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                        Piiblication-on-Demand Program
            Polycyclic  Aromatic Hycarbons
               -'BRARY. AWBERC, CINCINNA-,
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.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.
    The National Research Council was established by the National
Academy of Sciences in 1916 to associate the broad community of science
and technology with the Academy's purposes of furthering knowledge and
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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.
                                   ES-1

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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.
                                   ES-2

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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
                                   ES-3

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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
                                   ES-4

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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.
                                   ES-7

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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


                                     1-1

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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
                                    1-1

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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
                                     1-2

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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
                                    1-3

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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
                                    1-4

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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.
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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.
                                    1-8

-------
     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

-------
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

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                                   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

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                         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

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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

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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

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                            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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         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.
                                  2-1

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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

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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

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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

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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

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                                   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

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                                   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

-------
                                REFERENCES

 1.  Bagchi, N. J., and R. E. Zimmerman.   An industrial hygiene evalua-
     of chimney sweeping.  Amer. Ind.  Hyg. Aasoc.  J.  41:297-299,
     1980.
 2.  Bennett, R. L.,  K. T. Knapp, P.  W.  Jones,  J.  E.  Wilkerson,  and P. E,
     Strup.  Measurement of polynuclear  aromatic hydrocarbons  and  other
     hazardous organic compounds in stack gases, pp.  419-428.   In  P. W.
     Jones and P. Leber, Eds.  Polynuclear Aromatic  Hydrocarbons.  3rd
     International Symposium on Chemistry and Biology—Careinogenesis
     and Mutagenesis.  Ann Arbor, Mich.:   Ann Arbor  Science
     Publishers, 1979.
 3.  Bingham, E., R.  P. Trosset, and  D.  Warshawsky.   Carcinogenic  poten-
     tial of petroleum hydrocarbons.   J.  Environ.  Pathol.  Toxicol.
     3:483-563, 1980.
 4.  Bjorseth, A.  Determination of polynuclear aromatic hydrocarbons in
     the working environment, pp. 371-381.  In P.  W.  Jones and P.  Leber,
     Eds.  Polynuclear Aromatic Hydrocarbons.  3rd International
     Symposium on Chemistry and Biology—Carcinogenesis and
     Mutagenesis.  Ann Arbor, Mich.:   Ann Arbor Science Publishers,
     1979.
 5.  Buchanan, M. V., C.-H. Ho, M. R.  Guerin, and B.  R.  Clark.  Chemical
     characterization of mutagenic nitrogen-containing  polycyclic
     aromatic hydrocarbons in fossil  fuels,  pp. 133-144.   In M. Cooke
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 6.  Burlingame, J.  0., J. E. Gabrielson, P.  L. Langsjoen, and W.  M.
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 7.  Coleman, H. J.,  J. E. Dooley, D.  E.  Hirsch,  and  C.  J. Thompson.
     Compositional studies of a high-boiling 370-535° distillate from
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 9.  DeAngelis, D. G., D. S. Ruffin,  J. A. Peters, and  R. B. Reznik.
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     Environmental Protection Agency,  1980.   91 pp.
10.  DeAngelis, D. G., D. S. Ruffin,  and  R.  B.  Reznik.   Preliminary
     Characterization of Emissions from Wood-Fired Residential
     Combustion Equipment.  Report prepared by Monsanto Research
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     Agency, 1980.   146 pp.
                                 2-39

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11.  Duncan, J. R.,  K. M. Morkin, and M. P. Schmierbach.  Air Quality
     Impact Potential from Residential Wood-Burning Stoves.
     Presented at 73rd annual meeting, Air Pollution Control Assoc.,
     June 22-27, 1980.  Paper 80-7.2.  15 pp.
12.  Guerin, M. R.,  J. L. Epler, W. H. Griest, B. R. Clark, and T. K.
     Rao.  Polycylic aromatic hydrocarbons from fossil fuel conversion
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     Carcinogenesis.  Vol. 3.  Polynuclear Aromatic Hydrocarbons.  New
     York:  Raven Press, 1978.
13.  Guerin, M. R.,  I. B. Rubin, T. K. Rao, B. R. Clark, and J. L.
     Epler.  Distribution of mutagenic activity in petroleum and
     petroleum substitutes.  Fuel 60:282-288, 1981.
14.  Hangebrauck, R. P., D. J. von Lehmden, and J. E. Meeker.  Source
     of Polynuclear Hydrocarbons in the Atmosphere.  Cincinnati, Ohio:
     U.S. Department of Health, Education, and Welfare, 1967.
15.  Hoffmann, D., and E. L. Wynder.  Environmental respiratory carcino-
     genesis, pp. 324-365.  In C. E. Searle, Ed.  Chemical Carcinogens.
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     1976.
16.  Hubble, B. R.,  J. R. Stetter,  E. Gebert, J. B. L. Harkness, and R.
     D. Flotard.  Experimental measurements of emissions from
     residential wood-burning stoves, pp. 79-104.  In J. A. Cooper and
     D. Malek, Eds.   Residential Solid Fuels—Environmental Impacts and
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17.  Katz, E., and K. Ogan.  The use of coupled-column and high
     resolution liquid chromatography in the analysis of petroleum and
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     Battelle Press, 1981.
18.  Lao, R. C., and R.  S. Thomas.   The gas chromatographic separation
     and determination of PAH from industrial processes using glass
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     Mutagenesis.  Ann Arbor, Mich.:  Ann Arbor Science Publishers,
     1979.
19.  Lao, R. C., R.  S. Thomas, and J. L.  Monkman.  Computerized gas
     chromatographic-mass spectrometric analysis of polycyclic  aromatic
     hydrocarbons in environmental samples.  J.  Chromatog.  112:681-700,
     1975.
20.  Lee, M. L., K.  D. Bartle, and M. V.  Novotny.  Profiles of the poly-
     nuclear aromatic fraction from engine oils obtained by capillary
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     Anal. Chem. 47:540,-543, 1975.
21.  Masek, V.  Benzo(a)pyrene in the workplace atmosphere of coal and
     pitch coking plants.  J. Occup. Med. 13(4):193-198, 1971.
22.  McKay, J. F., and D. R. Latham.  Polyaromatic hydrocarbons in high-
     boiling petroleum distillates.  Isolation by gel permeation
     chromatography  and  identification by fluorescence spectrometry.
     Anal. Chem. 45:1050-1055, 1973.
                                  2-40

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23.  McMahon, C. K., and S. N. Tsoukalas.  Polynuclear aromatic
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24.  Miguel, A. H., and L. M. S. Rubenich.  Submicron size distribu-
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25.  Murphy, D. J., R. M. Buchan, and D. G. Fox.  Ambient particulate
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30.  National Research Council, Committee on Research Needs on the
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32.  National Research Council, Ocean Affairs Board.  Assessing Potential
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33.  Neff, J. M.  Polycyclic Aromatic Hydrocarbons in the Aquatic
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     Applied Science Publishers, 1979.
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     assessment of PAH from coal combustion and gasification,   pp.
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     Hydrocarbons:  5th International Symposium—Chemical Analysis and
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35.  Peake,  E., and K. Parker.  Polynuclear aromatic hydrocarbons and
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                                  2-41

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36.  Peters, J. A.  POM emissions from residential wood burning:  An
     environmental assessment, pp. 267-288. In J. A. Cooper  and  D.
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     on Ground Water Contamination:  Environmental Protection Agency
     Oversight.  House Report No. 96-1440.  1980.   137 pp.
44.  U.S. Department of Energy.   Environmental Readiness Document.
     Wood Combustion.  Report DOE/ERD-0026.  Washington, D.C.:   U.S.
     Department of Energy, 1979.  36 pp.
45.  U.S. Department of Energy.   Health Effects of Residential Wood
     Combustion:  Survey of Knowledge and Research.  Report
     DOE/EV-0114.  Washington, D.C.:  U.S. Department of Energy,  1980.
     29 pp.
46.  U.S. Environmental Protection Agency.  Coke Oven Emissions  from
     By-Product Coke Oven Charging, Door Leaks, and Topside Leaks on
     Wet-Coal Charged Batteries—Background Information for Proposed
     Standards.  Appendix E.  Summary of cancer-risk assessment.
     Washington, D.C.:  U.S. Environmental Protection Agency, 1981.
     45 pp.
47.  U.S. Environmental Protection Agency, Industrial Environmental
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     Protection Agency, 1981.
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     Washington, D.C.:  U.S. Environmental Protection Agency, 1978.
     17 pp.
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49.  White, C. W., A. G. Sharkey, Jr., M. L. Lee, and D. L.
     Vassilaros.  Some analytical aspects of the quantitative deter-
     mination of polynuclear aromatic hydrocarbons in fugitive
     emissions from coal liquefaction processes, pp. 261-275. In
     P. W. Jones and P. Leber, Eds.  Polynuclear Aromatic
     Hydrocarbons:  3rd International Symposium—Chemistry and
     Biology.  Ann Arbor, Mich.:  Ann Arbor Science Publishers,  1979.
                                 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

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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

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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

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    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

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    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
                                   3-8

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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.

                                  3-10

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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
                                  3-13

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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.
                                  3-14

-------
    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
                                4-2

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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
                                4-3

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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
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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
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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
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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


                                  4-7

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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—
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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
                                  4-14

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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.
                                  4-15

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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
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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
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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.
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    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

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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

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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

-------
                                 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

-------
                                 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

-------
                                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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       Inst. 56:1237-1242,  1976.
174.   Wallcave, L.   Gas chromatographic analysis of polycyclic aromatic
       hydrocarbons in soot samples.  Environ. Sci. Technol. 3:948,
       1969.

                                 4-61

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175.   Wallcave, L., D. L. Nagel,  J.  W.  Smith, and R. D. Waniska.  Two
       pyrene derivatives of widespread  environmental distribution:
       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

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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

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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

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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

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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

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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

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                                                                             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

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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

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                                                                     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

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                                 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

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                                   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

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  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

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                                                       H  OH
                                                                     B*nxothloph*n«dlonc
Bcnsothiophanc
                           5-15

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   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
                                  5-17

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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
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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
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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.
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    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.
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      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

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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
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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

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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

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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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           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.
                                   6-2

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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.
                                  6-3

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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."
                                   6-4

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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.
                                   6-6

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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.
                                  6-10

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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
                                  6-11

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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
                                  6-12

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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
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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
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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.
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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 
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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
                                  6-28

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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

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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

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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

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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

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    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

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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

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                                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

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                                             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.

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                                                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

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                          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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159.   Seidel, K., and H. Happel.  Effect of  refuse  compost  on 3,4-
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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

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     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

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     •   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

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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

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     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

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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

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                                      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

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   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

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                     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.                               ~~

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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

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                                                            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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                                 REFERENCES

 1.  Ambre, J., D. Graeff, F. Bures,  D. Haupt,  and  K.  Deason.   Antipyrine
     metabolism and bronchogenic carcinoma.   J.  Med.  8:57-70,  1977.
 2.  Andrews, A. D., S. F. Barrett, and J. H.  Robbins.   Relation  of D.N.A
     repair processes  to pathological aging  of the  nervous  system in xero'
     derma pigmentosum.  Lancet  1:1318-1320,  1976.
 3.  Arlett, C. F., and A. R. Lehmann.  Human  disorders  showing increased
     sensitivity to the induction of  genetic  damage.   Ann.  Rev. Genet.
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203. Yuspa, S. H., E. F. Spangler, R. Donohoe, S. Geusz, E. Ferguson,
     M. Wenk, and H. Hennings.  Sensitivity to two-stage carcinogenesis of
     SENCAR mouse skin grafted to nude mice.  Cancer Res. 42:437-439, 1982.
204. Zack, M., S. Cannon, D. Loyd, C. W. Heath,  J. M. Falletta,
     B. Jones, J. Housworth, and S. Crowley.  Cancer in children of
     parents exposed to hydrocarbon-related industries and occupations.
     Amer. J. Epidemiol. 111:329-336, 1980.
205. Zielske, J. V., and S. H. Golub.  Fetal calf serum-induced blasto-
     genic and cytotoxic responses of human lymphocytes.  Cancer Res.
     36:3842-3846, 1976.
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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
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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
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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.
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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.
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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

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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.
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                                     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
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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?
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    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.
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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"?

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    •  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.
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                     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

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                                   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

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                                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

-------
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

-------
     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

-------
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

-------
                        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

-------
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

-------
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

-------
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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25.   National Research Council.  Health Effects of Exposure to Diesel
      Exhaust.  The Report of the Health Effects Panel of the Diesel Impacts
      Study Committee.  Washington, D.C.:  National Academy Press, 1981.
      169 pp.
26.   National Research Council, Committee on Biologic Effects of
      Atmospheric Pollutants.  Particulate Polycyclic Organic Matter.
      Washington, D.C.:  National Academy of  Sciences, 1972.   361 pp.
27.   National Research Council, Safe Drinking Water Committee.  Drinking
      Water and Health.  Washington, D.C. :  National Academy of Sciences,
      1977.  939 pp.
28.   Neal, J., and R. H. Rigdon.  Gastric tumors in mice fed benzo(a)-
      pyrene:  A quantitative study.  Tex. Rep. Biol. Med.  25:553-557,
      1967.
29.   Peto, R.  Epidemiology, multistage models, and short-term mutagenicity
      tests, pp. 1403-1428.  In H. H. Hiatt,  J. D. Watson, and J. A.
      Winsten, Eds.  Origins of Human Cancer.  Book C.  Human Risk
      Assessment.  Cold Spring Harbor, N.Y.:  Cold Spring Harbor Laboratory,
      1977.

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30.    Pike, M. C., and B. E. Henderson.  Epidemiology of  polycyclic  hydro-
      carbons:  Quantifying the cancer risk from cigarette  smoking and
      air pollution, pp. 317-334.  In H. V. Gelboin and P.  0.  P.  Ts'o, Eds.
      Polycyclic Hydrocarbons and Cancer.  Vol. 3.  New York:  Academic
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31.    Raffle, P. A. B.  The health of the worker.  Brit.  J.  Indust.  Med.
      14:73-80, 1957.
32.    Redmond, C.  K., A. Ciocco, J. W. Lloyd, and H. W. Rush.  Long-term
      mortality study of steelworkers.  VI.  Mortality  from malignant neo-
      plasms among coke oven workers.  J. Occup. Med. 14:621-629, 1972.
33.    Royal College of Physicians of London.  Smoking or  Health.  The Third
      Report from the Royal College of Physicians of London.   London:
      Pitman Medical Publishing Co. Ltd.,  1972.  128 pp.
34.    Santodonato, J., P. Howard, and D. Basu.  Health and  ecological
      assessment of polynuclear aromatic hydrocarbons.  J.  Environ.  Pathol.
      Toxicol. 5:1-364, 1981.
35.    Sawicki, E.   Airborne carcinogens and allied compounds.  Arch. Environ.
      Health 14:46-53, 1967.
36.    Sawicki, E.  , W. C. Elbert, T. R. Hauser, F. T. Fox, and  T.  W.  Stanley.
      Benzo(a)pyrene content of the air of American communities.  Amer. Ind.
      Hyg. Assoc.  J. 21:443-451, 1960.
37.    Segi, M., M. Kurihara, and T. Matsuyama.  Cancer Mortality  for Selected
      Sites  in 24 Countries.  No. 5 (1964-1965).  Department of Public
      Health.  Sendai,  Japan:  Tohoku University School of  Medicine, 1969.
      174 pp.
38.    Stocks,  P.  Cancer in North Wales and Liverpool regions.  Supplement to
      British  Empire Cancer Campaign Annual Report, 1957.
39.    Stukonis, M. K.  Cancer incidence cumulative rates.   IARC Internal
      Technical Report No.  78/002.  Lyon,  France:  International  Agency for
      Research on Cancer, 1978.
40.    U.S. Department of Health, Education, and Welfare.  Office  on  Smoking
      and Health.   A Report of the Surgeon General.  DHEW  Publication No.
      (PHS)79-50066.  Washington, D.C.:  U.S. Department  of Health,
      Education,  and Welfare, 1979.   1196  pp.
41.    Waller,  R.  Trends in lung cancer  in London in relation  to  exposure to
      diesel  fumes, pp.  1085-1099.  In Health Effects of  Diesel Engine
      Emissions:  Proceedings of an International Symposium.
      EPA-600/9-80-057b.  Cincinnati:  U.S. Environmental Protection Agency
      Office of Research and Development,  1980.
42.    Wynder,  E.  L., and D. Hoffmann.  Experimental tobacco carcinogenesis.
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      Experimental Carcinogenesis.  New York:  Academic Press,  1967.  730 pp.
44.    Wynder,  E.  L., K. Mabuchi, and  E. J. Beattie.  The  epidemiology of
      lung  cancer.  Recent  trends.  J.A.M.A.  213:2221-2228, 1970.
45.   Wynder,  E.  L.,  and S. D.  Stellman.   Impact of  long-term  filter
      cigarette usage  on lung and  larynx cancer  risk:   A  case-control  study.
      J.  Natl. Cancer  Inst. 62:471-477,  1979.
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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
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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
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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;   >
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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

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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
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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
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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
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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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