the affected persons benefit from the reduction in risk and
because virtually all people expose themselves to risks in their day-to-day
behavior (whether they acknowledge it or not), the benefit of the risk
reduction should be roughly comparable with the value of the risks that
they incur or avoid (at the margin) in their day-to-day behavior. In
essence, if they are asked, "What would you be willing to pay in return for
a reduction in risk?" or "What would you need to receive to compensate you
for an increase in risk?" their responses should be roughly consistent with
their private behavior. In a market economy, the prices of goods and
services reflect (at the margin) a willingness to pay for those goods and
services. Public projects, to maximize the societal value that can be
achieved from society's resources, should also use willingness-to-pay
measures for valuation purposes wherever possible. Accordingly, the
risk-valuation approach is consistent for assessing pollution-reduction
programs.
There are no specific markets in the private sector where one could
directly observe a person's willingness to pay for risk reduction. But
people do choose to incur or avoid risk, gaining or giving up other things
D-5
-------
in return, in most aspects of their lives: They choose jobs that have
higher or lower risks of accidental death or injury, in return for explicit
or indirect wage premiums; they choose to use or not to use seatbelts
in automobiles, trading off time and convenience against reduced risk of
death or injury in the event of a crash;4 they choose to live in
neighborhoods with higher or lower air-pollutant concentrations, trading
off housing costs against the extra risks of mortality or morbidity from
the pollutants; and so on. Economists have been able to provide models
of individual behavior and, with actual data and econometric estimation
techniques, estimate the implicit value that people have placed on the
risks that they have incurred or avoided. For example, to estimate the
wage premium that accompanies extra risk, a researcher could collect a
sample of wage rates for various occupations, the actuarial data on
accidental deaths for those occupations, and data on the various influences
on wage rates (e.g., degree of unionization, amount of education, extent of
experience). The econometric techniques allow the researcher to control
for the other influences and thus to infer the implicit wage premium that
accompanies extra risk.
Placing a value on reducing the risk of death is very difficult and
controversial. Different values can be assigned. However, the values
discussed here fairly represent the research that has been done in this
field, and they provide a useful reference guide for decision-making with
regard to pollution control.
Studies of the value of risk do not yield identical estimates, but, as
Bailey showed, they can be grouped (after appropriate adjustments and
corrections) into a range of $170-715 (in 1978 dollars) in annual payment
per 0.001 (i.e., 0.1%) additional annual risk of death. A study by
Portney yielded an additional estimate that is in the middle of this
range. Freeman argued that the most likely value is $1,000 (in 1978
dollars) per 0.001 additional risk. This same figure was used in the NRC
study (pp. 244-245) of the costs of removing chloroform and other
no
trihalomethanes from drinking water. °
Some problems of using these studies and the estimates they yield for
evaluating public pollution-control programs should be noted. First, as
with the use of any econometric model, one needs to be satisfied that the
model has been properly specified and all important influences properly
accounted for. Second, the models assume that the persons involved were
aware of the risks they were incurring or avoiding. Third, use of the
models' estimates for public-policy purposes assumes that the persons in
the sample (and hence the estimates of the value of risk) are typical of
the general population. If a wage study included only or mostly high-risk
occupations, the resulting estimate of the value of risk might be an
underestimate of the value that applies to most of the population, since
persons with less fear of risk would likely gravitate toward high-risk
occupations or housing locations — i.e., self-selection might bias the
results. Fourth, people may feel differently about (value differently)
risks over which they have more control (e.g., job choice) and risks over
0-6
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which they have less control (e.g., the general Level o€ pollution in the
air they breathe). Finally, even if the models' estimates are representa-
tive of the general population's valuation of risk, individual persons will
have different values of risk and hence different perceptions of, say, the
concentrations for which a pollution-control prograa should aiau Within a
locality, however, all persons will have to be exposed to roughly the sasae
pollutant concentrations.
The last problea is an unresolvable dileaaa that is inherent in the
public—goods aspects of aaost pollution problems, which cause ttaea to be the
proper concern of nonindividualistic, government action in the first
place. This dilesama is present for ail public goods (e.g., national
iefense and local police protection) that people consume generally
automatically and equally as part of a community. As Sawielson31 has
deaonstrated, the proper procedure for deciding on the appropriate level of
a public-goods project is to saa the valuations of all affected persons aad
extend the project to the point at which the sum of the saarginal valuations
(benefits) equals the Marginal cost of the extension—exactly the criterion
stated ia the discussion of cost-benefit analysis.
Respite the possible probleras, the range of estiaates yielded by the
risk-valuation studies does appear to be reasonable whea compared with the
incone of a typical family and the safety-related expenditures it would
find worthwhile.~
One aspect of the risk-valuation estimates is worth emphasizing. If
one finds that people appear to be willing to pay §500 per year each to
avoid a 0.001 risk of death in a given year, the proper ase of this
estimate is as follows: Suppose a government pollution-control prograa can
reduce the risk of death in a coanunity of 1 million by a factor of 0.001.
Then, because each person, on average, should be willing to value this
improvement at about §500 per year, the 1 million people in the cosraamity
should be willing to pay about $500 aillion per year for these benefits,
and this aggregate value couli ?>e coaparsd with the anticipated cost of the
prograa. In essence, the aggregate cost of the benefit is estimated by
aultiplying the typical person's valuation of the risk reduction by the
naaber of persons involved (reduction in risk per person).
In contrast, the value of risk is soiaetises extrapolated to a value af
avoiding (or, in reality, delaying) a death or "the value of [extending] a
life"; i.e., the §500 per 0.001 risk would be extrapolated to 1500,Odd as
the value of avoiding a death. It is true that, if the goversssenc
implements the hypothetical prograa jusc aentioned, there will i>e 1,000
fewer deaths per year; and, because the prograa was valued at 5500 aillion
per year, this iaplies a value of $500,000 per avoided death. Further-
more, for soae purposes, it is sometimes convenient to speak or write in
terms of "the value of a life* (or the value of a statistical life). Sat
there is nothing in the statistical or conceptual procedures that leads to
the conclusion that any person would, could, or should pay §500,000 to
avoid a certain death. Rather, before the fact, the goverajaent project
-------
promises a change in risk, not a change in Che certainty of death for any
person. People behave toward and implicitly value risk in their everyday
life, so risk valuation is the consistent conceptual procedure to use.
The discussion thus far has focused entirely on valuing mortality
changes. In principle, the same procedures could be applied to valuing
changes in morbidity—i.e, willingness-to-pay measures could be inferred
from persons' behavior. There do not appear to be any studies that have
tried to generate such estimates. Instead, estimates of the medical costs
and lost productivity related to illness and accidents are usually used to
estimate these societal costs (and hence the societal benefits from their
reduction). These estimates may not be too far away from what the
appropriate willingness-to-pay measures, if they existed, would indicate,
except that the former probably underestimate the latter by excluding the
value of avoiding pain and suffering.
Finally, the limitations of cost-effectiveness and cost-benefit
analysis must be acknowledged. Knowledge about costs and benefits is never
perfect; in some cases, it may be quite imperfect. Risks and uncertainties
often pervade analyses. Society has multiple goals. But, in the end,
society's resources have to be allocated, and those resources are scarce
and have alternative uses. Cost-effectiveness and cost-benefit analysis,
imperfect though they may be, can be aids to effective societal decision-
making.
IMPLEMENTATION
Regardless of the target level of control desired, a number of choices
with respect to the implementation of an emission-control program are
possible. A useful dichotomy is provided by the division between fiat
methods (frequently called "command and control") and methods that rely on
the use of economic incentives.
At one extreme, after a desired reduction in emissions (or a desired
level of remaining emissions) has been ascertained, a central regulatory
control agency can attempt to specify to each emitter (or class of
emitters) the reduction or allowable emissions that will be required. If
the agency wished to minimize the societal cost of achieving the emission
reduction, it would try to have complete information about the total and
marginal cost schedules for each of the various emitters and allocate
reduction or emission appropriately, following the precepts of cost-
effectiveness analysis.
At the other extreme, the agency could set an effluent fee that would
require an emitter to pay a specified amount per unit of the pollutant that
was emitted. In the presence of rising marginal costs of control, emitters
would find it worthwhile to reduce emissions to the point at which the
marginal cost per unit of pollutant reduction was equal to the effluent
fee. The same knowledge of cost schedules assumed above would allow the
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agency to set an effluent fee that would achieve the same reductions as
those achieved by the fiat method.
As the previous paragraphs indicated, under conditions of complete
certainty, the two methods can achieve the same outcome. But knowledge
about the costs of control is rarely complete. With incomplete knowledge,
the^control agency is likely to make socially costly mistakes by improperly
assigning excessive emission reductions to emitters with high marginal
costs of control. The effluent-fee system has an important advantage Ln
this respect, in that it allows the high-cost and low-cost emitters to sort
themselves out and achieve che lowest overall cost of control through their
own behavior. Incomplete knowledge of costs may also lead to effluent-fee
schedules chat are too high or too low, with consequent emission reductions
that are off target. But the schedules can be readjusted by continuing to
observe emissions; incorrect assignments under the fiat method may never be
corrected, because correct cost information is not automatically revealed.
An alleged advantage of the fiat method is its apparent certainty of
outcome. Emitters will be told to reduce their emissions by a specified
amount, and that reduction "will" be achieved. The effluent-fee method
appears to be more indirect; one has to rely on the cost-reduction
consciousness of firms and individuals to recognize that reducing emissions
(up to a point) is less costly than paying effluent fees. But experience
with pollution-control programs has shown that even the expected certainty
of the fiat method often does not materialize. ' Many emission-
control programs are intended to be "technology-forcing"; they try to set
emission standards that are beyond the economical range of current
technology, thus attempting to force the development of advanced
technology. The ostensible sanctions for failure to meet emission
standards are usually severe fines or closure of offending companies. But
if the technology appears not to be available, the sanctions are not
credible or enforceable. Furthermore, regulators may have difficulty in
ascertaining whether the necessary technology is or is not available or
economical or whether a good-faith effort has been made to develop the
needed technology.
As a consequence of these uncertainties, the emitters (especially in an
industry with a relatively small number of large firms) have an incentive
to slow down their own technology development. Thus, the apparent
certainty of success of the fiat programs is not necessarily reflected in
actual practice, as the delays in the implementation of many
pollution-control programs have revealed.
Even if the sanctions behind them are thought to be credible, fiat
methods can lead to the development of inefficient techniques.
Technologies that are low in cost but that may fall short of the standards
are unlikely to be pursued; technologies that can, at low cost, reduce
emissions beyond the point set by the standards will be pursued only to the
point set by the standards; technologies that are expected to be low in
cost but have an uncertain likelihood of probability of success will be
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discarded in favor of high-coat, more certain technologies. An effluent-
fee system would not have these inefficiency properties.
Another method of control that retains most of the incentive properties
of effluent fees, but also has some of the possible quantitative certainty
of a fiat system, is a system of marketable emission permits. Under
this system, the central regulatory agency sets a target of maximal total
emissions of a given pollutant. It then creates a set of permits equal to
this total. The permits are, in essence, a property right in a given
amount of emissions. No one is allowed to emit without a permit; thus,
each emitter must control emissions down to the point for which permits
have been received. The agency could auction off the permits to the
highest bidder (thus lodging the property right in clean air initially with
the government), or it could initially assign the permits among emitters,
or even among the population generally, in some manner (thus initially
assigning the property rights in the manner chosen). If the permits are
auctioned or can be traded, emitters will again sort themselves into an
efficient, least-cost pattern, with low-cost emitters choosing to control
emissions more and buying relatively fewer permits and high-cost emitters
doing the opposite.
It is clear that, with appropriately chosen targets (costs and
emissions), an effluent-fee system and a marketable-permit system can
achieve the same outcome with comparable incentive effects. One difference
between them is that the effluent-fee system always implicitly lodges the
property right with the government, whereas the marketable-permit system
may lodge the property right with the government (if the auction method is
used) or in the private sector (if some assignment scheme is used).
Another difference is in the identity of the group that bears the risk in
the event of uncertainty about or variation in emitters' marginal-cost
schedules. In an effluent-fee system, variation in marginal-cost
schedules will mean that variation can be expected in the,quantities of
emissions; thus, the risk is borne by those who are exposed to the
emissions. In a marketable-permit system, variation in marginal-cost
schedules will mean variation in the prices paid for the permits; the risk
is borne by the emitters. The choice between the two systems on these
grounds should be determined by examining the societal costs of lodging the
risk with one group or the other. If, for example, the health consequences
of small variations in emissions could be severe, a marketable-permit
scheme would be preferred; if, however, the health consequences of small
variations in emissions are not severe and the price variance of permits
would cause firms to take relatively costly offsetting actions, the
effluent-fee system would be preferred.
Even within the context of a fiat system, there are measures that
increase the scope of economic incentives and efficiency. For stationary-
source emissions, a "bubble" strategy that allows individual firms to trade
off pollutant emission from different sources (e.g., different smokestacks)
at the same geographic location provides the possibility of reducing the
cost of controlling emission by a given amount. '" In essence, an
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individual firm can exchange emission permits for its emission sources
within the firm at the same location; this is a half-way step to a full
marketable-permit system, which would allow firms to trade permits among
different firms. Similarly, for motor vehicles, a fleet-averaging policy
that would allow a manufacturer to satisfy emission standards if the
sales-weighted average of its vehicles were at or below the standard,
rather than every vehicle's being required to meet the standard, would
allow the manufacturer to trade off low-cost ways (e.g., smaller vehicles)
of meeting the standard against high-cost ways (e.g., larger vehicles)
(White,'4 Chapter 7). Allowing vehicle manufacturers to trade (or even
to "bank" for future use) any margin between allowed and actual emissions
would convert a fleet-averaging scheme to a form of marketable-permit
scheme. It should be noted, however, that fleet averaging, even with
trading, is not identical with the standard marketable-permit scheme. The
latter sets a limit on the overall amount of emissions, whereas the former
sets a limit on the average per vehicle, but does not set a limit on the
number of vehicles that can be sold. There are some circumstances in which
a tightening of the standards in a fleet-averaging scheme could lead,
perversely, to an increase in total emission. '
Overall, if regulatory schemes to control PAHs are put into effect, it
appears desirable that the implementation tools chosen emphasize economic
incentives and efficiency, regardless of the control levels that are
selected as targets.
OBTAINING A BENCHMARK
A convenient way to start is to try to determine a societal value to
place on a reduction by 1 ton/yr in PAH emissions from at least one
important category of sources. If this benchmark figure can be estab-
lished, the assessment of the likely costs and benefits of controlling PAHs
from other categories of sources will be easier. The data from the
analyses of Chapters 1, 2, and 3 and Appendix C, plus the risk valuation
discussed earlier in this appendix, provide the basis for such a benchmark
calculation. Again, we use BaP as a representative of PAHs. Thus,
although emissions and concentrations are expressed in terras of BaP, they
really represent a far larger "soup" of PAHs for which BaP is, in essence,
the "tracer," or surrogate. Differences in particle size or other factors
that might affect respirability or bLoavailability are largely ignored.
Linearity is assumed in most models.
The information in Chapter 1 shows that in 1979 the amount of BaP
emitted into the atmosphere from urban road motor vehicles was sufficient
to cause an urban commuter to inhale a calculated dose of up to 20.1 ng of
BaP in the course of 24 h. By the year 2000, it is estimated that the same
commuter would inhale only 9.1 ng/d. These estimates are based on
inhalation rates of 15 m3/d. Thus, for 1979, the average concentration
of BaP from motor-vehicle emission in the air breathed by the "worst-case"
person was 1.34 ng/m3 —i.e., (20.1 ng/d)/(15 m3/d)—and for 2000 it
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would be 0.61 ng/m^. In the first year, 13.46 metric tons of BaP was
estimated to be emitted by motor vehicles, and in 2000, 10.14 metric tons.
If we adopt a rough linear model relating the gross BaP emissions per year
to average concentration, we estimate that 1 ton =0.1 ng/m for the
first year and 1 ton = 0.06 ng/m for the second year. In later
calculations, the first figure is used as a conservative estimate. Thus,
it is assumed that a reduction of 1 ton of BaP emissions per year from
motor vehicles is likely to reduce average urban BaP ambient concentra-
tions by 0.1 ng/m for that same year.
Next, this change in BaP concentration should be converted to an
equivalent mortality risk. In Appendix C, it is estimated that exposure to
BaP at an average of 1 ng/m for an entire lifetime leads to a cumulative
excess risk of lung cancer by the age of 70 of 0.02-0.06*. Again, the
latter figure is used as a conservative estimate. Thus, if it is also
assumed that a ton of BaP (representing a larger quantity of PAHs) from
motor vehicles has the same health consequences as a ton of BaP (also
representing a larger quantity of PAHs) from another atmospheric source,
then breathing BaP from motor vehicles at an average of 0.1 ng/m would
have a cumulative excess risk of 0.006%, or 0.6 x 10~ . This would be
the same risk generated by 1 ton of BaP emission per year for 70 yr. But
it is necessary to determine the risk generated by 1 ton of BaP in 1 yr.
As an approximation, this extra risk can be "smeared" equally over all 70
yr. Thus, the extra risk of premature death per year is 0.00009%, or a
0.9 x 10~° probability of a premature death in each year. (Although the
original 0.006? is a cumulative risk to age 70, with the risk of premature
death in each year rising exponentially, the absolute numbers are small
enough so that "smearing" equally makes little difference in the results.)
Finally, a value can be put on this probability. The range of the
annual value of avoiding a 0.001 annual probability of a premature death,
reported above, was $170-1,000 (in 1978 dollars). To be conservative, the
upper limit will be used and translated into a 1982 dollar figure of about
$1,500 per 0.001 risk. This figure, then, indicates that the reduction in
annual risk of 0.9 x 10 would be worth about $1.35/person. It should
be recalled that the atmospheric-concentration data apply to urban areas.
Approximately 75? of the U.S. population of 225 million live in urban
areis, or about 170 million. Accordingly, these calculations indicate that
the reduction in BaP emissions by 1 ton in 1 yr. which would lower urban
ambient concentrations of BaP by about 0.1 ng/ra and thus lower the
annual risk of death per person exposed by 0.9 x 10 , would be worth
about $225 million in the aggregate, or $225,000/kg, $225/g, or $0.22/mg of
BaP.
It should be noted again that BaP is being used here as a tracer to
represent a larger collection of PAHs and that the $0.22/mg of BaP really
represents the value of controlling this larger "soup" of PAHs that has a
potency that can be measured and represented by 1 mg of BaP. Also, the
risk valuation applies only to the lung-cancer consequences of exposure to
PAHs; other possible mortality and morbidity consequences of exposure have
been ignored.
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Furthermore, it should be emphasized that each of the key components of
the value estimate is an estimate that has a substantial range of
uncertainty. The risk of death associated with breathing BaP at a given
concentration has an uncertainty range of approximately a factor of 3; the
societal value of avoiding a premature death has a range of approximately a
factor of 6; and the likely atmospheric concentration from a ton of BaP has
a range of 1.5. Because these estimates are used multiplicatively, the
overall range of uncertainty on the final value estimate is approximately a
factor of 25. For the present analysis, in each case the most conservative
estimate of each component — the figure that would indicate the greatest
societal benefit from controlling PAHs—was used. Alternative methods
would have been to use the most likely value for each component and to
carry the range throughout. But information for choosing the most likely
values is not available; and, as noted, carrying the range throughout leads
to an uncertainty range of a factor of 25 downward from the estimate of a
societal value of $225 million per ton of BaP removed. Thus, at the other
end of the range, those who prefer to be less conservative could use a
value as low as $9 million per ton of BaP removed. In matters of public
decision-making concerning the societal value of actions that involve
avoiding premature deaths—a highly controversial subject—a conservative
approach seems warranted. Accordingly, the figure of $225 per ton is used
for the remainder of this analysis.
CONTROL OF PAH EMISSIONS FROM VARIOUS SOURCES
Although PAHs are the product of virtually every burning process, it
makes sense to focus on the quantitatively important sources. Chapters 1-4
and other studies indicate that the following sources are important
(not necessarily in order of quantitative importance):
• Road motor vehicles.
Other mobile sources (e.g., trains, planes, and ships).
Fireplaces.
Wood-burning stoves.
Residential coal-fired heating.
Industrial coal-fired boilers.
Coke production.
Industrial-commercial incinerators.
Agricultural open burning.
Land-clearing waste burning.
Prescribed burning of underbrush in forests.
Forest and prairie fires.
Structural fires.
Coal-refuse fires.
Volcanoes.
The control of PAH emissions from some of these sources can be ruled
out, because they cannot be controlled (such as volcanoes). In principle,
reductions in PAH emissions could probably be achieved by applying more
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resources to concrolling forest fires, structural fires, and coal-refuse
fires (largely in abandoned coal mines). But the other societal costs from
these sources probably bulk so large in comparison with their PAH-related
costs that greatly increased efforts to combat these fires could not be
justified solely on the basis of their PAH emissions.
Other sources may offer some promise of worthwhile control. The
discussion here begins with road motor vehicles, because the data on them
are best, and then examines stationary sources. In each case, only the
reductions in PAH emissions are valued and compared with the costs of the
reductions. In many instances, the effort to reduce PAH emissions will
reduce other harmful pollutants as well (e.g., particulates in general).
The reductions in these other pollutants may have additional societal
value; but that value is not calculated or considered here. Also, in some
instances, efforts to reduce PAH emissions may increase the emissions of
other pollutants. These additional effects are ignored. Thus, the
discussion here focuses on whether reductions in PAH emissions, valued
alone, justify (or come reasonably close to justifying) control efforts.
ROAD MOTOR VEHICLES
As noted earlier, gasoline-powered vehicles without exhaust catalytic
converters (i.e., all pre-1975 cars and light-duty trucks and all
heavy-duty trucks and buses of any vintage) and diesel cars, trucks, and
buses constitute the major sources of PAH emission from road motor
vehicles. An additional category of "problem" vehicles would include cars
and light-duty trucks of the 1975 and later model years that have emission
control systems that are no longer functioning properly. The categories of
gasoline and diesel vehicles are addressed separately. Unless otherwise
indicated, urban-rural driving distinctions are ignored, and emission
reductions in rural areas are valued at the same rate as urban reductions.
Gasoline Vehicles
At the beginning, it is useful to establish a relationship between
total hydrocarbon (HC) emission per mile and BaP emissions per mile for
gasoline-powered vehicles. The data in Chapters 1-3 and in Williams and
Swarin indicate an approximate relationship—HC at 1 g/mi * BaP at 2
ug/mi—that is used in the discussion that follows.
Once motor vehicles are manufactured and on the road, there are three
major ways to reduce emissions (including that of PAHs) from them:
retrofitting them with further controls, inducing better maintenance and
slower deterioration of their control systems, and inducing owners to junk
them in favor of newer vehicles.
Retrofitting does not seem to be a practical method; it would probably
achieve only modest emission reductions, and it is quite unpopular. Only
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one state (California) has a program for requiring retrofitting of older
cars, and that program applies only when cars change owners.
Gruenspecht *-° has analyzed the prospect of providing subsidies to
owners of older cars to junk them in favor of newer ones and finds it a
worthwhile strategy, compared with the costs of the tighter standards
imposed for the 1981 and later model cars. The societal benefits from PAH
reductions, not included by Gruenspecht, would add to his results. An
older car that emitted HC at 4 g/mi (and hence BaP at 8 ug/mi) more than a
new car and that was expected to last for another 40,000 mi of operation
would emit 320 mg of "extra" BaP during this period. The risk-valuation
calculations of this chapter have shown that a reduction of this amount of
BaP would be worth $70 to society. Thus, the bonus or subsidy paid to
owners of old cars to induce them to junk the vehicles could be increased
by this amount, to induce yet more turnover of the fleet.
Better maintenance of emission control systems can be induced by
inspection and maintenance (I&M) programs by states and locales. White^
examined these programs and concluded that they could be worthwhile under
some circumstances, especially if linked to safety-inspection programs.
The societal value of the reduction in emissions of all pollutants achieved
by such programs was estimated to be $23/vehicle, with $5 of this coming
from the 5 kg of HC reduction per year per vehicle that would be achieved.
The values were based on the comparative costs of achieving the equivalent
reductions in emissions from other sources. To the extent that vehicle HC
emissions contained appreciably more BaP than the HC emissions from other
sources, this might raise the societal value of the reduction. The limit
of this increase would be $2/vehicle [(10 mg)($0.22) » $2]. This figure is
well within the margin of error of the original calculations and hence does
not appear to make I&M programs appreciably more attractive than they
otherwise would be.
One other source of improved maintenance can be examined. The U.S.
Environmental Protection Agency (EPA) reported that 5-10% of 1975 and later
cars have used leaded fuel, which, after five or six tankfuls, permanently
poisons and renders useless the catalytic converters on these cars. If
only unleaded gasoline were sold, this poisoning would not occur. The
societal value, from the perspective of PAH emissions, of this change to
the production and sale of only unleaded gasoline can be calculated.
The effect on HC (and hence PAH) emissions of the loss of effectiveness
of the converter depends on the way the manufacturer has tuned the
remainder of the control system. If we use a change in HC of 2 g/mi as a
likely estimate, this implies additional emissions of BaP at 4yg/mi.
Suppose that 10Z of the catalytic-converter fleet (cars and light-duty
trucks) has poisoned catalysts and that this fleet accounts for 70% of the
140 x lO1^ mi driven annually by gasoline vehicles. Then the extra BaP
emissions from the poisoned-catalyst vehicles come to 392 kg of BaP per
year (4 ug/mi x 0.1 x 0.7 x 140 x 101 tni/yr). The risk-valuation
procedure indicates that the elimination of these emissions would be worth
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about $86 million. (This counts reductions in both urban and rural BaP
emissions as worth $0.22/mg; if only urban-emission reductions are
considered valuable, because rural emissions are more dispersed and are
relatively less harmful, the value would be only 60Z as great, or $52
million.)
To achieve this gain, the 40 x 10^ gal of leaded gasoline would have
to be replaced by unleaded. A conservative estimate of the cost of this
conversion would be $0.05/gal—a low estimate of the difference in retail
price between leaded and unleaded gasoline. (There appear to be no good
reasons why the difference in retail price should not fully reflect the
difference in the complete cost of production, distribution, and sale of
the two types of gasoline.) Thus, conservatively, the conversion would
cost at least $2 billion/yr. Because unleaded gasoline tends to have a
higher aromatic-compound content than leaded and a higher aromatic-
compound content tends to yield greater PAH emissions, the conversion would
tend to yield greater PAH emissions from older vehicles and from all
heavy-duty trucks—but lead emissions would decrease. In sum, the benefits
from PAH reductions alone appear to be far smaller than the costs of
converting the U.S. gasoline supply entirely to unleaded.
The HC emission standards for heavy-duty gasoline trucks currently
limit HC emissions to the equivalent of about 5 g/mi, or about 0.5 ton over
the life of a truck. The heavy-duty truck regulations likely to be
promulgated for 1984 and after will decrease this to about 0.23 ton of HC
over the life of the vehicle. Further tightening of standards (as
originally promulgated for 1984) could decrease lifetime HC emissions by an
additional 0.08 ton; this would yield a decrease of roughly 0.16 g of BaP.
The risk-valuation method indicates that this decrease would be worth $35.
EPA has estimated that the hardware (largely, a catalytic converter) for
this further tightening of a standard would cost about $300; some
manufacturers have indicated higher values. In any event, the value of the
reduction in the amount of BaP is unlikely to make an appreciable
difference in assessing the value of the change in HC over the life of the
vehicle.
In sum, for gasoline vehicles, the likely benefits from reductions in
PAH emissions appear to fall far short of the costs of the measures that
would be necessary to achieve them.
Diesel Vehicles
As noted in Chapter 1, diesel vehicles are important sources of PAHs,
with most of these compounds adsorbed on the surface of carbon particles.
Also given in Chapter 1 are estimates of BaP emissions of 13 ug/rai and
54 ug/mi for light-duty and heavy-duty diesel vehicles, respectively, and
they are used initially here.
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Reductions in particulate emissions from light-duty diesels have thus
far been achieved through engine modifications; because no regulations have
been promulgated for particulate emissions from heavy-duty diesels, it is
unlikely that any reductions in emissions from these vehicles have occurred,
Further reductions in emissions from light-duty diesels appear to be
possible from two sources: trap-oxidizers in the exhaust and fuel
modifications.
Assume that a trap-oxidizer, in reducing particulates, reduces PAH
emissions by a comparable percentage. EPA regulations currently mandate a
standard of a particulate maximum of 0.6 g/mi for 1982-1984 light-duty
diesels. For 1985 and after, current EPA regulations require standards of
0.2 g/mi for diesel cars and 0.26 g/mi for diesel light-duty trucks. It
appears unlikely that the required trap-oxidizer technology will be
available, and these regulations are likely to be modified. Nevertheless,
because the trap-oxidizer technology has been pursued in the context of the
0.2-g/mi standard, it is useful to analyze emissions in the same context.
The 0.2-g/mi standard would imply a roughly two-thirds reduction in
particulate emissions. If BaP emissions fall by the same proportional
amount, this implies a reduction of 8 ug/mi. Over the typical life of the
vehicle (100,000 mi), there would be 0.8 g less BaP emitted from the
vehicle. The risk-valuation procedures indicate that this reduction would
be worth $176. (In principle, a discount rate should be used for benefits
after the first year; Because of the pattern of use of these vehicles, a
real discount rate of 3Z would reduce the net present value of the benefits
by only 10%.) General Motors currently estimates that a trap-oxidizer, if
it is made practicable, will cost around $500; EPA has estimated the
cost as appreciably lower. ° There may also be fuel-economy penalties
and driver-satisfaction costs. The societal value of the reduction in PAH
emission alone would offset only a modest fraction of the cost of control.
As to heavy-duty diesel trucks, in late 1980 EPA proposed a set of
regulations that would have reduced particulate emissions from heavy-duty
diesels built in 1986 and later to one-third the emissions from unregulated
vehicles. EPA has taken no further action to make these rules final,
and, because they too required trap-oxidizers, it appears likely that they
will be modified. But they are useful as a benchmark. A two-thirds
reduction in BaP emissions would mean a reduction of 36 ug/mi. Over the
typical life of a heavy-duty diesel vehicle (475,000 mi), this implies a
reduction in BaP emissions of 17.1 g. The valuation method indicates that
this reduction, if it all occurred in urban areas, would be worth $3,760.
(These estimates do not incorporate a discount rate.) Here, however, the
rural-urban distinction cannot be ignored. The data in Chapter 1 indicate
that only 20Z of heavy-duty diesel mileage is likely to occur in urban
areas. Thus, these estimates are upper limits of their likely societal
value. If BaP emissions in rural areas were assumed to have little or no
serious human-health consequence (i.e., it is transformed or otherwise
removed in an ultimately harmless fashion before an appreciable number of
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people are exposed), Chen Che socieCal value (based only on urban exposure)
would be only $752. This lasc value is still relatively large. The
reduccion in PAH emissions from heavy-duty vehicles seems Co be societally
important (in essence, because of Che relatively heavy emissions from and
the high mileage accumulated by these vehicles). Even if Che large
reductions attempted by the regulations proposed in late 1980 cannot be
achieved, it appears that smaller reductions (which might be achieved
through relatively low-cost engine modifications, analogous to those
already achieved in light-duty diesels), although also promising smaller
benefits, would be societally worthwhile on the basis of PAH emissions
alone. In this respect, this appendix can echo the recommendation of the
recent NRC study of light-duty diesels: "Regulate particulate exhaust from
such large sources of emissions in road transport as heavy diesel trucks
and buses; this may be more cost-effective than tightening the emission
levels of diesel cars and light trucks."^'
The substitution of No. 1 diesel fuel for the currently used No. 2
diesel fuel can reduce particulate emissions and PAH emissions. >°>^
The results of Hare and Baines indicate that BaP emissions from
light-duty vehicles may be reduced by about 25%. An experiment on
Washington, D.C., buses suggests that the reduction might be even greater
for heavy-duty vehicles. The estimate of 25Z is used here.
The BaP emissions from both light- and heavy-duty diesels is 320
ug/gal. A 252 reduction would mean a reduction of 80 yg/gal. The risk-
valuation procedures place a value of $0.018 on this reduction. The
current retail-price difference between the two fuels is around $0.08 per
gallon. It does not appear that the benefit from the PAH reduction alone
would exceed the costs of the substitution of No. 1 for No. 2 diesel fuel.
In sum, efforts to achieve engine modifications in heavy-duty vehicles
appear to be the most cost-effective way to achieve net societal gains.
A few caveats should be mentioned with respect to the discussion of
diesel vehicles. First, the diesel analyses assume that the PAH emissions
from diesel vehicles, as represented by BaP, have the same health
consequences per ton as the PAH emissions from other sources. But a recent
NRC study failed to find any definite association between diesel-
exhaust emissions and carcinogenesis in humans, despite the presence in the
exhaust of PAHs that are known to be carcinogenic in other animals. The
study's authors suggested that there may be something special about the
bioavailability of these compounds to humans when they are present in
diesel exhaust. At the other extreme, however, the data in Appendix C
indicate that the extracts from some diesel exhaust may be as much as 89
times more potent mutagenically when measured on a BaP-equivalent basis.
The discussion in this section, steers a middle course and assumes that the
BaP in vehicle exhaust represents the same carcinogenic potential for
humans as does the BaP from the sources covered in the review of
epidemiologic studies in Appendix C.
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Second, the results of Springer,33 Hare and Raines,20 and Williams
and Swarin45 provide estimates for light-duty diesel BaP emissions of
around 3 ug/mi, with some vehicles achieving emissions below
1 pg/mi; these estimates should be compared with the figure of 13 Pg/«i
used in Chapter 1 and in this chapter. Thus, BaP emissions from light-duty
diesel vehicles may be an order of magnitude lower than the figure used
here, and the same qualification may apply to heavy-duty diesels.
OTHER TRANSPORTATION VEHICLES
Chapter 1 indicated that airplanes and ships are the major sources of
BaP emissions in this category. Little other information appears to be
available on emissions or possible avenues of control. Because most of
these emissions occur outside urban areas, it is probably safe to neglect
them in this analysis.
WOOD-BURNING STOVES
As the prices of other fuels have risen, burning wood for residential
heating has become more popular. Wood stoves have 2 or 3 times the thermal
efficiency of fireplaces and have become increasingly popular. ^It is
estimated that a million wood-burning stoves were sold in 1979, and
sales have been increasing. Much of this wood-burning occurs in rural
areas, but a substantial amount occurs in or affects urban areas. For
example, Cooper et al.10 found that approximately SOX of the respirable
particles in the ambient air of Portland, Oregon, in January 1978 came from
residential wood combustion. Cooper also estimated that residential woodg
combustion emitted 1.4 tons of BaP in Portland's ambient air during 1978.
It appears that wood-burning stoves emit BaP at about 2.5 mg/kg of wood
burned.9 Chapter 2 cited data that indicate that households burning wood
as the primary source of heat each used an average of 5.6 metric tons of
wood. It is likely that such a household used a wood stove. Thus, it
would emit 14 g of BaP per year. (Thus, in terms of BaP emissions, one
wood stove is the equivalent of over 100 diesel cars each emitting 13
ug/mi.) If we assume that emissions from a wood stove in an urban area
hfve about the same effects on ambient BaP concentration as do vehic: e
e
emissions, we can use the valuation method to indicate that ^e complete
elimination of these emissions would have a societal value of $3,080 £er
/ear.
Thus, it appears that the benefits from the control of PAH emissions
from wood-burning stoves, especially in urban areas, are quite large.
Unfortunately, research on emissions from wood-burning stoves s stil at a
relatively early stage. It appears that eh. design and struc ure of stoves
can make some difference in PAH emissions.14 Even more important y,
catalytic corabustors (similar to the catalytic converters on cars; in-
stalled in the chimneys of wood stoves may be capable of Cueing organic
Compound emissions by up to 95Z.5t These control devices, if they become
D-19
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practicable, are expected to cost, installed in a new stove, around
$125-150. It is unclear how durable they are, but even if they lasted only
a year, it appears that their likely benefits greatly outweigh their likely
costs.
It appears unrealistic to expect any feasible program for requiring
retrofitting of existing wood stoves in residential use. But many current
owners of wood stoves may be "environment-conscious" and might be prepared
to retrofit their wood stoves voluntarily if retrofit devices were
available and there were sufficient publicity. And a program to require
(or induce) manufacturers of new wood stoves to incorporate changes in
design or technology that would reduce PAH emissions appears to have great
societal benefits and relatively small societal costs.
RESIDENTIAL FIREPLACES
Residential fireplaces are a large source of PAHs; they emit about
one-third as much BaP as wood-burning stoves per kilogram of wood.
It is unlikely that any emission-control program aimed at fireplaces
would be feasible. It is unclear whether any technology is currently
available for controlling emissions from fireplaces; even the catalytic
cotnbustors, which appear promising for wood stoves, are unlikely to be
practicable for fireplaces, because the combustors require a higher
temperature than most fireplace chimneys are likely to provide. Further-
more, any retrofitting program would be highly unlikely to be put into
effect, and installation of any technologic device in new residences would
deal with only a tiny fraction of the problem.
Nevertheless, because the aggregate amount of PAH emissions from
fireplaces is large, and likely to grow, this appears to be a fruitful
direction for research.
RESIDENTIAL COAL-FIRED HEATING
There appears to be little available information on this subject.
Some residential heating units are capable of burning both wood and coal.
Because the emissions from these units are potentially large, more
information should be collected and research encouraged.
INDUSTRIAL COAL-FIRED BOILERS
Comparatively little appears to be known about the properties of these
boilers. EPA still appears to be in the data-collecting stage with respect
to these devices. » Because these boilers are still being manufactured
and are in the hands of relatively larger and more sophisticated users
(compared with households), control efforts would probably be feasible if
D-20
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the cost-benefit ratios were favorable. Retrofitting, inducement to use
gas and oil, and improved design and technology (e.g., possibly catalytic
corabustors or precipita-tors) appear to be possible. Clearly, more research
is necessary.
INDUSTRIAL-COMMERCIAL INCINERATORS
Municipal incinerators do not appear to be a serious source oE PAH
emission, but industrial-commercial incinerators are.1^ One possible
reason is that municipal incinerators operate at higher, more efficient
temperatures.
Unfortunately, little other information is available about these
sources. Two control strategies seem to be possible. One would be to
focus on the technology of industrial incineration itself—i.e., focus on
retrofitting, improving the design of new devices, exploring the
possibility of corabustors or precipitators, etc. Again, one would want to
make sure that the cost-benefit ratios were favorable before embarking on
such an effort.
A second control strategy would be to require that industrial-
coramercial trash be hauled to municipal incinerators for burning. A rough
estimate of the costs and benefits of such a strategy can be provided. If
industrial-commercial incinerators have BaP emission rates of 120-570
yg/kg of refuse burned, burning a ton of refuse would yield 120-570 rag of
BaP. If this occurred in urban areas and the emissions had a dispersion
pattern similar to that of motor-vehicle emissions, the risk-valuation
method would indicate that the complete elimination of these emissions
would be worth $26-125. It appears that municipal incinerators have BaP
emissions rates 2-3 orders of magnitude lower than industrial-commercial
incineration rates. Thus, the use of municipal incineration would mean
the virtual elimination of the BaP emission.
In 1975, the cost of refuse collection was about $25/ton. * This
figure should probably be doubled to bring it to 1982 dollars, for an
estimate of $50/ton. This cost estimate is within the range of the likely
benefits from the reduction in industrial-commercial incineration.
Accordingly, efforts to reduce industrial-commercial incineration or the
emissions from industrial-commercial incineration may be worthwhile and
bear further investigation.
COKE-OVEN EMISSIONS
Coke ovens are well-known sources of PAH and BaP. Coke manufacturers
are currently in the process of implementing emission reductions under EPA-
supervised state implementation plans, consent decrees, and Occupational
Safety and Health Administration plans. EPA has recently proposed further
standards that would control emissions to a greater degree. EPA
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estimates chat its standards will reduce emissions by 880 tons of
benzene-soluble organic compounds (BSOs) per year. EPA estimates the cost
at $46 million/yr. The ratio of BaP emissions to BSO emissions appears to
range from 1:500 for wet-coal charging to 1:133 for battery stacks.5
This yields a range of 1.76-6.62 tons by which BaP emissions would be
reduced by the proposed regulation. The risk-valuation procedure indicates
that even the lower estimate is worth $387 million. This allows for
substantial error in the estimates of costs and benefits that would
nevertheless leave the proposed regulations cost-effective. Unfortunately,
although EPA discussed yet more stringent regulations in its proposal, no
cost figures were presented, so no evaluations can be made. The large
margin between the likely benefits and the likely costs of control,
however, indicates that further restrictions in coke-oven emissions couLd
well be worthwhile (up to the point at which marginal costs equal marginal
benefits).
PRESCRIBED BURNING
The burning of underbrush in forests appears to be a standard
practice. It is claimed that such burning reduces the incidence of
wildfires and achieves a substantial (5- to 10-fold) net reduction in
particulate emissions. If PAH emissions follow the same pattern, there
appears to be little or no necessity for any corrective action.
AGRICULTURAL AND LAND-CLEARING WASTE BURNING
These are standard practices and take place in rural areas. The
alternatives to burning appear to be collection and either central burning
(at high, more efficient temperatures) or disposal in landfills. The BaP
emissions from waste burning appear to be 190-430 ug/kg of waste
burned. This is roughly the same range as for industrial incinerators,
and a similar analysis can be applied. It is not clear whether the cost
per ton of rural collection of waste is higher or lower than the urban
cost. Landfill disposal could add another $10/ton. Consequently, the same
range of cost-benefit uncertainty that applied to industrial-commercial
incineration appears to apply to rural waste burning—with the added
element that these emissions are in rural areas and hence may be less
socially damaging. Further study is warranted.
SUMMARY
This appendix has discussed the principles of public decision-making
with respect to PAH emissions and has provided illustrative risk-valuation
analyses co examine the societal costs and benefits of reducing PAH
emissions from the major sources.
D-22
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