United States
Environmental Protection
Agency
Environmental Research
Laboratory
Gulf Breeze FL 32561
EPA-600/9-82-013
July 1982
Research and Development
&EPA
Symposium:
Carcinogenic Polynuclear
Aromatic Hydrocarbons in
the Marine Environment
oio
0:010
-------
EPA-600/9-82-013
July 1982
SYMPOSIUM:
CARCINOGENIC POLYNUCLEAR AROMATIC HYDROCARBONS
IN THE MARINE ENVIRONMENT
Pensacola Beach, Florida
14-18 August 1978
Edited by
N.L. Richards and B.L. Jackson
Environmental Research Laboratory
Gulf Breeze, Florida 32561
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
GULF BREEZE, FLORIDA 32561
-------
DISCLAIMER
Although some of the research described in this publication has been
funded wholly or in part by the United States Environmental Protection
Agency, it has not been subjected to the Agency's required peer and
administrative review and, therefore, does not necessarily reflect the
view of the Agency and no official endorsement should be inferred.
-------
CONTENTS
Foreword vi
Overview vii
Acknowledgments ix
Keynote
Review of Species-Specific Metabolic Pathways of Carcinogenic
Polynuclear Hydrocarbons in Marine Organisms 1
H.E. Kaiser and E.K. Weisburger
A. Fate and Detection
Negative Chemical lonization Mass Spectra of Some Polynuclear
Aromatic Hydrocarbons 14
R.C. Dougherty, S.V. Howard, and J.D. Wander
A Characterization of the Polycyclic Aromatic Hydrocarbon
Content of Tars, Tarballs, and Sediments from the Marine
Environment 26
J.L. Lake, C.B. Norwood, and C.W. Dimock
Toxic Photoxygenated Products Generated under Environmental
Conditions from Phenanthrene 36
J.R. Patel, J.A. McFall, G.W. Griffin, and J.L. Laseter
B. Biological Indicators
The Monitoring of Substances in Marine Waters for Genetic
Activity 58
J.M. Parry, M.A.J. Al-Mossawi, N. Danford, and J. Ballantine
Biphenyl Hydroxylase Activity and the Detection of Carcinogens. .81
N. L. Couse, J.J. Schmidt-Collerus, J. King, and L. Leffler
Petroleum and Petroleum Combustion Byproducts as Potential
Sources of Marine Environmental Mutagens 102
J.F. Payne, R. Maloney, A. Rahimtula, and I. Martins
Chemical Carcinogenesis in Fish: Induction of Hepatic Drug
Metabolizing Enzymes and Bacterial Mutagenesis with Polycyclic
Aromatic Hydrocarbons (PAH) 110
D. E. Hinton, J.E. Klaunig, M.M. Lipsky, R. Jack, M. Kahng,
H. Sanefuji, R.T. Jones, and B.F. Trump
Induction of Benzo(a)pyrene Monooxygenase in Fish after I.P.
Application of Water Hexane Extract—A Prescreen Tool for
Detection of Xenobiotics 124
B. Kurelec, M. Protic, M. Rijavec, S. Britivic, W.E.G. Muller,
and R.K. Zahn
iii
-------
C. Metabolism
In Vivo and In Vitro Studies on the Metabolism of Polycyclic
Aromatic Hydrocarbons by Marine Crabs 137
R.F. Lee and S.C. Singer
Techniques for the Waterborne Administration of Benzo(a)pyrene
to Aquatic Test Organisms 148
S.P. Felton, W.T. Iwaoka, M.L. Landolt, and B.S. Miller
Activation and Uptake of Polynuclear Aromatic Hydrocarbons by
the Marine Ciliate, Parauronema acutum 163
D.G. Lindmark
Effect of Polynuclear Aromatic Hydrocarbons and Polyhalogenated
Biphenyls on Hepatic Mixed-Function Oxidase Activity in Marine
Fish 172
M.O. James and J.R. Bend
Metabolism of Benzo(a)pyrene by Ciona intestinal is 191
W.M. Baird, R.A. Chemerys, L. Diamond, T.H. Meedel, and
J.R. Whittaker
Bioactivation of Polynuclear Aromatic Hydrocarbons to Cytotoxic
and Mutagenic Products by Marine Fish 201
J.J. Stegeman, T.R. Skopek, and W.G. Thilly
D. Genotype
Hypersensitivity for Carcinogenesis Resulting from Species
Hybridization Impairing Control of Cellular Oncogenes as Tool
towards Tailoring Test Animals Suitable for Monitoring
Carcinogens. 212
M. Schwab, S.S. Abdo, and G. Kollinger
The Use of Genetically Modified Fish in the Detection and
Measurement of Carcinogens in Water 233
L.S. Shelton, M.L. Bellamy, and D.G. Humm
E. Field Studies
The Distribution of Benzo(a)pyrene in Bottom Sediments and of
Neoplasms in Bottom-dwelling Flatfish Species of the Pacific and
Atlantic Oceans, North, China, Bering, and Beaufort Seas, and
Sea of Okhotsk 244
H.F.Stich, B.P. Dunn, A.B. Acton, F. Yamaski, K. Oishi, and T. Harada
Polynuclear Aromatic Hydrocarbons in Estonian Water, Sediments,.
and Aquatic Organisms. . 260
P. Bogovski, I. Veldre, A. Itra, and L. Paalme
IV
-------
F. Uptake
Bioaccumulation and Toxicity in English Sole, Parophrys vetulus,
following Waterborne Exposure to Benzo(a)pyrene 268
M.L. Landolt, S.P. Felton, W.T. Iwaoka, B.S. Miller, D. DiJulio,
and B. Miller
Accumulation and Release of Polycyclic Aromatic Hydrocarbons from
Water, Food, and Sediment by Marine Animals 282
J.M. Neff
Some Aspects of the Uptake and Elimination of the Polynuclear
Aromatic Hydrocarbon Chrysene by Mangrove Snapper, Lutjanus
griseus and Pink Shrimp, Penaeus duorarum , .321
D.L. Miller, J.P. Corliss, R.N. Farragut, and H.C. Thompson, Jr.
Accumulation, Tissue Distribution, and Depuration of
Benzo(a)pyrene and Benz(a)anthracene in the Grass Shrimp,
Palaemonetes puqio 336
F.R. Fox and K.R. Rao
G. Food Web Transfer
An Ecological Perspective on Human Food Webs 350
Rufus Mori son
The Cellular Fate of Benzo(a)pyrene 367
V. Ivanovic and I.E. Weinstein
Polycyclic Aromatic Hydrocarbons in the Aquatic Environment
and Cancer Risk to Aquatic Organisms and Man 385
J.M. Neff
-------
FOREWORD
The protection of our estuarine and coastal areas from damage caused
by toxic organic pollutants required that regulations restricting the
introduction of these compounds into the environment be formulated on a
sound scientific basis. Accurate information describing dose-response
relationships for organisms and ecosystems under varying conditions is
required. The EPA Environmental Research Laboratory, Gulf Breeze (ERL.GB),
contributes to this information through research programs aimed at
determining:
the effects of toxic organic pollutants on individual species and
communities of organisms;
the effects of toxic organics on ecosystem processes and
components;
the significance of chemical carcinogens in the estuarine and
marine environment.
This publication is a compilation of papers contributed by scientists
who participated in the Symposium on "Carcinogenic Polynuclear Aromatic
Hydrocarbons in the Marine Environment" sponsored by ERL, Gulf Breeze and
the Environmental Protection Agency (EPA) Office of Energy, Minerals, and
Industry August 14-18, 1978, at Pensacola
number of questions related to the impact
ecosystem: their physical, chemical, and
into the aquatic environment; the current
Beach. Participants addressed a
of these compounds on the marine
biological fate after release
state-of-the-art for their
detection
food.
and identification; and their potential for transfer to human
Henry
Director
Environmental Research Laboratory
Gulf Breeze, Florida
-------
OVERVIEW
by
Norman L. Richards
Symposium Convener
Environmental Research Laboratory
Gulf Breeze, Florida 32561
Development of an EPA research program to assess the potential for
mutagenic or carcinogenic polynuclear aromatic hydrocarbons (PAHs) to reach
man through the marine food web was a formidable undertaking. An
understanding of the complex marine sources and sinks of PAHs and the
processes involved in the evaporation, dissolution, biological uptake, and
transformation at different trophic levels of food webs is only beginning
to emerge in the literature. Our effort required a multidisplinary team
approach to a rapidly expanding field of environmental science.
This symposium was convened primarily to summarize current knowledge
about the potential for carcinogenic PAHs and their metabolites to
accumulate in seafood organisms. Although speakers presented papers in
highly specialized areas, the interrelationship of their research will be
obvious to readers. For example, chemists with expertise in
carcinogen/mutagen separation and identification from complex environmental
mixtures can now team up with biologists who have sensitive biological
carcinogen/mutagen detection methods and study a wide variety of problems.
Another breakthrough involves a better understanding of the induction of
PAH metabolizing enzymes in marine animals as a possible monitoring tool.
Topics for invited papers ranged from the development of sensitive new
techniques for analytical chemical detection and fate of PAHs to their
detection through the use of biological indicators. Recent developments in
metabolism by marine organisms, effects of genotype on responses, and the
cellular fate of PAHs these compounds were discussed. Speakers included
internationally known authorities in such disciplines as analytical
chemistry, photochemistry, enzymology, physiology, molecular biology,
genetics, aquatic toxicology, microbiology, and fishery biology.
Research results supported the belief that additional breakthroughs
will be forthcoming that will improve our understanding of the potential
for PAHs to reach man through the marine environment. Interspecies
comparisons illustrate both similarities and differences in PAH uptake,
metabolism and detoxification. Biological methods with potential for
vi i
-------
improved detection of mutagens/carcinogens in complex environmental
mixtures were described. An understanding of the role of heredity,
exposure route, immunology, metabolism, and detoxification mechanisms in a
diverse range of marine organisms was recognized as essential to predict
the outcome of exposure of marine organisms to PAHs.
The release of PAHs into the marine environment is likely to increase
in response to increased petroleum transportation and offshore oil
exploration and production. The possibility also exists that in the
distant future a shift to fuels derived from shale oil may expose the
marine environment to PAHs of different chemical composition. Because the
photochemistry of PAHs is now better understood, we may acquire new
insight into the weathering of PAHs in the marine environment and thus into
the selection of appropriate model compounds for toxicity and
bioaccumulation studies.
We now have many of the research tools required to assess the
potential for mutagenic/carcinogem'c PAHs and their degradation products to
accumulate in seafood organisms. An intensified research and monitoring
program would help resolve questions regarding the human health risk when
mtitagenic/carcinogenic PAHs enter the marine environment.
vm
-------
ACKNOWLEDGMENTS
This publication is a result of the patience and cooperation of
contributors who prepared and edited their texts. Mrs. Valerie Caston and
Mrs. Maureen Stubbs are recognized for their work in the preparation of
camera copy and arrangement of illustrations.
IX
-------
REVIEW OF SPECIES-SPECIFIC METABOLIC PATHWAYS OF CARCINOGENIC
POLYNUCLEAR AROMATIC HYDROCARBONS IN MARINE ORGANISMS
by
H. E. Kaiser
Department of Pathology
School of Medicine
University of Maryland
Baltimore, MD 21201
and
E. K. Weisburger
National Cancer Institute
National Institutes of Health
Bethesda, MD 20205
ABSTRACT
The problem of neoplastic growth is a fundamental one to man
and to organisms in the surrounding environment as our
environment becomes more and more polluted. The marine
environment, the largest on earth, has great import for the
future, affecting aspects of nutrition, transportation, resource
recovery, energy, and recreation.
We know from Aristotle, Organon, Part IV, that we can
attack a problem by its general feature, known in philosophy as
katholou, in contrast to kathekaston, which is the specific. It
is a pleasure to speak in this symposium on the general
comparative aspects of the metabolic pathways of carcinogenic
polynuclear aromatic hydrocarbons in marine and other organisms.
Other presentations will deal with specific topics of these
compounds in marine organisms. It is necessary for our review
to include current knowledge of nonmarine organisms, including
man and nonmarine plants, to trace briefly the action of
carcinogenic polynuclear aromatic hydrocarbons and their
species-specific metabolic pathways. Finally, we shall point to
the necessity for stimulating new research for protecting the
marine environment. The polycyclic aromatic hydrocarbons, a
most important group of chemical carcinogens, are the concern of
this symposium. Improvements in the methods for the detection
Shortened version of address
-------
of the carcinogenic polynuclear aromatic hydrocarbons and other
compounds have facilitated studies concerning their fate in the
oceans, as well as comparative evaluations of the species-
specific, organ-specific, tissue-specific, and pathway-specific
processes leading to carcinogenesis. The recent successes in
fusing cell structures of animals (including man) and plants
give new impetus to a broad comparison of species-specific
metabolic pathways of the carcinogenic polynuclear aromatic
hydrocarbons in the organisms of the marine environment. The
pathways of chemical compounds can be compared with respect to
the following topics: contact of the carcinogen(s) with
tissue(s), absorption, storage, effect of circadian and other
rhythms, and metabolic interactions. These may include
species-specific differences in enzyme systems involved in the
metabolism of carcinogens, differences in the chemical binding
of carcinogens or their metabolites to cell constituents,
differences in biochemical constituents in the cell during
carcinogenesis, variable pathways of the same carcinogen, the
excretion of carcinogens, and the influence of the altered
metabolism of the neoplastic tissues on the total organism.
These changes express themselves simultaneously in morphologies
of histogenesis and in metabolism.
A full understanding of the broad range of the topic
requires a consideration of the framework of comparative
pathology in general. Unfortunately, the information available
on metabolic and other pathways of carcinogenic polynuclear
aromatic hydrocarbons in marine organisms is still rather
scarce. Historically, basic discoveries in this area have
involved primarily nonmarine organisms.
INTRODUCTION
Abnormal growth is a complex but fundamental problem of life.
Neoplastic growth is the most pervasive type of abnormal growth for man and
the surrounding environment. (The marine environment of concern to this
symposium is important for the future, affecting aspects of nutrition,
energy, resource recovery, transportation, and recreation.) The ocean as
the source of nutrition for the future concerns us indirectly with regard
to the life change of many inhabitant groups that serve as food for fishes
and for marine mammals, such as whales, sea cows, etc.; marine flora also
may undergo deleterious effects. These developments have serious economic
significance for man. Directly, we are concerned with the chain reactions
of toxicants in the ocean because of the possible contamination of food
that enters our own bodies. Many years ago it was noted that molluscs are
able to store carcinogenic hydrocarbons, for example, benzo(a)pyrene
(Cahnmann and Kuratsune, 1957). (Later we shall see differences in the
reaction of molluscs and mammalian integument to hydrocarbons.) Certain
organisms may only store hydrocarbons, others may metabolize them toxico-
logical ly with higher carcinogenic power than the parent substance, and
others may simply detoxify them. Thus we need to observe species-specific
-------
interaction and differences in this regard, as well as species reactions
with nonliving matter.
Regarding energy, v/e may distinguish between clean energy as that
produced from the tides, the waves, and different temperatures in varying
layers of the waters of the oceans and dirty energy, which refers to
resource recovery as it concerns deep-sea and shelf mining. The primary
sources of energy in all organisms come from the sun by organismic
conversion of sun rays and inorganic matter through the process of photo-
synthesis, making the plants the basic elements of all life functions.
It is impossible to understand pathological processes without a
preliminary understanding of normal processes. Comparative pathology is
based on three pillars: namely, comparative histology, comparative embry-
ology, and, of special concern to us, the normal distribution of chemical
compounds in the organisms.
SPECIES SPECIFICITY OF METABOLISM
In considering species specificity of metabolism, we shall see that
diseases can be seen from three aspects: (1) the organisms they attack,
(2) the causative agents or types of a disease, and (3) the life functions
impaired by a disease. Neoplasms, the most important diseases caused by
carcinogenic hydrocarbons that may also be mutagenic, are common (at least
theoretically) to all organisms with true tissues. If we look at the
comparability across taxonomic borders, we can state that the compounds of
interest to us in this symposium are able to produce neoplasms in a large
number of organisms in animals and plants alike. It is of interest to note
that we can compare morphological structures of animals to those of plants
if, for example, we equate the sex organs of the human female to such a
spring flower as the snow drop, more specifically in this regard, the
uterus in the human to the ovary in the plant. The glands in plants as
well as in animals are composed of secretory and supporting cells. Tissues
of both kingdoms in the whole sense are comparable, especially as concerns
the lining membranes, but there is also incomparability of certain tissue
groups. It has been well-known for many years that animal and plant cells
show certain differences but can be compared easily in general character-
istics. [It is of special importance to our understanding of the organisms
tissues we deal with today and their biochemical pathways in the cell that
it was possible for Jones and coworkers (1976) to implant plant plastids
into human Hela cells and that Lima-de-Faria (1981) was similarly able to
fuse human and plant cells or their parts.]
Carcinogenesis, and Chemical Carciogenesis in Particular, a Multistep
Process with or without Species-specific, Organ-specific, Tissue-specific
Variation
Berenblum (1974) stated: "A clue to the biochemical mechanism of
carcinogenic action can sometimes be derived from indirect evidence, e.g.,
by studying differences in species and organ response to carcinogenic
-------
action when these are associated with distinctive patterns of metabolic
changes in the causative agents."
This statement explains the intention to view comparative pathology on
as wide a background as possible. We can state as a rule "the wider
scientific valuable frame of comparative pathology, the deeper the under-
standing we gain." Today the attention of pathology has been shifted from
a science dealing with diseased organs to a science dealing with
biochemical, biophysical, and ultrastructural changes in diseased cells.
But there are not only the primary factors, such as a particular carcinogen
and cell constituents shaping the process of carcinogenesis,- but also many
factors, until now little known, which may be called secondary factors. In
experiments undertaken several years ago it could be shown that, for
example, a carcinogenic solution of 1% benzo(a)pyrene in an organic solvent
was fast expelled from different species of coelenterates, such as Tealia
felina, Sargatiogeton sp., Actina equina, and Metridium senile, and led to
an immediate secretion of the compound into other invertebrates, such as
the pulmonate, Arion subfuscus. It resulted in the building of resins and
simply a foreign body reaction. In the case of implantation of the
compounds, toxic results only could be attained in different species of
starfish; however, granuloma production was observed after two months of a
one-time implantation in the starfish, Sol aster papposus. It is our
opinion that temperature in the different body fluids and the solubility of
the carcinogen and different types of body fluids and circulatory systems
may play a more or less important role leading to direct negplastic
transformation, in addition to the primary factors in the interaction of
initiators, promoters, and the cell (Kaiser, 1965).
In the comparison of plants and animals, we should remember that the
two sources where most spontaneous neoplasms and growth anomalies have been
observed are in man himself and in the angiosperms. There are over 14,000
growth abnormalities, known as galls, limited and unlimited in the growth
potential, described in plants. There are tissues in plants, such as the
men'stems, not present in the vertebrate. These are the tissues with at
least theoretically unlimited growth potential during the life span of a
vascular plant. There are no muscular or nervous tissues in plants. To go
back once more to the meristematic tissues, we must state that the most
comparable tissue structures among animals occur in cirripe.d crustaceans
(larval tissues) or imaginal discs of such creatures as the seventeen-year
locust. Invertebrates also exhibit such developmental round-abouts as
placentae. Further, we know that embryonal development and tumor develop-
ment can run in parallel ways. The role of ontogenetic transformations in
larval developments, including normal histolysis and biochemical processes
and their changes during embryonal development, is a widely neglected field
of oncology, particularly in the area of biochemical pathways of
carcinogens. Marine organisms with placenta development are Sal pa primata
and Thalia democratica of the class Thaliacea, phylum Tum'cata.
-------
Contact of carcinogen with tissue--
Epithelium in animals and lining membranes in the plants are the
tissues in the organisms susceptible to contact with hydrocarbons, which
act in contrast to the aromatic amines, at the site of application.
The integuments and body coverings permit a broad comparison of how
contact occurs between the carcinogens and the organism investigated and
affected by the compound(s). Special emphasis must be put on the distri-
bution of the different tissue types which make certain variations in this
regard. Special emphasis has to be put on excretion of the tissues, such
as mucous layers, cuticles and exoskeletons, on one hand, and surface
specialization, such as villi and cilia, on the other. A direct effect of
the carcinogenic process on these structures can be distinguished from an
indirect one. A direct effect may play a role in the process or prevention
of carcinogenesis (Arcadi, 1977): 7,12-dimethylbenz(a)anthracene, a potent
carcinogen of papillary tumors in mice, did not produce skin tumors in the
mollusc, Lehmannia poireri. Such different effects may have something to
do with the mucous layer of the gastropod. We used the gill epithelium of
the bivalves, Um'o or Mytilus. which are ciliated, to check ciliastatic
components in cigarette smoke condensate. The phenolic compounds have been
found to be not only promoters in carcinogenesis but also potent
ciliastatic compounds. This side effect can be seen as an indirect
influence of an otherwise cocarcinogenic compound. The comparison is
possible because nearly all cilia in both kingdoms of the two-kingdom
approach are built similarly (Wynder et aj_., 1963).
Absorption—
In the second step, absorption, important general differences are
exhibited by animals and plants. Animals, with few exceptions, lack
nonliving tissues. However, plants have nonliving tissues, as well as
rigid cell walls, whereas the animal cell can be considered nude. Cuticles
occur in both groups, for example, the exoskeletons of crustaceans or the
wax cuticle on the leaf of red cabbage. The typical excreted cell wall of
the plant is composed of cellulose; in the case of fungi (especially the
higher ones), of chitin. Some cells of flagellates, lower fungi, and sex
cells are also nude in plants {in the two-kingdom approach). This is of
little importance because by definition such lower organisms exhibit no
true tissues necessary for neoplastic development. There is a significant
difference between animals and plants in that the plant needs a wound in
most cases for absorption of chemical carcinogens for reasons mentioned
previously. This is true for autonomous plant neoplasms, such as crown
gall disease, the wound tumor virus disease, or galls and neoplasms
produced by chemicals. The only exceptions are the genetic tumors in
plants. In addition to the species-specific differences of absorption,
regional differences exist also, for example, in the skin of mice. Further
complications are occasioned by such factors as circadian and other
rhythms.
The surface area of different species of plants and animals was
treated with a 1% solution of benzo(a)pyrene in acetone or DMS. The
treated skin area was removed at various time intervals. The surface area
was then exposed to 2 ma of acetone and injected in the 881 Perkin Elmer
5
-------
Gas Chromatograph and the Gary Model 15 Spectrophotometer. The absorption
rate of carcinogenic polycyclic hydrocarbons varies: (1) in animals a
stepwise absorption occurs, whereas in plants the main absorption takes
place at once; (2) differences in absorption do not occur in areas where
the strongest carcinogenic effect is produced. These facts suggest that
the fastest rate of absorption in plants, as in animals, does not
necessarily accompany the strongest carcinogenic effect. The metabolic
interaction of the carcinogen with the treated species at specific time
intervals appears to play a more important role (Theisz ^t _§!_., 1966).
Storing in tissue--
In the storing of chemical compounds in tissues, the time element
becomes a significant factor. The metabolic interaction between a
chemical compound and cell organelle of a particular species requires a
specific time elapse. For example, it is very hard to produce neoplasms
with chemical carcinogens in the frog, because the frog has a large
lymphatic system that eliminates carcinogens very rapidly (because of
enzyme influence, etc.). Diffusion rates can be species-specific in the
same site of injection, as shown with the carcinogenic hydrocarbon
7,12-dimethylbenz(a)anthracene. It remained for a long time in the
subcutis of mice and rats where it was carcinogenic, but was eliminated
very fast in rabbits (Berenblum, 1954).
Equally important is the time at which a carcinogen or a chemother-
apeuticum is applied to a particular species. The activity, amount, and
presence of such constituents as enzymes vary, thus playing a role in the
interaction, enabling the formation of pathways of the carcinogens, and
determining which compound may be actively carcinogenic in one species and
not in another, or only during a particular time interval (Edmunds and
Hal berg, 1981).
Time Factor of Application—
The biochemical reactions of different species can be affected by
chronobiological influences. This holds true in that the various types of
rhythmicity of species differ because of variations in life span, life
habits, food supply, environment, and other aspects. It need only be
realized that the mouse, for example, is a nocturnal animal, whereas man
is a day creature. However, direct comparison from chronobiology alone is
impossible if rhythmic differences are not included in the calculation.
Duration of Stay--
in general the period of chemical carcinogenesis can be divided thus:
the time interval during which the neoplastic transformation occurs and a
second phase, the progression from the first cancer cells to the tumor
development. Concerning the organismic background of our environment, the
chain reaction of different organisms and the interaction of morphological
and biochemical/biophysical aspects may be distinguished. The chain reac-
tions between different organisms can be positive or negative regarding
aflatoxin (Ferrando et al_., 1977): Diet including 20 p. 100 of lyophilized
milk produced by a goat that consumed peanut meals containing 1530, 79, and
-------
54 yg/kg aflatoxin with 1136, 64, and 54 yg/kg aflatoxin Bl was fed to
duckling for 23 days. Direct consumption of peanut meals, with the highest
level of aflatoxins, produced 18 p. 100 mortality and characteristic
injuries of aflatoxicosis.
Fischer and Horvath (1977) found that microvilli of Tubifex tubifex
Mii'll. of the supporting cells of the epidermis play a significant role in
repairing cuticular injuries. Acid mucopolysaccharides of the cuticle and
epidermis may function as traps for heavy metals, as proven by their
significantly heavy metal content. The cytosols of the epidermal cells
possess considerable DAB-reactivity. The enzyme, responsible for the
DAB-reaction, may be transported by the microvilli toward the cuticular
surface and can play a central role in the detoxication of organic foreign
compounds.
Species-specifc differences in enzyme systems involved in the metabol-
ism of carcinogens—The different possibilities of metabolic pathways of
chemical carcinogens exhibit more or less pronounced species specificity.
Weisburger and Weisburger (1958) found that the urinary excretion of the
various ring hydroxy metabolites of i4C-labelled AAF in the highly
responsive rat differed greatly from the nonresponsive guinea pig. The
7-OH metabolite dominated in the guinea1 pig and while the levels of the 1-,
3-, 5-, and 8-OH metabolites were greater in the rat. Hydroxylases convert
the polycyclic aromatic hydrocarbons into phenolic derivatives in the
microsome and endoplasmic reticulum of liver cells. A few biochemical
characteristics of hydroxylation of different aromatic hydrocarbons have
been demonstrated. Methylcholanthrene, administered by daily intraperi-
toneal injection to rat liver (40 rug/kg body weight), resulted in elevated
hepatic levels not only of malic enzymes but also of the pentose phosphate
pathway dehydrogenases (Shas and Pearson, 1978). A few metabolic
mechanisms of the biotransformation by direct oxidative alkyl chain
cleavage in the vicinity of the hydroxy group of (2-hydroxybutyl)-N-butyl-
nitrosamine were demonstrated by Blattman (1977). Acetaminophen is
metabolized through a variety of pathways, as shown by Thorgeirsson and
coworkers (1978). The prevention of benzo(a)pyrene-induced mutagenicity by
homogeneous epoxide hydratase was described by Oesch and coworkers (1976).
Differences in enzyme content and chronobiological rhythms also seem to
play a role in racial variations of breast cancer in women of different
countries as shown by Hal berg (1981).
Species-specific differences in chemical jruiding of carcinogens or
their metabolites with cell constituents—Carcinogens themselves or their
metabolites seem to act by a chemical mutation in DNA genes of the nucleus.
Differences exist among species regarding the binding capacity of chemical
carcinogens and also of the binding site with regard to varying
carcinogens. Variations also occur concerning the binding of carcinogenic
and noncarcinogenic compounds of the same chemical type.
It is not always compelling that the largest amount of a carcinogen in-
a particular organ may cause the largest amount of neoplasms. 1,2-dimethyl
hydrazine administered to rats showed that 96% brought with bile to the
intestine was less effective than the 4% received by the intestinal wall
7
-------
through circulation. The small amount of the compound metabolites played a
leading role in the genesis of intestinal tumor according to Pozharisski
and coworkers (1976). Comparable to two separate components in a
methanol-sodium borate solution gradient, as in the case of benzo(a)pyrene,
2-stereoisomeric diolepoxides are involved in the binding of
7,12-dimethylbenz(a)anthracene to DNA in culture cells. The binding of the
carcinogen does not decrease the overall metabolism of the carcinogen in a
remarkable way. Comparative studies of Roebuck and coworkers (1978)
demonstrated that no specific pattern of toxicity or carcinogenicity was
connected with the liver metabolism of aflatoxin by the duck, rat, mouse,
monkey, and humans.
Species-specific differences in biochemical changes in the cell durings
carcinogenesis—According to the large number of organisms with true tissue,
the species-specific variations of the action cell components with
carcinogens are very limited, especially for organisms of the marine
environment. We know absolutely nothing about the biochemical cellular
changes of most of the minor phyla, but it would be easy to do more experi-
mental work and gain interesting new results with these organisms and their
physiology. It would be necessary only to apply to these organisms the
methods used for terrestrial animals, such as the mouse and rat, or aquatic
animals, such as fish.
A few years ago, Harshbarger and coworkers (1971) investigated the
effects of carcinogen-contaminated water on crustaceans. More recently,
Khudoley (1977) studied the effect of diethyl- and dimethylnitrosamines
dissolved in tank water (200-400 ppm), which induced basophilic cell
neoplasms in 16 of 95 and in 6 of 17 molluscs of the species, Um'o
pictorum. Neoplasms occurred in the hepatopancreas.
Methylnitrosoguanidine produced only inflammatory reactions at the
injection site. It is suggested that these invertebrates be used as
indicators of hydrospheric pollution with chemical carcinogens. A later
paper reported the cellular transformations of polycyclic aromatic
hydrocarbons. The interaction of N-acetoxy-N-2-acetyl- aminofluorene and
its binding to the regions of chromatin sensitive to enzymes in duck
erythrocytes were reported by Metzger and coworkers (1976); tumor growth in
molluscs was reviewed by Khudoley and coworkers (1977).
Variable pathways of same carcinogen in the same tissue of the same
organism—Similar to normal chemical compounds, carcinogens are also
metabolized in different ways, sometimes in the same species. Two
cytochromes, for example, from liver microsomes of the rabbit treated with
2,3,7,8-tetrachlorodibenzo(p)dioxin appear to be distinct entities and
function in different catalytic pathways (Johnson and Muller-Eberhard,
1977).
The main urinary metabolite of safrole in the rat and man was. excreted
in a conjugated form. Small amounts of eugenol or its isomer l-methoxy-2-
hydroxysafrole, approximate carcinogen of safrole, and S'-hydroxyisosafrole
8
-------
were detected as conjugates in the urine of the rat. However, Benedetti
and coworkers (1977) were unable to demonstrate the presence of the latter
metabolites.
Species-specific ways of detoxication and excretion of carcinogens--
In a more general way, the chemical carcinogens can be considered cell
toxins. There exist, of course, different ways of detoxification, particu-
larly among the large number of species in the marine environment. Also
the smaller phyla have been neglected here despite the fact that their use
could reveal quite valuable information.
It is the main belief of the public that the cancer or the neoplasm in
general is one type of disease, and the tumor itself is its focal point.
In one respect this is true, but in another it is not, because the primary
tumor is only one part of the process known as neoplastic disease.
MORPHOLOGICAL CHANGES AS A PARALLEL PROCESS OF CARCINOGENESIS IN DIFFERENT
SPECIES USING SELECTED EXAMPLES
To complete our review, it is necessary to take a brief look at the
morphological changes during carcinogenesis. We must keep in mind that
biochemical changes and morphological changes are only two versions of the
same process that run simultaneously but are observed with different
methods. It was necessary to describe each separately.
The Change from Normal, Sometime Via Metaplastic, to Neoplastic Tissue in
situ
The integument of many invertebrates is characterized by a simple
columnar epithelium which is very often ciliated. The largest group of
invertebrates, the insects, lack cilia. One of us (Kaiser) participated
for many years in the investigation of the transformation of normal
columnar ciliated epithelium of the human bronchus to its metaplastic
stage to nonciliated squamous cell epithelium, and finally to the neoplasm
in situ. As stated before, cilia are built similarly in nearly all
organisms, plants and animals alike. We therefore selected the change in
the columnar ciliated epithelium, which occurs during carcinogenesis.
The Development of the Neoplastic Tissue in situ to the Primary Neoplasm
The development of the neoplastic cells in situ is the first stage of
the primary tumor and a crucial phase in the whole process of tumor
development. It is the stage in which the barrier of basement membranes,
in some cases, and the border to other tissues have to be broken. This
means chemically a new adaptation. Simultaneously one of the main charact-
eristics of the malignant neoplasms is established, that of infiltrative
growth. It is also a turning point if compared with the characteristics of
benign neoplasms, which lack the capability of infiltrative growth, and are
generally characterized by a surrounding stroma capsule instead. On the
other hand, it must be stated that the neoplastic cell in situ already has
the cellular malignant characteristics.
-------
Metastatic Developmenmt of Animal Neoplasms Only
For comparative purposes, we should take a short view of metastatic
development and the most important group-specific differences. This is
necessary to study and understand the various groups of marine organisms.
(1) All malignant plant tumors lack metastatic growth. This needs to
be associated with two characteristic facts of plant cells and
tissues:
(a) The plant cell is characterized by a rigid cell wall.
(b) No floating cells occur or are able to survive in the body
fluid of vascular plants.
(2) Each metastatic growth in an organism is dependent on a circula-
tory system. Von Albertini (1974) distinguished between
hematogenic and lymphogenic and three special types of metastasis
in the human. In marine organisms, the hematogenic metastasis is
the one of importance which, of course, will be modified by the
type of circulatory in the different animal groups. Our knowledge
of the manner in which tumors spread has increased (Kaiser, 1981,
1982). We also know more today about the multistep processes.
Our expanded knowledge is also reflected in a better understanding
of the sequence of the events in the metastatic process and its
heterogeneity (Folkman, 1982; Fidler and Hart, 1982).
SUMMARY AND CONCLUSIONS
As we look backward and also forward, we see that neoplastic growth in
organisms is a complex process which can be separated into several phases,
seen from a morphological as well as a biochemical (metabolic) point of
view. These different phases run over a species- and tumor-specific length
of time. It is possible to distinguish the external and internal
environments of living matter. The different portions of the external and
internal environments interact with each other, as can be seen in certain
characteristic chemical processes. Among the most important and best
studied groups of chemical carcinogens are the polynuclear aromatic
hydrocarbons. The pathways of their metabolism were studied mainly in
terrestrial animals. The marine organisms invite additional study because
the diversification of the specific structures and biochemistry could shed
tremendous new insight. In addition, marine organisms are and can be
excellent indicators of pollution in the largest environmental unit on
earth, which is so important to the future of man and his environment. But
to reach this goal, we must increase our knowledge of toxicants in the
marine environment. The compounds of concern in this symposium are among
the most significant and challenging groups. Of highest importance are the
species-specific ways of transformation of these compounds in marine
organisms as seen under comparative illumination. Participants in this
symposium will inform us of the newest scientific research in the field, in
the sense of Aristotle's kathekaston.
10
-------
REFERENCES
Arcadi, J.A. 1977. Apparent resistance of the integument of the
invertebrate, Lehmanm'a poireri to production of papillary tumors by a
known chemical carcinogen. J. Surg. Oncol. 9(1):87-91.
Benedetti, M.S., E.A. Malno, and A.L. Broillet. 1977. Absorption,
metabolism and excretion of safrole in the rat and man. Toxicology
7(l):69-83.
Berenblum, I. 1954. Carcinogenesis and tumor-pathogenesis. Adv. Cancer
Res. Vol. II, pp. 129-175.
Berenblum, I. 1974. Carcinogenesis as a biological problem. Amsterdam:
North-Holland Publishing Co., American Elsevier Publishing Co., Inc.,
New York.
Blattmann, L. 1977. Direct alkyl chain cleavage after C-hydroxylation of
dialkylnitrosamines in rats. A new pathway of the oxidative
biotransformation. Z. Krebsforsch. 88(3):315-322.
Cahnmann, H.J., and M. Kuratsune. 1957. 'Determination of polycyclic
aromatic hydrocarbons in oysters collected in polluted water. Anal.
Chem. 29:1312-1317.
Edmunds, L.N., and F. Halberg. 1981. Circadian time structure of euglena:
a model system amenable to quantification. In: Neoplasms—comparative
pathology of growth in animals, plants, and man. H.E. Kaiser, Ed.,
Williams and Wilkins, Baltimore, MD. pp. 105-134.
Ferrando, R., A. Parodi, N. Henry, J. Delort-Lavel, and A.L. N'Diaye.
1977. Milk aflatoxin and transfer of toxicity. C. R. Acad. Sci. (D)
Paris 284(10):855-858.
Fidler, I.J., and T.R. Hart. 1982. Principle of cancer biology: biology
of breast cancer metastasis. In: Cancer principles and practice of
oncology. V.T. di Vita, Jr., S. Hellman, and S.A. Rosenberg, Eds.,
J.B. Lippincott, Philadelphia-Toronto, pp. 80-92.
Fischer, E., and I. Horvath. 1977. Cytochemical studies on the cuticle
and epidermis of Tubifex tubifex Mull. With special regard to the
localization of polysaccharides, heavy metals and the DAB-reactivity.
Histochemistry 54(3):259-271.
Folkman, J. 1982. Tumor invasion and metastasis. In: Cancer medicine,
2nd edition. J.F. Holland and E. Frei III, Eds., Lea and Febiger,
Philadelphia, PA. pp. 167-177.
11
-------
Halberg, F. 1981. International geographic studies of oncological
interest on chronobiological variables. In: Neoplasms— comparative
pathology of growth in animals, plants, and man. H.E. Kaiser, Ed.,
Williams and Wilkins, Baltimore, MD. pp. 553-604.
Harshbarger, J.C., G.E. Cantwell, and M.F. Stanton. 1971. Effects of
N-nitrosodimethyl amine on the crayfish, Procambams clarkii. Proc.
4th International Colloquium on Insect Pathology, Society for
Invertebrate Pathology, Aug. 25-28, 1970.
Johnson, E.F., and U. Muller-Eberhard. 1977. Resolution of two forms of
cytochrome P-450 from liver microsomes of rabbit treated with 2,3,7,
8-Tetrachlorodibenzo(p)dioxin. J. Biol. Chert. 252(9):2839-2845.
Jones, C.W., I.A. Mastrangelo, H.H. Smith, H.Z. Liu, and R.A. Meek. 1976.
Interkingdom fusion between human (Hela) cells and tobacco hybrid
(GGLL) protoplasts. Science 193:401-403.
Kaiser, H.E. 1965. Artspezifische Untersuchungen ueber die Carcinogenese.
1. Mitt.: Untersuchungen ueber die Reaktion nach Injektion,
Implantation and Verfuetterung von polycyclischen Kohlenwasserstoffen
bei Coelenteraten und Echinodermen (Studies on species-specific
carcinogenesis in different phyla). 1. Studies regarding the reaction
on coeleneterates and echinoderms after implantation and feeding of
polycyclic hydrocarbons). Arch. f. Geschwulstforsch, Berlin
25(2):118-121.
Kaiser, H.E. 1980. Species-specific potential of invertebrates for
toxicological research. University Park Press, Baltimore, MD 224 p.
Kaiser, H.E., Ed. 1981. Neoplasms—comparative pathology of growth in
animals, plants, and man. Williams and Wilkins, Baltimore, MD 908 p.
Kaiser, H.E., Ed. 1982. Progressive stages of neoplastic growth (in
preparation).
Khudoley, V.V. 1977. Tumor induction by carcinogenic agents in anuran
amphibian, Rana temporaria. Arch. Geschwulstforsch. 47(5):385-399.
(Rus.)
Lima-de-Faria, A. 1981. Fusion of human cells with plant protoplasts and
its implications for cell differentiation. In: Neoplasms—comparative
pathology of growth in animals, plants, and man. H.E. Kaiser, Ed.,
Williams and Wilkins, Baltimore, MD. pp. 441-450.
Metzger, G., F.X. Wilhelm, and M.L. Wilhelm. 1976. Distribution along DNA
of the bound carcinogen N-acetoxy-N-2-acetylaminoflurene in chromatin
modified in vitro. Chem. Biol. Interact. 15:257-265.
Oesch, P., P. Bentley, and H.R. Glatt. 1976. Prevention of benzo(a)pyrene
induced mutagenicity by homogeneous epoxide hydratase. Int. J. Cancer
18(4)448-452.
12
-------
Pozharisski, K.M., I.A.D. Shaposhnikov, S.A. Petrov, and A.I.A. Likhachev.
1976. Distribution and mechanism of the carcinogenic action of
1,2-dimethylhydrazine in rats. Vopr. Onkol. 22(5):48-53. (Rus.)
Roebuck, B.D., W.G. Siegel, and G.N. Wogan. 1978. In Vitro metabolism of
aflatoxin B2 by animal and human liver. Cancer Res. 38(4):999-1002.
Shas, S., and D.J. Pearson. 1978. The effect of phenobarbitone on
cytoplasmic NADP-1inked dehydrogenase activities in rat liver.
Biochem. Biophys. Acta 539(1):12-18.
Theisz, E, H.E. Kaiser, and J.C. Bartone. 1966. Species-specific
differences of absorption as a variation of stage one (contact) of
multi-phase carcinogenesis. Presented at the annual meeting of the
Virginia Academy of Sciences. (Unpublished.)
Thorgeirsson, S.S., S. Sakai, and R.H. Adamson. 1978. Induction of mono-
oxygenases in rhesus monkeys by 3-methylcholanthrene: metabolism and
mutagenic activation of N-2-acetylaminofluorene and benzo(a)pyrene.
J. Natl. Cancer Inst. 60(2):365-369.
Von Albertini, A. 1974. Histologische Geschwulstdiagnostik, 2nd edition,
Georg Thieme Verlag, Stuttgart.
Weisburger, E.K., and J.H. Weisburger. 1958. Chemistry, carcinogenicity
and metabolism of 2-fluorenamine and related compounds. Adv. Cancer
Res. 5:331-431.
Weisburger, E.K. 1981. Species-specific biochemical pathways of malignant
growth. In: Neoplasms—comparative pathology of growth in animals,
plants, and man. H.E. Kaiser, Ed., Williams and Wilkins, Baltimore,
MD. pp. 335-350.
Wynder, E.L., H.E. Kaiser, D.A. Goodman, and D. Hoffman. 1963. A method
for determining ciliastatic components in cigarette smoke.
Cancer pp. 1222-1225.
13
-------
NEGATIVE CHEMICAL IONIZATION MASS SPECTRA OF SOME
POLYNUCLEAR AROMATIC HYDROCARBONS
by
Ralph C. Dougherty and Stephanie V. Howard
Department of Chemistry, Florida State University
Tallahassee, Florida 32306
and
Joseph D. Wander
Department of Chemistry, The University of Georgia
Athens, Georgia 30602
Abstract
Negative chemical ionization (NCI) mass spectra of 9
ortho-fused, 8 ortho and peri-fushed, and 13 methylated
polynuclear aromatic hydrocarbons (PAHs) were obtained using
isobutane as the reagent gas. The isobutane-mediated NCI
(NCI-IB) spectra of all 30 PAHs were characterized by a small
number of abundant, large-mass ions, suitable for qualitative
or quantitative detection. In general, initial processes of
resonance electron capture or anion attachment account for the
principal ions formed by the unsubstituted PAHs, whereas
dissociative capture resulting in loss of a hydrogen atom
produces the most abundant ion in the NCI-IB mass spectrum of
the alkylated PAHs. For most of the compounds studied, NCI-IB
mass spectra produced in a hot (235°) ion source were better
suited for high-sensitivity measurements than the corresponding
spectra produced in a cooler (125°) source. NCI mass spectra
do not provide information concurring isomeric structures,
e.g., it would not be possible to distinguish benz(a)pyrene
from benz(e)pyrene; however, the spectra reported indicate
substantial potential for use of NCI mass spectrometry in
screening environmental extracts for contamination with PAHs
and their derivatives.
INTRODUCTION
Negative chemical ionization (NCI) mass spectra are obtained, by
operating a mass spectrometer with a high pressure (approximately one
torr.) ion source in the negative ion mode. Spectra obtained under these
conditions appear to be uniquely suited for screening extracts of
14
-------
environmental substrates for contamination with toxic substances. The
reasons for this suitability stem from the fact that most man-made toxic
substances are either oxidizing agents or alkylating agents, whereas most
biomolecules are highly reduced and have large numbers of high energy
electrons. Oxidizing or alkylating agents universally produce intense
negative ion mass spectra, either through electron capture or through
molecule-anion association. With the exception of free carboxcyclic acids
and some of the prosphetic groups from the electron transport chain,
biomolecules in contrast produce only very weak negative ion spectra. Thus
it is possible to obtain molecularly specific information concerning the
nature and the abundance of toxic substances in a matrix that contains a
significant quantity of biomolecules in the extract (Dougherty and
Piotrowska, 1967a, 1967b; Dougherty and Hett, 1978).
Negative chemical ionization mass spectra using methylene chloride as
the reagent gas have been used to screen human urines and seminal fluid
(Dougherty and Piotrowska, 1976a) and items from the food chain (Dougherty
and Piotrowska, 1976b) for contamination with polychloroinated organics.
NCI mass spectra can reliably detect polycyclic insecticides such as
Mirex (Dougherty et jtl_., 1976), industrial chemicals like polychloro-
biphenyls [Dougherty et^al_., 1974), and polychlorodioxins (Mass et a!.,
1978) in extracts of environmental substrates.
Virtually all PAHs have positive electron affinities (Compton and
Huebner, 1970) and should produce intense, negative ion mass spectra. In
order to increase the utility of NCI screening for toxic substances, we
have obtained the NCI mass spectra of a series of polynuclear aromatic
hydrocarbons using isobutane as the reagent gas. The use of hydrocarbon
reagent gases is a simple extension of pressure enhanced negative ion mass
spectra which was first applied to aromatic hydrocarbons (Mass et al.,
1978).
The capture of thermal electrons occurs in low pressure mass
spectrometers under conditions of electron impact ionization. This is
generally a minor process at low source pressure because the reactions are
sharply peaked at very low electron energies. Introduction of an inert gas
or hydrocarbon gas to increase the pressure in the source can considerably
enhance the population of low energy electrons and thus the formation of
negative ions by resonance capture processes.
The capture of slow electrons is not only the process by which
negative ions can be formed in a mass spectrometer. Electrons with
energies between thermal energy and roughly ten electron volts are often
captured by molecules with resulting fragmentation; the process is called
"dissociative electron capture" (Reaction 2). The products of this
reaction is an even-electron ion and a radical. Ionization of the reagent
gas by resonance capture or disassociative capture may afford a population
of anions within the source, and these anions may combine with the sample
molecule to produce a negatively charged adduct (Reaction 3).
15
-------
AB + e~ + AB-* (1)
slow
AB + e" -> A- + 8* (2)
AB + C" + ABC- (3)
Anions or molecule anions which contain considerable excess energy
usually will be converted rapidly to neutrals by ejection of a electron.
This means that the anions that one observes in NCI mass spectra must have
low internal energies, which severely restricts the amount of fragmentation
that can occur within the molecule. Thus negative ion mass spectra are
essentially dominated by molecule anions, the products of disassociative
capture, and anion-molecule reactions.
The ability of NCI mass spectra to specifically detect traces of
halo-organics or PAHs in the presence of biomolecules is exemplified by the
camparison of the isobutane-mediated NCI mass spectra of methylstearate and
hexahelicene (Figure 1). Under our instrumental conditions, the isobutane
mediated NCI mass spectrum of methylstearate contains no ions whatsoever at
the highest sensitivity of the instrument. The NCI mass spectrum of
helicene, on the other hand, contains essentially only one ion, the
molecule anion, and the sensititivy of the technique is such that only one
nanogram of hexahelicene is sufficient to produce a recognizable spectrum.
In this paper we report the hydrocarbon mediated NCI mass spectra for
30 different PAHs at two temperatures.
EXPERIMENTAL
NCI mass spectra were recorded using an AEI MS-902 double focussing
mass spectrometer fitted with an SRIC chemical ionization source and an
external-8 kV power supply. Research grade isobutane (Matheson) was
introduced into the source through a needle valve at a rate adjusted to
provide a pressure in the source of one torr (130 Pa); no effort was made
to exclude adventitious oxygen. The source was maintained at 125 _+ 5° for
the first series of determinations and at 235 +; 5° for the second. PAHs
were placed in freshly fired quartz cuvettes and introduced into the source
on a direct insertion probe; heat was applied to those samples that did not
vaporize spontaneously. Mass spectra were recorded using a light-beam
oscillograph.
RESULTS AND DISCUSSION
Table 1 presents the principal ions in the isobutane-mediated negative
chemical ionization mass spectra of the following polynuclear hydrocarbons
at the source temperature: anthracene (1); phenanthrene (2); benzo(ghi)-
fluoranthene (3); benz(a)anthracene (4); benzo(c)phenanthrene (5); chrysene
(6); benzo(a)pyrene (7); benzo(e)pyrene (8); perylene (9); benzo(ghi)-
perylene (10); benzo(b)triphenylene (11); dibenz(a,h)anthracene (12);
coronene (13); dibenzo(def,p)chrysene (14); dibenzo(a,i)pyrene (15);
hexahelicene (16); and dibenzo(g,p)chrysene (17).
Ib
-------
TABLE 1 RELATIVE INTENSITIES OF PRINCIPAL IONS IN THE ISOBUTANE-MEDIATED NEGATIVE CHEMICAL IONIZATION
(NCI-IB) MASS SPECTRA OF PAHs 1-17, DETERMINED AT SOURCE TEMPERATURES OF 125 ± 5° and 235 ± 5°
PAH Mol. Formula
/ 1 \ X^^*"-^^*^* 'i^>i (* U
OOJOJ 14 10
Anthracene
(2) pJ© C14H10
•prnnnthrrn*
©lr
benzo[B]iiJflgoranthene
b«nc[u]ant.hr«ccne
(5) (Ql C.-.H.,
\-X^ lO \.£
if M-H " MC-H ~ M07T (M-H,)~
(m/z; I1258*) (m/z; I125"a) (m/z; I125°a) (m/z; I125°a) (m/z; I125°a)
I235° I235° I235° T235° $235°
178; 100 177; 3 193; 19 210;10
100 19
178; 177; 193; 210;
55 55 100
226; 100 225; - 241; 6 258; -
100 - 2
228; 100 227; - 243; 3 260; 42
100 4 18 2
228; 2 227; 10 243; 100 260: 55 226; 27
11 34 19 3 100
btn*o[c}ptienanthr«ne
-------
TABLE 1. (CONTINUED)
(13)
C24H12
300;
100
299;
315} 332;
21
(14)
C24H14
302; 100
100
301; -
317; 3
2
334; -
(15)
(16)
fllbrnt|>,l)pyrin>
24 "14
C26H16
302;
100
328; 100
100
301;
327; -
3
317;
343; 10
12
334;
360; 2
(17)
C26H16
d 1 tamo 11, j>J ehrr»n t
328; 100
100
327; 2
12
343; 8
9
360; 2
1
'Relative intensity at source temperature of 125 ± 5" and at 235 ± 5% respectively; a dash indicates that the ion is not
observed at significant intensity, and number represent percent of the intensity of the most-abundant ion (base peak).
-------
(6)
TABLE 1. (CONTINUED)
C18H12
228;
14
227;
28
243;
100
260;
226;
(7)
©S)
Iftlplol
C20HI2
252; 100
100
251; -
267; -
3
284; -
^
©
C20H12
252; 100
30
251; -
28
267; 3
36
284; 92
4
250; -
100
(9)
C20H12
252; 100
251; -
267; 7
284; -
(10)
$s
B3&
C22H12
276; 100
100
275; -
291; 5
3
308; -
(11)
bcnt(i|s,h.ljpcryl.nc
bcnto(h]crlphonylene
C22H14
C22H14
278; 100
100
278; 100
100
277; -
34
277; 13
4
293; 2
72
293; 19
25
310; 54
310; 45
3
41benc|a,h)ant;hrac«n«
-------
328
M
360
MO-
343
MO-H
Figure 1. Isobutane mediated NCI mass spectrum of hexahelicene (16),
measured at a source temperature of 120°.
The spectrum of hexahelicene (16, Figure 1) is typical of the members
of_this class, resonance captre (Reaction 1) produced the most abundant ion
(M*) at m/z 328; less-abundant ions form by capture of an electron and an
oxygen molecule (Reactions 1 and 3) to give M02-, m/z 360, which
subsequently loses the elements of an hydroxyl radical to form a fragment
ion that may be represented as an oxy analogue of the parent PAH,
(M-H+0)", m/z 343. Dissociative capture (Reaction 2) contributes
negligibly to the net ionization, even at the higher source temperature.
Compounds 4,5,8,11, and 12 exhibit a much stronger tendency to form
M02T in the cool source, thereby decreasing the amount of M* in the
spectrum. Compounds 2,11, and 13 appear to be relatively inefficient at
capturing electrons and gave relatively weak responses. Compounds 2,5, and
6 did not produce major molecule anions, although each did form one
principal ion which was characteristic of the particular PAH and suitable
for quantitative measurement. The mass spectrum of 12 also contained a
metastable ion at m/z 249.5, suggesting that at least a fraction of the
M~ ion (m/z 278) formed by loss of an oxygen molecule from M02* (m/z
310), which presumably formed by capture of Q^^ (Reaction 3).
Table 2 presents the masses and relative intensity data for the
principal ions in the isobutane NCI mass spectra of the following
methylated PAHs: 4,5-dimethylphananthrene (18); 4-methylpryene (19),
3,4,5,6-tetramethylphenanthrene (20); 1-methylbenzo(c)phenanthrene (21);
2-methylbenzo(c)phenanthrene (22); 3-methylbenzo(c)phenanthrene (23);
4-methylbenzo(c)phenanthrene (24); 5-methylbenzo(c)phenanthrene (25);
6-inethylbenzo(c)phenanthrene (26); 7,12-dimethylbenz(a)anthracene (27);
-------
TABLE 2. RELATIVE INTENSITIES OF PRINCIPAL IONS IN THE ISOBUTANE-MEDIATED NEGATIVE CHEMICAL IONIZATION
(NCI-IB) MASS SPECTRA OF PAHs ig-jO. DETERMINED AT SOURCE TEMPERATURES OF 125 * 5° amd 235 +. 5°
PAH
Mol. Formula (m/z; I125°a)
M-H
(m/z; I125"a)
'235
MO-H
MO,
(m/z ;
(m/z ;
Other Ions
235
(19)
(20)
(27)
C16H14
Col:
206; 13
205; 49
221; 100
238; 21
IOl8l C17H12 216; 46
©©;
. xjx C H 234; 39
1 1O] 22
3,* , 5 ,6-t«t raaethytptManthrtn*
xX-^s. c H 242-23
Try) 19 14 tnt,*-3
S~^S^y 27
Ȥk C19H14 M2; 12
TO] ly " 40
215; 100
100
233; 100
100
241; 100
100
241; 57
100
231; 40
10
249; 50
10
257; 47
32
257; 26
29
248; 76
4
266; 37
274; 34 297; -
9 53
274; 100 297; -
56
272; -
231
2-»«thjrl benzo (c) pheo*athren«
-------
TABLE 2. (CONTINUED)
242; 62
69
242; 25
19
241; 74
100
241; 100
100
257; 25
53
257; 82
25
274; 100 297; 20 272; 1
5 163
274; 97 297; 31 272; -
7-20
256; 26
23
ro
242; 25
21
241; 100
100
257; 42
12
274; 48 290; 18 256; 10
11 - 14
(26)
C19H14
242; 21
24
241; 100
100
257; 90
19
274; 63 297; 40 256; 19
13 - 22
256; 100
100
255; -
71
271; 1
13
288; -
2
C21H16
e hoi *o tbr«n«
268; 14
267; 100
27
283; -
2
300; - 266; 92
-------
TABLE 2. (CONTINUED)
(29)
C22H20
i.5,t,12~t«tri«thrlb«iio[e)phm«iithr*
284; 28
31
283; 100
100
299; 27
23
316; 39
20
(30)
C27H18
342; 32
100
341; 100
20
357; 11
14
374; 2
3
340; 21
20
IN:
OJ
See footnote a, Table 1.
Line(s) projecting from the aromatic nucleus indicate the location of the methyl substituent (s).
-------
3-methlcholanthrene (28); l,5,8,12-tetramethylbenzo(c)phenanthrene (29);
and 7-methylhexahelicene (30), at source temperatures of 125 _+ 5° and 235 +_
5°. The spectrum of 7-methylhexahelicene (30) is typical of that of most
members of this group, having as the base peak the product of dissociative
capture (Reaction 2) with loss of hydrogen atom to give the ion represented
as [M-H]~, at m/z 341. Most of-the signal observed at m/z 342 derives
-from [M-H]~ ions containing a C aton and only a small amount of the
molecular anion (M7) is formed. M02~ and (M-H+0)" are also minor ions
in the spectrum. The mass spectrum of 30 was relatively insensitive to the
temperature of the ion source.
Conversely, the spectra of the isomeric methylbenzo(cjphenanthrenes
(21-26) illustrate a striking dependence on source temperature. At 235°,
the spectra of all six isomers are substantially identical, whereas the
corresponding spectra measured at 125° exhibit gross variations in relative
intensities for different isomers. The latter observation is of potential
interest for structural studies, but for analytical purposes, a spectrum in
which a single, characteristic ion preponderates offers greater sensitivity.
The (M-H)~ ion was comparatively minor in the NCI spectra of 18, 27,
and 28, although each gives one major ion. Loss of Ho from M" occured
extensively in 28, the (M-2H)~ ion being the base peak in the hot source.
Dehydrogenation of 28 would yield a PAH containing a cyclopentadieneide ring
which should be a very stable anion. The higher temperature spectrum of 27
exhibited a large increase in the intensity of (M-H)~, which would be
expected if the disassociative capture reaction were thermally activated.
The relative amount of MO?1 was observed to decrease as the sample
remained in the ion source. Trie use of a reagent gas mixture containing a
fixed proportion of oxygen stabilized the relative proportions of products,
e.g. M02-, formed subsequent to ionization.
Two other molecules examined in this study of these picene failed to
produce ions presumably because of their involatility. At either source
temperature, the NCI spectrum of triphenylene (M.W. 228) produced only m/z
252, which cannot be accounted for as simply as the ions formed from 1-30,
but is nonetheless quite suitable for quantitative detection of trace
amounts of triphenylene.
CONCLUSIONS
The isobutane NCI mass spectra of 30 PAHs show real promise for the
development of screening methods of these compounds in environmental
substrates. The spectra were uniformly intense-roughly nonogram sample
requirements and molecularly specific. The spectra of alkylated PAHs were
generally dominated by (M-H)~ ions while those of the parent hydrocarbons
were dominated by molecule anions.
24
-------
ACKNOWLEDGEMENTS
Support for this work was provided by the National Institutes of
Health and the U.S. Environmental Protection Agency (R806334).
REFERENCES
Compton, R.N., and R.H. Huebner. 1970. Collisons of low-energy electrons
with molecules: threshold exitation and negative ion formation. Adv.
Rad. Chem. 2:281.
Dougherty, R.C., and C.R. Weisenberger. The negative ion mass spectra of
benzene, naphthalene and anthracene: a new technique for obtaining
relatively intense and reproducible negative ion mass spectra. J.
Am. Chem. Soc. 90:6570.
Dougherty, R.C., and J.D. Roberts, H.P. Tannenbaum, and P.O. Bivos. 1974.
Positive and negative chemical ionization mass spectra of polychlorinated
pesticides. In: Mass spectrometry and MMR spectroscopy in pesticide
chemistry. F.J. Biros and R. Hague, Eds., M. Dekker, New York, p. 33.
Dougherty, R.C., and K. Piotrowska. 1976a. Screening by negative chemical
ionization mass spectrometry for environmental contamination with toxi
residues: application to human urines. Proc. Nat. Acad. Sci., USA
73:1777.
Dougherty, R.C., and K. Piotrowska. 1976b. Multiresidue screening by
negative chemical ionization mass spectrometry: polyhal-organics.
J.A.O.A.C. 59:1023.
Dougherty, R.C., A. Bergner, P. Levonowich, and J.D. Roberts. 1976c.
Positive and negative chemical ionization mass spectra for pesticide
screening. Adv. Mass Spectrom. Biochem. Med. 1:181.
Dougherty, R.C., and E.A. Hett. 1978. Negative chemical ionization mass
spectrometry: applications in environmental analytical chemistry.
Environ. Sci. Res. 12:339.
Hass, J.R., M.D. Friesen, D.J. Harran, and C.E. Parker. 1978.
Determination of dibenzo-p-dioxins in biological samples by negative
chemical ionization mass spectromety. Anal. Chem. 50:1474.
25
-------
A CHARACTERIZATION OF THE POLYCYCLIC AROMATIC HYDROCARBON
CONTENT OF TARS, TARBALLS, AND SEDIMENTS FROM THE MARINE ENVIRONMENT
by
James L. Lake, Curtis B. Norwood, and Crandall W. Dimock
U.S. Environmental Protection Agency, Environmental Research Laboratory
Narragansett, RI 02882
ABSTRACT
Glass capillary column gas chromatography (GC) and gas
chrornatography-mass spectrometry (GC-MS) were used to
characterize the hydrocarbons present in samples of tars,
tarballs, and sediments from the marine environment. Emphasis
was directed toward the analysis of polycyclic aromatic hydro-
carbons (PAHs) because of their known toxicity. The relative
abundances of alkylated v^s_ non-alkylated PAHs and GC patterns
obtained from the tarballs indicated that tarball samples
collected on the Brittany Coast of France were
petroleum-derived, whereas those obtained on a Rhode Island
beach were formed from coal tar.
Analyses of PAHs in sediments and in coal tar from docks
included determinations of (1) the relative content of
non-alkylated PAH parent molecules, i.e. parent compound
distributions (PCDs); (2) alkylation patterns of these PAH
molecules, i.e., alkyl homolog distributions (AHDs); and (3)
phenanthrene/anthracene (P/A) ratios. Comparisons of these
measurements demonstrated that the sediments surrounding tarred
piers were contaminated by coal tar used to coat the pilings;
however, the PAH assemblages in samples of sediment from the
middle of Narragansett Bay reflected a different source.
INTRODUCTION
Concern about the toxicity of polycyclic aromatic hydrocarbons (PAHs)
has resulted in research to determine the sources and fates of these com-
pounds in the environment. Researchers have characterized PAH assemblages
and investigated the spatial distribution of PAH compounds in samples of
sediments from the marine environment (Youngblood and Blumer, 1975;"Hites
and Biemann, 1975; Hase and Hites, 1976; Hites, 1976; Hites et a\_., 1977;
26
-------
Farrington et_ jil_., 1977; LaFlamme and Kites, 1978; Windsor and Hites,
1979). These studies found that the PAH assemblages in marine sediments
contained numerous non-alkylated (parent) PAH compounds and the alkyl
homologs of these parent compounds. The PAH alkyl homolog distributions
found in marine sediments have been compared with those in airborne
particulate material (Hase and Hites, 1976), combustion products of fossils
fuels (Lee et _al_., 1977; Hase and Hites, 1976), petroleum (Youngblood and
Blumer, 197*5]", and water (Hase and Hites, 1976). Studies have investigated
the probable origins of PAHs found in marine sediments in New England
(Windsor and Hites, 1978; Lake et_ jj]_., 1979) and in other areas (LaFlamme
and Hites, 1978). The determination that the PAH content of marine
sediments rapidly decreased with increased distance from population centers
(Windsor and Hites, 1979; Lake _et _§]_., 1979) indicated that most of the
PAHs in these sediments came from anthropogenic inputs.
While the majority of PAHs present in marine sediments were believed
to result form the deposition of aeolian transported fossil fuel combust-
ion products (LeFlamme and Hites, 1978), or in some instances, from a com-
bination of inputs from combustion and petroleum (Lake et a1_., 1979),
smaller localized inputs of PAH may have environmental significance. To
further characterize PAH inputs to the marine environment, we examined the
hydrocarbon contents of tarballs, coal tar used in marine construction, and
marine sediments.
METHODS
Samples of sediments taken near piers in the Narragansett Bay and in
New Harbor on Block Island, RI, were collected with a Ponar grab sampler.
The top 10 cm of an "undisturbed" portion of the sediment were retained for
analysis. A sediment sample was obtained near the north end of Jamestown
Island in mid-Narragansett Bay by a diver using a plexiglas core tube.
All sediment samples were returned to the laboratory as quickly as
possible and analyzed immediately or frozen at -20° C until analysis.
Prior to analysis of the top 10 cm section of the core, the sediment con-
tacting the core liner was scraped off.
The extraction and clean-up procedures used on sediment samples have
been described (Lake et a]_., 1980). Briefly, this method consists of a me-
thylene chloride reflux extraction followed by separation of the extract
into F-l (aliphatic hydrocarbons, including some olefins) and F-2 (aromatic
hydrocarbons, including some olefins and PAH compounds) fractions on a
silica gel column.
Samples of tarballs were collected on the Narragan.sett Town Beach
during May of 1978 and on the Brittany Coast of France in March 1978.
Samples of coal tar used to coat docks were scraped from piers.
The tarballs from the Brittany Coast were well-aged and did not
originate (or were not contaminated ) during the AMOCO CADIZ oil spill
27
-------
: 3
Figure 1. Gas
(B) tarball
chromatograms of F-2 fraction of: (A)
sample from Brittany Coast of France.
tarball sample from Narragansett Beach,
-------
TABLE 1. PAH COMPOUNDS IN FIGURE 1
Compound Molecular
Number Weight
1 178
2 178
3* 192
/•
4 202
5 202
6 228
7 228
8 252
9 252
10 252
11 252
Z
Number
-18
-18
-18
-22
-22
-24
-24
-28
-28
-28
-28
Tentative
Compound
Identification
Phenanthrene
Anthracene
C, -Phenanthrenes
+ C, -Anthracenes
Fluoranthene
Pyrene
Benzanthracene
Chrysene
Benzofluoranthene
Benzo(e)pyrene
Benzo(a)pyrene
Perylene
* Other compounds may also be present under this bracket.
29
-------
Tars and tarballs were dissolved in CH2 C12, solvent exchanged to
hexane and separated into F-l and F-2 fractions on silica gel columns.
The column chroinatographic fractions were analyzed on glass capillary
columns by both gas chromatography and gas chromatography-mass spectromet-
ry (GC-HS). The method of analysis has been described in detail (Lake
et_ al_., 1980). This analysis determined the concentrations of selected PAH
compounds by integrating the GC-MS extracted ion current profiles (EICPs)
corresponding to the molecular ions of the compounds of interest. Specific
peaks in the EICPs were not integrated if their spectra did not correspond
to those of the compounds of interest or if they did not have characteris-
tic retention times. The integrated values were normalized and displayed
in a semi-logarithmic format as parent compound distributions (PCDs) and
alkyl homolog distributions (AHDs). PCDs represent the relative
concentrations of the parent compounds of interest; i.e., PAHs with
molecular weights of 178, 202, 228, 252, 276, and 278. These distributions
were obtained by correcting the raw data for instrument response with the
aid of response factors calculated from known PAH standards. In the few
instances where a standard was not available, raw data were corrected by
the use of another PAH standard with the same molecular weight. AHDs show
the concentration of the parent compound in relation to its Cj through
£3 alkyl homologs. The AHDs were not corrected for instrument response
because of a lack of necessary alkylated standards. Z numbers were calcu-
lated from CnH2n+z. Phenanthrene/anthracene (P/A) ratios were
calculated by integrating separately the areas under the two peaks in the
EICP for m/e of 178.
The GC analyses were performed on a 20 m glass capillary column coated
with SE-54 in a Hewlett-Packard 5840A GC with splitless injection. The
GC-MS analyses were performed on a similar column in a Shimazdu Model
GC-4CM GC connected to a Finnegan 1015 mass spectrometer equipped with a
Systems Industries data system with Riber D-8 software. The mass spectro-
meter was operated in the El (electron impact) mode at 70 eV. Reagent
blanks and standard samples were processed with samples to ensure continued
satisfactory performance of the methods.
RESULTS
The gas chromatograms (GCs) of fraction 1 (F-l) of the tarballs
collected from Narragansett Beach showed only small amounts of hydrocarbon
material whereas, GCs from the F-l fraction of the French tarball showed
a much higher proportion of normal and branched alkanes in addition to a
large area under the resolved peaks called the "unresolved complex
mixture." GCs from the F-2 of the Narragansett tarballs were well-resolved.
The major peaks in these fractions were non-alkylated PAH molecules (Figure
1(A); Table I). In contrast, GCs of F-2 fractions from the French tarballs
were not well resolved (Figure 1(B)). GC-MS analyses of the F-2 fractions
from the French tarballs revealed that many of the compounds were alkylated
compounds of the Z=-12 (naphthalene) and the Z=-18 (anthracene/phenanthrene)
homolog series.
30
-------
CO
s i.o
Z
o
z
GO
UJ
S O.I
UJ
-------
The gas chromatograms of extracts from sediments surrounding docks
from Narragansett Bay and Block Island, Rhode Island, showed relatively
small amounts of material in the F-l fractions, but relatively large
amounts of well resolved compounds in the F-2 fractions. Similar GCs were
obtained from extracts of coal tar collected from docks. GC-MS analyses of
F-2 fractions from these samples showed that most of the largest resolved
peaks in these chromatograms were non-alkylated PAHs. The PCDs and Z=-22
AMDs of representative samples are shown (Figures 2 and 3). The
phenanthrene/anthracene (P/A) ratios of these samples were relatively high
(Table 2).
The GCs from the sediment sample collected near Jamestown Island
showed a large unresolved complex mixture in the F-l fraction, but larger
relative amounts of resolved components in the F-2 fraction. GC-MS
analysis of the F-2 fraction showed some of the resolved compounds were
PAHs. The PCDs and Z=-22 AMDs calculated from this sample are shown
(Figures 2 and 3). The P/A ratio of this sample was relatively low (Table
2).
DISCUSSION
Analyses of tarballs from Rhode Island and from the Brittany Coast of
France revealed major differences in chemical composition. Gas chroma-
tograms (GCs) of the F-l fractions showed that the tarballs from
Narragansett Beach contained relatively small amounts of both resolved
normal alkanes and unresolved complex mixture. In contrast, French
tarballs contained relatively large amounts of normal alkanes, the
isoprenoids pristane and phytane, and a large amount of unresolved complex
material.
GCs from the F-2 fraction from the Narragansett tarball showed well
resolved non-alkylated PAH compounds (Figure 1(A), Table I), and were very
similar to GCs obtained from coal tar collected from dock coatings. The
GCs from the F-2 fractions of French tarballs were not well resolved
(Figure 1(B)). These fractions contained many alkylated compounds in the
Z=-12 (naphthalene) and Z=-18 (anthracene/phenanthrene) homolog series.
The relatively low amount of material in F-l fractions, the abundance
of well-resolved, non-alkylated PAH compounds in F-2 fractions, and the
similarities of GCs from Narragansett tarballs and those from coal tar used
to coat pilings indicates that the tarball samples from Narragansett Beach
were probably coal tar. Comparisons of the GCs of F-l fractions from the
French tarball samples with GCs from some petroleum derived tarballs
(Morris and Butler, 1973) showed many similarities in the patterns of
alkane peaks and in the relative amount of unresolved complex mixture.
These similarities and the French tarballs1 high content of alkylated PAH
compounds in the Z=-12 and Z=-18 homolog series, which are the most abund-
ant PAHs found in petroleum (Pancirov and Brown, 1975; Youngblood and
Blumer, 1975), combined to indicate that the tarball samples from the
Brittany Coast probably originated from petroleum.
32
-------
An examination of the PAH assemblages in extracts of sediments
obtained near piers in the Narragansett Bay was undertaken to determine if
the coal tar used to coat these piers was a significant source of PAHs to
Narragansett Bay sediments. Comparisons of PCDs, AMDs, and P/A ratios
from the sediments near the docks with those obtained from the coal tar
used to coat the docks showed a definite correspondence between the PAHs in
the dock tar and in the sediments. The PAH assemblages in the dock tar and
in these contaminated sediments showed characteristic PCDs (Figure 2)
steeply sloping Z=-22 AHDs (Figure 3), and relatively high (>20) P/A ratios
(Table 2). To ensure that other inputs of PAH material were not
influencing the observed measurements, samples of dock tar and sediments
were obtained from a relatively pristine location-Block Island, Rhode
Island (14 miles off the coast of Rhode Island). The PAH measurements from
these samples matched each other and corresponding tar and sediment samples
from Narragansett Bay. These data indicate that the use of coal tar in
marine construction resulted in the input of PAH compounds to sediments
adjacent to docks.
The PCD, AMD, and P/A ratio of the North Jamestown sediment sample are
shown in Figures 2 and 3 and in Table 2. The PAH assemblage of this sample
has been shown to be quite similar to those in other Narragansett Bay
sediments which were not collected near docks (Lake jjib _§]_., 1979).
Comparisons of the above measurements with those from coal tar used to coat
the docks showed many differences and thereby indicated that dock tar was
probably not the major PAH contaminant of bay sediments. Rather, as
described in detail elsewhere (Lake et_ aj_., 1979), the PAH assemblages in
these sediment samples appeared to result from a combination of inputs.
1.0
UJ
o
CD
UJ
UJ
01
O.I
.01 -
NARR.
N. JAMES DOCK TAR
NARR. BLOCK I. BLOCK I.
DOCK SEP. TAR SED.
NUMBER OF ALKYL
CARBON ATOMS
Figure 3. Z=-22 Alkyl Homolog Distributions (AHDs) from samples of tars
and sediments.
33
-------
TABLE 2. PHENANTHRENE/ANTHRACENE RATIOS OF DOCK TARS AND SEDIMENTS
Description
of Sample P/A Ratio
Narragansett Dock Tar 38
Block Island Dock Tar 39
Narragansett Dock Sediment 26
Block Island Dock Sediment 23
North Jamestown Sediment 4.3
REFERENCES
Farrington, J.W., N.M. Frew, P.M. Gschwend, and B.W. Tripp. 1977.
Hydrocarbons in cores of northwestern Atlantic coastal and continental
marine sediments. Estuar. Coast. Mar. Sci. 5:793-808.
Hase, A., and R.A. Hites. 1976. On the origin of polycyclic aromatic
hydrocarbons in the aqueous environment. In: Identification and
analysis of organic pollutants in water. Keith, L.H., Ed., Ann Arbor
Publications, Inc., Ann Arbor, MI. pp. 205-214.
Hites, R.A. 1976. Sources of polycyclic aromatic hydrocarbons in the
aquatic environment. In: Sources, Effects and Sinks of Hydrocarbons
in the Aquatic Environment. American Institute of Biological Science,
Washington, DC. pp. 325-333.
Hites, R.A., and W.G. Biemann. 1975. Identification of specific organic
compounds in a highly anoxic sediment by gas chromatographic-mass
spectrometry and high resolution mass spectrometry. Adv. Chem. Ser.
147:188-201.
Hites, R.A., R.E. LaFlamme, and J.W. Farrington, 1977. Sedimentary poly-
cyclic aromatic hydrocarbons: The historical record.
Science 198:829-831.
34
-------
LaFlamme, R.E., and R.A. Hites. 1978. The global distribution of
polycyclic aromatic hydrocarbons in recent sediments. Geochim. Cosmo-
chim. Acta 42:289-303.
Lake, J.L., C.B. Norwood, C.W. Dimock, and R. Bowen. 1979. Origins of
polycyclic aromatic hydrocarbons in estuarine sediments. Geochim.
Cosmochim. Acta 43:1847-1854.
Lake, J.L., C.W. Dimock, and C.B. Norwood. 1980. A comparison of methods
for the analysis of hydrocarbons in marine sediment. ACS Advances in
Chemistry Series. No. 185 Petroleum in the Marine Environment,
L. Petrakis and F. Weiss, Eds., pp. 343-360.
Lee, M.L., G.P. Prado, J.B. Howard, and R.A. Hites. 1977. Source
identification of urban airborne polycyclic aromatic hydrocarbons by
gas chromatography mass spectrometery and high resolution mass
spectrometry. Biomed. Mass Spectrom. 4:182-186.
Morris, B.F., and J.N. Butler. 1973. Petroleum residues in the Sargasso
Sea and on Bermuda beaches. Proceedings Conference on Prevention and
Control of Oil Spills, sponsored by API, EPA, USCG, March 13-15, 1973,
Washington, DC.
Windsor, J.G., and R.A. Hites. 1979. Transport of polycyclic aromatic
hydrocarbons across the Gulf of Maine. Geochim. Cosmochim. Acta
43:27-33.
Youngblood, W.W., and M. Blumer. 1975. Polycyclic aromatic hydrocarbons
in the environment: homologous series in soils and recent marine
sediments. Geochim. Cosmochim. Acta 39:1303-1314.
35
-------
TOXIC PHOTOOXYGENATED PRODUCTS GENERATED UNDER ENVIRONMENTAL
CONDITIONS FROM PHENANTHRENE
by
Jayanti R. Patel, Jo Ann McFall, Gary W. Griffin,
and John L. Laseter
Center for Bio-Organic Studies, University of New Orleans,
New Orleans, LA 70122
ABSTRACT
The photooxidation of phenanthrene under simulated
environmental conditions to 9,10-epoxy-9,10-dihydrophenan-
threne, among other oxygenated products, serves as a possible
model for the conversion of polycyclic aromatic hydrocarbons
(PAHs) to potentially mutagenic and/or carcinogenic products in
the environment. The reaction was carried out in a hexane-
aqueous phase illuminated by a lamp whose output is similar to
that of sunlight. These toxic photoproducts were isolated and
identified by glass capillary gas chromatography-mass spectrom-
etry, through comparison of GC retention times and mass spectral
fragmentation patterns with data observed for authentic samples
obtained independently through synthesis or commercial sources.
Some of these products were found to be soluble in water which
suggests the possibility of the intrusion of oxygenated PAHs
into natural waters. These results are discussed in terms of
the environmental oxidative processes of various mutagenic and
carcinogenic PAHs.
INTRODUCTION
A recent study of 22 PAHs reveals that the PAHs themselves are not
mutagenic towards Salmonella typhimurium, but their irradiated reaction
mixtures are found to be mutagenic (Gibson et aK, 1978). In general, PAHs
that are mutagenic or carcinogenic require metabolic activation before they
can induce their effect (Asby and Styles, 1978). Although the mutagenicity
of some PAHs appears to be involved by both photooxidation by air and
oxidation by rat liver S9 microsomes, they may operate by different or
identical oxidative mechanisms to generate oxygenated products (Gibson
eit ^1_., 1978). The biotransformation of PAHs to mutagens or carcinogens
needs further study as well as chemical identification of active metabol ites.
The oxygenated products of various PAHs such as epoxides, phenols, diols,
and epoxydiols present a threat to human health as evidenced in several
Present Address: BASF Wyandotte Corporation
P.O. Box 181, Parsippany, NJ 07054
36
-------
hundred toxicological reports on their carcinogenicity and mutagenicity
(DePierre and Ernster, 1978; Yang et aj_., 1978).
Among the widely claimed environmental sources containing PAHs are
urban air (Bartle ^t al_., 1977; Bjorseth, 1977; Giger and Schaffner, 1978;
Smillie et a\_., 1978), tobacco smoke (Bartle et _§]_., 1977), coal tar
(Lijinsky etal_., 1963), coal (Schabron et a\_., 1977; Woo et _al_., 1978),
soot of electrolysis furnaces (Tausch and Stehlik, 1977), marijuana smoke
(Bartle £t £[., 1977), mineral oils (Lijinsky ^t al_., 1963), petroleum
(Lijinsky et _§]_., 1963), sediments (Giger and Schaffner, 1978; Jungclaus
et al_. 1978T, river particulates (Giger and Schaffner, 1978), sludge
TGrimmer eit atK, 1978), tissues of marine organisms (Pancirov and Brown,
1977), industrial waste waters (Jungclaus ^t al., 1978), street dust (Giger
and Schaffner, 1978, and even drinking waters~TBasu and Saxena, 1978).
Recently, the ocean has been contaminated by oil spills which deposit tons
of petroleum hydrocarbons (Boylan and Tripp, 1971). It has been noted,
however, that photooxidative processes presumably occurring at the surface
of the oil films can lead to other highly toxic and water soluble materials
during such weathering processes (Scheier and Gominger, 1976). The leached
water soluble oxygenated products can be toxic to algae, fish, and other
marine organisms. Recent evidence (Lacaze and de Naide, 1976; Payne
et _§]_., 1978; Scheier and Gominger, 1976) indicates increased toxicity of
the water soluble fraction after irradiation of crude oils. Thus
photooxidation of petroleum hydrocarbons in the environment increases the
toxicity of petroleum and may present a threat to human health and have an
effect on marine organisms. The natural seepage of oil and PAHs onto the
continental shelf has been estimated to release 0.2 x 10 to 6 x 105
metric tons of PAHs per year into the sea (Wilson et jil_., 1974). The
finding that the ratio of aromatic hydrocarbons to cycloalkanes decreases
in non-volatile dispersed oil on the sea surface suggests that aromatic
hydrocarbons are destroyed by sunlight. Polycyclic aromatic hydrocarbons
(1 to 3.5% of crude oils) may be oxidized in nature by the molecular oxygen
in air and sunlight or by singlet oxygen present in natural waters (Zepp
et al_., 1977), or produced by molecular oxygen and naturally occurring
hydrocarbons which act as photosensitizers (Patel et _al_., 1978c). Phenan-
threne and several other aromatic and heteroaromatic hydrocarbons commonly
found in petroleum have been observed to be decomposed by irradiation
(Nagata and Kondo, 1977).
The ultimate mutagenic or carcinogenic activity of PAHs is coupled
through the metabolic activation to primary and secondary metabolites.
Dihydroepoxides of PAHs have been the subject of extensive studies of
metabolic carcinogens, and their interaction with DNA has also been
investigated (Sims and Grover, 1974). Attention has been focused on
7,8-dihydroxy-9,10-epoxy-7,8,9,10-tetrahydrobenzo(a)pyrene and its role as
the ultimate carcinogen of benzo(a)pyrene (Yang et a]_., 1978). However,
there is widespread evidence that other polycyclic aromatic hydrocarbon
epoxides behave similarly (Sims and Grover, 1974). It is of interest to
know that photo- oxidation is the most significant mechanism by which the
carcinogenic benzo(a)pyrene degrades in aqueous systems in the presence of
oxygen (Suess, 1977). This strongly suggests a need to investigate the
37
-------
effects of photooxidation on other polycyclic aromatic hydrocarbons
commonly found in crude oils. The formation of proximate carcinogens in
the environment through photooxidation or biological pathways is an
important though disturbing possibility which merits further study.
We have carried out the photooxidation of such PAHs under simulated
environmental conditions and have isolated some of the major photoxygen-
ated products which are known to be toxic, carcinogenic, and mutagenic.
This report presents the isolation and the characterization of such
products from phenanthrene, a model PAH contaminant. Moreover, phenan-
threne is a relatively simple PAH containing both a "K-region" and a
"bay-region"; molecular features of certain PAHs have been the object of
considerable investigation with respect to the carcinogenicity of those
compounds.
METHODS AND MATERIALS
(A) Photooxidation and Separation of Products
A two-phase system composed of solutions of phenanthrene in hexane on
synthetic seawater (pH 8.2) was irradiated at the surface as previously
described (Dowty et ^1_., 1974) under conditions simulating those found in
the environment at air-sea interfaces (sea-water droplets formed by wind
and waves at the sea surface [Bezdek, and Carlucci, 1974]). A visible
source (Sylvania EHC 500 watt tungsten lamp shielded by a Uranium glass ^
filter) whose spectral output approximates that of sunlight was employed .
The reactions were conducted in the presence of oxygen and perylene, a
naturally occurring PAH (Lijinsky ^t a\_., 1963; Schabron et _al_., 1977;
Giger and Scaffner, 1978; Smille et a].., 1978; Woo et a]_., 1978), which
acts as a singlet oxygen sensitizer.
The reaction mixtures obtained upon photooxidation were separated into
two fractions (hexane and aqueous) (Figure 1). The hexane fraction was
dried over anhydrous sodium sulfate and concentrated under vacuum on a
rotary evaporator, while the aqueous fraction was extracted with
dichloromethane. The remaining aqueous layer was made acidic by the
addition of 1 M hydrochloric acid (pH 2) and then extracted with
dichloromethane. The dichloromethane extract containing the acidic
components (mainly phenols and carboxylic acids) was methylated with
methyl iodide and sodium hydroxide in dimethyl sulphoxide (Q^SOQ^)
according to the method described by Gill is (1968). The photooxidation of
9,10-epoxy-9,10-dihydrophenanthrene and other oxygenated products were
conducted under conditions similar to those mentioned for phenanthrene.
The obvious control experiments conducted in the dark, in the absence of
sensitizer and without oxygen, were also carried out under otherwise
The absorbance spectra of uranium glass probe measured on a Model Gary
17 spectrophotometer cuts out at 339.5 nm, i.e., ultraviolet radiation was
blocked out.
38
-------
REACTION MIXTURE
HEXANE LAYER
AQUEOUS LAYER
EXTRACTION C CH2Cl_2
"NEUTRAL LAYER"
v —
s
\
AQUEOUS
+ 1 M HCL
EXTRACTION C CH2lL2
"ACIDIC FRACTION"
^v
\
AQUEOUS
\f
+ 1 M NAOH _ ru
EXTRACTION C CH2l!_2
"BASIC FRACTION"
*x
— \
AQUEOUS
Figure 1.
Y
DISCARD
Experimental diagram of the photooxygenated reaction mixture
into compound-class fractions.
39
-------
identical conditions. Finally the hexane, dichloromethane, and methylated
acidic fractions were concentrated and analyzed by GC-MS techniques.
(B) Analytical Procedures
Gas Chromatographic (GC) Analysis-
Products were identified by the comparison of GC retention times.
All analyses used a Hewlett-Packard 5711A gas chromatograph coupled to a
Hewlett-Packard 3354A Laboratory Data System. The injection port and
detector temperatures were maintained at 250° and 300° C, respectively.
The compounds were separated in a 28 m x 0.3 mm (i.d.) glass capillary
column coated with SE-52 (in dichloromethane) (Lawler^tjjK, 1977).
Helium was used as the carrier gas at a flow rate of approximately 1.9
nu/min. The air and hydrogen flow rates were maintained at 240 and 30
mJi/min, respectively.
Gas Chromatographic-Mass Spectrometric (GC-Ms) Analysis--
All products and reactants were subjected to analysis using the same
chromatrographic conditions described above with the exception that a glass
capillary column (13 m) was utilized and the carrier gas flow rate was
adjusted correspondingly (3 nu/min). The gas chromatograph was coupled to
a Hewlett-Packard 5980A mass spectrometer operated with an ionizing voltage
(70 eV) and 160 yamp current. The source temperature and the pressure were
maintained at 150° C and 2 x W Torr. A Hewlett-Packard 5933A Data
System was interfaced with the mass spectrometer.
Mass Spectral Analysis—
The mass spectral fragmentation patterns of authentic reference
standards were obtained by the direct insertion probe method. The probe
temperature was varied between ambient temperature and 150° C during the
course of the analyses, while the source temperature and pressure were
maintained at 155° C and 6.4 x 10"° Torr.
Preparation of Reference Standards-
Authentic samples of the following products were prepared by methods
described in the literature: 2'-formylbiphenyl-2-carboxylic acid (Bailey,
1956), 2,2'-diformylbiphenyl (Bailey and Erickson, 1961), 9,10-epoxy-9,
10-dihydrophenanthrene (Newman and Blum, 1964; 2,3:4,5-dibenzoxepin
(Brightwell and Griffin, 1973), 9-pnenanthrol (Newman and Blum, 1964),
diphenic acid anhydride, and 3,4-benzocoumarin. (Samples of the latter two
products were provided by Dr. G.W. Griffin.) The physical properties of
the synthetic samples compared favorably with those reported in the
lieterature. Additional reference standards were obtained from commercial
sources (Aldrich Chemical Company, Milwaukee, WI).
RESULTS
The air-sea interface model created by oxygen bubbled through a two-
phase system (composed of a phenanthrene solution in hexane on synthetic
seawater) was found to be appropriate for the photooxidation of PAHs under
-------
simulated environmental conditions. The direct exposure of sunlight and
atmospheric oxygen in the marine environment is most likely to occur at the
air-sea interface where tiny water droplets are formed by the action of
wind and waves (Bezdek and Carlucci, 1971). The UV-visible spectra
measured for the visible source in our study (Sylvania EHC 500 watt
tungsten lamp covered by a Uranium glass filter) approximates that of
sunlight.
Among the various photosensitizers evaluated, the naturally occurring
hydrocarbon, perylene (Lijinsky et al_., 1963; Schabron et al., 1977; Giger
and Schaffner, 1978; Smillie et al_., 1978; Woo et _al_., 197FJ, was found to
be an effective photosensitizer for the generation of the different classes
of oxygenated products from phenanthrene (Patel et _§]_., 1978c). The
results of these reactions, i.e., the product ratios, were compared with
those observed in the photooxidation of phenanthrene conducted with known
photosensitizers such as methylene blue and rose bengal. The results were
identical, which indicates that a naturally occurring organic compound can
lead to the sequences of events involving singlet oxygen, resulting in the
photodecomposition of aromatic hydrocarbons.
The disappearance of phenanthrene was monitored by analyzing aliquots
collected at intermittent time intervals' (0 to 22 hr) as the reaction
progressed (Figure 2). During the course of the experiment, 32% of the
phenanthrene was consumed. The halflife of the phenanthrene under these
conditions was determined to be approximately 80 hr.
To improve GC resolution and facilitate identification of the major
products obtained upon photooxidation of phenanthrene, the acidic products
(mainly carboxylic acids and phenols) were effectively separated initially
by liquid-liquid extraction from the aqueous phase. The acidic components
were subsequently methylated utilizing methyl iodide prior to analysis
(Gillis, 1968)(Figure 3D). Peaks are identified in Table 1, and the
structures of corresponding compounds are shown in Figure 4.
The structures of the individual components were established by
comparing GC retention times and mass spectral fragmentation patterns with
the corresponding data obtained with authentic samples. The precise mass
spectral fragmentation patterns of the primary and secondary photoproducts
of phenanthrene were determined (Patel ^t a]_., 1978b).
The GC runs from the aqueous phase (Figures 3C and 3D) show the
presence of oxygenated products which are soluble in water. Among the
water-soluble products, 2-formylbiphenyl-2'-carboxylic acid, diphenic acid,
and 9-phenantrol were found in the acidic fraction (Figure 3D). They were
separated and detected as corresponding methylated derivatives by
methylation prior to analysis, while other neutral water soluble products
were extracted by dichloromethane prior to acidification (Figure 3C).
Under identical conditions, various suspected intermediates such as
9,10-epoxy-9,10-dihydrophenanthrene, 9,10-phenanthrenequinone, 9-fluorenone,
41
-------
00
CO
(0
ro
C
0
C
fll
-c
a
-------
n
u
CO
o
CL
CO
UJ
a
a
LU
a
a
o
o
UJ
a
phenanthrene
A. Total Run
1.0
B. Hexane Fraction
1.0 ul/3ml
pnenanthrene
phenanthrene.
^henantlrene
C. Dlchlerome thane Fraction
1.0 «1/300 al
0. Acidic Methylated
1.0 wl/150 UJ
jLJjLu.
JUU
1,3
^-
38
TIME CMINUTES)
4.0 50 60
70
8,0
70
90
ia
130
213 230240
150 170 190
TEMPERATURE CDEG. C5
Figure 3. Gas chromatograms of photooxidation of phenathrene.
43
-------
TABLE 1. PHOTOOXIDATION PRODUCTS FROM PHENANTHRENE AFTER 9 HR IRRADIATION
NO.*
1
2
3
4
** 5
7
4.8
.v-.. g
JO
—
NAME OF THE PRODUCTS
Fluorene
Fluorenone
2,3:4,5-dibenzoxepin
2 ,2 ' -di formyl bi phenyl
2'-formylbiphenyl -2-carboxyl ic acid
3,4-benzocoumarin
diphenic acid
9-phenanthrol
9 , 1 0-phenanthrenequi none
diphenic acid anhydride
Formula
r w
^1 3nl 0
C13H80
Ci,H100
CnHloOa
Cx.Hx.0,
Cj 3H802
Ci,H100,
C1UH100
C^HaO,
ClttH803
Mol. Wt.
165
180
194
210
226
196
242
194
208
224
Numbers refer to the peaks in Figure 3 and the structures illustrated
in Figure 4.
These compounds were observed as their corresponding methylated
derivatives.
44
-------
U1
10
Figure 4. Photooxygenated products from phenanthrene.
-------
and fluorene were irradiated and products identified by gas chromatographic
analysis. The combined gas chromatograms of the photooxidation reaction
mixtures from these substrates are depicted in Figure 5.
DISCUSSION
The involvement of singlet oxygen in the photooxidation of PAHs in the
environment is a very significant observation since singlet oxygen, an
excited molecule, is a reactive form of oxygen responsible for various
environmental oxidation (Coomber ert jj]_., 1970; Politizer et al_., 1971).
Furthermore, the generation of singlet oxygen from molecular oxygen using a
naturally available hydrocarbon and sunlight suggests that photooxidation
of various PAHs under actual environmental conditions would produce large
numbers of products, some of which may be toxic. The water soluble PAHs
(Frankenfeld, 1973) (phenanthrene found to be soluble in water, [May et
al», 1978]) also may be subjected to oxidation by singlet oxygen available
in natural waters (Zepp^t j*U, 1977) or by singlet oxygen generated by
water soluble photosensitizer and transmitted sunlight. It is known that
photosensitization generally accelerates the decomposition of petroleum
(Klein and Pilpel, 1974). The proposed consideration of the possible role
of singlet oxygen for the oxidation of PAHs with a K-region double bond
(Khan and Kasha, 1970) in chemical carcinogenesis is also justified by our
study. All products identified from the photooxidation of phenanthrene
seem to originate after the initial attack of singlet oxygen at the
9,10-double bond.
To achieve the actual complete conversion of phenanthrene to its
oxygenated products on the air-sea interface of the ocean may require
exposure to bright sunlight and sufficient oxygen. However, it is shown
that the photodegradation rate constant for highly condensed PAHs increases
with temperatures over the range of 5 to 31° C (Suess, 1971). Thus, at
ambient temperatures, the generation of oxygenated products may be enhanced
many fold.
As an integral part of the product identification phase of our
program, we have studied the mass spectral fragmentation patterns of
various oxygenated compounds. The structural rearrangements of several
2,2'-disubstituted biphenyls induced upon electron impact and the mass
spectral comparrison of 9,10-epoxy-9,10-dihydrophenanthrene and certain
other structural isomers were investigated (Patel jit jj]_., 1978b).
During the preliminary studies conducted in our laboratory on
dye-sensitized photooxidation of phenanthrene (Dowty et aj_., 1974),
complete resolution of the reaction was achieved after the proper
analytical techniques were developed. We were successful in achieving our
goal by using modified high resolution gas chromatography.. The constit-
uents derived from phenanthrene upon oxidation with singlet oxygen include
fluorene, fluorenone, 2,2'-diformylbiphenyl, 3,4-benzocoumarin,
9,10-phenanthrenequinone, 9-phenanthrol, 2-formylbiphenyl-2'-carboxylic
acid, and diphenic acid. The GC peak due to phenanthrene is off-scale and
46
-------
A. Phenanthr».i« oxide reaction
B. Phenanthrer.equlnone reaction
C. Fluorene reaction
/luorene
0. Fluorenone reaction
-Fluorenone
10 20
30
TIME CMINUTES)
40 50 60
, phenanthrenequl none
80 SO
72 90 110 130 150 173 190 210 230 240
TEMPERATURE OEG C)
Figure 5. Gas chromatograms of photooxidaticn of intermediates.
47
-------
unfortunately obscures a broad and significant region of the chromatogram
(Figure 3). Figure 6 shows possible pathways for the formation of
phenanthrene photooxidation products.
9,10-Epoxy-9,10-dihydrophenanthrene was thought to be one of the
precursors for certain other photooxidation products derived from
phenanthrene because its photooxidation under identical conditions gave
several identical products such as 9-phenanthrol, fluorenone,
3,4-benzocoumarin, 2,3:4,5-dibenzoxepin, diphenic acid, and phenanthrene
(Figures 5 and 6). It is noteworthy that during direct photooxidation
reactions (Shudo and Okamoto, 1973; Dowtyjrt al^, 1974; Jerina et al.,
1974) in organic solvents, 9,10-epoxy-9,10-dihydrophenanthrene gave mainly
9-phenanthrol and 2,3:4,5-dibenzoxepin, while solvolysis was a competing
reaction in the aqueous phase. At pH 8.2 phenanthreneoxide converts to
9-phenanthrol (20%) and trans_-9,10-dihydroxy-9,10-dihydrophenanthrene (80%)
due to solvolysis in aqueous solution (Bruice et^ al_., 1976). Moreover, the
conversion of the oxide to two other products had occurred in the gas
chromatograph because of thermal sensitivity (Patel et. aK, 1978a).
Despite these complexities, the presence of 9-phenanthrol from the
photooxidation products of phenanthrene, and the comparable results from
the independent photolysis reaction of the oxide under identical
conditions, led us to believe that the oxide may be formed as a primary
product. Goto and co-workers (1978) reported that the radio!ysis of liquid
C02 with phenanthrene leads to several oxygenated products including
9,10-epoxy-9,10-dihydrophenanthrene (35%), 9-phenanthrol (46%), 2,3:4,5-
dibenzoxepin (1.3%), and 2,2'-diformylbiphenyl (0.9%).
9,10-Phenanthrenequinone may have formed through a dioxetane
intermediate ji which is known to occur in the photooxidation of
9,10-dimethoxyphenanthrene or the solvoysis of trans-9,10-dihydroxy-9,10-
dihydrophenanthrene ^. The ready conversion of 9,10-phenanthrenequinone to
several other products: fluorenone, 3,4-benzocoumarin, 2-formylbiphenyl-
2'-carboxylic acid, and diphenic acid, upon photooxidation under identical
conditions, may explain the occurrence of minor amounts of the quinone from
the photoreaction of phenanthrene (Figures 3 and 6). Moreover, to under-
stand the formation of fluorene as a minor product during the photoisomer-
ization (Brightwell and Griffin, 1973; Dowty ^t al_., 1974) of the oxide and
to gain some insight on the stability of 9-fluorenone, the photoreactions
of these two compounds have been tried under identical conditions (Figure 5),
0-0
48
-------
Solvolysis
+ 9-phenanthrol
+ 9,10-phenanthrenequinone
+ various 2,2'-disubstituted
biphenyls
R = R' = CHO
R = R' = COOH
R = CHO, R1 = COOH
Figure 6. Proposed pathways for the formation of phenanthrene
photooxidation products.
49
-------
Most of the oxygenated products are found to be completely or
partially soluble in water (Figure 3). These compounds may have great
impact on marine organisms. Thus the occurrence of acidic and phenolic
compounds (Guard ^t al_., 1975; Winters et ^1_., 1976; McFall et al_.,1978)
and aromatic aldehydes and ketones (Guard et jiK, 1975; Winters and Parker,
1977) in the environment suggests that the oxidation of hydrocarbons is
responsible for the formation of oxygenated water soluble compounds.
Moreover, phenolic compounds have been found after an oil spill in water
extracts (Burwood and Speers, 1974).
The photooxidation of aromatic hydrocarbons by sunlight (Calder
j?t jil_., 1978; Freegarde et aU, 1971), a natural process by which crude
oils are decomposed during oil spills on the ocean, is a relatively slow
process which generates water soluble oxygenated compounds including
carbonyl compounds (Burwood and Speers, 1974; Reed, 1977). Very recent
studies (Epler jet _al_., 1978; Payne et_ _al_., 1978) on the mutagenicity of
crude oils reveal that the fraction containing polycyclic aromatic
hydrocarbons is mainly responsible for total mutagenic activity; however,
the mutagenic activity was increased after irradiation of the crude oils
(Payne £t a±., 1978). It is very interesting to know that such an increase
in activity was attributed to the formation of oxygenated products. In
similar studies, water soluble oxygenated compounds toxic to Baker's yeast
were formed upon the environmental irradiation of a No. 2 fuel oil (Larson
et^aK, 1977). Even the phytotoxicity of crude oils (Kuwait crude oil) was
also increased about two to three times after exposure to artificial
illumination (Lacaze ^t al_., 1977), and it was thought that oxygenated
compounds after photooxidation are responsible for such changes. Thus the
facts accumulated to date are in accord with the conclusion that the
photooxidation of petroleum hydrocarbons in the environment increases the
toxicity of petroleum and may present a serious threat to human health and
catastrophic effects on marine organisms. Our studies involving a single
PAH indicate clearly that carbonyl and other oxygenated compounds are
generated upon environmental photooxidation. The toxicity of some of the
products obtained from this study is summarized in Table 2. Moreover,
compounds analogous to 3,4-benzocoumarin, 9-phenanthrol, 9,10-epoxy-9,10-
dihydrophenanthrene, and diols of phenanthrene are found to have an adverse
effect on biological systems (Acros, 1978).
It is significant that 9,10-epoxy-9,10-dihydrophenanthrene is believed
to be the primary product among the photooxidation products of phenanth-
rene. A comparison of the metabolites formed from 9,10-epoxy-9,10-dihydro-
phenanthrene with those formed at the 9,10-bond of phenanthrene in rats
showed that both compounds yielded trans-9,10-dihydro-9.10-dihydroxyphenan-
threne (Jerena et jfL» 1974). Both compounds also yield a mercapturic acid,
N-acetyl-S-(9,10-dihydro-9-hydroxy-10-phenanthryl) cysteine. Almost all
investigations of the metabolism of polycyclic aromatic compounds conclude
that the formation of an epoxide is the first step from the parent compound
(Sims and Grover, 1974). Arene oxides are primary metabolites of PAHs
(Levin et^ a\_., 1976), and many are carcinogenic in mammalian cells. They
exhibit significant mutagenic (McCann jit aU, 1975), carcinogenic (Levin
et jil_., 1976), anti-viral (Hsu jet jil_., 1977), and cellular transformation
50
-------
(Marquardt et jjl_., 1972) activity coupled with the ability to bind
covalently to nucleic acids (Baird et aj[., 1975; Blobstein j^t jj]_., 1975).
The identification of an arene oxide by the environmental photooxidation of
polycylic aromatic hydrocarbons is a highly significant observation for the
linkage of environmental processes to carcinogen!city or mutagenicity.
TABLE 2. TOXIC PHENANTHRENE PHOTOOXYGENATION PRODUCTS*
COMPOUND
9,10-Phenanthrene
-quinone
9-Fluorenone
BIOLOGICAL
ACTIVITY
CAR
TOX
CAR
TEST
ANIMAL
Mouse
(skin)
Mouse
(intraperit-
oneal )
Rat
(subcutaneous)
DOSE
2000 mg/kg
165 mg/kg
360 mg/kg
DURATION
28 weeks
(continuous)
single dose
26 weeks
(intermitant)
Christensen et jil_., 1977
CAR = Carcinogenic
TOX = Toxic
ACKNOWLEDGEMENTS
The authors are indebted to S.W. Mascarella and V. Warren for techni-
cal assistance, and D. Trembley for aid in preparation of this manuscript.
This work has been supported by the U.S. Environmental Protection Agency,
Grant R804646-01-1 and in part by the NIH (Grant CA 18346).
51
-------
REFERENCES
Arcos, J.C. 1978. Criteria for selecting chemical compounds for carcino-
genic testing: an essay. J. Environ. Path. Toxicol. 1:433-458.
Asby, J., and J.A. Styles. 1978. Does carcinogenic potency correlate with
mutagenic potency in the Ames assay Nature 271:452-455.
Bailey, P.S. 1956. The ozonolysis of phenanthrene in methanol. J. Am.
Chem. Soc. 78:3811-3817.
Bailey, P.S., and R.E. Erickson. 1961. Diphenaldehyde (biphenyl, 2-2'-
diformyl). Org. Synth. 41:41-45.
Baird, W.M., R.G. Harvey, and P. Brookes. 1975. Comparison of the
cellular DNA-bound products of benzo(a)pyrene with the products formed
by the reaction of benzo(a)pyrene-4,5-oxide with DNA. Cancer Res.
35:54-57.
Bartle, K.D., M.L. Lee, and M. Novotny. 1977. Identification of environ-
mental polynuclear aromatic hydrocarbons by pulse fourier-transform 'H
nuclear magnetic resonance spectroscopy. Analyst 102:731-738.
Basu, O.K., and J. Saxena. 1978. Polynuclear aromatic hydrocarbons in
selected U.S drinking waters and their raw water sources. Environ.
Sci. Technol. 12:795-798.
Bezdek, H.F., and A.F. Carlucci. 1974. Concentration and removal of
liquid microlayers from a seawater surface by bursting bubbles.
Limmol. Oceanogr. 19:126-132.
Bjorseth, A. 1977. Analysis of polycyclic aromatic hydrocarbons in
particulate matter by glass-capillary gas chromatography. Anal. Chem.
Acta 94:21-27.
Blobstein, S.H., I.B. Weinstein, D. Grunberger, J. Weisgros, and R.G.
Harvey. 1975. Products obtained after an in vitro reaction of
7,12-dimethylbenz(a)anthracene-5,6-oxide with nucleic acids. Biochem.
14:3451-3457.
Boylan, D.B., and B.W. Tripp. 1971. Determination of hydrocarbons in sea-
water extracts of crude oil and crude oil fractions. Nature 230:44-47.
Brightwell, N.E., and G.W. Griffin. 1973. Photorearrangement of
9,10-epoxy-9,10-dihydrophenenthrene synthesis of 2,3:4,5-dibenzoxepin.
J. Chem. Soc. Chem. Commun. pp. 37-38.
Bruice, P.Y., T.C. Bruice, P.M. Dansette, H.G. Selander, H. Yagi, and D.M.
Jerina. 1976. Comparison of the mechanisms of solvolysis and
rearrangement of K-region vs. non-K-region arene oxides of phenanthrene.
52
-------
Comparative solvolytic rate constant of K-region and non-K-region
arene oxides. J. Am. Chem. Soc. 98:2965-2973.
Burwood, R., and G.C. Speers. 1974. Photo-oxidation as a factor in the
environmental dispersal of crude oil. Estuarine Coast. Mar. Sci.
2:117- 135.
Calder, J., J. Lake, and J.L. Laseter. 1978. Chemical composition of
selected environmental and petroleum samples from the Amoco Cadiz oil
spill. In: The Amoco Cadiz oil spill. NOAA/EPA special report.
W.H. Hess, Ed., Government Printing Office, Washington, DC. pp.
21-83.
Christensen, H.E., and T.T. Luiginbyhl. 1977. Suspected Carcinogens.
U.S. Department of Health, Education, and Welfare, Washington, D.C.
Coomber, J.W., D.M. Hebert, W.A. Kummer, D.J. Marsh, and J.N. Pitts, Jr.
1970. Singlet oxygen in environmental sciences. Possible production
of '02 by energy transfer following oxygen enhanced absorption.
Environ. Sci. Toxicol. 4:1141-1147.
DePierre, J.W., and L. Ernster. 1978. The metabolism of polycyclic
hydrocarbons and its relationship to cancer. Biochem. Biophys. Acta
473:149-186.
Dowty, B.J., N.E. Brightwell, J.L. Laseter, and G.W. Griffin. 1974. Dye-
sensitized photooxidation of phenanthrene. Biochem. Biophys. Res.
Commun. 57:452-456.
Epler, J.L., J.A. Young, A.A. Hardigree, T.K. Rao, M.R. Guerin, J.B. Rubin,
C.H. Ho, and R.R. Clark. 1978. Analytical and biological analyses of
test materials from the synthetic fuel technologies. I. Mutagenicity
of crude oils determined by the Salmonella typhimuriurn/microsomal act-
ivation system. Mutat. Res. 57:265-276.
Frankenfeld, J.W. 1973. Factors governing the fate of oil at sea. Variat-
ions in the amounts and types of dissolved or dispersed materials
during the weathering process. In: Proc. Joint Conference on
Prevention and Control of Oil Spills, March 13-15, 1973, Washington,
DC. pp. 485-498.
Freegarde, M., C.G. Hatchard, and C.A. Parker. 1971. Oil spilt at sea:
its identification, determination and ultimate fate. Lab. Pract.
20:35-40.
Gibson, T.L., V.B. Smart, and L.L. Smith 1978. Non-enzymic activation of
polynuclear aromatic hydrocarbons as mutagens. Mutat. Res. 49:153-161.
Giger, W., and C. Schaffner. 1978. Determination of polycyclic aromatic
hydroarbons in the environment by glass capillary gas chromatography.
Anal. Chem. 50:243-249.
53
-------
Gill is, R.G. 1968. Trideuteromethylation in dimethyl sulfoxide. Tetrahe-
dron Lett. pp. 1413-1414.
Goto, S., A. Hori, S. Takamuku, and H. Sakurai. 1978. The reaction of
0( p)atoms with indene and phenanthrene induced by the Y-Radiolysis
of liquid carbon dioxide. Bull. Chem. Soc. Jpn. 51:1569-1570.
Grimmer, G., H. Bohnke, and H. Borwitzky. 1978. Gas chromatographische
profilanalyse der polycylischen aromatschen kohlenwasserstoffe in
klarschlammproben. Fresenius Z. Anal. Chem. 289:91-95.
Guard, H.E., L. Hunter, and D.H. Disalvo. 1975. Identification and
potential biological effects of the major components in the seawater
extract of a Bunker Fuel. Bull. Environ. Contam. Toxicol. 14:395-400.
Hsu, W.T., R.G. Harvey, E.J. Lin, and S.B. Weiss. 1977. Mechanism of
phase X174 DNA inactivation by benzo(a)pyrene-7,8-dihydrodiol-9,10-
epoxide. Proc. U.S. Natl. Acad. Sci., 74:3335-3339.
Jerina, D.M., H. Yagi, and J.W. Daly. 1973. Arene oxides-oxepins.
Heterocycles 1:267-320.
Oerina, D.M., B. Witkop, C.L. Mclntosh, and O.L. Chapman. 1974.
Photolysis of arene oxides at low temperature. Oxygen walks and keto
tautomers of phenols. J. Am. Chem. Soc. 96:5578-5580.
Jerina, D.M., H. Selander, H. Yagi, M.C. Wells, J.F. Davy, V. Mahadevan,
and D.T. Gibson. 1976. Dihydrodiols from anthracene and
phenanthrene. J. Am. Chem. Soc. 98:5988-5996.
Jungclaus, G.A., V. Lopez-Avila, and R. A. Hites. 1978. Organic compounds
in an industrial wastewater: a case study of their environmental
impact. Environ. Sci. Technol. 12:88-96.
Khan, A.U. and M. Kasha. 1970. An optical residue singlet oxygen theory
photocarcinogenicity. Anal. N.Y. Acad. Sci. 171:24-33.
Klein, A.E., and N. Pilpel. 1974. The effects of artificial sunlight upon
floating oils. Water Res. 8:79-83.
Lacaze, J.C., and 0. Villedon de Naide. 1976. Influence of illumination
on phytotoxicity of crude oil. Mar. Pollut. Bull. 7:73-76.
Larson, R.A., L.L. Hunt, and D.W. Blankenship. 1977. Formation of toxic
products from a No. 2 fuel oil by photooxidation. Environ. Sci.
Technol. 11:492-496.
Lawler, G.C., W.A. Loong, B.J. Fiorita, and J.L. Laseter. 1977. An auto-
mated glass capillary gas chromatographic system for routine
quantitative analysis. J. Chromatogr. Sci. 15:532-526.
54
-------
Levin, W., A.W. Wood, H. Yagi, P.M. Dansette, D.M. Jerina, and A.H. Conney.
1976. Carcinogem'city of benzo(a)pyrene 4,5-, 7,8-, and 9,10-oxides
on mouse skin. Proc. U.S. Natl. Acad. Sci. 73:243-247.
Lijinsky, W., I. Domsky, G. Mason, H.Y. Ramahi, and T. Safavi. 1963. The
chromatographic determination of trace amounts of polynuclear
hydrocarbons in petrolatum, mineral oil, and coal tar. Anal. Chem.
35:952-956.
Marquardt, H., T. Kuroki, E. Huberman, O.K. Selkirk, C. Heidelberger, P.L.
Grover, and P. Sims. 1972. Malignant transformation of cells derived
from mouse prostate by epoxides and other derivatives of polycyclic
hydrocarbons. Cancer Res. 32:716-720.
May, W.E., S.P. Wasik, and D.H. Freeman. 1978. Determination of the
solubility behavior of some polycyclic aromatic hydrocarbons in water.
Anal. Chem. 50:997-1000.
McCann, J.E. Choi, E. Yamaski, and B. Ames. 1975. Detection of
carcinogens as mutagens in the Salmonel1 a/microsome test: assay of
300 chemicals. Proc. U.S. Natl. Acad. Sci. 72:5135-5139.
McFall, J., W.Y. Huang, and J.L. Laseter. 1978. Organics at the air-water
interface of Lake Pontchartrain. Bull. Environ. Contam. Toxicol.
22:80-87.
Nagata, S., and G. Kondo. 1977. Photo-oxidation of crude oils. In:
Proc. Oil Spill Conference (Prevention, Behavior, Control, Cleanup),
March 8-10, 1977, New Orleans, LA. pp. 617-620.
Newman, M.S., and S. Blum. 1964. A new cyclization reaction leading to
epoxides of aromatic hydrocarbons. J. Am. Chem. Soc. 86:5598-5600.
Pancirov, R.J., and R.A. Brown. 1977. Polynuclear aromatic hydrocarbons
in marine tissues. Environ. Sci. Technol. 11:989-992.
Patel, J.R., G.W. Griffin, and J.L. Laseter. 1978a. Determination of
arene oxides by a gas chromatography-mass spectrometry system:
Thermal reactions of 9,10-epoxy-9,10-dihydrophenanthrene. Anal. Lett.
Bll(3):239-247.
Patel, J.R., I.R. Politizer, G.W. Griffin, and J.L. Laseter. 1978b. Mass
spectra of the oxygenated products generated from phenanthrene under
simulated environmental conditions. Biomed. Mass Spectrom.
5(12):664-670.
Patel, J.R., G.W. Griffin, and J.L. Laseter. 1978c. The reaction of
polycyclic aromatic hydrocarbons with naturally occurring singlet
oxygen: photooxidation of phenanthrene. Thirty-fourth ACS Southwest
Regional Meeting, Corpus Christi, TX, November 29-December 1, 1978.
55
-------
Payne, J.F., I. Martina, and A. Rahimtula. 1978. Crankcase oils: Are
they a major mutagenic burden in the aquatic environment Science
200:329-330.
Politizer, I.R., G.W. Griffin, and J.L. Laseter. 1971. Singlet oxygen and
biological systems. Chem. Biol. Interactions 3:73-93.
Reed, W.E. 1977. Molecular compositions of weathered petroleum and
comparison with its possible source. Geochem. Cosmochem. Acta 41:237-
247.
Rio, G., and J. Berthelot. 1972. Photooxydation sensibilise du dimethoxy-
9,10-phenanthrene: un dioxetane rapidement dissociable a basse temper-
ature. Bull. Soc. Chem. Fr. pp. 822-824.
Schabron, J.F., R.J. Hurtubise, and H.F. Silver. 1977. Separation of
hydroaromatics and polycyclic aromatic hydrocarbons and determination
of tetralin and napthalene in coal-derived solvents. Anal. Chem.
48:2253-2260.
Scheier, A., and D. Gominger. 1976. A preliminary study of the toxic
effects of irradiated vs non-irradiated water soluble fractions of No. 2
fuel oil. Bull. Environ. Contam. Toxicol. 16:595-603.
Shudo, K., and T. Okamoto. 1973. Wavelength-dependent photolysis. Chem.
Pharmacol. Bull. 21:2809-2810.
Sims, P., and P.L. Grover. 1974. Epoxides in polycyclic aromatic
hydrocarbon metabolism and carcinogenesis. Adv. Cancer Res. 20:165-
274.
Smillie, R.D., D.T. Wang, and 0. Meresz. 1978. The use of a combination
of ultraviolet and fluorescence detection for the selective detection
and quantitation of polynuclear aromatic hydrocarbons by high pressure
liquid chromatography. J. Environ. Sci. Health A13:47-59.
Suess, M.J. 1971. 3,4-benzpyrene in aqueous systems. Environ. Lett.
2:131-133.
Tausch, H., and G. Stehlik. 1977. Bestimmung polycyclischer aromaten in
ruB mittels gas chromatographie-massenspektrometerie. Chromatographia
10:350-357.
Wilson, R.D., P.M. Monoghan, A. Osanik, L.C. Price, and M.A. Rogers. 1974.
Natural marine oil seepage. Science 184:857-864.
Winters, K., R. O'Donnell, J.C. Batterton, and C. Van Baalen. 1976. Water
soluble components of four fuel oils: chemical characterization and
effects on growth of microalagae. Mar. Biol. 36:269-276.
56
-------
Winters, K., and P.L. Parker. 1977. Water soluble components of crude
oils, fuel oils, and used crank-case oils. In: Proc. Oil Spill
Conference (Prevention, Behavior, Control, Cleanup), 8-10 March, 1977,
New Orleans, LA. pp. 579-581.
Woo, C.S., A.P. D'Silva, V.A. Fassel, and G.J. Oestreich. 1978.
Polynuclear aromatic hydrocarbons in coal: identification by their
x-ray excited optical luminescence. Environ. Sci. Technol.
12:173-174.
Yang, S.K., P.P. Roller, and H.V. Gelboin. 1978. Benzo(a)pyrene
metabolism: Mechanism in the formation of epoxides, phenols,
dihydrodiols, and the 7,8-diol-9,10-epoxides. In: Carcinogens!s,
Vol. 3: Polynuclear Aromatic Hydrocarbons, P.W. Jones and R.I.
Freudenthal, Eds., Raven Press, New York. pp. 285-301.
Zepp, R.G., N.L. Wolfe, G.L. Banghaman, and R.C. Hollis. 1977. Singlet
oxygen in natural waters. Nature 267:421-423.
The following related information was published after submission of this
manuscript:
Patel, J.R., E.B. Overton, and J.L. Laseter. 1979. Environmental
photooxidation of dibenzothiophenes following the Amoco Cadiz oil
spill. Chemosphere 8:557-561.
Patel, J.R., S.W. Mascarella, E.B. Overton, and J.L. Laseter. 1979.
Monitoring the photooxidation of Crude Oil by a gas chromatography -
mass spectrometry system: extructed ion current profiles of
dibenzothiophenes and their sulfoxides. Annual Conference on Mass
Spectrometry and Allied Topics, Seattle, WA., June 3-8, 1979, MAMOA9.
Overton, E.B., J.R. Patel, and J.L. Laseter. 1979. Chemical
characterization of mousse and selected environmental samples from the
Amoco Cadiz oil spill. In: Proc. Oil Spill Conference, American
Petroleum Institute, Publ. 79, 4308. (Prevention, Behavior, Control,
Cleanup) pp. 169-174.
Patel, J.R., S.W. Mascarella, G.W. Griffin, and J.L. Laseter. 1979.
Identification and differentiation of isomeric methylated derivatives
of oxygenation products of polycyclic aromatic hydrocarbons by gas
chromatography - mass spectrometry. Anal. Lett. 12(B11):1179-1188.
57
-------
THE MONITORING OF SUBSTANCES IN MARINE WATERS FOR GENETIC ACTIVITY
by
James M. Parry, M.A.J. Al-Mossawi, and N. Danford
Department of Genetics
University College of Swansea
Singleton Park, Swansea SA2 8PP, United Kingdom
and
J. Ballantine
Department of Chemistry
University of College of Swansea
Singleton Park, Swansea SA2 8PP, United Kingdom
ABSTRACT
Extracts of tissue from the edible mussel, Mytilus edulis.
were screened for mutagenic activity using Salmonella
typhimuriym and a variety of strains of Escherichia coli.
Genetically active material was found in the mussel extract
from six of eight sites sampled in the United Kingdom. An
assay of the individual tissue types from mussels at the
Mumbles, Wales, site indicated that the mantle tissue was the
major site of the genetically active chemical. This chemical
was provisionally identified as di-2-ethylhexyl phthalate and
was shown to induce mutation predominantly by insertion or
deletion of nucleotide bases.
INTRODUCTION
The use of the marine environment as a sink for the disposal of
chemicals both deliberately and accidently suggests that at least a
fraction of the living organisms found in the seas and oceans may be
exposed to potentially mutagenic agents. Such exposure may result in
changes in the genetic architecture of marine populations; if such agents
enter food chains, they may lead to the unwitting exposure of human
populations.
A variety of screening systems have been developed for the use of
biological indicator organisms in the detection of the possible mutagenic
activity of environmental chemicals. Such systems involve the use of
58
-------
organisms varying from relatively simple bacteria to mammalian cells in
tissue culture. The systems utilize the procedure of exposing cells of
known genotype to test agents and the plating of treated cells upon
selective medium to detect those cells which have been mutated by the
agent. In a large number of cases, mutagem'c activity in the screening
systems has been shown to be correlated with the ability of a chemical to
produce tumours either in experimental animals or humans (McCann et al.
1975).
The assay of the seas and oceans for the presence of mutagenic
chemicals can be performed in two fundamentally different ways. Chemical
analysis can be performed with ocean samples to identify constituent
chemicals, which can be individually tested for mutagenic activity. This
is a formidable task because of the very large number of chemicals both
natural and man-made that are detectable and must therefore be screened
for mutagenic activity. However, such studies performed on the
constituents of some samples of drinking water in the U.S. have revealed
the presence of mutagenic chemicals (Simmons ert jil_., 1977).
The second approach, involves the collection of ocean samples,
concentration of the constituents, and the exposure of the resulting
concentrate to a range of mutagenic screening systems. However, this
procedure has at least two fundamental limitations: the problem of
developing a suitable concentration procedure and, perhaps more
importantly, the difficulty of detecting mutagens in the presence of toxic
chemicals. For example, a mutagenic chemical might be undetectable if the
assay is performed in the presence of a chemical which is highly toxic to
the test organism.
We attempted to overcome the problems of the latter procedure by using
marine organisms as concentrators of potentially mutagenic chemicals. By
the use of such living organisms, we have also selected against the
possible masking effects of toxic chemicals on the assumption that such
toxicity would lead to the death of the marine species in question. We
have developed an assay system based upon the extraction of tissues
derived from the mussel, Mytilus edulis, and the screening of these
extracts for mutagenic activity, using a number of microbial indicator
species. We were able to identify the presence of an industrial chemical
capable of inducing mutation in bacteria.
MATERIALS AND METHODS
Strains
The strains of bacteria used in these studies included cultures of
Salmonella typhimurium, auxotrophic for histidine (kindly provided by Dr.
Bruce Ames) and a variety of Escherichia coli strains (kindly provided by
Dr. G. Mohn, Dr. M. Green, and Dr. D. Tweats or our laboratory). They
have been described in the literature (Parry ot_ jil_., 1976) and are capable
of detecting both frameshift and base-substitution mutagens.
59
-------
The yeast strain JD1 was auxotrophic for both histidine and tryptophan
and produced prototrophic colinies by the process of mitotic gene
conversion, a process of genetic change that responds to treatment by
mutagens and carcinogens in an essentially non-specific manner. The use
of this yeast strain has been described in detail elsewhere (Davies et
al_., 1975).
Preparation of Mussel Extracts
Sanples of the mussel, Mytilus edulis, were collected from a variety
of sites, washed with running water and either extracted immediately or
kept frozen at -20° C until required. (There was no evidence in our work
of any reduction in genetic activity during storage.)
For the preparation of extracts of whole mussel tissue, 50 shelled
mussels were extracted in 100 mis of 95% ethanol after disintegration of
the tissue in a Uaring blender or Atomix. The resulting tissue homogenate
was left to stand overnight at 4° C, resuspended and centrifuged at 5000 g
for 15 nin. The supernatant was sterilized by passing through a membrane
filter and stored at 4° C for use within one week and at -20° C for
long-term use. All the samples obtained were tested for the presence of
radioactive material. None of the samples described here showed levels of
radioactivity above the background.
Detection of Genetic Activity
Fluctuation tests were carried out, with minor modifications, as
described by Green et_ ^1_. (1976). Overnight cultures of E. col i and _S^
tryphimurium tester strains were prepared in supplemented Davis-Mingioli
minimal medium. After two rinses in saline, the cells were resuspended in
saline at a concentration of 5.0 x 10 cells per m£; 100 ml Davis-Mingioli
basal salts were combined with 0.7 m£ 40% glucose, 0.1 ma tryptophan
solution (200 ug/nu) (for fluctuation tests involving tryptophan
auxotrophs only), or 0.1 m lysine solution (200 pg/ntt) (for lysine
auxotrophs only), or 0.1 RU of histidine solution (200 mg/m£) (for
histidine auxotrophs only), 0.1 mi washed cells, and, in the case of the
test treatments, either methyl methane sulphonate (final concentration,
1 yg/nfc) as a positive control, or 0.1 mi of alcoholic mussel extract.
For experiments involving the use of multiple auxotrophs, the supplements
required by the non-selected markers were added in excess. Control
experiments involving mussel extracts also contained 0.1 mi 95% ethanol as
a negative control. Each treatment was dispensed in mi samples into 50
test tubes or in some cases 1 n\i samples into 100 test tubes. After the
auxotrophic bacteria exhausted the small amount of supplement present,
only prototrophic revertants continue to grow. From 2 days and thereafter,
tubes in which mutation had occurred became turbid, while other tubes
remained relatively clear. The number of turbid tubes was routinely
scored after 3 days. The significance of the response of each set of 50
tubes to the presence of mussel extract was determined by the use of
Chi-square analysis as described by Green et. al_. (1976). The arginine-56
mutation used in some experiments is leaky, and as a result some residual
60
-------
growth was observed even in the absence of arginine. Therefore, trace
amounts of arginine were not added when this marker was used.
In a number of experiments, the fluctuation test was modified
according to the procedure developed by Dr. David Gatehouse (personal
communication): samples were added to 96-well microtitre plates (Sterilin,
Ltd.) in the form of 0.2 mi aliquots per well. All other steps in the
procedure were as described above.
In the yeast assays, we used stationary phase cultures of the strain
JD1 suspended in saline at a concentration of 10 cells/nu. Samples of
the mussel extracts (up to 4%) were added to the yeast minimal medium at
45°C together with yeast cells at a concentration of 35 x 10 cells/m&
for the detection of histidine-independent prototrophs and 35 x 10
cells/mi for the detection of tryptophan-independent prototrophs and
poured into 9-cm Petri dishes. The culture medium was supplemented with
20 ug/mz tryptophan and 0.1 ug/m£ histidine for the detection of histidine-
independent prototrophs and with 20 yg/mji histidine and 0.1 ug/nu
tryptophan for the detection of tryptophan-independent prototrophs. These
supplements enabled the auxotrophic cells to undergo three cell divisions
in the presence or absence of mussel extracts. All plates were grown in
the dark at 28° C and scored after 9 days of incubation. In all
experiments involving the yeast cultures, we used at least five replicate
plates per treatment.
CHEMICAL ANALYSIS OF MUSSEL EXTRACTS
Fifty-gram samples of mussel mantle tissue were homogenized in 1 a 2:1
chloroform/methanol in a blender. The homogenized material was left
overnight and then filtered and treated with 200 m£ 0.05M KC1. The
resulting organic layer was dried with 50 gm of sodium sulphate and
filtered; the whole extract was evaporated to dryness in a rotary
evaporator. The residual material was taken up in 5 mi pentane and
applied to a 200 ma silica-gel column. Solvents then were run through the
column in the following order: 1. pentane, 2. 95% pentane + 5% ether,
and 3. chloroform. The individual ellutants were collected and rotary
dried; the residual material from each sample was taken up in either
ethanol or DMSO and tested for mutagenicity.
After the major fraction of the detectable mutagenic activity was
verified in the chloroform extract, the sample was separated further. A
portion of the chloroform extract was applied to a preparative thin layer
chromatography plate. The plate was then run in one dimension, using
chloroform. After the solvent front reached the end of the plate, the
plate was dried and examined. Of the 9 labelled zones, 3, 6, and 8 were
fluorescent when observed in UV light. Each zone was scraped from the
plate and shaken up with 95% ethanol; the sample obtained was filtered
through a membrane filter and tested individually for mutagenicity. All
samples were coded by a colleague and tested blind.
61
-------
RESULTS
Geographic Variation in the Presence of Genetically Active Chemicals
Alcoholic extracts of the whole tissue of mussels collected from a
variety of geographical locations were screened for the presence of
genetically active chemicals by the measurement of cells prototrophic for
histidine and tryptophan produced by induced mitotic gene conversion in
the yeast, Saccharomyces cerevisiae. The sites sampled were Swansea,
Mumbles, Caswel1 Bay, Solva, Llangranog, and Milford Haven in South Wales,
Anglesey in North Wales, and Plymouth in Southwest England. The samples
were screened in two experimental series as shown in Table 1, using
ethanol as a negative control and the alkylating agent ethyl methane
sulphonate, as a positive control.
The results (Table 1) demonstrate the presence of genetically active
material in the samples collected from Plymouth, Caswell Bay, Mumbles,
Swansea, Milford Haven, and Llangranog; no such activity was detectable in
samples collected from Anglesey and Solva.
Further confirmation of the site variation in genetic activity was
obtained by examining alcoholic extracts of mussels collected on the same
day at a series of sites westward from Mumbles. These samples were
screened for genetic activity by the measurement of mitotic gene
conversion, using the yeast strain JD1. The results obtained from these
assays (Table 2) demonstrate that genetic activity decreased in the
samples collected at increasing distances from the Mumbles site. The
decreased genetic activity correlates with reduced visual pollution and
increased distance from the industrial area surrounding Swansea Bay.
Tests for the Presence of Nutrients in the Mussel Extracts
The majority of the tests to detect genetic activity in mussel tissues
are based upon the measurement of the production of prototrophic colonies
in auxotrophic microbial cultures. Assays of this type may lead to
inconclusive results if the samples contain significant amounts of the
specific nutrient deleted from the selective media. In such cases,
additional growth of the test culture would lead to a spurious positive
result. Thus, it was necessary for us to ensure that all the samples did
not contain significant amounts of nutrients.
In the case of the bacterial fluctuation tests, superficial estimates
of the presence of nutrients can be made by visual and colorimetric
examination of the clear tubes, i.e., those that show only background
growth due to the presence of the original auxotrophic cells. When
significant amounts of nutrients were present in the mussel extracts,
increased background growth in the tubes was readily observable. If free
nutrients were present, the samples were rejected for assay, or used in a
mutagenicity assay that did not require the contaminating nutrient was
used.
62
-------
TABLE 1. THE INDUCTION OF MITOTIC GENE CONVERSION IN YEAST IN THE PRESENCE OF A
VARIETY OF MUSSEL EXTRACTS
Ol
CO
Treatment
Experimental series 1
Control 1
Control 2
4% Alcohol control 1 Negatiye Controls
4% Alcohol control 2
Plymouth extract 4%
Mumbles extract 4%
Caswell Bay 4% extract
Anglesey 4% extract
Ethyl methane sulphonate (lug/mS.)
- Positive Control
Experimental series 2
Control 1
Control 2
Swansea (St. Helens) 4%
Solva (S. Wales) 4%
Milford Haven (S. Wales) 4%
Llangranog (S. Wales) 4%
4% Alcohol control 1
4% Alcohol control 2
Ethyl methane sulphonate (lug/mj,)
Mean no. of
his4
prototrophs
per plate
8.6
7.8
6.5
4.2
202.6
71.2
69.0
4.7
284.7
21.4
28.7
91.0
19.4
53.6
83.1
17.3
15.9
317.2
his"'
prototrophs
per 10s
survivors
±s.e.
26.8 ± 4.6
32.5 ± 5.8
25.6 ± 5.0
11.8 ± 3.0
493.6 ±21.3
323.6 ±19.2
276.0 ±16.6
18.3 ± 4.2
900.9 ±26.7
62.2 ± 8.4
84.9 ± 9.7
256.7 ±12.3
53.2 ± 9.1
175.4 ±12.1
284.7 ±11.9
65.4 ± 8.4
47.3 ± 7.1
1,050.7 ±29.6
Mean no. of
trp+
prototrophs
per plate
7.4
5.2
9.4
7.6
154.8
85.3
69.2
7.6
346.2
11.4
9.3
68.2
8.7
61.3
114.1
5.9
7.2
214.9
trpr
prototrophs
per 10-'
survivors
±s.e.
23.1 ± 4.3
21.7 ± 4.8
37.0 ± 6.0
21.4 ± 3.9
463.5 ± 18.6
387.7 ± 20.6
276.8 ± 16.6
29.6 ± 5.4
1,095.0 ± 29.4
33.1 ± 5.1
27.6 ± 4.8
192.6 ± 11.6
23.9 ± 8.4
200.4 ± 15.6
390.9 ± 19.7
22.3 ± 4.3
21.1 ± 3.1
712.3 ± 27.0
% Viability
91.4
68.6
73.1
101.4
95.4
62.9
71.4
73.4
90.3
98.3
96.5
101.3
104.1
87.3
83.4
75.6
97.6
86.2
-------
TABLE 2. THE INDUCTION OF MITOTIC GENE CONVERSION IN YEAST IN THE PRESENCE
OF MUSSEL EXTRACTS COLLECTED IN THE GOWER AREA OF SOUTH WALES
Control I Ufo ethanol
Control 2 k% ethanol
Mumbles 4$
Caswell Bay k%
Oxwich k%
Port Eynon t\%
Rhossilli 4$
his
prototrophs
per 1O6
survivors
is. e.
19.7 i 1.9
24.2 - 2.2
621.4 i 11.1
492.9 - 7.0
34.2 - 1.8
29.0 i 1.7
31.3 - 2.5
trp*
prototrophs
per 1O5
survivors
is.e.
7.1 i 1.2
8.9 i 1.3
247.5 i 7.O
132. O i 3.6
15.1 i 1.2
9.2 i 0.9
6.4l U.I
$ cell
viability
84.2
73.3
65.4
78.5
82.2
74. 0
68.6
-------
Samples that showed positive genetic activity were further
investigated for the presence of free nutrients by making viable counts of
the numbers of bacterial cells present in both the clear and turbid tubes
in the fluctuation tests. A typical example of the results of such an
assay (Table 3) shows the viable cell counts made from individual
fluctuation test tubes, using the Escherichia coli culture uvrAtrp(R46) in
both the presence and absence of Mumbles mussel extract. The results
demonstrate that the median and mean numbers of viable bacterial cells per
clear tube, that is, in tubes that do not contain a significant number of
trp revertants, were unaffected by the presence of mussel extracts.
Therefore, we can conclude that the particular sample of Mumbles mussel
extract did not enhance the growth of trp-cells when the added tryptophan
supply had become exhausted. We can further conclude that the observed
increase in the frequency of turbid tubes in the bacterial fluctuation
test was not due to increased cell division in the presence of free
nutrients, which would have allowed the production of an increased
frequency of spontaneous mutants.
The possible nutrient effects of the mussel extracts upon auxotrophic
yeast cultures were determined in two ways:
1. After the incubation period, the plates of selective agar were scored
for the frequency of prototrophic colonies and then discs of agar 9 mm
in diameter were removed from areas of the plates showing background
growth, but not large prototrophic colonies. Agar discs sampled from
10 plates per treatment were added to 20 rn sterile saline; yeast
cells contained in the discs were suspended by sonication. After
appropriate dilutions, the saline suspensions were plated upon yeast
complete agar, and the colonies produced were counted after 5 days
incubation at 28°. Comparisons of the viable cells contained in the
background agar of both the treated and untreated cultures then could
be used to determine if any further growth of auxotrophic yeast cells
had taken place in the solid medium in the presence of the mussel
extract. The results of such an assay using the mussel extract
derived from the Plymouth site (Table 4) demonstrate that there are no
significant differences in the growth of auxotrophic cells between the
plates with and without mussel extract.
2. Samples of the mussel extracts were added to liquid minimal medium
containing either excess histidine or excess tryptophan together with
10 yeast cells/mJi of strain JD1 auxotrophic for both histidine and
tryptophan and the cultures were aerated for periods of up to 48 hr.
During this period cell growth was determined by counting cell numbers
with a heamocytometer. Under these conditions, samples that contain
nutrients showed increases in cell numbers during the first 6 hr
incubation. In samples lacking nutrients, no increases in cell
numbers were observed until after approximately 24 hr when the growth
of pre-existing prototrophic cells become significant. The results of
a typical experiment of this type are shown in Figure 1.
65
-------
Cell
number!2 -
xlO5
is + trp
1% mussel extract
control
Time of incubation at 28 in hours
Figure 1. The effects of the addition of mussel extract upon the growth of yeast cells of strain ,1DI in
Yeast Minima Medium; A -medium supplemented with 20 ug/nu histidine and tryptophan- o
-medium supplemented with 20 ug/mji histidine only; . -medium supplemented with 20 vg/mi
histidine and 1% ethanol; D -medium supplemented with 20 ug/m* histidine and 1% mussel
extract;A -medium supplemented with 20 yg/m* trytophan and 1% mussel extract.
-------
TABLE 3. VIABLE CELL COUNTS OF TURBID AND CLEAR TUBES FROM A FLUCTUATION TEST
INVOLVING THE USE OF ESCHERICHIA COLI CULTURE UVR A TRP (R46) WITH
AND WITHOUT THE ADDITION OF MUSSEL TISSUE EXTRACT FROM THE MUMBLES
AREA
Control
Control + O.I m#- ethanol
O.I mi Mussel extract
Mean value
turbid tubes
5.9* x 108
6.0 x 108
5.3 x 1O8
Mean value
clear tubes
cells/mJl
5.3 x 1O6
5.O x 1O6
5.6 x 1O6
Median value
clear tubes
cells/mS.
1.27 x 10?
1.95 x 1O?
9.O x 1O6
-------
TABLE 4. ASSAY OF THE BACKGROUND GROWTH OF YEAST CELLS FROM THE PLATES OF SELECTIVE
AGAR WITH AND WITHOUT THE PRESENCE OF ADDED MUSSEL EXTRACT (PLYMOUTH)
Treatment
Control
Background growth, mean
of discs removed from
selective plates used
for the detection of
his prototrophs
cells/m£
2.9 - 0.5 x 105
Background growth, mean
of discs removed from
selective plates used
for the detection of
trp prototrophs
cells/mil
.1 - O.8 x 1C/*
Control + 1% ethanol
3.2 - 1.1 x 105
3.7 - 1.2 x 1C/*
CT>
CO
1% Mussel extract
2.4 - O.7 x 1O5
3.5 - 1.1 x
2% Mussel extract
1.9 - 0.8 x 105
4.3 - O.8 x
14% Mussel extract
1.7 0.6 x 10
4.0 - 1.3 x
-------
None of the mussel samples reported here to be mutagenic were
demonstrated to contain significant amounts of the specific nutrient
required by the test microbes. However, we detected such nutrients in
many samples; in particular we found that extracts of fish tissue also
studied in our laboratory contained significant amounts of free amino
acids, particularly histidine.
The Assay of Specific Mussel Tissues
The experiments described above demonstrate the presence of genet-
ically active chemicals in alcoholic extracts derived from all the tissues
of the organism. We also determined whether this activity was distributed
throughout all the tissues of the organism or concentrated within specific
tissues.
Initially we separated the body tissues into (1) the hepatopancreas
and (2) all the remaining tissues; 10 gm of sample (1) and sample (2) were
then extracted in 50 mi 95% ethanol. These extracts of hepatopancreas and
the remaining tissues were screened for genetic activity with a range of
bacterial strains in a series of fluctuation tests. The results of these
fluctuation tests (Table 4) demonstrate, that mutagenic activity was not
detectable in the hepatopancreas but was found in the remaining tissue of
mussels from the Mumbles site.
Further subdivisions of the mussel tissue were made and extracts
prepared from mantle, foot, muscle, and the reproductive system at a
concentration of 10 gm of tissue in 50 ma 95% ethanol. Samples of each of
the tissue extracts were assayed for the presence of genetically active
chemicals by the measurement of induced mutation in a series of fluc-
tuation tests and by the measurement of induced mitotic gene conversion in
yeast.
The results of the assays of the various mussel tissues for genetic
activity are shown in Table 5 (bacterial tests) and Table 6 (yeast tests).
Both the mutation tests with bacteria and those of induced mitotic gene
conversion in yeast demonstrate that the major site of genetically active
chemicals in mussels collected from the Mumbles site was the mantle
tissue.
Identification of the Molecular Nature of Mutagenic Present in the Mussel
Extract
Mutagenic chemicals may be classified on the basis of the type of
change that they produce in cellular DNA. The strains of bacteria used in
this work made it possible to determine whether a mutagenic chemical
produces predominantly base change mutations leading either to missense or
nonsense changes at the level of protein synthesis or to the insertion or
deletion of nucleotide bases leading to a change in the reading frame of
at the level of protein synthesis (frameshift).
69
-------
TABLE 5. EFFECTS OF MUSSEL EXTRACTS FROM A VARIETY OF DIFFERENT TISSUES UPON
INDUCED MUTATION, MEASURED IN BACTERIAL FLUCTUATION TESTS USING
ESCHERICHIA COLI
No. of tubes positive
Strain
341/113
341/113
(R46)
uvr A (R46)
343/113
343/113
(R46)
uvr A (R46)
Locus Treatment
arg 56 MMS
Hepatopancreas
Remaining tissue
arg 56 MMS
Hepatopancreas
Remaining tissue
trp MMS
Hepatopancreas
Remaining tissue
arg 56 MMS
Mantle
Foot
lys 60 MMS
Mantle
Foot
trp Mantle
Foot
Reproductive system
No. of
tubes
sampled
50
50
50
50
50
50
100
100
100
50
50
50
50
50
50
50
50
50
Control
24
24
24
28
35
35
54
48
48
21
24
24
28
28
28
15
15
15
Treated
41
34
41
48
41
45
87
59
70
39
44
34
40
40
35
28
18
18
Signi ficance
(probabil i ty)
< .01
NS
< .01
X .001
• MS
< .01
< .001
NS
< .05
< .01
> .05
NS
< .02
< .02
NS
< .02
NS
NS
MMS = Positive control methyl methane sulphonate
NS = not significant
-------
TABLE 6. EFFECTS OF MUSSEL EXTRACTS FROM A VARIETY OF DIFFERENT TISSUES UPON
THE INDUCTION OF MITOTIC GENE CONVERSION IN YEAST STRAIN JD1
Treatment
Control
Negative control 2% ethanol
Muscle 1%
2%
Mantle 1%
2X
Hepatopancreas 2%
Foot 2%
Reproductive system 2%
Viability
100.0
101.3
97.3
94.4
101.9
93.2
85.3
92.4
76.8
trp+
prototrophs
per 105
survivors
is.e.
19.1 + 1.8
22.4 ± 1.7
16.6 ± 1.4
29.3 + 1.8
94.5 ± 3.9
135.7 ± 8.2
27.9±+ 1.5
17.6 + 0.9
14.3 ± 2.3
Ms+
prototrophs
per 106
survivors
is.e.
36.7 ± 2.3
32.4 i 1.9
37.5 i 2.3
41.0 i 2.9
170.3 ill. 6
589.6 i23.4
41.6 i 2.7
39.4 i 2.1
20.7 ± 1.9
-------
Table 7 summarizes a number of experiments that were performed with
the mussel extract from the Mumbles site, using a variety of bacterial
cultures in fluctuation tests. The cultures of Escherichia coli used in
the tests have been described in Material and Methods and are capable of
detecting the presence of both frameshift and base change mutagens by the
induction of prototrophs at specific genetic markers.
As seen in Table 7, the extract of mantle tissue from mussels
collected at Mumbles was capable of inducing prototrophs only at those
loci, i.e. lys 60, arg 56. and gal Rs, and of reverting by frameshift
events. It was without activity at the trp locus, which is capable of
reversion by base change mutation. Thus we conclude that the particular
chemical agent present in the mantle tissue of mussels collected from the
Mumbles site is capable of inducing mutation predominantly by the mech-
anism of insertion or deletion of nucleotide bases.
Chemical analysis of the contents of mussel extracts
To determine the nature of those chemicals present in the mussel
extracts which show genetic activity, we prepared the samples in 2:1
chloroform/methanol as described in Materials and Methods. After
fractionation on a silica-gel column extracts made up in pentane, 95%
pentane + 5% ether and chloroform were obtained and tested for mutagenic
activity with the Salmonella typhimurium strain TA98 in a series of
fluctuation tests.
The results obtained for pentane, pentane and ether, and chloroform
extracts of all mussels from Mumbles (Table 8) demonstrate that the
chlorofrom extract contains the major activity as shown by an increase in
mutation at the his D locus of Salmonella strain TA98 measured in
fluctuation test experiments, using three separate extractions of mussel
tissue from the Mumbles site.
The chloroform extract was further separated by thin layer chroma-
tography to produce nine separate samples labelled 1 to 9, three of which
(namely 3,6, and 8) were fluorescent when observed on a chromatography
plate. Each zone was removed from the plate and shaken with 95% ethanol
and the resulting extracts tested for the presence of mutagenic chemicals.
The results of the mutagenicity tests of the extracts of the zones using
bacterial fluctuation tests involving the Salmonella strain TA98 (Table 9)
demonstrate that the zones 3, 6, and 8 contain mutagenic chemicals that
produce a significant increase in the number of positive tubes in the
fluctuation tests.
The residue from the UV fluorescent band 3 (Rf 0.74) was examined by
proton magnetic resonance with a high sensitivity F.T. instrument (Varion
XL-100). The presence of a long chain phthalate ester was stabilized and
identified by its characteristic mass spectrum, as di-2-ethylhexyl
phthalate, with a high resolution MS-9 mass spectrometer. The nmr and
mass spectra were identical to a commercial sample of di-2-ethylhexyl
phthalate. A quantitative estimation of the phthalate by nmr indicated
the presence of approximatley 11 yg per mussel.
72
-------
TABLE 7. SPECIFICITY OF ACTION OF ALCOHOLIC EXTRACTS OF MANTLE AND FOOT
TISSUES OF MUSSELS COLLECTED FROM MUMBLES UPON MUTATION
INDUCTION IN THE BACTERIAL FLUCTUATION TEST USING CULTURES OF
ESCHERICHIA COLI
343/113 (R46) lys 60
predominantly
frame shift
343/113 arg 56
base change
and frames hi ft
343/113 gal RS
base change
and frames hi ft
WP2 uvrA trp
base change
Expt 1 MMS
Mantle
Foot
Expt 2 Mantle
t" GG v.
Exot 3 Mantle
Foot
Expt 4 MMS
Mantle
Foot
Expt 1 MMS
Mantle
Foot
Expt 2 Mantle
Foot
Expt 1 Mantle
Foot
Expt 1 KMS
Mantle
Foot
Expt 2 MMS
Mantle
Foot
Expt 3 Mantle
Foot
50
100
100
£0
50
50
50
50
50
50
50
50
50
50
50
75
75
50
100
100
50
50
50
50
50
28
- 42
42
28
28
16
16
28
28
28
21
21
21
19
19
33
33
15
24
24
15
15
15
9
9
40
74
61
42
31
34
15
40
40
33
39
34
24
30
19
51
39
32
32
26
31
16
16
16
10
<.C2
<.0i
fiS
<.:i
v-
iO
<.G1
NS
< _Q2
<'.02
MS
<.01
<.02
MS
<.05
KS
<.G1
NS
<.C01
NS
NS
<.01
MS
KS
NS
NS
MMS = positive control mutagen methyl methane sulphonate.
NS = no significant difference in the frequency of positive tubes in the
control and treated series of tubes.
-------
TABLE 8. GENETIC EFFECTS OF SOLVENT EXTRACTS AS MEASURED IN BACTERIAL
FLUCTUATION TESTS USING THE STRAIN SALMONELLA TYPHIMURIUM TA 98
Strain Location Solvent No.of tubes No.of tubes Significance
(°i solution) tested positive
Control Treated
TA 98
Mumbles
.1 Pentane
.1 Pentane +
ether
.1 Chloroform
50
50
50
10
10
10
11
16
26
NS
NS
< .01
TA 98
Mumbles
.1 Pentane
.1 Pentane +
ether
.1 Chloroform
50
50
50
8
8
8
9
7
28
NS
NS
< .001
TA 98 Mumbles .1 Pentane 50
.1 Pentane +
ether
.1 Chloroform
50
50
10
10
10
13
12
23
NS
NS
< .01
74
-------
TABLE 9. GENETIC EFFECTS OF EXTRACTS OF EACH ZONE MEASURED IN A
BACTERIAL FLUCTUATION TEST USING BACTERIAL STRAIN
SALMONELLA TYPHIMURIUM TA 98
Strain Zone No.of tubes No.of tubes oositive Significance
sampled Control Treated
TA 98
1 50 10 9 NS
2 50 10 15 NS
* 3 50 10 22 < .02
4 50 10 13 NS
5 50 10 12 NS
* 6 50 10 23 < .01
7 50 10 11 NS
* 8 50 10 21 < .05
9 50 10 20 NS
Zone 6 was identified as containing 2 - ethylhexyl phthalate
75
-------
Because phthalate esters are known to be found in biological extracts
due to the extraction of residual plasticisers in solvents, plastic
utensils, and tubing, it was considered necessary to carry out a blank
extraction procedure as far as the silica gel column stage with the same
volumes of solvents and the same apparatus used in the mussel extraction.
The residue at the end of this blank extraction procedure was shown by
nmr and mass spectra to contain only negligible quantities of phthalates,
thus demonstrating that the phthalate isolated from the mussels was
genuinely present in the mussel tissue.
To confirm the mutagenic potential of phthalate esters, we
investigated the activity of a commercial sample of phthalate esters
di-iso-octyl phthalate (B.D.H. Poole, Dorset), which was known to contain
high levels of di-2-ethylhexyl phthalate. The results (Table 10) obtained
with a variety of strains of bacterial cultures in a series of fluctuation
tests demonstrate that significant increases in mutation were detected
with concentrations of 1 and 5 yg/mi di-iso-octyl phthalate in Salmonella
strain TA98 and in some experiments with the Escherichia coli strain
343/113.
Further confirmation of the mutagenic activity of di-iso-octyl
phthalate was provided by a series of fluctuation tests, using the
Salmonella strain TA98 in the presence of a rat liver microsome extract
(S-9 mix) and appropriate co-factors. The results (Figure 2) demonstrate
that concentrations of di-iso-octyl phthalate of 2 to 4.5 yg/nu produce
significant increases in the frequency of turbid tubes produced by
mutation to prototrophy. The results also indicate that at higher
concentrations of di-iso-octyl phthalate, the number of positive turbid
tubes was reduced presumably due to cell toxicity.
DISCUSSION
The results presented here demonstrate the practical value of
microbial assay systems for the detection of genetically active chemicals
in the tissue of the edible mussel, Mytilus edulis. We have found from
selected sites around the coast of the United Kingdom that mantle tissues
of mussels may contain chemicals which are extractable in ethanol and are
capable of inducing mutation in bacteria and mitotic gene conversion in
yeast. The presence of nutrients, such as free ami no acids in the tissues
of many marine species, prevents the uncritical use of the standard
microbial mutagenicity assay systems with extracts of all marine
organisms. However, with appropriate technical modifications, we have
been able to utilize the techniques with a diverse range of marine
organisms such as plankton, oysters, crabs, and a number of fish species.
The genetic techniques described here demonstrate only the presence of
genetically active materials. Although we were able to show that at least
one sample contains a predominantly frameshift mutagen, these techniques
did not provide direct information of the chemical nature of the
contaminating agent.
76
-------
TABLE 10. GENETIC EFFECTS OF DI-ISO-OCTYL PHTHALATE AS MEASURED IN A
SERIES OF FLUCTUATION TESTS USING CULTURES OF ESCHERICHIA
COLI AND SALMONELLA TYPHIMURIUM
Strain Concentration
of phthalate
in ygm/mi
Salmonella
typhimurium
TA98
TA98
TA1538
TA100
Escherichia coli
343/113
lys 60
343/113
lys 60
1
5
10
1
5
10
1
5
10
1
5
10
1
5
10
1
5
10
No. of tubes
sampled
50
50
50
96
96
96
96
96
96
96
96
96
96
96
96
96
96
96
No. of tubes positive
Control Treated
13
13
13
2
2
2
2
2
2
19
19
19
6
6
6
14
14
14
18
26
18
9
12
6
9
6
6
18
27
17
10
23
8
19
18
10
Significance
NS
< .02
NS
< .05
< .01
NS
< .05
NS
NS
NS
NS
NS
NS
< .001
NS
NS
NS
NS
77
-------
100-1
90-
Fluctuation Test using
Salmonella typhimuriur
80-
70-
«60-
0)
.a
•4-1
o>50-
+j
'35
Q-40-
o
o
Z30-
20-
10-.
•5% Probability value
Figure 2.
I
2
3
i
4
i
5
I
6
fjgms/roP
diso octyl phthalate + 89 mix
The effects of di-iso-octyl phthalate in the presence of S9 mix
and co-factors upon the induction of turbid tubes produced by
mutation to prototrophy in Salmonella typhimurium strain TA98.
The test was performed with 96 well microtitre plates. The
figure shows the number of positive tubes produced by
concentrations of up to 5 ug/nui di-iso-octyl phthalate
dissolved in 95% ethanol. All results above the dotted line
are significant at the 5% level.
78
-------
The individual mussel samples from the different sites probably
contain a diverse collection of chemicals; it is unlikely that the
activity we detected derives either from a single or group of chemicals.
It is likely that each of the sites represent a different pool of
genetically active chemicals that will require characterization. It is
also unlikely that extraction of samples in only a single solvent (in our
case, ethanol) would provide a complete extraction of all chemicals
capable of inducing genetic change. However, extraction of mussel tissue
with a variety of solvents demonstrated that the major proportion of
detectable genetic activity was extractable with chloroform. Chemical
analysis of the chloroform extract indiciated the presence of phthalate
esters at concentrations as high as 11 ug per mussel.
Assays of the mutagenic potential of the one of the identifiable
phthalate esters, i.e., di-2-ethylhexyl phthalate, demonstrated that the
chemical was active in bacterial fluctuation tests both in the presence
and absence of a rat liver microsome fraction. Experiments are still in
progress to determine the mutagenic potential of the remaining chemicals
detectable in the mussel extracts.
Our work demonstrated the presence in mussels of chemicals which
cause mutation in microbes, but did not provide information on the
potential of these chemicals to produce similar changes in the cells of
higher organisms. However, our techniques do provide useful information
on the chemicals in the marine environment which have the potential to
produce genetic damage in organisms through direct exposure or via
marine food chains. After the identification of the active chemicals by
techniques we have described, the evaluation of their potential hazards
will require the use of a range of test systems of more direct relevance
to the organisms of concern. In this context, it is of interest that the
extracts of mussels collected in the Mumbles area show predominantly
frameshift mutagenic activity. In the case of such mutagens, studies of
enzymatic variants of indicator species would produce only limited
information on the amount of genetic change taking place because variants
would be produced by predominantly base-change mutagens.
79
-------
REFERENCES
Davies, P.O., W.E. Evans, and J.M. Parry. 1975. Mitotic recombination
induced by chemical and physical agents in the yeast Saccharomyces
cerevisiae. Mutat. Res. 29:301-314.
Green, M.H.L., W.J. Muriel, and B.A. Bridges. 1976. Use of a simplified
fluctuation test to detect low levels of mutagens. Mutat. Res.
38:33-42.
McCann, J., E. Choi, E. Yamasaki, and B.N. Ames. 1975. Detection of
carcinogens as mutagens in the Salmonella/microsome test: assay of
300 chemicals. Proc. Nat. Acad. Sci. 72:5135-5147.
Parry, J.M., D.J. Tweats, and M.A.J. Al-Mossawi. 1976. Monitoring the
marine environment for mutagens. Nature 264:538-540.
Scott, 0., B.A. Bridges, and F.H. Sobels. 1977. Progress in genetic
toxicology. Elsevier/North Holland, Amsterdam.
Simmons, V.F., K. Kauthanen, and R.G. Tarditt. 1977. Mutagenic activity
of chemicals identified in drinking water. In: Progress in genetic
toxicology. Elsevier/North Holland, Amsterdam.
80
-------
BIPHENYL HYDROXYLASE ACTIVITY AND THE
DETECTION OF CARCINOGENS
by
Nancy L. Couse
Department of Biological Sciences,
University of Denver, Denver CO. 80208
and
Josef J. Schmidt-Collerus, Jeannette King, and LaRose Leffler
Denver Research Institute, Chemical Division,
University of Denver, Denver CO. 80208
ABSTRACT
The in vitro stimulation of biphenyl 2-hydroxylase activity
by chemical carcinogens was examined for selected plant and
animal microsomes by spectrophotofluorometric methods. An
apparent increase in 2-hydroxybiphenyl was observed after pre-
incubation of microsomes with carcinogens. In each case, the
increase could be attributed to metabolites of the carcinogen
which fluoresced at the same wavelength used to measure
2-hydroxybiphenyl. Quantification of biphenyl metabolite
production using high pressure liquid chromatography to separate
the compounds showed that preincubation of microsomes with
chemical carcinogens had no effect on 2-hydroxybiphenyl formation
and caused a decrease in 4-hydroxybiphenyl production. By using
terphenyl as a substrate, a minimum of three different metabolites
were formed in vitro by hamster microsomes as determined by high
pressure liquid chromatography. One of these metabolites was not
formed after preincubation of the microsomes with benzo(a)pyrene.
INTRODUCTION
Biphenyl and polychlorinated biphenyls are present in marine, fresh-
water, and terrestrial environments. The microsomal mixed function
oxidases of a number of marine (Willis and Addison, 1974), freshwater
(Creaven et al_., 1965; Willis and Addison, 1974), and terrestrial (Creaven
et al_., 1965; Basu j|t a\_., 1971) animals detoxify biphenyl by formation of
4-hydroxybiphenyl as the major metabolite and 2-hydroxybiphenyl as a minor
metabolite. Depending on the species examined, the ratio of 4-hydroxybi-
phenyl to 2-hydroxybiphenyl ranges from 26:1 to 2:1. Microbial metabolism
-------
of biphenyl has also been examined, and it has been shown that bacteria
forn dihydroxybiphenyls (Catelini et^ aj_., 1970, 1971, 1973, loannides and
Parke, 197G; Cernlglia et al., 1978), whereas fungi form monohydroxy-
biphenyls (McPherson et flU, 1976; Dodge et al-,1978).
Pretreatment of animals with chemical carcinogens such as safrole,
benzo(a)pyrene, or methylcholanthrene results in a selective stimulation of
biphenyl 2-hydroxylase. Metabolite production is measured in vitro with
purified microsomes from treated and untreated animals, biphenyl as a
substrate, and an NADPH-regenerating system. The quantity of 2-hydroxybi-
phenyl and 4-hydroxybiphenyl formed, in most cases, is determined spectro-
photofluorometrically (Creaven jrt _aj_., 1965). The amount of 2-hydroxybi-
phenyl formed by microsomes from treated animals is increased 2- to 20-fold
over the untreated controls (Friedman _et _al_., 1972; Burke and Bridges,
1975; Mebert et al., 1975; Atlas and Nebert, 1976; Burke and Prough, 1976;
Tredger and Chhabra, 1976.) Pretreatment with phenobarbitone stimulates
biphenyl 4-hydroxylase with no effect on biphenyl 2-hydroxylase (Burke and
Bridges, 1975; loannides and Parke, 1975). The stimulation of biphenyl
2-hydroxylase by chemical carcinogens occurs in two phases: an initial
activation of the enzyme followed by enzyme induction (McPherson et al.,
1976; Parke, 1976). It has been suggested that biphenyl 2-hydroxylase is
associated with cytochrome P 448, whereas biphenyl 4-hydroxylase is assoc-
iated primarily with cytochrome (Burke and Bridges, 1975; Burke and Prough,
1976; Parke, 1976).
Preincubation of animal and plant microsomes with chemical carcinogens
has also been reported to selectively increase 2-hydroxybiphenyl forma-
tion. This in vitro stimulation results in a 60 to 300% increase in the
amount of 2-hydroxybiphenyl (Burke and Bridges, 1975; McPherson et al.,
1974a, 1974b, 1975a, 1975b, 1975c, 1976). Non-carcinogens have no effect
(McPherson _ejt _aj_., 1974a), and 4-hydroxybiphenyl production is not affected.
Storage destroys the ability of the microsomes to respond to the.carcino-
gens (McPherson jit j»]_., 1975a). Initial experiments employed ( C)
biphenyl and used radioisotopic methods to measure metabolite production
(Burke and Bridges, 1975; McPherson _e_t _a_K, 1975a). In subsequent work,
the spectrophotofluorometric assay (Creaven et a]_., 1965) was used exclus-
ively.
The main objective of our work was to examine the feasibility of using
the in vitro stimulation of biphenyl 2-hydroxylase by chemical carcinogens
as a rapid prescreen for mutagenic and carcinogenic compounds to complement
the presently available biological testing methods. Initial experiments
were designed to duplicate exactly the work of others. We therefore
examined the effect of known chemical carcinogens and non-carcinogens on jji
vitro biphenyl 2-hydroxylase, using microsomes from the mouse, rat,
hamster, and avocado, Persea americana. Microsomes from the cauliflower
and apple were also tested. In addition, we began a preliminary study of
the in vitro metabolism of m-terphenyl by purified hamster microsomes.
82
-------
MATERIALS AMD METHODS
Chemicals
Renzo(a)pyrene (BaP) and safrole (SA) were purchased from Aldrich
Chemical Co., a-naphthylamine (aNA) and p-naphthylamine (pNA) were from
Sigma, and 20-methylcholanthrene (MC) was from K K Chemicals. All were of
the highest purity available. These compounds were dissolved in peanut oil
(Planters) to provide stock solutions at ImM. Biphenyl (99.9 %pure, Ultrex)
was obtained from J.T. Baker. It was dissolved in 1.5% (w/v) Tween 80 and
1.15% (w/v) KC1 to make a 13 mM stock solution. The hydroxylated standards,
2-hydroxybiphenyl (99+%) and 4-hydroxybiphenyl (97%) were obtained from
Aldrich Chemical Co. Stock solutions were made in 5% (v/v) aqueous ethanol
at the following concentrations: 4-hydroxybiphenyl, 0.146 y moles/nu; 2-
hydroxybiphenyl, 0.0342 y moles/nu; 2-hydroxybiphenyl and 4-hydroxybiphenyl
mixture at 0.0342 and 0.146 y moles/nu, respectively. A thin-layer chroma-
togram of the substrate and standard compounds at 0.1 mg material per spot
showed no impurities in the biphenyl or 2-hydroxybiphenyl, but trace
amounts of biphenyl and 2-hydroxybiphenyl were found in the 4-hydroxybi-
phenyl standard. Meta-terphenyl was recrystallized and a 13 mH stock solut-
ion in Tween 80 and 1.15% KC1 made as for biphenyl. The n-heptane was
"distilled in glass" from Burdick and Jackson Laboratories, and the
succinic acid (99.4%) was obtained from J. T. Baker. All other chemicals
were reagent grade.
Animals
Swiss-Webster mice (mean weight 37.9 g), Sprague-Dawley rats (mean
weight 172.5 g), and Syrian hamsters (mean weight 118.7 g) were obtained
from commerical breeders. Water and food (Wayne Lab Blocks) were provided
ad libitum. Animals were sacrificed between 0830 and 1030 hr by decap-
itation.
Preparation of Hepatic Microsomes
Microsomes were prepared by the method of McPherson jrt jil_. (1976).
Livers were rapidly removed into cold buffered KC1 (1.15% w/v KC1, 0.3 ^
NaH2P04, pH 7.6), blotted, weighed, and placed in fresh cold buffered
KC1. The pooled, weighed livers were homogenized with a motor-driven
teflon pestle using 10 strokes of 10 seconds each at 1200 rpm. The homog-
enate was diluted with cold buffered KC1 to 250 mg tissue per m£ of
homogenate and centrifuged (2° C) for 10 min at 15,000 g in an IEC B20
centrifuge. In one case, this low speed pellet was resuspended in buffered
KC1 at 25 mg protein/nu and used in a hydroxylation experiment. The low
speed supernatant was decanted and centrifuged (2° C) for 60 min at 104,000
g in a Beckman L2-65B ultracentrifuge. The supernatant was discarded and
the pellet washed with cold buffered KC1, resuspended in cold buffered KC1,
and again centrifuged (2° C) for 60 min at 104,000 g. The final pellets
were resuspended in cold buffered KC1 at a protein concentration of 10
mg/mJl. Protein was determined by the method of Lowry et al. (1951).
83
-------
Preparation of Plant Microsomes
Plant microsomes were prepared according to the method of McPherson
jit a\_. (1975b). Plant material was obtained 24 hr prior to use and stored
in the cold. Cauliflower heads were soaked in cold water for 1 hr before
storage to rehydrate the tissue. The mesocarp portion of both avocado and
apple was used; rosettes of cauliflower were shaved from the head. The
tissue was weighed, placed in cold phosphate buffer (0.1 M^ NaH2PC>4 pH 7.4),
and homogenized in either a Virtis homogenizer or a Waring blender.
Tissues were homogenized at 0.5 to 2 g tissue per mi of phosphate buffer.
The homogenate was filtered through muslin and centrifuged (2° C) for
20 min at 13,500 g in an IEC B20 centrifuge. The supernatant was decanted
and centrifuged {2° C) for 90 min at 80,000 g in a Beckman L2-65B
ultracentri- fuge. The pellets were resuspended in cold phosphate buffer
and adjusted to 1 to 10 mg/nu protein with cold buffered KC1. Protein was
determined by the method of Lowry et jjl_. (1951).
Hydroxylation Reactions
Hydroxylation reactions were performed according to the method of
McPherson j2t a\_. (1975b, 1975c, 1976). All reactions were carried out at
37° C in a shaking water bath at 100 cpm. The microsomal mixtures were
warmed for 60 sec after addition of the NADPH-regenerating system. The 10-
fold concentrated NADPH-regenerating system consisted of: glucose-6-phos-
phate dehydrogenase, 20 lU/rrtt; glucose-6-phosphate, 25 mM; NADP, 5 mM;
MgS04, 0.5 ntl dissolved in buffered KC1.
In the case of the crude homogenates, low speed supernatants and
pellets, and plant microsomes, 1.8 mi of the preparation was used directly
in the reaction mixture. The final protein concentration in the 2 mi
reaction mixture for these preparations is given in Table 1. An aliquot
(0.4 mi] of the first high speed pellets and purified animal microsomes
(second high speed pellets) at 10 mg protein/ma was added to 1.4 mi of cold
buffered KC1 to provide a final protein concentration of 2 mg/mi in the 2
mi reaction mixture. Each tube received 0.2 mi of the 10-fold concentrated
NADPH-regenerating system. Test compound (0.5 mi) in oil, or oil alone was
added, and incubated for 10 min. Biphenyl (0.3 mi of 13 mM stock solution)
was added, and incubation continued for an additional 5 min. The reaction
was terminated by the addition of 1 nut of 4 M_ HC1 to each tube. The incu-
bation mixtures used in each hydroxylation experiment are given in Table 2.
In one experiment using purified hamster microsomes, 0.3 nu pf 13 mM
m-terphenyl was added as substrate in place of biphenyl.
Metabolite Extraction
The entire incubation and heptane extraction procedure was carried out
in 20 mi glass tubes with teflon lined screw caps. Following addition of
HC1 and standards where indicated, the tubes were immediately extracted
with 10 mi of ji-heptane by mechanically shaking for 5 min (Creaven, et al.,
1965). The tubes were centrifuged at 2000 rpm for 15 min and stored
overnight in the cold. Tubes were allowed to return to ambient temperature,
84
-------
TABLE 1. PROTEIN CONCENTRATIONS IN HYDROXYLATION EXPERIMENTS
Preparation
Protein Concentration
(mg/nu)
Mouse homogenate
Rat homogenate
Hamster homogenate
Mouse low-speed pellet
Hamster low-speed supernatant
Avocado microsomes
Apple mlcrosomes
Cauliflower microsomes
Cauliflower low-speed
supernatant
29.7
27.0
29.7, 35.9
22.5
13.3
5.8
1.0
5.0, 9.1
2.2
Two numbers indicate the concentrations in two separate experiments.
Protein is given as the final concentration in 2 ma of reaction mixture.
TABLE 2. INCUBATION MIXTURES USED IN HYDROXYLATION REACTIONS
Microsomal
system
+ Oil
+ Oil
+ Test compound
in oil
+ Test compound
in oil
+ Oil
+ Biphenyl
•+• Biphenyl
+ Biphenyl
+ HC1
+ Test compound
in oil
•+• 2- or 4-Hydroxy-
biphenyl +
biphenyl
Materials are listed in order of addition from left to right. The
microsomal system and HC1 were added to all tubes. A blank space
^ indicates no addition, but continued incubation.
The microsomal syustem consists of the nicrosomes or homogenate
***fractions plus the NADPH-regenerating system.
The HC1 was added at the end of the incubation period to terminate
the reaction.
85
-------
and ?. ru of the heptane layer in each tube was shaken for 5 min with 10 nu
of 0.1 _N_ NaOH. The mixture was centrifuged for 10 min at 2000 rpm, 2 ma of
the "laOH layer transferred to a cuvette, and adjusted to pH 5.5 by
addition of 0.5 nu of 0.5 N_ succinic acid (Creaven elt ^1_., 1965).
Analytical Methods
Fluorescence of the 2- and 4-hydroxbiphenyl compounds was deter-
mined by the method of Creaven j?t aj_., (1965) using an Aminco-Bowman
spectro photofluoroineter with a high-pressure xenon lamp. The fluorescence-
intensity of each sample was measured first for 4-hydroxybiphenyl by excit-
ing at 270 to 272 nn and measuring emission at 335 nm, and then for 2-
hydroxybiphenyl by exciting at 290 nm and measuring emission at 412 nm.
Standard curves were constructed for each biological preparation for every
experiment with dilutions of the tubes containing known concentrations of
2- and 4-hydroxybiphenyl. The fluorescence at wavelengths used to measure
2-hydroxybiphenyl was corrected for the contribution of 4-hydroxybiphenyl
according to the method of Creaven et a]_., (1965).
Metabolite separation was accomplished by a Perkin-Elmer
sure liquid chromatograph (HPLC) with a Partisil mPAC (Re
220 high
pressure liquid chromatograph (HPLU) with a Partisil "'PAU (Keeve
Angel) column of 25 cm x 46 mm inside diameter, The solvent was 85%
ji-hexane, 15% tetrahydrofuran (no preservatives) at a flow rate of 2 mi per
min. Column pressure was 300 psi, and the temperature was ambient. The
chromatograph was attached to the spectrophotofluorometer by means of 150
\ii flowthrough cell having a 2 mm path length. Fluorescence was measured
for biphenyl metabolites by exciting at 300 nm and recording emission a 335
nm. Terphenyl metabolites were examined at several different
excitation-emission wavelengths. The location of compounds in the
chromatographic fractions was recorded by a Linear Instrument Co. strip
chart recorder at a chart speed of 16 inches per hr. All chromatograms
were recorded at range 0.33.
The ji-heptane layer (10 y£) from each sample was injected directly
into the HPLC. Quantitation of peak heights was accomplished by con-
structing standard curves with three different concentrations of 2-hydroxy-
biphenyl and 4-hydroxybiphenyl ranging from 1 to 5 ng per injection.
Calibration curves were based on peak height because repeated injections
showed sufficient reproducibility (in retention time and peak width) to
alleviate the need for area measurements. The limit of detection was
determined to be approximately 1 ng for 4-hydroxybiphenyl and 0.5 ng for
2-hydroxybiphenyl .
RESULTS
Measurement of Biphenyl Metabolites by Fluorometric Analysis
The results (Tables 3 and 4) were obtained from several experiments
designed to either duplicate previously reported work (Burke and Bridges,
1975; McPherson et al . , 1974a, 1974b, 1975a, 1975b, 1975c, 1976) or to
86
-------
examine microsomes from other sources with respect to activation of
biphenyl 2-hydroxylase. The quantity of 2-hydroxybiphenyl and 4-hydroxy-
biphenyl in each of the reaction mixtures was determined fluorometrically
by the method of Creaven &t_ aj_. (1965). The known carcinogen, BaP, was
included in all experiments as a positive control, and where possible, aNA
was included as an example of a non-carcinogen. All other test compounds
are known carcinogens (McCann jrt jj_., 1975) and were used when the supply
of microsomes was sufficient.
It was apparent that the oil used as the solvent for test compounds
contributed fluorescence at the wavelengths used to measure both
hydroxylated biphenyls. However, the oil was present in all reaction
tubes, including those containing the 2- and 4-hydroxybiphenyl standards.
Fluorescence contribution of the oil was incorporated into the standard
curves used to obtain quantitative data; therefore subtraction from values
obtained for the experimental incubation mixtures was not required.
Table 3 shows the results of three different hydroxylation experiments
using various fractions of hepatic homogenates from mice, rats, and
hamsters. Comparison of the amount of 2-hydroxybiphenyl produced in the
oil plus biphenyl reaction mixture with that produced in the carcinogen
plus biphenyl reaction mixture indicated that there was an apparent
stimulation of 2-hydroxybiphenyl production of 123 to 3500% in the presence
of carcinogen. However, extracts of incubation mixtures containing test
compound alone showed a significant fluorescence at the wavelength used to
determine 2-hydroxybiphenyl, and a smaller fluorescence at the wavelength
used to determine 4-hydroxybiphenyl. In nearly all cases, the apparent
increase in the amount of 2-hydroxybiphenyl produced in the presence of
carcinogen could be eliminated by subtracting the fluorescence contributed
by the metabolites of the test compound.
The effect of various test compounds on hydroxybiphenyl production by
plant microsomes was also examined because McPherson jet_ _al_. (1975b, 1975c)
u=»d reported that 3,4-benzopyrene stimulates 2-hydroxybiphenyl formation by
avocado microsomes. The results (Table 4) using plant microsomes are
similar to those obtained with animal microsomes. In this case, metabolites
of the test compound contributed significantly to the determination of both
hydroxybiphenyl compounds.
The fluorescence excitation and emission spectra of extracts from some
of the incubation mixtures were examined by fluorometric assay (Creaven
et _§]_., 1965). The spectra for the two hydroxylated standards are shown in
Figure 1. The 2-hydroxybiphenyl standard showed an emission peak at 412 nm,
and 4-hydroxybiphenyl had a major emission peak at 335 nm with a minor peak
in the vicinity of 412 nm in agreement with the data of Creaven et aj^.,
(1965).
Material extracted from incubation mixtures containing purified
hamster microsomes and oil plus biphenyl showed (Figure 2A) a shift in the
emission peak for 2-hydroxybiphenyl to 418 nm and a broad shoulder in the
region used to measure 4-hydroxybiphenyl (335 nm). Incubation mixtures to
87
-------
TABLE 3. EFFECT OF TEST COMPOUNDS ON PRODUCTION OF 4-HYDROXBIPHENYL AND
2-HYDROXYBIPHENYL BY LIVER FRACTIONS
Fraction Animal
.Crude Mouse
homogenate
Rat
Hamster'*"
Low -speed Hamster
supernatant
Low -speed Mouse
pellet
First high- Hamster
speed pellet
Second high" Mouse""
speed pellet
Reaction
mixture
Oil
BaP
Oil + biphenyl
Biphenyl + BaP**
BaP + biphenyl
Oil
BaP
Oil + biphenyl
Biphenyl + BaP
BaP + biphenyl
Oil
BaP
SA
Oil + biphenyl
Biphenyl + BaP
Bipheny] + SA
BaP + biphenyl
SA + biphenyl
Oil
BaP
Oil + biphenyl
Biphenyl + BaP
BaP + biphenyl
Oil
BaP
Oil + biphenyl
Biphenyl + BaP
BaP + biphenyl
Oil
BaP
SA
MC
Oil + biphenyl
Biphenyl + BaP
Biphenyl + SA
Biphenyl + MC
BaP + biphenyl
SA + biphenyl
MC + biphenyl
Oil
BaP
SA
MC
aNA
BNA
Oil + biphenyl
Biphenyl + BaP
Biphenyl + SA
Biphenyl + MC
Biphenyl + aNA
Biphenyl + bNA
BaP + biphenyl
SA + biphenyl
MC + biphenyl
aNA + biphenyl
BNA + biphenyl
n mole
4— Hydroxybiphenyl
0.007
0.014
0.026
0.033
0.005 ( 19%:***
0.016
0
0.018
0.024
0.011 ( 61%)
0.004
0
0.006
0.022
0.018
0.034
0.007 ( 32%)
0.021 ( 96%)
0.016
0.011
0
0.142
0.055 ( — )
0.005
0.009
0.032
0.045
0.032 (100%)
_Jf
0
0
0
0.47
—
1.58
1.32
0.51 (1082)
0.66 (140%)
0.56 (1192)
0.21
0.09
0.82
0.29
0.26
0,07
2.54
2.23
3.85
2.68
3.12
3.38
2.86 (1132)
1.15 ( 45Z)
4.00 (1582)
2.97 (1172)
1.10 ( 432)
, -1 J -1
min mg protein
2-Hydroxybiphenyl Corrected*
1.05
1.41
2.55
1.03
1.58 ( 62%) 0.17
0.19
0.54
0.34
0.86
0.92 (271%) 0.38
0.19
0.37
0
0.29
0.22
0.005
0.78 (269%) 0.41
0 ( — ) £
0.008
0.020
0.020
0.122
0.075 (375%) 0.055
0.003
0.008
0.005
0.006
0.008 ( 160%) 0
—
0.56
0.07
0.37
0.26
—
0.85
1.43
0.54 ( 208%) 0
0.32 ( 123%) 0.25
0.65 ( 250%) 0.28
0.06
0.10
0
0.12
0.07
0.29
0.008
0.13
0
0.26
0.03
0.04
0.034( 4252) 0
0.07 ( 875%) 0.07
0.03 ( 3752) 0
0 ( - ) 0
0.28 (3500%) 0
88
-------
TABLE 3. (CONTINUED)
Second high- Rat Oil
speed pellet BaP
Oil + biphenyl
Biphenyl + BaP
BaP + biphenyl
Hamster5 Oil
BaP
SA
MC
aNA
0NA
Oil + biphenyl
Biphenyl + BaP
Biphenyl + SA
Biphenyl t MC
Biphenyl + aNA
Biphenyl t 3NA
BaP + biphenyl
SA + biphenyl*
MC + biphenyl''
aNA + biphenyl*
BNA + biphenyl*
0.54
0.13
1.25
1.10
1.01 ( 81%)
0.12
0.15
0.62
0.25
0.21
0.71
0.93
1.46
2.08
1.29
1.57
1.72
0.78 ( 84%)
1.07 (115%)
1.03 (111%)
1.04 (112%)
0.78 ( 84%)
0
0.10
0.015
0.044
0.074 (493%)
0.008
0.20
0.06
0.42
0.13
0.58
0.17
0.40
0.23
0.44
0.68
0.65
0.32 (188%)
0.18 (106%)
0.30 (176%)
0.28 (165%)
0.56 (329%)
0
0.12
0.12
0
0.15
0
**
***
The contribution of test compound metabolites as determined in
reaction mixtures containing test compound alone was subtracted from
the quantity of 2-hydroxybiphenyl apparently present in the test
compound plus biphenyl reaction mixtures.
When the test compound follows the substrate, it was added to the
reaction mixture after the addition of HC1.
'Numbers in parentheses are results expressed as percent relative
to control values.
"^Results are the average of two experiments for the oil, oil and
biphenyl, and BaP mixtures.
fA dashed line means that the results were not determined because the
tube was lost.
§ Results are tha average of two experiments, 2 to 3 replicates for the
oil, oil and biphenyl, and BaP mixtures.
#Results are the average of two replicates in one experiment.
-------
TABLE 4. EFFECT OF TEST CQMPOUNDS Ofl PRODUCTION OF 4-HYDROXYRIPHuMYL
PLANT HICROSOMES
Aril) 2-HYQROXYUIPHENYL BY
Fraction
Low speed
supernatant
Purified
mlcrosomes
n mole min rag protein"^
Reaction
Plant mixture 4-Hydroxybiphenyl Corrected 2-Hydroxybiphenyl
Cauliflower Oil
BaP
SA
Oil + btphenyl
Biphenyl + BaP
BaP + biphenyl
SA + biphenyl
Cauliflowet** Oil
BaP
SA
MC
Oil + biphenyl
Biphenyl + BaP
Biphenyl + SA
Biphenyl + MC
BaP + biphenyl
SA + biphenyl
MC + biphenyl
Avocado Oil
BaP
Oil + biphenyl
Biphenyl + BaP
BaP + biphenyl
Apple Oil
BaP
Oil + biphenyl
Biphenyl + BaP
BaP + biphenyl
0.53
0.80
0.80
0.27
0.49
0.43 0
0.61 0
0.18
0.21
0.25
0.036
0.15
0.065
0.081
0.048
0.12 0
0.21 0
0.048 0.012
0.102
0.082
0.077
0.113
0,087 0.005
0.34
0.31
0.54
0.57
0.43 0.12
0.27
0.39
0.18
0.18
0.66
0.31
0.12
0.085
0.075
0.032
0.039
0.003
0.026
0
0.27
0.032
0,012
0.036
0.010
0.031
0.005
0.031
0.005
0
0.31
0
0.20
0
Corrected
0
0
0
0
0
0
0
**
Symbols and procedures are described in Table 3.
All except MC are the average of results from two experiments.
-------
SO
Em 335
i
1
,
" /
X
\
i ^~X
\
v
u\
\~s ^^fj
N.
N.
^
i
Ei 274
v
^
250
350 450
WAVELENGTH (NANOMETERS)
550
Figure 1.
Excitation
standards.
measures at
standards,
excitation
ively. The
is shown in
figures.
and emission spectra of 2- and 4-hydroxybiphenyl
Dashed lines are excitation spectra with emission
412 nm and 335 nm for the 2- and 4-hydroxybiphenyl
respectively. Solid lines are emission spectra with
at 290 nm and for 2- and 4-hydroxybiphenyl, respect-
2-hydroxybiphenyl spectrum is vertically offset and
the upper portion as in this and all subsequent
91
-------
r\
t-.
'_._/ 1
/ 1
' 1
/ 0'
I"
1
*.'*
Figure ?. Excitation and emission spectra of material extracted from
incubation mixtures containing purified hamster microsomes and
(A) oil plus biphenyl; (B) BaP in oil; (C) Bap in oil plus
biphenyl. See Figure 1 for a description of symbols. Dashed
vertical lines designate 335 and 412 nm, the v.'avelengths at
v/hich emission is measured in the fluorornetric assay.
A'
no, ,
/ ;
\
V
\J
.^i
-I
ro 3. Excitation and emission spectra of material extracted .from
incubation mixtures containing purified cauliflower microsomes
and (A) oil plus biphenyl; (C) BaP in oil; (C) BaP in oil plus
biphenyl. For a description of the symbols, see Figures 1 and
2.
-------
which BaP in oil had been added (Figure 2R) contained material which had a
broad emission peak from approximately 380 to 470 nm after excitation at
290 nm. The emission at 335 nm (excitation at 272 nm) was part of the
excitation scatter peak. In mixtures containing BaP in oil plus biphenyl
(Figure 2C), material produced also had the 418 nm peak, but lacked the 335
nm shoulder.
Plant microsomal (Figures 3, 4) material extracted from the incubation
mixtures containing oil plus biphenyl showed two emission optima (Figure
3A). Excitation at 290 nm gave an emission peak at 405 nm, and at 272 nm,
an emission peak at 367nm. A similar pattern was oberved in material from
microsomes incubated with BaP in oil (Figure 3B) or MC in oil (Figure 4A),
Incubation mixtures containing both carcinogen and biphenyl yielded
material which had an increased amount of fluorescence in the 405 to 410 nm
region. The emission spectrum obtained by exciting at 272 nm showed
essentially no discrete peaks in the case of BaP, and a broad 350 to 450 nm
shoulder in the case of MC.
20
0
j
Em335 i
*
_^--
1
i
\
\
1 "f 4IZ!
/
*''
1
1
£.272
250 3SO 450 55
TEmJiS '
20
/
/
/
/
X
f "
0 250
! E«272
i -1
333 412'
n "]
350 *5O 5-
•AVELENGTX (NANOMETERS)
WAVELENGTH (NANOMETERS I
Figure 4. Excitation and emission spectra of material extracted from
incubation mixtures containing purified cauliflower microsomes
and (A) MC in oil; (B) MC in oil plus biphenyl.
-------
Analysis of Rlphenyl Metabolites using HPLC
An alternative method which involved separation of metabolites by high
pressure liquid chromatography and detection by fluorometry was used to de-
termine the amount of 2- and 4-hydroxybiphenyl produced in incubation
mixtures containing hamster microsones. Figure 5 shows that this method
allows complete separation of the two hydroxylated biphenyls in a mixture
of pure compounds. The average retention time was 9.0 min for 2-hydroxybi-
phonyl and 11.7 min for 4-hydroxybiphenyl.
Figure 6 shows chromatograms of material obtained from incubation mix-
tures which contained purified hamster microsomes. Microsomes incubated
with oil plus biphenyl produced easily detectable amounts of 2- and 4-hy-
droxybiphenyl (Figure 6A). The material eluting immediately after inject-
ion was derived from the oil (Figure 6B) and was present in all samples.
Figure 6C shows a typical chromatogram of a sample from an incubation
mixture containing RaP in oil plus biphenyl.
Table 5 presents quantitative data obtained from these chromatograms.
It can be seen that the test compounds did not contribute fluorescent
material to the 2- and 4-hydroxybiphenyl peaks under the conditions of this
assay. The amount of 2- and 4-hydroxybiphenyl produced in the reaction
containing substrate alone agrees with the quantity determined using the
fluorometric method (Table 3). However, in the presence of test compound,
the amount of 4-hydroxybiphenyl apparently decreased in all but one case,
and the amount of 2-hydroxybiphenyl remained constant or decreased
slightly.
Terphenyl Metabolism
Purified hamster microsomes were used in an experiment designed to
examine the metabolites of terphenyl produced in vitro. The effect of BaP
on ni-terphenyl metabolite production was also investigated. Incubation
conditions were identical to those used for biphenyl. The n-heptane
extract from these mixtures was chromatographed by HPLC as described for
biphenyl. A number of different excitation and emission wavelengths were
used (Table 6) to detect possible metabolites of terphenyl. In each case
(Figure 7), unaltered terphenyl was observed at an average retention time
of 3.1 min. A total of three different metabolites were detected at two
different excitation and emission wavelengths (Table 6). A chromatogram
showing the first two metabolites is presented in Figure 7A. A reaction
mixture containing terphenyl as a substrate which had been preincubated for
10 min with BaP was also examined at an excitation wavelength of 270 nm and
and an emission wavelength of 350 nm. The metabolite with a retention time
of 14.2 to 14.4 min was absent, and the metabolite with a retention time of
16.2 to 16.3 min had a reduced peak height (Figure 7B).
94
-------
TABLE 5. EFFECT OF TEST COMPOUNDS ON PRODUCTION OF 4- and
2-HYDROXYBIPHENYL AS DETERMINED BY QUANTITATIVE HPLC
Reaction aixtura
n racle r.in " r.s protein "
i-KvdroxvbiDhenvi 2-Hvdrcxvbinner
First high speed pellet
Oil -r- biphenyl
Ba? -t- biphenyl
SA -r biphenyl.
MC - biphenyl
Ba?
Second high speed pellet
Oil -r biphenyl
3aP -r biphenyl
SA 4- biphenyl
MC -r biphenyl
aNA •+• biphenyl
3NA 4- biphenvl
Oil
3ap
0.64
0.36
0.26
0.55
0.00
0.59
0.26
0.19
0.24
0.30
0.14
0.00
0.04
0.16
0.14
0.20
0.21
0.00
0.22
0.15
0.17
0.16
0.18
0.09
0.00
0.02
* Hamster microsomes were used.
TABLE 6. METABOLITES OF TERPHENYL PRODUCED BY PURIFIED HAMSTER MICROSOMES
Peak heights (nun)* at Rt of:
Wavelength
(nm)
Ex
270,
Ex250,
Ex
Ex
Ex
270,
300,
300,
Em
Em
Em
Em
Em
14.2-14.4 min
360 5.8
360 2.2
350 5.4
360
335
16.2-
14
5
15
7
3
16.3 min 17.8 min
.8 — **
.5
.0
.0 9.0
.0 4.0
Peak widths as half-height were identical for each metabolite.
Symbols used are: R^ = retention time, Ex = excitation, Em = emission.
**
Not Present.
-------
Figure 5. Separation of 2- and 4-hydroxybiphenyl standard by HPLC. The
sample contained 2.4 ng of each pure compound. For details of
the procedure, see Materials and Methods.
Fi gure 6.
Material separated by HPLC and obtained from reaction mixtures
containing purified hamster microsomes and (A) oil plus
biphenyl; (B) oil alone; (C) BaP in oil plus biphenyl. For
experimental details, see Materials and Methods.
96
-------
«3OO
ON'
X300
a
I KIN I
Figure 7. Material separated by HPLC obtained from reaction mixtures
containing purified hamster microsomes and (A) oil plus
m-terphenyl; (B) BaP in oil plus m-terphenyl. The excitation
wavelength was 270 nm and the emission wavelength was 350 ntn.
97
-------
Discussion
As reported hy others (Burke et _al_., 1977; Tong et a]_., 1977a), the
data presented in our paper demonstrate that the fluorometric determination
of 2-hydroxybiphenyl production in vitro in the presence of carcinogens may
result in spuriously high values. The data (Tables 3, 4; Figures 2,3, and
1) indicate that much, if not all, of the increase in fluorescence at 412
rm (the wavelength used to measure 2-hydroxybiphenyl) may be attributed to
fluorescence of test compound metabolites. In addition, the emission peak
for 2-hydroxybiphenyl changed from the expected 412 nm to 418 nm in the
case of animal nicrosomes, and 405 nm in the case of plant microsomes.
Fluorescence of the 4-hydroxybiphenyl compound at 335 nm became part of the
excitation scattering peak in all microsomal extracts. Measurements made
at 412 and 335 nm, therefore, are subject to considerable error.
Using a high pressure liquid chromatography system which permits
unequivocal identification and quantification of the hydroxylated biphenyls,
we were not able to demonstrate an in vitro stimulation of biphenyl 2-hy-
droxylase in contrast to the findings of others (Creaven jrt aj_., 1965;
McPherson et al_., 1974a, 1974b, 1975a, 1975b, 1975c, 1976; Tredger and
Chhabra, 1976; Tong and Parke, 1977b). The data in Table 5 show that the
amount of 4-hydroxybiphenyl decreased in the presence of test compounds.
The amount of 2-hydroxybiphenyl remained constant or decreased ($NA). The
discrepancy between these results and those of Burke and Bridges (1975) and
McPherson elt a\_. (1975a) who reported stimulation of biphenyl 2-hydroxylase
by chemical carcinogens in vitro, using ( C)biphenyl as a substrate is
presently unexplained.
Terphenyl was metabolized in vitro to at least three products by
purified hamster microsomes. The formation of one of these products is
prevented by preincubation of the microsomes with BaP. Further work is
necessary to determine whether or not this substrate may prove useful in
the development of an in vitro microsomal screen for chemical carcinogens.
ACKNOWLEDGEMENTS
We are greatly indebted to F. Krohlow for invaluable technical
assistance. We also thank Dr. N. Richards, Project Monitor, EPA Research
Laboratory Sabine Island, Gulf Breeze, Florida, for his comments and
suggestions. The work was supported by the Environmental Protection
Agency, R8-05671010: Development of Enzymatic Systems for Screening of
Mutagens and Carcinogens in Environmental Pollutants (J.J. Schmidt-Collerus,
N.L. Couse, «]. King, and L. Leffler).
98
-------
REFERENCES
Atlas, S.A., and D.W. Nebert. 1976. Genetic association of increases in
naphthalene, acetanilide, and biphenyl hydroxylations with inducible
aryl hydrocarbon hydroxylase in mice. Arch. Biochem. Biophys.
175:495-506.
Basu, T.K., J.W.T. Dickerson, and D.V.W. Parke. 1971. Effect of
development on the activity of microsomal drug-metabolizing enzymes in
rat liver. Biochem. J. 124:19-24.
Burke, M.D., and J.W. Bridges. 1975. Biphenyl hydroxylations and
spectrally apparent interactions with liver microsomes from hamsters
pre-treated with phenobarbitone and 3-methylcholanthrene. Xenobiotica
5:357-376.
Burke, M.D., and R.A. Prough. 1976. Some characteristics of hamster liver
and lung microsomal aryl hydrocarbon (biphenyl and benzo(a)pyrene)
hydroxylation reactions. Biochem. Pharmacol. 25:2187-2195.
Burke, M.D., D.J. Benford, J.W. Bridges, and D.V. Parke. 1977.
High-pressure liquid chromatographic and other assays for biphenyl
hydroxylation compared. Biochem. Soc. Trans. 5:1370-1372.
Catelani, D., G. Mosselmans, J. Nienhaus, S. Sorlini, and V. Treccani.
1970. Microbial degradation of aromatic hydrocarbons used as reactor
coolants. Experientia 26:922-923.
Catelani, D., C. Sorlini, and V. Treccani. 1971. The metabolism of
biphenyl by Pseudomonas putida. Experientia 27:1173-1174.
Catelani, D., A. Columbi, C. Sorlini, and V. Treccani. 1973. Metabolism
of biphenyl. 2-Hydroxy-6-oxo-6phenylhexa-2,4-dienoate: the meta-
cleavage product from 2,3-dihydroxybiphenyl by Pseudomonas putida
Biochem. J. 134:1063-1066.
Cerniglia, C.E., R.L. Hebert, R.H. Dodge, P.J. Szaniszlo, and D.T. Gibson.
1979. Some approaches to studies on the degradation of aromatic
hydrocarbons by fungi. In: Microbial degradation of pollutants in
marine environments. A.W. Bourquin and H.P. Pritchard, Eds., U.S.
Environmental Protection Agency, EPA-600/9-79-012, Gulf Breeze, FL.
pp. 360-369.
Creaven, P.J., D.V. Parke, and R.T. Williams. 1965. A fluorimetric study
of the hydroxylation of biphenyl in vitro by liver preparations of
various species. Biochem. J. 96:879-885.
Dodge, R.H., C.E. Cerniglia, and D.T. Gibson. 1978. Fungal metabolism of
biphenyl. Abstr., Annual Meeting Am. Soc. Microbiol. 179 p.
99
-------
Friedman, M.A., E.J. Greene, R. Csillag, and S.S. Epstein. 1972.
Paradoxical effects of piperonyl butoxide on the kinetics of mouse
liver microsomal enzyme activity. Tox. Appl. Pharmacol. 21:419-427.
loannides, C., and D.V. Parke. 1975. Mechanism of induction of hepatic
microsomal drug metabolizing enzymes by a series of barbiturates. J.
Pharmacol. 27:739-746.
Lowry, O.H., N.J. Rosenbrough, A.L. Farr, and R.J. Randall. 1951. Protein
measurement with the folin phenol reagent. J. Bio. Chem.
193:265-275.
Lunt, D., and W.C. Evans. 1970. The microbial metabolism of biphenyl.
Biochem. J. 118:54-55.
McCann, J., E. Choi, E. Yamasaki, and B.N. Ames. 1975. Detection of
carcinogens in the Salmonella/microsome test: assay of 300 chemicals.
Proc. Nat. Acad. Sci., U.S. 72:5135-5139.
McPherson, F., J.W. Bridges, and D.V. Parke. 1974a. In vitro enhancement
of hepatic microsomal biphenyl 2-hydroxylation by carcinogens. Nature
252:488-489.
McPherson, F.J., J.W. Bridges, and D.V. Parke. 1974b. The enhancement of
biphenyl 2-hydroxylation by carcinogens in vitro. Biochem. Soc.
Trans. 2:618-619.
McPherson, F.J., J.W. Bridges, and D.V. Parke. 1975a. Studies on the
nature of the in vitro enhancement of biphenyl 2-hydroxylation pro-
voked by some chemical carcinogens. Biochem. Pharmacol.
25:1345-1350.
McPherson, F.J., A. Markam, J.W. Bridges, G.C. Hartman, and D.V. Parke.
1975b. Effects of preincubation in vitro with 3,4-benzopyrene and
phenobarbital on the drug-metabolism systems present in the microsomal
and soluble fraction of the avocado pear (Persea americana). Biochem.
Soc. Trans. 3:283-285.
McPherson, F.J., A. Markham, J.W. Bridges, G.C. Hartman, and D.V. Parke.
1975c. A comparison of the properties in vitro of biphenyl 2- and
4-hydroxylase in the mesocarp from avocado pear (Persea americana) and
Syrian-hamster hepatic tissue. Biochem. Soc. Trans. 3:281-283.
McPherson, F.J., J.W. Bridges, and D.V. Parke. 1976. The effects of
benzopyrene and safrole on biphenyl 2-hydroxylase and other drug-
metabolizing enzymes. Biochem. J. 154:773-780.
Nebert, D.W., J.R. Robinson, A. Niwa, K. Kumake, A.P. Poland. 1975.
Genetic expression of aryl hydrocarbon hydroxylase activity in the
mouse. J. Cell. Physiol. 85:393-414.
100
-------
Parke, D.V. 1976. The activation and induction of biphenyl hydroxylation
and chemical carcinogenesis. In: Microsomes and drug oxidations, V.
Ullrich, Ed., Pergamon Press, New York. pp. 721-729.
long, S., C. loannides, and D.V. Parke. 1977a. Possible pitfalls of the
biphenyl test for chemical carcinogens. Biochem. Soc. Trans.
5:1372-1374.
Tong, S., C. loannides, and D.V. Parke. 1977b. Enhancement of biphenyl by
organochlorine insecticides. Biochem. Soc. Trans. 5:1374-1377.
Tredger, J.M., and R.S. Chhabra. 1976. Preservation of various microsomal
drug metabolizing components in tissue preparations from the livers,
lungs, and small intestines of rodents. Drug Metab. Dispos.
4:451-459.
Willis, D.E., and R.F. Addison. 1974. Hydroxylation of biphenyl in vitro
by tissue preparations of some marine organisms. Comp. Gen. Pharmac.
5:77-81.
101
-------
PETROLEUM AND PETROLEUM COMBUSTION BYPRODUCTS AS POTENTIAL
SOURCES OF MARINE ENVIRONMENTAL MUTAGENS
by
Jerry F. Payne and R. Maloney
Research and Resources Services
Department of Fisheries and Oceans
P.O. Box 5667
St. John's, Newfoundland
Canada, A1C 5X1
and
A. Rahimtula and I. Martins
Department of Biochemistry
Memorial University of Newfoundland
St. John's, Newfoundland
Canada, A1C 5S7
ABSTRACT
We previously reported that polycyclic organic enriched
fractions of used engine oil were mutagenic in the Ames
Salmonella assay. Recent studies on a variety of oil samples
obtained from automobile service stations in the St. John's area
has demonstrated that used engine oils consistently contain
mutagenic material, presumably polycyclic organic compounds. A
mutagenic response has also been obtained with extracts of a
sample of used, but not with extracts of a comparable brand of
unused, non-petroleum base, synthetic engine oil. Circumstan-
tial evidence is presented to suggest that the mutagens found in
used engine oils originate from fuel combustion. In contrast to
the positive results obtained with used engine oils, mutagenic
activity has not been obtained with any of several different
types of petroleum or unused engine oils. The possibility that
polycyclic organic compounds originating from combustion
processes could be a major anthropogenic source of marine
environmental mutagens is briefly discussed.
102
-------
INTRODUCTION
Environmental carcinogenesis is not only a human health concern, but
has become an area of prime interest in all environmental studies,
including aquatic toxicology. Polycyclic organic compounds (POC) are major
air and water pollutants and are receiving considerable attention because
a few have been shown to be both mutagenic and carcinogenic. An increase
in the number of oil tanker spills has focused interest on the possible
effects of elevated concentrations of POC on marine organisms. Major new
energy technologies including coal combustion, liquefaction, and
hydrogenation, as well as extensive tar sand extraction, will likely result
in an increased contamination of the aquatic environment by polycyclic
organic compounds.
Urban-associated wastewater and atmospheric fallout are probably the
chief routes by which complex pollutant mixtures originating from
industrial, domestic, and natural sources enter the aquatic environment.
Used engine oil is a common pollutant in municipal wastewater (Farrington
and Quinn, 1973; Tanacredi, 1977) and we have recently demonstrated that
used oil contains mutagenic compounds (Payne eit ^1_., 1978). This report
summarizes further studies on a variety of sources of potentially mutagenic
POC which may enter waterways via oil spills, land runoff, or municipal
wastewater.
Petroleum Hydrocarbon Mutagenicity
The method used to assess mutagenic activity in POC enriched fractions
of petroleum hydrocarbons was previously reported (Payne ^t aj^., 1978).
Oil was extracted with an equal volume of dimethyl sulfoxide (DMSO) and
titers of this extract were used directly, with and without liver enzyme
activation (9000 x g supernatant fraction of S9), in the Ames Salmonella
assay (Ames jrt ^1_., 1975). DMSO is a good solvent for extracting
polycyclic organic compounds from aliphatic material (Natusch, 1978), and a
number of compounds found in POC enriched fractions of a variety of
petroleum oils are metabolized by fish liver S9 preparations (Table 1).
Crude petroleum likely contains hundreds or thousands of different
polycyclic compounds and we have not yet begun to fractionate any oils into
chemical "classes" for mutagenesis testing. There are limitations to
testing complex mixtures of similar chemicals, such as the possibility for
producing synergistic or antagonistic effects, but such effects may be of
key biological importance. For instance, substrate competition has been
suggested for the apparent inability of the potent carcinogen,
benzo(a)pyrene, to form tumors when applied in combination with other
hydrocarbon compounds (Schmeltz jrt jil_., 1978). Conversely, supraadditive
effects can also be postulated when two or more chemicals are being tested
in combination (Fingl and Woodburg, 1975).
103
-------
TABLE 1. METABOLISM OF PETROLEUM BY FISH LIVER HOMOGENATES
Source
Kuwait crude
Louisiana crude
No. 2 fuel oil
Sable Island crude
Band
1
3
2
3
2
1
1
3
2
3
1
2
Number of
Metabolites
Detected
3
1
3
3
3
2
3
2
2
2
2
4
RF. value
.30
.40
.30
.25
.16
.30
.17
.40
.24
.23
.30
.20
.40
.50
.50
.40
.50
.40
.76
.50
.50
.50
.30
.50
.60
.55
.50
.50
.44 .50
Polycyclic aromatic compounds were extracted from petroleum with DMSO.
These extracts were streaked on silica-gel plates which were developed in
hexane:benzene (18:2). Fluorescent bands were eluted with methylene
chloride which was evaporated under nitrogen, and the residues were
dissolved in 0.1-0.5 mi of methanol. These residues were substituted for
benzo(a)pyrene in our regular aryl hydrocarbon hydroxylase assay (Payne
and Penrose, 1975). Hexane extracts were evaporated under nitrogen to
approximately 100 vi and this volume was streaked on silica-gel plates
which were developed in benzenermethanol (18:1) in the dark. Fluorescent
metabolites were detected under UV.
104
-------
Crude petroleum obtained from a variety of world sources, as well as
some refined oils, including several brands of unused engine oil have now
been assessed for mutagenic activity. Assays producing at least twice the
background numbers of revertant colonies were considered to be positive.
It is of special interest that mutagenicity has not been reliably
demonstrated in POC extracts of any crude or refined oils assessed to date
(Table 2). Table 3 presents some representative values obtained with four
different types of petroleum including Pembina, Atkinson's Point, Norman
Wells, and Alaskan crudes. In contrast to the negative results obtained
with petroleum and refined oils, used engine oil collected from a number of
automobile service stations in the St. John's area has consistently
demonstrated a significant level of mutagenic activity. Most of the
activity found in the DMSO extracts of used engine oil is dependent on
enzyme activation, and this can be effected with either fish or rat liver
tissue preparations. Enyzme/mediated mutagenicity has also been
demonstrated with various Salmonella strains including TA 98, TA 100, TA
135, and TA 1538. Approximately one-fifth to one-quarter of the activity
found in some samples of used oil does appear, however, to be due to the
presence of direct acting mutagens.
To date, mutagenic activity has only been observed in extracts of used
engine oil. This suggests that the mutagenic principal(s) originates from
gasoline combustion or is generated from specific engine oil compounds
within the crankcase. Mutagenic activity has also been detected in POC
extracts of a sample of used, non-petroleum base, synthetic engine oil, but
not in extracts of a comparable brand of unused oil (Table 4). Since
petroleum-base and synthetic engine oils differ in composition, it is
likely that the source of the mutagens found in both types is gasoline
combustion. Further supportive evidence has come from analysis of samples
of soot obtained from a car carburetor. This soot was observed to be
mutagenic and POC-enriched fractions of soot, as well as both types of used
engine oils, displayed similar fluorescent profiles on thin layer
chromatograms. Also, most of the mutagenic activity associated with all
three has been shown to be found in a band of material which migrates with
a common Rf.
DISCUSSION
Used engine oils which may enter the sea via runoff and municipal
wastewater could be an important source of marine environmental mutagens.
It would appear that engine oil mutagenesis is a combustion-derived
phenomenon and petroleum hydrocarbons (excellent sources of POC) which
enter the ocean via accidental discharge could represent a minor source of
mutagenic material (even if some proportions were mutagenic) compared with
the potential offered by various combustion processes. For instance,
approximately 8 million barrels of oil are lost yearly via accidental
discharge into the world's oceans (calculated from a review by McAuliffe,
1976), whereas 6 million barrels of fuel are combusted daily by vehicular
highway traffic in the United States (calculated from NAS, 1972). Also,
vehicular oil consumption likely includes only a small proportion of the
total combustion "budget" of any industrialized country.
105
-------
It is of special interest that studies on hydrocarbon cores from the
northwest Atlantic have demonstrated that urban air fallout is the most
likely source of complex hydrocarbon mixtures found in coastal and
continental sediments (Farrington et_ aK, 1977). All in all, the present
evidence suggests that polycyclic aromatic material originating from
combustion processes could be a major anthropogenic source of marine
environmental mutagens.
TABLE 2. PETROLEUM HYDROCARBON MUTAGENICITY
Crude oils Refined oils
Kuwait Fuel Oil No. 2
Louisiana Venezuelan Bunker
Western Canada Texaco Motor Oil
Venezuelan Esso Motor Oil
Norman Wells Irving Motor Oil
Atkinson's Point Gulf Motor Oil
Sable Island Veedol Motor Oil
Alberta Tar Sands
Alaskan
Pembina
Samples were extracted with DMSO in a 1:1 ratio and the polycyclic aromatic
fraction obtained was used directly in the Ames assay with Salmonella
strain TA 98. Extracts were tested in volumes ranging from 10 to 100 vi or
more.
106
-------
TABLE 3. PETKOLFUM HYDROCARBON MUTAGENICIFY
Petroleum TA 98 Revertants
Pembina
20 P£ + 59 68, 55
100 -,]£ + 59 40, 41
Atkinson1s Point
20 y£ + S9 53, 57
100 y£ +59 51, 51
Norman Hells
20 y£ + S9 67, 62
100 »j£ + 59 41, 46
Alaskan
20 M£ + S9 84, 76
100 y£ +59 34, 37
Used Engine Oil (Positive Control)
10 pi + S9 770, 729
Bacteria + 59 (Control) 47, 50
Petroleum was extracted with DMSO in a 1:1 ratio and the polycyclic
aromatic fraction obtained was used directly in the Ames assay with
Salmonella strain TA 98. Duplicate plate counts for a typical experiment
are presented.
107
-------
TABLE 4. MUTAGENICITY OF USED (SYNTHETIC) MOTOR OIL
Test Conditions TA 98 Revertants
Bacteria + S9 49, 47
Unused Oil (100 y*) 27, 33
Unused Oil (100 y*) + S9 40, 41
Used Oil (10 vi) 53, 43
Used Oil (10 yi) + S9 612, 637
Used Oil (20 y*) + S9 803, 845
Samples of used and unused (same brand) were extracted with DMSO in a 1:1
ratio, and the polycyclic aromatic fraction obtained was used directly in
the Ames assay with Salmonella strain TA 98. Duplicate plate counts for a
typical experiment are presented.
Acknowledgements
Oil samples were generously supplied by the following: Drs. Jerry
Neff, Jon Percy, Howard Sargeant, Bruce McCain, and John Starr. We also
acknowledge the cooperation of Golden Eagle Refining Company, Imperial Oil
Company, and the Alberta Oil Sands Research Center.
REFERENCES
Ames, B.N., J. McCann, and E. Yamasaki. 1975. Methods of detecting
carcinogens and mutagens with the Salmonella/mammalian microsome
mutagenicity test. Mutat. Res. 31:347.
Farrington, J.W., and J.G. Quinn. 1973. Petroleum hydrocarbon and fatty
acids in wastewater effluents. J. Water Pollut. Control Fed.
45:704.
Fingl, E., and D.M. Woodburg. 1975. General principles. In: Pharmacol-
ogical Basis of Therapeutics. L.S. Goodman and A. Gilman, Eds.,
Macmillan Publishing Co., New York. p. 1.
McAuliffe, C.D. 1976. Surveillance of the marine environment for
hydrocarbons. Mar. Sci. Commun. 2:13.
108
-------
National Academy of Sciences. 1972. Sources of polycyclic organic matter.
In: Particulate Polycyclic Organic Matter. National Academy of
Sciences/National Research Council, Washington, DC. p. 13.
Natusch, D.F.S., and B.S. Tonkins. 1978. Isolation of polycyclic organic
compounds by solvent extraction with dimethyl sulfoxide. Anal. Chem.
50:1429.
Payne, J.F., and W.R. Penrose. 1975. Induction of aryl hydrocarbon
[benzo(a)pyrene] hydroxylase in fish by petroleum. Bull. Environ.
Contam. Toxicol. 14:112.
Payne, J.F., I. Martins, and A. Rahimtula. 1978. Crankcase oils: are
they a major mutagenic burden in the aquatic environment Science
200:329.
Schmeltz, I., J. Task, J. Hi Ifrich, N. Hi rota, D. Hoffman, and E.L. Wynder.
1978. Bioassays of naphthalene and alkylnaphthalenes for
co-carcinogenic activity: relation to tobacco carcinogenesis. In:
Carcinogenesis, Vol. Ill: polynuclear aromatic hydrocarbons. J.W.
Jones and R.I. Freudenthal, Eds., Raven Press, New York. p. 47.
Tanacredi, J.T. 1977. Petroleum hydrocarbons from effluents: detection
in marine environments. J. Water. Pollut. Control. Fed. 49:216.
109
-------
CHEMICAL CARCINOGENESIS IN FISH: INDUCTION OF HEPATIC DRUG
METABOLIZING ENZYMES AMD BACTERIAL MUTAGENESIS WITH
POLYCYCLIC AROMATIC HYDROCARBONS (PAH)
by
David E. Hlnton, James E. Klaunig, Michael M.
Myong Kahng, Hayato Sanefuji,*** Raymond T. Jones,
Department of Pathology, University of Maryland.
Baltimore, MD 21201
and
Department of Anatomy, West Virginia University.
Morgantown, WV 26506
**
Lipsky, Rhona Jack,
and Benjamin F. Trump,
School of Medicine,
School of Medicine,
ABSTRACT
**
***
Channel catfish, Ictalurus punctatus, and rainbow trout,
Sal mo gairdeneri, were exposed to the PAHs benzo(a)pyrene (BaP)
and 3-methylcholanthrene (3-MC) and to polychlorinated biphenyls
(PCBs) (channel Catfish). Livers were studied by biochemical
and morphological methods. Fish injected once daily for a 7-day
period with 50 mg PCBs/kg b.w. showed increased levels of cyt-
ochromes P-450 and b5, and increased activity of NADPH cyto-
chrome c reductase. Longer exposure (21 days) caused greater
increase in the above and increased the activity of aminopyrine
demethylase. Catfish exposed to BaP or 3-MC by i.p. injection
or gastric intubation showed increased liver arylhydrocarbon hy-
droxylase (AHH) activity. Maximum induction (8-fold) followed 3
treatments totaling 75 mg 3-MC/kg b.w. Livers of trout actuely
exposed to 3-MC showed a 3-fold increase in cytochrome P-450 and
a 2-fold increase in NADPH cytochrome c reductase activity.
Aminopyrine demethylase activity was also enhanced. Increased
enzyme activity correlated with increased smooth endoplasmic
reticulum of hepatocytes. Appropriate liver fractions from non-
induced channel catfish and from 5-day induced (Aroclor 1254)
rats were compared for ability to convert BaP to a mutation
inducing compound in a bacterial mutagenesis assay. The number
of revertants in both systems was nearly identical. Data
suggest suitability of fish for carcinogen bioassay using PAH.
Present address, Department of Pathology, Medical College of Ohio,
C.S. 10008, Toledo, OH, 43699.
Present address, Ell 14th Avenue, Spokane, WA, 99202.
Present address, Department of Pathology, Universtiy of Occupational
and Environmental Health, Yahatanish-KU, Kitakyushu-City, Japan 807.
110
-------
INTRODUCTION
The aquatic environment ultimately becomes exposed to virtually every
pollutant entering the biosphere (Revelle, 1968) and neoplasms have been
reported within nearly all organic systems of feral fishes (Mawdesley-
Thomas, 1975; Wellings, 1969). Thus, analysis of tumor incidence within
fish populations might serve as a bioindicator of carcinogens in the
aquatic environment. Some application of this approach has been made
(Stitch and Acton, 1976). However, as Stewart (1977) has suggested,
controlled laboratory exposure of different fish species to known
carcinogens is needed to provide information concerning suitable species
for bioassay and for those conditions of exposure necessary for tumor
production. Such data would form a more precise basis for subsequent
environmental monitoring of aquatic carcinogens.
Except for the aflatoxin-induced liver neoplasm of rainbow trout, in
which data on metabolism of the carcinogen, histogenesis of the lesion, co-
carcinogenic and synergistic factors, and metastasis of the tumor have been
reported (Sinnhuber jrt jj]_., 1977), limited data on laboratory chemical
carcinogenesis studies are available for fish. The objective of our
research group has been to test the feasibility of using fish as a monitor
of aquatic carcinogens. To establish this system we used a correlated
morphological/biochemical approach to analyze specific steps in the
carcinogenic process, i.e., induction of drug metabolizing enzymes, activ-
ation of procarcinogens to mutagens, development of cellular culture
systems for carcinogen binding studies, and long-term exposure to various
carcinogens for the purpose of producing tumors. At each step, careful
comparison between fish and mammals, primarily the rat, was undertaken to
more carefully characterize the data in fish.
The polycyclic aromatic hydrocarbons (PAHs) are an important class of
environmental pollutants, some of which have proven carcinogenic activity
(Saffiotti ^t aU, 1968; Albert, 1976; Kraybill, 1976). PAH enrichment of
sediments in lake and ocean has recently been linked to the utilization of
coal (Muller e* j|l_., 1977) and petroleum (Kites et aU, 1977). From the
mammalian literature, it is known that the microsomal metabolism of PAH is
a necessary prerequisite for binding of these components to cellular macro-
molecules (Harris, 1976). Once activation of PAH has taken place and
binding to cell macromolecules including DMA has been demonstrated,
mutations occur in bacterial mutagenesis systems (Ames et a]_., 1975). PAH
have been shown to be metabolized by fish (Pedersen and Hershberger, 1974;
Ahokas et^ aj_., 1977). This paper relates our findings on the induction of
the fish microsomal mixed-function oxidative system (MFOS) by the PAH
benzo(a)pyrene and 3-methylcholanthrene. The data for PAH are compared to
the data for induction using the PCB mixture Aroclor 1254. Morphologic
studies are correlated to biochemical findings and compare the effects by
the inducing agents PCB and 3-methycholanthrene. In addition, our data
comparing the fish MFOS with that of the rat in the Ames (Ames et jil_.,
1975) mutagenic system, using benzo(a)pyrene (BaP), are presented.
Ill
-------
MATERIALS AND METHODS
Animals
Channel catfish of both sexes, weighing 100 to 150 g were obtained
from a local hatchery. Rainbow trout of similar weight were donated by the
State of West Virginia Department of Natural Resources. Fish were
acclimated and maintained in 100-gal living stream aquaria (Frigid Units,
Inc.) and fed a pelleted ration (Purina trout chow) daily throughout the
study.
At the start of the experiment, random fish were transferred to 20-gal
"tanks with aerated water at temperature of 12 to 15° C; 25% of the water
from control and treated tanks was changed daily.
Treatment of Animals
Fish were exposed to either BaP, 3-MC, or PCB in corn oil via intra-
peritoneal injection or gastric intubation (Table 1). Control fish
received identical treatment with corn oil only.
TABLE 1. SUMMARY OF ANIMAL TEATMENTS
Number
Fish Compound of Fish
Channel
Catfish
Rainbow
trout
PCB 15
PCB 5
BP 4
BP 4
3-MC 5
3-MC 12
Treatment Administration
50 mg/kg b.w. i.p.*
6 daily doses
1000 mg/kg b.w. g.i.**
single dose
100 rag/kg b.w. i.p.
single dose
25 rag/kg b.w. i.p.
6 daily doses
20 mg/kg b.w. i.p.
50 mg/kg b.w. i.p.
4 daily doses
Time of sacrifio
(post-exposure)
24 hours
21 days
48 hours
24 hours
24 hours
24 hours
* intraperitoneal injection
** gastric insertion
112
-------
Preparation of Microsomal Fractions
Catfish were killed by severing the spinal cord and the livers were
rapidly removed, weighed, and placed in 0.5M Tris-1.5% KC1 buffer, pH 7.4,
on ice. Minced tissues were rinsed repeatedly to remove blood, and placed
in fresh buffer. Trout were anesthetized in a 1:4000 solution of tricaine
methane sulfonate (MS-222) in water. Blood was removed from the livers by
whole body perfusion (Hinton, 1975) with a fish ringer solution at 10° C.
Following perfusion, livers were rapidly removed, blotted, weighed, and
placed in ice cold 0.3M sucrose. A 25% homogenate was made by four
complete passes of a glass Potter-Elvehjem tube and teflon pestle at 260 rpm.
Cells, nuclei, mitochondria, and lysosomes were removed by centri-
fugation at 9,000 x g for 15 min. In order to obtain a microsomal pellet,
the resulting post mitochondrial supernatant (S9) was sedimented at 100,000
x g for 60 min at 0 to 4° C. The pellet was resuspended in 0.5M Tris-
1.15% KC1 (catfish) or 0.3M sucrose (trout) to a final concentration of 0.5
grams liver/mi of microsomal suspension. Microsomal protein concentration
was determined by the method of Lowry et _al_. (1951).
Enzyme Assays
Cytochromes P-450 and b5 were determined according to the method of
Omura and Sato (1964). In order to estimate the amount of hemoprotein
present, extinction coeffients of 91 mm cm at 450 -490 nm
(P-450) and 170 mM cm at 424 -409 nm ^5) were used. Aminopyrine
demethylation was determined according to the method of Gnosspelius et a!.
(1969). The amount of formalydehyde formed was determined according to the
procedure of Nash (1953). NADPH cyt-c reductase was determined by the
method of Dallner j|t jl_. (1966), using an extinction coefficient of
19.1 mM cm at 550 nm for reduced cytochrome c. The previously
determined optimal incubation temperature for fish microsomal assays (27 to
29° C) was used. Aryl hydrocarbon hydroxylase (AHH) was determined
according to the method of Nebert and Gelboin (1968). Determination of
optimum procedures for this assay was performed and a temperature of 29° C,
a pH of 7.5, an incubation time of 20 min, and a microsomal protein
concentration between 0.1 and 0.2 mg were used. The AHH activity was ex-
pressed as nanomoles of 3-OH BaP formed per 20 min per mg of microsomal
protein.
AMES Mutagenic Bioassay
Liver homogenates (25%) from control channel catfish and Aroclor 1254
induced rats were centrifuged at 9,000 x g for 15 min. The S9 fractions
were used as the microsomal source for the bioassay. Specially constructed
mutants of Salmonella typhimurium were obtained from Dr. Bruce Ames
(University of California).Varying concentrations of BaP were added to
varying amounts of S9 from either catfish or rat. These mixtures were
incubated with one of the five modified strains of Salmonella typhimurium
according to Ames et_ al_. (1975) and McCann and Ames (1976). The number of
revertants were counted and means were expressed per mg S9 protein.
113
-------
Electron Microscopy
Pieces of liver from control and treated fish were minced in 4% form-
aldehyde-1% glutaraldehyde in phosphate buffer, pH (Mawdesley-Thomas, 1975),
for 24 hr. The tissue was post-fixed in 1% 0564, dehydrated in graded
ethanol solutions, and embedded in Epon (Luft, 1961). Semi-thin sections
(0.5-lym) were stained with toluidine blue (Trump jrt ^1_., 1961) and viewed
under a light microscope. These sections were used as a histologic survey
to determine lobular zones. Thin sections were stained with uranyl acetate
and lead citrate and viewed under an AE1-6B or a Jeol 100B electron micro-
scope.
RESULTS
MFOS Activity
The data on the MFOS response to PCB were reported in previous papers
(Klaunig e* al_., 1978; Lipsky jit £l_., 1978). Table 2 summarizes these
results as % of controls. Acute PCB exposure caused a moderate induction
of MFOS enzymes. Subacute PCB exposure (21 days) resulted in a significant
increase in induction of MFOS activity over both control and acute PCB
treatment. Both cytochrome P-450 and aminopyrine demethylase were
increased three fold over control levels. Induction of aryl hydrocarbon
hydroxylase (AHH) by two PAHs is shown in Table 2. The data are also
presented in Table 2. 3-MC caused a 10-fold increase in AHH activity, while
g.i. and i.p. exposure to BaP caused a 4- to 5-fold induction of hepatic
AHH. Table 3 summarizes the response of the rainbow trout hepatic MFOS to
acute 3-MC exposure. Cytochrome P-450 was induced to 331% of control
amounts. NADPH-cyt c reductase was also induced to over two times the
control activity. The largest degree of induction occured in AHH which was
elevated to almost 14 times over the control activity.
In all of these studies there was little to no effect on the liver to
body weight ratios or amount of microsomal protein.
Electron Microscopy
The ultrastructure of control channel catfish liver (Figure 1) was
identical to that previously reported (Hinton and Pool, 1976). The acute
and subacute effects of PCB on hepatocyte morphology were previously
reported by this laboratory (Klaunig £t al_., 1979; Lipsky e^t aj_., 1978).
In summary, acute PCB exposure resulted in proliferation of rough endoplas-
mic reticulum (RER) and formation of vesicular profiles. Little smooth
endoplasmic reticulum (SER) response was seen in the acute study. However,
subacute PCB exposure caused extensive SER proliferation in the form of
vesicles and membrane whorls (Figure 2). Increased lipid, in the form of
cytoplasmic vacuoles and as rounded aggregates inside membranes of endo-
plasmic reticulum (Figure 2), occurred in all affected hepatocytes.
Control rainbow trout hepatic morphology was identical to that previously
reported by Scarpelli (1976). Acute 3-MC treatment caused changes similar
to that seen in channel catfish after acute exposure to PCB. An apparent
114
-------
TABLE 2. SUMMARY OF INDUCTION STUDIES ON CATFISH HEPATIC MFOS COMPONENTS
P-450 05
1. PCB, 7 days*
2. PCB, 21 days*
3. 3-MC**
4. BP i.p.**
g.i.**
132%
294%
296%
223%
NADPH-cytc
Redact ase
136%
156%
Aminopyrine
Demethylase
N.C.
362%
AHH
1000%
380%
513%
All values = % of control
N^C. = No change
from Lipsky et _al_. (1978)
^from Klaunig et a].. (1978)
new data this laboratory
TABLE 3. EFFECT OF ACUTE EXPOSURE TO 3-MC UPON MFOS OF TROUT LIVER
Microscroal
protein
XADPH cycc
reductase
Aryl*
hydrocarbon
hydroxylase
Liver/bodv
wt. x 10"^
Control
3-MC
Jo induction
0.175 + .05
0.579 + .20
331%
21.03 j- 10.2
16.42 + 1.03
No change
11.24 + 2.47
25.61 + 2.61
228%
0.54 jf .06
7.37 + 1.38
1365%
1 . 40 + .2
1.29 i .10
No change
All values based upon 3 "pools" of 4 livers each +; standard deviation of
£he mean.
values based upon 1 pool of 4 livers in control and in treated fish.
115
-------
*».-
*•
g'y
••/ - - • " •>"-'
-.'/lumt '('
"^ K.I -• •
, : *
«
S . >j
• •"." *. -
Figure I. Portion of two hepatocytes from control channel catfish.
Extensive glycogen deposits (gly) occupy large portions of
cytoplasm. Mitochondria and parallel stacks of RER (arrow)
exist together near nucleus (N) and in partitions between
glycogen depot sites,
^ 1 i
i«C:
a/*". = -.» * '•*-.'• ".-"•-, •-..-• '
_____________________
Figure 2. Smooth menbrane "whorls" (SER) in catfish hepatocyte following
subacute PCB exposure. Note lipid (L) droplet and membrane-bound
lipid (arrows) in center of whorl.
116
-------
increase in RER content and disruption of paralled cisternae, in some
areas, with the appearance of vesiculated and dilated RER was noted
(Figure 3). Dilated vesicles contained a material of low to moderate
electron density. Scattered vesicular profiles of smooth membranes were
also seen (Figure 3). An electron micrograph of a microsomal pellet from a
control fish is shown in Figure 4. The pellet was composed of vesicular
forms of rough and smooth-surfaced membranes. Some free ribosomes and
glycogen particles were also seen. Pellets were free of mitochondria and
lysosomes.
Bacterial Mutagenic Bioassay
The results of the Ames mutagenic bioassay are summarized in Table 4.
The mean number of revertants per microgram of benzo(a)pyrene was
calculated. BaP produced revertants in all four Salmonella typhimurium
strains tested using rat or fish S9. Channel catfish S9 was approximately
74% and 39% as effective in producing revertants as rat S9 in the strains
1535 and 1537, respectively. However, in strains TA98 and TA100 S9 from
channel catfish proved to be over 90% as effective as S9 from PCB-induced
rat. Toxicity was noted at high benzo(a)pyrene concentrations in strain
1537 with both rat and fish.
DISCUSSION
Data presented here and in other papers in this symposium confirm
earlier reports that the fish liver contains a microsomal mixed-function
oxidase system, which is inducible by a variety of important environmental
pollutants (Pohl et jj]_., 1974; Bend et al_., 1977a; 1977b) including the
PAH, 3-MC, and BaP (Gruger et al_., 1977; Statham et ^1_., 1978). With
respect to chemical carcinogenesis, perhaps the most important feature of
the MFOS system is the induction of aryl hydrocarbon hydroxylase. This
enzyme system has been shown to be responsible for converting procarcinogenic
PAH into epoxides (Grover and Sims, 1968), more proximate carcinogenic
forms (Harris, 1976). With all of the compounds that we have tested so
far, biochemical induction of MFOS enzymes is accompanied by morphologic
change, particularly in the endoplasm reticulum of hepatocytes (Klaunig
et al.., 1978; Lipsky et ^1_, 1978). The acute response with PCB and 3-MC
was identical. The paralled RER cisternae were maintained in some areas
while dilated, vesicular profiles appeared in others. These data correlate
well with the studies by Scarpelli (1976), using aflatoxin and with changes
reported in mammalian liver (Smuckler and Arcosoy, 1969). The next step in
our investigation, following the demonstration of induction of fish MFOS
enzymes by PAH, was to determine whether PAH metabolites would cause
mutations in a bacterial mutagenesis test. Since the bacterial screen
utilizilng microsomal frac- tions from fish has been done in only a few
labs (Ahokas et al_., 1977; Stott and Sinnhuber, 1978), we ran parallel
tests with rat S9 fractions, a system which has been studied in more detail
(Ames et ^]_., 1975). The fish microsomal fraction converted BaP into
metabolites, which subsequently caused mutations in the bacterial tester
strains. The number of mutants was nearly equal to those caused by similar
fractions from the rat. These data are even stronger when it is noted that
117
-------
. ,-, a
RER
rp
,.
'
• r or*. * - -it
*>'
A •«^""- <••
V J -
.
•
'
* '•—
#&.
;0
> •
R S&
. r-1
-« liim 3
' ^ '
Figure 3. Portion of rainbow trout hepatocyte following acute 3-MC
treatment. Extensive dilated RER vesciles (R) fill large area
of cytoplasm. Parallel cisternae (RER) remain in some areas.
Smooth surfaced vesciles are also apparent (arrows).
Figure 4. Typical microsomal pellet from control fish liver.
rough (R) and smooth (arrow) surfaced membranes.
Note both
118
-------
TABLE 4. COMPARISON OF FISH AND RAT USING BaP IN AMES MUTAGENIC BIOASSAY
Salmonella
Strain
TA 98
TA 100
L537
L535
S9 source*
Rat
Fish
Rat
Fish
Rat
Fish
Rat
Fish
0
1.6
1.7
7.7
9.8
0.4
0.5
1.1
1.3
5
24.5
17.9
20.6
23.3
8.6
2.5
10.2
6.2
Micr«
10
50.0
43.8
25.8
27.6
12.7
12.2
25.8
18.8
ograms (
15
63.3
60.9
39.8
46.2
11.9
11.8
39.8
30.7
Hq) BaP
20
66.7
65.7
54.6
51.9
15.5
12.4
58.2
46.2
25
85.2
78.5
74.0
60.9
- 0 -
20.1
82.7
63.9
30
129.0
121.8
87.8
85.8
- 0 -
- 0 -
104.1
91.1
Mean
3.35
3.60
2.44
2.25
1.92
0.76
3.20
2.38
50 \ii of S9 per plate.
All values represent the mean number of revertants per milligram S9
protein per plate (minus background).
-------
the rat was induced with Aroclor 1254 for 5 days prior to the preparation
of the S9 fractions. Thus, we can state that fish not only have an MFOS
system containing enmzymes similar to those in the rodent liver but also
hu*e a capacity for making mutagenic metabolites of procarcinogens.
Subsequent studies are needed to determine whether MFOS induction and
metabolic activation of PAH will actually lead to tumor formation in fish.
In this regard, it is interesting to note that 3-MC and BaP caused
epitheliomas when painted onto skin of fish (Ermer, 1970). The wide
geographic distribution of PAH and their induction of
carcinogenesis-associated enzymes makes this system warrant further
investigation. In our laboratory, we are now ready to compare changes seen
in human cells with those in fish and rat cells and to continue long-term
feeding protocols in an effort to induce tumors in vivo. It is through the
use of these in vitro assays that interspecies comparisons and
extrapolation to man are possible. These studies are necessary (Stitch and
Acton, 1976) so that some assessment of the applicability of findings in
fish to potential human health problems can be made.
ACKNOWLEDGMENTS
The authors acknowledge the excellent technical assistance of Steven
Fidler and Michael Baladi. Rainbow trout were supplied by Raymond
Menendez, State of West Virginia, Department of Natural Resources, El kins,
WV. Research was supported in part by the Environmental Protection Agency,
Grant R-804866-01-0, and the Water Research Institute, West Virginia
University, project A-037 WV, allocated under the Water Resources Act of
1964 (PL88-379) administered by the Office of Water Resources and
Technology, Department of the Interior.
REFERENCES
Ahokas, J.T., R. Paakkonen, K. Ronnholm, V. Raunio, N. Karki, and 0.
Pelkonen. 1977. Oxidative metabolism of carcinogens by trout liver
resulting in protein binding and mutagenicity. In: Microsomes and
drug oxidations. V. Ullrich, Ed., Pergamon Press, New York. pp.
435-441.
Albert, R.E. 1976. Skin carcinogenesis. In: Cancer of the skin:
biology, diagnosis, management. R. Andrade, S.L. Gumport, G.L.
Popkin, and T.D. Rees, Eds., W.B. Saunders Co., Philadelphia.
Ames, B., J. McCann, and E. Yamasaki. 1975. Methods for detecting
carcinogens and mutagens with the Salmonella-mammalian-microsome
mutagenicity test. Mutat. Res. 31:347-363.
Bend, J.R., M.O. James, and P.M. Dansette. 1977a. In vitro metabolism of
xenobiotics in some marine animals. Ann. N.Y. Acad. Sci. 298:505-521
Bend, J.R., R.J. Pohl, E. Arinc, and R.M. Philpot. 1977b. Hepatic
microsomal and solubilized mixed-function oxidase systems from the
120
-------
little skate, Raja erinacea, a marine elasmobranch. In: Microsomes
and drug oxidations. V. Ullrich, Ed., Pergamon Press, New York.
pp. 160-169.
Dallner, G., P. Siekevitz, and G.E. Palade. 1966. Biogenesis of
endoplasmic reticulum membranes. 1. Structural and chemical
differences in developing rat hepatocyte. J. Cell. Biol. 30:73-96.
Ermer, M. 1970. Versuche mit cancerogenen mitteln bei Kurzlebigen
fischarten. Zool. Anz. 184:193-199.
Gnosspelius, Y., H. Thor, and S. Orrenius. 1969. A comparative study on
the effects of phenobarbital and 3,4-benzpyrene on the hydroxylating
enzyme system of rat liver microsomes. Chem. Biol. Interact.
1:125-137.
Grover, P.L., and P. Sims. 1968. Enzyme-catalysed reactions of polycyclic
hydrocarbons. Biochem. J. 110:159-160.
Gruger, E.H., M.M. Wekell, P.T. Numoto, and D.R. Craddock. 1977.
Induction of hepatic aryl hydrocarbon hydroxylase in salmon exposed to
petroleum dissolved in seawater and to petroleum and polychlorinated
biphenyls, separate and together, in food. Contam. Toxicol.
17:517-520.
Harris, C. 1976. Chemical carcinogenesis and experimental models using
human tissues. Beitr. Pathol. Bd. 158:389-404.
Hinton, D.E. 1975. Perfusion fixation of whole fish for electron
microscopy. J. Fish Res. Board Can. 32:416-422.
Hinton, D.E., and C.R. Pool. 1976. Ultrastructure of liver in channel
catfish, Ictalurus punctatus (Rafinesque). J. Fish Biol. 8:209-219.
Hites, R.A., R.E. Laflamme, and J.W. Farrington. 1977. Sedimentary
polycyclic aromatic hydrocarbons: the historical record. Science
198:829-831.
Klaunig, J.E., M.M. Lipsky, B.F. Trump, and D.E. Hinton. 1979.
Biochemical and Ultrastructural changes in teleost liver following
sub-acute exposure to PCB. J. Env. Path. Toxicol. 2:953-963.
Kraybill, H.F. 1976. Distribution of chemical carcinogens in aquatic
environments. Prog. Exp. Tumor Res. 20:3-34.
Lipsky, M.M., J.E. Klaunig, and D.E. Hinton. 1978. Comparison of acute
response to PCB in liver of rat and channel catfish: A biochemical
and morphological study. J. Toxicol. Environ. Health 4:107-121.
121
-------
Lowry, O.H., N.J. Rosebrough, A.L. Farr, and R.J. Randall. 1951. Protein
measurements with the Folin phenol reagent. J. Biol. Chem.
193:265-275.
Luft, J.H. 1961. Improvements in epoxy resin embedding methods. J.
Biophys. Biochem. Cytol. 9:409-414.
Mawdesley-Thomas, I.E. 1975. Neoplasia in fish. In: The anatomic
pathology of fishes. W.E. Ribelin and G. Migaki, Ed., University of
Wisconsin Press, Madison, WI. pp. 805-870.
McCann, J., and B. Ames. 1976. Detection of carcinogens as mutagens in
the Salmonella/microsome test: assay of 300 chemicals: discussion.
Proc. Nat. Acad. Sci. 73:950-954.
McDowell, E.M., and B.F. Trump. 1976. Histologic fixatives suitable for
diagnostic light and electron microscopy. Arch. Pathol. Lab. Med.
100:405-414.
Muller, G., G. Grimmer, and H. Bohnke. 1977. Sedimentary record of heavy
metals and polycyclic aromatic hydrocarbons in Lake Constance.
Naturwissenschaften. 64:427-431.
Nash, J. 1953. The colorimetric estimation of formaldehyde by means of
the Hantzsch reaction. Biochem. J. 55:416-421.
Nebert, D.W., and H.V. Gelboin. 1968. Substrate-inducible microsomal aryl
hydroxylase in mammalian culture. J. Biol. Chem. 243:6242-6249.
Omura.T., and R. Sato. 1964. The carbon monoxide binding pigment of liver
microsomes. I. Evidence for its hemoprotein nature. II.
Solubilization, purification, and properties. J. Biol. Chem.
239:2370-2385.
Pedersen, M.G., and W.K. Hershberger. 1974. Metabolism of 3,4-benzpyrene
in rainbow trout, (Salmo gairdneri). Bull. Environ. Contam. Toxicol.
12:481-486.
Pohl, R.J., J.R. Bend, A.M. Guarino, and J.R. Fouts. 1974. Hepatic
microsomal mixed-function oxidase activity of several marine species
from coastal Maine. Drug Metab. Dispos. 2:545-555.
Revelle, R. 1968. The ocean. Sci. Am. 221:55-65.
Saffiotti, U., F. Cefis, and L. Kolb. 1968. A method for the experimental
induction of bronchogenic carcinoma. Cancer Res. 28:104-124.
Scarpelli, D.G. 1976. Drug metabolism and aflatoxin-induced hepatoma in
rainbow trout, (Salmo gairdneri). Prog. Exp. Tumor Res. 20:339-350.
122
-------
Sinnhuber, R.O., J.D. Hendricks, J.H. Wales, and G.B. Putnam. 1977.
Neoplasms in rainbow trout, a sensitive animal model for environmental
carcinogenesis. Ann. N.Y. Acad. Sci. 298:389-408.
Smuckler, E.A., and M. Arcasoy. 1969. Structural and functional changes
of the endoplasmic reticulum of hepatic parenchymal cells. Int. Rev.
Exp. Path. 7:305-418.
Statham, C.N., C.R. Elcombe, S.P. Szyjka, and J.J. Lech. 1978. Effect of
polycyclic aromatic hydrocarbons on hepatic microsomal enzymes and
disposition of methyl-napthalene in rainbow trout in vivo.
Xenobiotica 8:65-71.
Stewart, H.L. 1977. Discussion paper: enigmas of cancer in relation to
neoplasms of aquatic animals. Ann. N.Y. Acad. Sci. 298:305-315.
Stitch, H.F., and A. B. Acton. 1976. The possible use of fish tumors in
monitoring for carcinogens in the marine environment. Prog. Exp.
Tumor Res. 20:44-54.
Stott, W.T., and R.O. Sinnhuber. 1978. Trout hepatic enzyme activation of
aflatoxin BI in a mutagen assay system and the inhibitory effect of
PCBs. Env. Contam. andToxicol. 19:35-41.
Trump, B.F., E.A. Smuckler, and E.P. Benditt. 1961. A method of staining
epoxy sectons for light microscipy. J. Ultrastruct. Res. 5:343-348.
Wei lings, S.R. 1969. Neoplasia and primitive vertebrate phylogeny:
echinoderms, prevertebrates and fishes - a review. Nat. Cancer Inst.
Mongr. 31:59-121.
123
-------
INDUCTION OF BENZO(A)PYRENE MONOOXYGENASE IN FISH AFTER
I.P. APPLICATION OF WATER HEXANE EXTRACT — A PRESCREEN TOOL
FOR DETECTION OF XENOBIOTICS
by
*t ^* * -*
B. Kurelec , M. Protic, M . Rijavec , S. Britivic ,
W.E.JS. Muller*1", and R.K. Zahn*f
*"Rudjer Boskovic" Institute, Center for Marine Research
Laboratory for Marine Molecular Biology, Zagreb and Rovinj,
41001 Zagreb, P.O. Box 1016, Yugoslavia
and
t Academy of Science and Letters,
65000 Mainz, Federal Republic of Germany (FRG)
ABSTRACT
Carps, Cyprinus carpio. benzo(a)pyrene monooxygenase (BaPMO)
is inducted after intraperitoneal (i.p.) application of benzo-
pyrene, methylcholanthrene, benzoanthracene, aflatoxin, or
Libyan crude oil. The kinetics of BaPMO induction in experi-
imental fish have been studied at various concentrations and at
various exposures. At a fixed time of exposure of 48 hr, the
dose-response was established after i.p. application of a series
of different concentrations of crude oil or benzo(a)pyrene
dissolved in and extracted from charcoal-treated seawater
samples. The induction-response could be traced in 1000 mi of
seawater extraction to concentrations of 0.5 parts per million
(ppm) of crude oil. This range of sensitivity lies well within
the range of naturally occurring concentrations in the marine
environment and is applicable in field monitoring of hydro-
carbons.
124
-------
INTRODUCTION
Benzo(a)pyrene monooxygenase (BaPMO) is induced in fish exposed to
petroleum (Payne and Penrose, 1975; Ahokas et a\_., 1976; Kurelec et al.,
1976; Gruger et^ ^]_., 1977). BaPMO measurements in local fish appear to be
a practical biological indicator for monitoring marine petroleum pollution
(Payne, 1976), enabling the identification of polluted marine areas and a
follow-up study of the consequences of an oil-pollution incident by
measuring the increased BaPMO levels in the livers of the local Blenniidae
(Kurelec ^t al_., 1977). It seems that presence of inducible microsomal
mixed function oxydases (MFO) activity in fish (Bend et jal_., 1977) could
serve as a useful biochemical diagnostic tool for environmental monitoring
and, at the same time, as a relevant parameter in evaluation of the effects
of acute or chronic pollution at a given site. However, from the
ecological and the environmental assessment point of view, it would be
highly desirable to detect xenobiotics present in water by means of this
early and relevant biochemical parameter. Such a method would fill the gap
which usually exist between the estimated concentration of xenobiotics in
water and corresponding biological effects. At the same time, this method
could be used as a prescreen tool to design cost-effective xenobiotic
monitoring programs using expensive analytical chemical methods.
We studied the consequence of intraperitoneal (i.p.) application of
extracts from polluted water on the liver BaPMO in young carp. The
inducibility and the basic characteristics of BaPMO in carp and the
dose-response of BaPMO after the i.p. application of polluted water
extracts are demonstrated in our first attempts to apply this method for
detection of xenobiotica in the marine environment.
MATERIAL AND METHODS
Chemicals
Benzo(a)pyrene (BaP) was from Roth, Karlsruhe, Germany; 9,10 dimethyl
1,2-benzanthracene and 20-methylcholanthrene were from Calbiochem, Lucerne,
Switzerland; NADPH from Serva, Heidelberg, Germany; hexane, fluorescent
grade, was from Merck, Darmstadt, Germany, and corn oil, commercial grade,
was from Oil Factory, Zagreb, Yugoslavia. "Sarir", Lybian crude oil, was a
gift from INA-Zagreb. Its paraffin content was 14 to 19% and the density
at 15° C was 0.843 g/cnr. All other chemicals were of analytical grade.
Seawater Samples
Samples were collected seasonally at 8 stations in the Rijeka Bay and
3 stations in the vicinity of Rovinj, within the framework of the project
"Ecological Study of Rijeka Bay" (Project Rijeka) and "Ecological Study of
Rovinj Area" (Project Rovinj), respectively. Samples were collected with a
sampler at a depth of 1 m. Experimentally polluted seawater samples were
prepared with charcoal treated seawater as the base. Quantities of Libyan
crude oil were dispersed in 50 ma of charcoal treated seawater with
Ultra-Turrax (Janke and Kunkel, Staufen, Germany) and then mixed with
125
-------
950 mi of charcoal treated seawater; l-t samples of test seawater or of
experimentally polluted seawater were extracted with hexane and the extract
prepared for i.p. application as described previously (Kurelec et al.,
1979). Aliquots of these extracts were screened for their content of
extractable substances by measuring their fluorescence at an activation
wavelength of 313 run and an emission wavelength of 360 nm in a Zeiss PMO 3
spectrophotofluorimeter calibrated with hexane on 1000 fluorescence units
(f.u.)-
Animals
One-year-old specimens of artificially hatched carp, Cyprinus carpio,
weighing 8 to 12 g were adapted for 1 month in 200-A basins, with 150 t of
dechlorinated, well-areated water, at a density of 400 specimens/m at a
flow of five total changes per day at 14° C and then used in the experi-
ments. Hexane extracts, dissolved in 0.1 mi of corn oil, were injected
i.p. to carp from 0800 to 1000 hr. Animals were given no food during the
experimental period.
Homogenate Preparation
Subcellular fractions and BaPMO assay with protein measurements were
performed as previously described (Kurelec jrt aj_., 1977).
RESULTS AND DISCUSSION
Some Properties of BaPMO in Carp
The induction of BaPMO was measured in carp subsequent to the i.p.
application of a single dose of different substances. The results of the
BPMO activity measurements after various exposure times are presented in
Table 1. In these experiments, the postmitochondrial liver fraction of a
carp which had been induced by a single dose of the test substance was
estimated at a optimal pH (7.4) and at a temperature of 29° C. The
velocity of benzo(a)pyrene hydroxylation was proportional to the amount of
postmitochondrial fraction from 25 vi to 150 \ii, or in the range from 0.04
to 0.23 mg of proteins per sample.
Time Course of BaPMO Induction
A group of carp was treated i.p. with a single dose of benzo(a)pyrene
dissolved in 0.1 mi of corn oil at a dose of 10 mg/kg. During the first
24 hr, 4 specimens were sacrificed every 2 hr. At 36, 48, 72, and 96 hr,
after injection additional fish were killed. The BaPMO activity was
quickly estimated and the rise in the activity was plotted (Figure 1).
Induction of BaPMO with Extracts of Seawater Experimentally Polluted with
Crude Oil
A series of 12 carp were treated i.p. by injection of 0.75, 1.5, and
15 mg of crude Libyan oil dissolved in 0.1 mi of corn oil. Another group
126
-------
TABLE 1. INDUCTION OF BaPMO ACTIVITY IN CARP AFTER I.P.
APPLICATION OF SINGLE TREATMENT WITH DIFFERENT
SUBSTANCES
Substance
Benzo(a)
pyrene
20-Methylcho-
lanthrene
Dose in mg/kg
body weight
5-7
40
4-5
0.15
1.0
3.0
6-7
5
2
6
8-11
5
7
4-7
9,10 Dimethyl
1-2 benzanthra- 5
cene
Phenobarbital
Control
4
100
Exposure
time in hr
24
48
72
72
72
72
. 120
24
24
24
72
72
120
24
120
72
BaPMO activity
in a. u.
171.1 ± 70.8
2600 + 1870
1920 + 20.2
73 + 56
296 + 145.6
820 ~
2463 + 886.4
4540
1749
2143 + 550
2035 + 1065
1398
2549
1132 + 60.8
1100
846
18.9 + 3.6
23.8 ± 19.9
(5)
(2)
(2)
(2)
(2)
(1)
(3)
(1)
(1)
(2)
(3)
(1)
(1)
(2)
(1)
(1)
(2)
(7)
127
-------
1200
1000
ro
co
VI
c
O
Q.
en
fO
c
o
Q.
CO
800
600
400
200
Control fish 10.2-14.6 (11)
_L
12
24
36
Figure 1.
Time course of BaPMO induction in carp after i.p. treatment with a
single dose 10/mg/kg benzo(a)pyrene.
48 72
exposure hours
96
-------
was injected with the hexane-extracted material from 1-t charcoal-treated
seawater samples to which 0.75, 1.5, and 15 mg of crude oil were added.
All fish were exposed for 48 hr. The dose-response of BaPMO-induced
activity is shown in Table 2. Prior to evaporation, fluorescence of hexane
extracts of the experimentally polluted seawater were measured. The
results are plotted in Figure 2.
Hexane extracts of seawater samples collected at 8 stations from the
Project Rijeka revealed of 232.7*315.6 fluorescence units, which is
equivalent to 8.0-10.9 yg/Ji crude oil. The range of fluorescence units
in 13 samples lies between 50 and 1022 f.u. (or the equivalent of 1.72 to
35.2 yg/Jt crude oil). As will be shown, these fluorescent materials re-
vealed higher biological activity than the material of the crude oil-origin
which had a 20-fold higher fluorescence.
Field Observations
In September and December 1977, and in March and June 1978, we
collected samples of seawater within the work on Rijeka Project at 8
stations (Figure 3). The samples (1 m depth) were extracted expeditiously
with hexane and the amount of fluorescent material estimated. BaPMO tests
were accomplished within a week. The results of a March 1977 excursion,
which was chosen as a representative,out of four similar results obtained
in other excursions, are presented in Table 3. Although the biological
activities of these samples were low, the discrepancy between the
fluorescence value and the BaPMO induction are obvious. Station A in Table
3 was a coastal site where we estimated the seasonal BaPMO level in a
territorial fish, Blennius pavo. There, 500 m from the petroleum
processing discharge outlet, these fish were highly induced relative to
fish in five other sites (see Figure 3). Seawater samples taken there
contained a large amount of extractable material that gave a high BaPMO
induction.
On March 29, 1978, samples of seawater collected at three stations
from the Project Rovinj (Figure 4) revealed an interesting phenomenon:
samples from sites R4 and R5, containing the "usual" concentration of
hexane extractable material (as estimated by fluorescence units) do contain
substances that are very biologically active substances (see Table 4).
It seems likely that this active material also caused extreme BaPMO
induction in the population of Mugil cephalus that existed for a long
period in the vicinity of a fish-cannery discharge. We have shown that
hexane-extracted material from seawater samples taken at the border of the
"mixing zone" of this pollution source possessed premutagens and/or
precarcinogens, since they caused an increase of the number of his+
revertants of Salmonella typhymurium strain T 100 after activation with
postmitochondrial fraction of liver homogenates from induced Mugil
specimens (Kurelec j?t £l_., 1979).
129
-------
TAHLE 2. INDUCTION OF HaPMO WITH EXTRACTS OF SEA WATER EXPERIMENTALLY POLLUTED WITH CRUDE OIL
Cnulo oil
1 loxnne
oxl racl s
Experiment
No
1
1
2
Dose in nig
0 75
23 8 i- 19.9 (7) 21 +
23. R + 19-9 (7) 38 +
44. 6 ± 9-9 (10) 63,8 +
of crude oil/kg body weight
150
2.8 (2) 50.5 + 17.7 (2)
11 (5) 50 ± 23.4 (6)
15 .1 (4) 61.8 + 12.8 (4)
1500
124 (I)
124.8+36.7 (4)
156.8 + 75.7 (4)
co
o
llnl'MO ACTIVITIES ARE EXPRESSED IN UNlTS/mg OF PROTEIN + STANDARD DEVIATION. NUMBER Ol>
SPECIMENS USED IN EXPERIMENTS ARE IN PARENTHESIS.
-------
TABLE 3. BaPMO INDUCTION IN CARP TREATED I.P. WITH HEXANE EXTRACTS OF
SEAWATER SAMPLES FROM RIJEKA BAY
March 8, 1978
Station
1
12
9 a
19
7
20a
11
8a
A
Fluorescence units of
hexane extracts
1252
1220
5088
1236
1771
1348
1200
1410
8240
BaPMO induction
in a.u./mg protein
22
28
14
5
34
17
30
8
925
TABLE 4. INDUCTION IN CARP TREATED I.P. WITH HEXANE EXTRACTS OF SEAWATER
SAMPLES COLLECTED IN THE VICINITY OF ROVINJ
March 29, 1978
Station
R4
R5
R6
Fluorescence units
of hexane extracts
1280
1140
2230
BaPMO induction
in a.u./mg protein
706
302
34
131
-------
m
o
n
400
ro
H
01
40 -
CO
ro
8
V)
EM
20
0.5
1.0
1.5
15.0
Figure 2. The fluorescence of hexane extracts of the seawater treated with
0.75, 1.5, and 15 mg/st, crude oil.
-------
- IPO * ^0 n 11
_ I w v-/ wW >-L • '—J «
Figure 3. Sampling sites in the Rijeka Bay where samples of
seawater (numbers) and specimens of Blennius pavo were
collected (capitals).
133
-------
Figure 4. Sampling sites in the vicinity of Rovinj,
134
-------
SUMMARY
In some cases, biological responses were concentration-dependent. In
others, fluorescent material was not biologically active, or considerable
biological activity was not correlated with the corresponding fluorescence.
These observations indicate that BaPMO induction could distinguish two
types of hexane extractable fluorescent material: biologically active
substances and biologically inactive ones. In addition, this system
monitors hexane extractable material that does not fluoresce as expected,
i.e., it detects xenobiotics where fluorescence monitoring methods fail to
signal their presence. In the present work, BaPMO was shown to be induced
by known carcinogens, crude oil, complex petroleum processing discharges,
domestic sewage, and discharges from a fish cannery. This demonstrates
that BaPMO induction in fish after i.p. application of hexane extractable
fractions could be used in a survey for xenobiotics in the marine
environment. A special value would represent, as in our field observations,
the use of the BaPMO induction test as a prescreen tool for developing an
appropriate cost-effective analytical method that should (or should not) be
used in specific sites. The method described here brings a new quality for
the application of MFO induction monitoring in environmental studies: it
sensitively detects xenobiotics in any given sample of water. I.p.
induction, was strongly correlated with naturally occuring induction in
fish living in the corresponding water. Therefore, substances that induce
BaPMO after i.p. treatment apparently would induce fish living in polluted
environment. In addition to the diagnostic value for one biochemical
parameter, the BaPMO induction also has a predictive value in the
assessment of the environmental hazard from toxic, mutagenic, or
carcinogenic substances (Kurelec et^ al_., 1979). BaPMO induction monitoring
may allow us to distinguish environmentally hazardous substances from
innocuous ones. This fact, therefore, could be a relevant basis for the
development of regulatory criteria in water quality control. In our
experiments, carp BaPMO was induced significantly only at crude oil
concentrations above 1 mg/z. These levels represent realistic
environmental levels near petroleum processing discharges that are limited
by U.S. regulations to an effluent level of less than 1.0 mg/£ freon
extractables (D.J. Baumgartner, personal communication).
We intend to improve the method further by both technical and biolog-
ical means. The latter include the use of either Fi-sister generation of
fish and the use of monoclonal fish species (Hart et al_., 1977), which will
decrease background deviation and give more uniform induction-response. We
hope to further increase the sensitivity of the method, but not circumvent
the necessity of the development of regulatory criteria based on the
rational use of early indicators for environmental contaminations by
compounds that can be bioactivated to cytotoxins, mutagens, and carcinogens.
135
-------
ACKNOWLEDGEMENTS
The authors are grateful to the Self-Management Community of Interest
for Scientific Research of S.R. Croatia for financial support. We also
acknowledge gifts and support from the Bundesministerium fur Forschung
und Technologie, Internationales Buro der KFA Julich, Germany and from the
Academy of Science and Literature, Mainz, Germany. The work has been
conducted as a FAO/UNEP Joint Coorindated Project on Pollution in the
Mediterranean.
REFERENCES
Ahokas, J., R. Paahhonen, K. Ronholm, V. Raunio, and 0. Pelkonen. 1976.
Oxidative metabolism of carcinogens by trout liver resulting in
protein binding and mutagenicity. Z. Physiol. Chem. 357:1028.
Bend, J.R., M.O. James, amd P.M. Dansette. 1977. In vitro metabolism of
xenobiotics in some marine animals. Ann. N.Y. Acad. Sci. 298:505.
Gruger, E.R., Jr., M.M. Wekell, P.T. Numoto, and D.R. Craddock. 1977.
Induction of hepatic aryl hydrocarbon hydroxylase in salmon exposed to
petroleum dissolved in seawater and to petroleum and polychlorinated
biphenyls, separate and together, in food. Bull. Environ. Contam.
Toxicol. 17:512.
Hart, R.W., S. Hays, D. Brash, F.B. Daniel, M.T. Davis, and M.J. Lewis.
1977. In vitro assessment and mechanism of action of environmental
pollutants. Ann. N.Y. Acad Sci. 298:141.
Kurelec, B., R.K. Zahn, S. Britvic, N. Rijavec, and W.E.G. Muller. 1976.
Benzopyrene hydroxylase induction - molecular response to oil
pollution. ACMRR/IABO Expert Consultation on Bioassays with Aquatic
Organisms in Relation to United Nations Food and Agriculture
Organization (FAO).
Kurelec, B., S. Britvic, M. Rijavec, W.E.G. Muller, and R.K. Zahn. 1977.
Benzo(a)pyrene monooxygenase induction in marine fish - molecular
response to oil pollution. Mar. Biol. 44:211.
Kurelec, B., Z. Matijasevic, M. Rijavec, I. Alacevic, S. Britvic, W.E.G.
Muller, and R.K. Zahn. 1978. Induction of Benzo(a)pyrene
monooxygenase in fish and the Salmonella test as a tool for detecting
mutagenic/carcinogenic xenobiotics in the aquatic environment. Bull.
Environ. Contam. Toxicol. 21:799.
Payne, J.F., and H.R. Penrose. 1975. Induction of aryl hydrocarbon
[benzo(a)pyrene] hydroxylase in fish by petroleum. Bull. Environ.
Contam. Toxicol. 14:112.
Payne, J.F. 1976. Field evaluation of benzopyrene hydroxylase induction
as a monitor for marine petroleum pollution. Science 191:945.
136
-------
IN. VIVO AND _IN VITRO STUDIES ON THE
METABOLISM OF POLYCYCLIC AROMATIC HYDROCARBONS
BY MARINE CRABS
by
Richard F. Lee and Sara C. Singer
Skidaway Institute of Oceanography
P.O. Box 13687
Savannah, Georgia 31406
ABSTRACT
Polycyclic aromatic hydrocarbons were taken up from the
food or water. The metabolism of these aromatics resulted in
hydroxylation of the parent compound followed by excretion of
the metabolites. In vitro assay for aryl hydrocarbon hydroxy-
lase showed high activity in the stomach tissues of both male
and female crabs. The green gland, an organ with functions
similar to the vertebrate kidney, was high in activity only in
the female crab. The aryl hydrocarbon hydroxylase activity in
the green gland varied during the molt cycle with large
increases after the final molt. These results suggest that the
green gland may function as a regulator of hormone levels. The
effects of various inhibitors, including detergents, phospho-
lipase, cytochrome c, carbon monoxide, piperonyl butoxide, and
benzoflavone indicated that aryl hydrocarbon hydroxylase system
in crabs was composed of cytochrome P-450, phospholipid, and
cytochrome P450 reductase.
INTRODUCTION
Effects of aromatic and other petroleum hydrocarbons on a wide variety
of marine benthic crustaceans (e.g., shrimp, crabs, and lobster) have been
investigated (Anderson et aj_., 1974; Atema and Stein, 1974; Caldwell et al_.,
1977; Eorns, 1977; Karinen and Rice, 1974). When oil spills reach in-
tertidal areas, the crabs are among the most severely affected animals. In
the San Francisco Bay spill and Buzzards Bay spill, many years were
required before shore crabs, Pachygrapsus, or fiddler crab Uca, recolon-
ized heavily oiled areas (Chan, 1975; Kreb and Burns, 1977). Crab larvae
are susceptible to the water soluble fractions of oil, possibly due to fre-
quent molts during this period of their life history (Mecklenburg et al_.,
1977; Katz, 1973).
137
-------
Crabs can take up aromatic hydrocarbons from food or water. These
hydrocarbons are oxidized to various products and the metabolites are later
excreted in the urine or feces (Lee et jjU, 1976; Corner ej; aU, 1973).
This paper summarizes studies on the metabolism of PAH polycyclic aromatic
hydrocarbons by anomuran and brachyuran crabs. The tissue and enzyme
systems involved in these reactions are discussed.
Metabolism of Polycyclic Aromatic Hydrocarbons by Living Crabs
Exposure of blue crabs, Callinectes sapidus. to a variety of radio-
labeled aromatic hydrocarbons including benzo(a)pyrene (BaP), fluorene,
naphthalene, methyl naphthalene, and methylcholanthrene) in food or water
resulted in metabolism of those compounds to various phenols, diols, and
their conjugates (Lee et _al_., 1976). The hepatopancreas contained highly
polar metabolites—predominantly diols and their conjugates, whereas the
blood had phenols and diols. The build-up of metabolites in the hepato-
pancreas suggested that this organ was important in hydrocarbon metabolism.
The green gland, which has excretory functions, had no hydrocarbons and all
radioactivity was in the form of highly polar metabolites.
Corner and coworkers (Corner ^t a\_., 1973) exposed the spider crab,
Maia squinado. to food containing naphthalene. After uptake the urine
contained unchanged naphthalene, l,2-dihydro-l,2-dihydroxynaphthalene, a
glucoside of this compound, 1-naphthyl sulphate, and 1-naphthyl glucoside.
The presence of glucosides and sulfate indicated that crabs have the
required conjugating enzymes. In mammals glucuronic acid is the main
glycosidic conjugate but in crustaceans and insects it appears that glucose
serves as the glycoside (Corner et al_., 1973; Kahn £t_aJL, 1974; Elmamlouk
and Gessner, 1977).
In Vitro Metabolism of Polycyclic Aromatic Hydrocarbons by Crabs
Extensive studies have been conducted on the metabolism of aromatic
hydrocarbons by microsomal preparations of mammalian livers (Houston,
1975). The initial reaction is an oxidation catalyzed by an oxygenase
system to form arene oxides. These reactive arene oxides can be nonenzy-
matically hydrated to phenols, enzymatically hydrated to diols by an epoxide
hydrase, or glutathione conjugates can be formed by the action of gluta-
thione-S-transferase (Bend et al_., 1976; Lu et £l_., 1976).
In an assay using formation of hydroxybenzo(a)pyrene from BaP,
significant mixed function oxygenase (MFO) activity has been noted in the
microsomes from the stomach and green gland of blue crabs, Callinectes
sapidus (Singer and Lee, 1977). However, only very low activity was detect-
ed in the hepatopancreas. This low activity in the hepatopancreas was at
least partly due to a MFO inhibitor in this tissue. A similar inhibitor
has been reported in the hepatopancreas of lobster (James ^t ^1_., 1977).
Using the amount of aldrin epoxide formed for the assay, Burns (1976) re-
ported MFO activity in the green gland of the fiddler crab, Uca pugnax.
Low MFO activity has been detected in gonadal and gill tissues of the blue
crab (Singer and Lee, 1977).
138
-------
Characterization of the MFO system in the stomach of blue crabs in-
dicated that all activity was in the microsomes and maximal enzyme acti-
vity was attained at 30° C and pH 7.5 in the presence of NADPH, oxygen and
magnesium (Singer et jj]_., 1978). The association of the enzyme activity
with the microsomes (collected at 100,000 x g) indicated that bacteria were
not responsible for the observed hydrocarbon degradation since bacteria
were removed by the preliminary low speed centrifugation (800 x g).
The components of the MFO system in mammals are cytochrome P-450,
NADPH cyctochrome P-450 reductase, and phospholipid (Lu jrt^l_., 1976). The
MFO system in crab stomach appeared to be composed of similar components
(Singer £t ^1_., 1978). P-450 inhibitors tested included SKF-525A at
10~4M(48% inhibition), 10~°M 7,8-naphthoflavone (84% inhibition),
and 10 M piperonyl butoxide (70% inhibition). The presence of NADFPH
cytochrome p-450 reductase activity was determined with cytochrome c as the
electron recipient. MFO activity was inhibited when cytochrome c was added
to the assay. The phospholipid requirement for crab stomach MFO was shown
by the inhibition of MFO activity by detergents and phospholipase C.
The primary product of BaP metabolism by crab stomach microsomes was
3-hydroxybenzo(a)pyrene. This was shown by the ultraviolet fluorescence
spectra of the isolated metabolites compared with the authenthic standards
(Figures 1 and 2). BaP metabolite had a retention time identical to
3-hydroxybenzo(a)pyrene and a small peak tentatively identified as
9-hydroxybenzo(a)pyrene (Lee and Gonsoulin, 1978). In addition to BaP the
crab stomach microsomal preparations were presented with four other
polycyclic aromatic hydrocarbons, which included phenanthrene, chrysene,
benz(a)anthracene and dimethylbenz(a)anthracene. The primary products of
dimethylbenz(a)anthracene and benzo(a)pyrene metabolism were the phenolic
metabolites while for the other hydrocarbons diol derivatives were
predominant (Table 1).
No induction of MFO activity in stomach or green glands was observed
when blue crabs were injected with either benz(a)anthracene or phenobarb-
ital. Food contaminated with benz(a)anthracene was fed to crabs but
resulted in no increase in stomach MFO activity. Fiddler crabs collected
from a salt marsh contaminated with oil showed no difference in MFO
activity compared with crabs from a clean area (Burns, 1976). In contrast,
insects, fish and mammals all showed higher MFO activity after exposure to
polycyclic aromatic hydrocarbons (Gelboin, 1972; Philpot et al_., 1976;
Payne and Penrose, 1975; Wilkinson and Brattsten, 1972).
Using ^C-styrene oxide as the substrate, James and coworkers
(1977) found high epoxide hydrase activity in microsomal preparations from
the hepatopancreas of the rock crab, Cancer ittorus. and the blue crab,
Callinectes sapidus. Hepatic glutathione-S-transferase activity was less
in these crabs than epoxide hydrase activity whereas glutathione-S-trans-
erase was higher in activity in mammals, suggesting that microsomal
hydrases in crabs may be relatively more important for epoxide detoxifica-
tion than in mammals. The presence of glucoside and sulfate conjugates of
naphthalene metabolites in the urine of naphthalene exposed spider crabs
139
-------
CO
•z.
LL)
LU
I
—I
LU
01
2a
CO
UJ
Ld
UJ
CC
2b
430 490
WAVELENGTH (nm)
300
WAVELENGTH (nm)
400
Figure 1. (a) Emission spectra of 3-hydroxbenzo(a)pyrene in ethanol
isolated from in vitro assay ( )of arylhdrocarbon
hydroxylase and from an authentic standard ( ).
Excitation wavelength set at 374 nm. (b) Excitation spectra
of 3-hydroxybenzo(a)pyrene in ethanol as in (a). Emission
wavelength set is 430 nm (From Singer and Lee, 1978).
-------
.8
3'hydroxy benzo(a)pyrene
in vitro product
.5
a>
o
c
o
_
O
lf>
JO
.2
240
260 280 300 320
Wavelength (nm)
340
360
380
400
Figure 2. UV spectra of 3-hydroxybenzo(a)pyrene in ethanol isolated from in vitro assay
( ) of arylhydrocarbon hydroxylase and from an authentic standard ( )
(from Singer and Lee, 1978).
-------
TABLE 1. MODIFICATION IN VITRO OF CARIOUS ARYLHYDROCARBONS BY STOMACH
MICROSOMES FROM BLUE CRAB
Substrate
Phenanthrene
Benz(a)anthracene
Dimethyl benz (a ) anthracene
Chrysene
Benzo(a)pyrene
Phenols
(pmole/hr)
0
10
187
7
88
Diols
(pmole/hr)
39
20
64
16
4
Specific
Activity
(pmole/hr mg)
65
50
410
38
154
Incubation mixture had 1 mg microsomal protein, 0.6 pmoles
and 0.02 ymoles radiolabeled aromatic hydrocarbons. The compounds were
(9- C) phenanthrene (11.3 yCi/ymole), (12- C)benz(a)anthracene
(49 yCi/ymole), 7, 12-dimethyl(12- C)benz(a)anthracene.(21 yCi/ymole),
[5,6(11,12-14C)] chrysene (6.3 yCi/ymole), and (7, 10"1\)benzo(a)pyrene
(51 yCi/mole). Assays were performed in triplicate and incubated for 60
min. The assay was terminated by adding 1 ma cold acetone and 3 mi hexane.
The entire organic phase was evaporated under nitrogen gas and quantitatively
spotted on precoated silica-gel thin-layer plates. The plates were developed
in benzene: ethanol (9:1) and allowed to dry. Areas of the plates
occupied by monohydroxylated (phenols) and dihydroxylated (diols) products
corresponding to the Rf's of authentic standards 3-hydroxybenzo(a)pyrene
and 5, 6-dihydroxybenz(a)anthracene were marked, scraped, and radio-
activity determined in a liquid scintillation mixture. All other areas of
the plates were pooled and determined for radioactivity in the same manner.
Units are picomoles per hour (pmol/hr).
142
-------
(Corner jjit al_., 1973) suggests the presence of UDP-glucose transferase and
sulfate transferase in these crabs. Elmamlouk and Gessner (1977) showed
glucosylation of p-nitrophenol by homogenates of hepatopancreas of the
lobster, Homarus americanus.
Interactions of Mixed Function Oxygenase, Steroid Hormones, and Polycyclic
Aromatic Hydrocarbons
The presence of MFO activity in the stomach of crabs suggests this
organ is important in the metabolism of aromatic hydrocarbons encountered
via the food. The role of the MFO system in the green gland is not clear
since this organ is considered to have primarily excretory functions.
Crabs exposed to hydrocarbon showed no buildup of hydrocarbons in the green
gland and the presence of only polar metabolites as expected for an organ
with excretory functions (Lee^t jil_., 1976). The MFO activity in the blue
crab green gland changes during development (Figure 3) in a manner similar
to that described for insects during their molting cycle (Perry and
Buckner, 1970; Yu and Terriere, 1971). The changes in green gland MFO
activity and levels of molting hormones can be correlated thus: molting in
crustaceans is controlled by a group of steroid hormones called ecdysones
produced by the y organs (Passano, 1960; Goad, 1976); during intermolt, a
molt inhibiting hormone produced by the x organ is present which suppresses
the y organ. Ecdysones have been measured in Callinectes sapidus during
three stages of the molt (Faux et _§]_., 1969). The hormones were lowest in
intermolt, higher during proecdysis, and highest just after ecdysis when
the crab was soft. The MFO activity in crab green glands was inversely
related to the levels of ecdysones, highest in intermolt, falling during
proecdysis, and lowest just after the molt when the crab was soft (Figure 1)
Immediately following ecdysis, MFO activity increased rapidly when the
levels of ecdysones undergo an opposing decrease. An important ecdysone in
crabs is crustecdysone, which is a 20-hydroxy derivative of a cholesterol-
like presursor. Thus, steroid hydroxylase activity by the MFO system in
the green gland could regulate levels of this hormone.
If mixed function oxygenases help to control molting hormone levels,
then foreign organic compounds, such as polycyclic aromatic hydrocarbons,
could compete with molting hormones at the active site of MFO and thus
alter the rate at which crabs pass through early molts. This competition
may explain why juvenile crabs are much more sensitive to pollutants than
aduts (Armstrong et al_., 1976; Karinen and Rice, 1974; Nimmo _et al_., 1971).
143
-------
12345
(stages)
T
fast
I
fast
12 16 20 24 48
(hours) (weeks)
IHh
INTERMOLT
PROECDYSIS ECDYSIS
POSTECDYSIS
Figure 3. Flucation of mixed function oxygenase activity during molting in the green gland
of female blue crabs. Intermolt stages were arbitrarily judged by appearance of
molting rings followed by a fast beginning 3 to 7 days prior to ecdysis. Post-
ecdysis was measured from the moment when newly molted crab was free of old carapace
(from Singer and Lee, 1977).
-------
REFERENCES
Anderson, J.W., J.M. Neff, B.A. Cox, H.E. Tatem, and G.H. Hightower. 1974.
Characteristics of dispersions and water-soluble extracts of crude and
refined oils and their toxicity to estuarine crustaceans and fish.
Mar. Biol. 27:75-88.
Armstrong, D.A., D.V. Buchanah, M.H. Mallon, R.S. Caldwell, and R.E.
Millemann. 1976. Toxicity of the insecticide methoxychlor to the
dungeness crab Cancer magister. Mar. Biol. 38:239-252.
Atema, J. and L.S. Stein. 1974. Effects of crude oil on the feeding
behavior of the lobster Homarus americanus. Environ. Pollut.
6:77-86.
Bend, J.R., Z. Ben-Zui, J. Van Anda, P.M. Dansette, and D.M. Jerina. 1976.
Hepatic and extrahepatic glutathione S-transferase activity toward
several arene oxides and epoxides in the rat. In: Carcinogenesis,
Vol. 1. Polynuclear aromatic hydrocarbons: chemistry, metabolism,
and carcinogenesis. R.I. Freudenthal and P.W. Jones, Eds., Raven
Press, New York. pp. 63-75.
Burns, K.A. 1976. Hydrocarbon metabolism in the intertidal fiddler crab
Uca pugnax. Mar. Biol. 36:5-11.
Caldwell, R.S., E.M. Caldarone, and M.H. Mallon. 1977. Effects of a
seawater-soluble fraction of Cook Inlet crude oil and its major
aromatic components on larval stages of the Dungeness crab, Cancer
magister Danna. JJK Fate and effects of petroleum hydrocarbons in
marine organisms and ecosystems. D.A. Wolfe, Ed., Pergamon Press,
New York. pp. 210-220.
Chan, G.L. 1975. A study of the effects of the San Francisco oil spill on
marine life part II: recruitment. In: Conference on prevention and
control of oil pollution. American Petroleum Institute, Washington,
DC. pp. 457-461.
Corner, E.D.S., C.C. Kilvington, and S.C.M. O'Hara. 1973. Qualitative
studies on the metabolism of naphthalene in Mai a squinada (Herbst).
J. Mar. Biol. Assoc. U.K. 53:819-932.
Elmamlout, T.H., and T. Gessner. 1977. Glucosylation of p-nitrophenol by
hepatopancreas of Homarus americanus. Xenobiotica 7:111-112.
Faux, A., D.H.S. Horn, E.J. Middleton, H.M. Fales, and M.E. Lowe. 1969.
Molting hormones of a crab during ecdysis. Chem. Commun., (J.Soc.
Chem. Sect. D) 1969:175-176.
Forns, J.M. 1977. The effects of crude oil on larvae of lobster Homarus
americanus. In: Proceedings 1977 Oil Spill Conference. American
Petroleum Institute, Washington, DC. pp.569-573.
145
-------
Gelboin, H.U. 1972. Studies on the mechanism of microsomal hydroxylase
induction and its role in carcinogen action. Rev. Can. Biol.
31:31-60.
Goad, L.J. 1976. The steroids of marine algae and invertebrate animals.
^n: Biochemical and biophysical perspectives in marine biology. D.C.
Mai ins and J.R. Sargent, Eds., Academic Press, London, pp. 213-218.
Hutson, D.H. 1975. Mechanism of biotransformation. In: Foreign compound
metabolism in mammals, Vol 3. D.E. Hathway, Ed., The Chemical
Society, London, pp. 449-549.
James, M.O., J.R. Fouts, and J.R. Bend. 1977. Xenobiotic metabolizing
enzymes in marine fish. In: Pesticides in the aquatic environment.
M.A.Q. Khan, Ed., Plenum Press, New York. pp. 171-189.
Karinen, J.F., and S.D. Rice. 1974. Effects of Prudhoe Bay crude oil in
molting tanner crabs, Chionoecetes bairdi. Mar. Fish. Rev.
36:31-37.
Katz, L.M. 1973. The effects of water soluble fraction of crude oil on
larvae of the decapod crustacean Neopanope texana (Sayi). Environ.
Pollut. 5:199-203.
Krebs, C.J., and K.A. Burns. 1977. Long-term effects of an oil spill on
populations of the salt-marsh crab Uca pugnax. Science.
197:484-487.
Lee, R.F., and F. Gonsoulin. 1978. Unpublished data.
Lee, R.F., C. Ryan, and M.L. Neuhauser. 1976. Fate of petroleum
hydrocarbons taken up from food and water by the blue crab Callinectes
sapidus. Mar. Biol. 37:363-370.
Lu, A.Y.H., W. Levin, M. Vore, A.H. Conney, D.R. Takker, G. Holden, and
D.M. Jerina. 1976. Metabolism of benzo(a)pyrene by purified liver
microsomal cytochrome P. 448 and epoxide hydrase. In:
Carcinogenesis, Vol. 1. Polynuclear aromatic hydrocarbons:
chemistry, metabolism, and carcinogenesis. R.I. Freudenthal and P.W.
Jones, Eds., Raven Press, New York. pp. 115-126.
Mecklenburg, T.A., S.D. Rice, and J.F. Karinen. 1977. Molting and
survival of king crab (Paralithodes camtschatica) and coon stripp
shrimp (Panda!us hypsinotus) larvae exposed to Cook Inlet crude oil
water-soluble fraction. Jji: Fate and effects of petroleum
hydrocarbons in marine organisms and ecosystems. D.A. Wolfe, Ed.,
Pergamon Press, New York. pp. 221-228.
Nimmo, D.R., R.R. Blackman, A.J. Wilson, and J. Forester. 1971. Toxicity
and distribution of Arochor 1254 in the pink shrimp Panaeus duorarum.
Mar. Biol. 11:191-197.
146
-------
Passano, L.M. 1960. Molting and its control. In: The physiology of
crustacae, Vol. 1. T.H. Waterman, Ed., Academic Press, New York, pp,
473-527.
Payne, J.F. and W.R. Penrose. 1975. Induction of arlhydrocarbon
(benzo(a)pyrene) hydroxylase in fish by petroleum. Bull. Environ.
Contain. Toxicol. 14:112-166.
Perry, A.S., and A.J. Buckner. 1970. Studies on microsomal cytochrome
P-450 in resistant and susceptible houseflies. Life. Sci.
9:335-350.
Philpot, R.M., M.O. James, and J.R. Bend. 1976. Metabolism of
benzo(a)pyrene and other xenobiotics by microsomal mixed-function
oxidases in marine species. In: Sources, effects and sinks of
hydrocarbons in the aquatic environment. American Institute of
Biological Sciences, Washington, DC. pp. 184-199.
Singer, S.C., and R.F. Lee. 1977. Mixed function oxygenase activity in
blue crab, Callinectes sapidus: tissue distribution and correlation
with changes during molting and development. Biol. Bull.
153:377-386.
Singer, S.C., P.E. March, and R.F. Lee. 1978. Mixed function oxygenase
activity in the blue crab, Calllnectes sapidus: characterization of
enzyme activity from stomach tissue. Unpublished manuscript.
Wilkinson, C.F., and L.B. Brattster. 1972. Microsomal drug metabolizing
enzymes in insects. Drug. Metab. Rev. 1:153-228.
Yu, S.J., and L.C. Terriere. 1971. Hormonal modification of microsomal
oxidase acticity in the house fly. Life Sci. 10:1173-1185.
147
-------
TECHNIQUES FOR THE WATERBORNE ADMINISTRATION OF
BENZO(A)PYRENE TO AQUATIC TEST ORGANISMS
by
Samuel P. Felton, W.T. Iwaoka, M.L. Landolt, and B.S. Miller
School of Fisheries, Fisheries Research Institute
University of Washington, Seattle, WA 98195
ABSTRACT
Benzo(a)pyrene (BaP) has been used in many experiments to
study its uptake by and effects on aquatic organisms. BaP has
been reported to be sparingly soluble in water (4 to 12 ug/*)«
However, even these low levels do not persist because of
adsorption to surfaces or to particulate matter in test systems.
Addition of BaP to water with small amounts of solubilizing
carriers (j_.e^. ethanol, methanol, benzene) does not appreciably
increase the amount in solution because BaP tends to precipitate
after contact with water.
Two techniques reported in this paper describe how to make
BaP more available to aquatic test organisms and increase its
water solubility. A technique has been developed in which the
compound is thinly coated onto sand particles to provide a large
surface area of BaP crystals in water contact. Although water
concentrations are very low, benthic test organisms are exposed
to large quantities of BaP while resting on or burrowing into
the sand.
Another technique has been developed in which BaP is made
water soluble by entrapping the compound in a water soluble
carrier. Bovine serum albumin (BSA) is used as a physical
carrier and concentrations of up to 200 ppb of BaP can be
effectively dissolved in water without precipitating any of the
BaP. Although the concentration of the BaP-BSA complex
decreases with time due to breakdown and adsorption onto
particulates or glass surfaces, this complex remains dissolved
in water longer than crystalline BaP. BaP levels can be
adequately maintained in the water by regular additions of the
BaP-BSA complex.
148
-------
INTRODUCTION
One of the more carcinogenic of the polycyclic aromatic hydrocarbons
(PAH) is 3, 4 benzo(a)pyrene (BaP) which occurs ubiquitously in the
environment (Andelman and Suess, 1970). BaP in the environment has been
associated to a large extent with man-controlled activities, such as coal
burning, open burning, or automobile exhausts, and to a lesser extent, with
synthesis by lower organisms. An important characteristic of BaP is its
extremely low solubility in water, ranging from 4 to 12 up to 53 parts per
billion (ppb), depending on the conditions (Davis et aU, 1942; Wolk and
Schwab, 1968).
BaP solubility in water can be increased by incorporating it into
detergent micelles or by introducing BaP into the water with fairly large
quantities of water-soluble organic solvents such as ethanol, acetone,
dioxane, or methanol (Suess, 1972a). The solubility of BaP in water can
also be increased by adding other compounds such as lactic acid, amino
acids, purine bases, and other PAHs (Bohon and Claussen, 1951; Andelman and
Suess, 1970).
Other investigators have used more complex organic molecules and
mixtures such as horse serum, albumin, or DNA to bind PAHs (Liquori et al.,
1962; Sahyun, 1964; Bothorel and Dismazes, 1974; Ceas, 1974). Serum
albumin's ability to complex small molecules is well-known and has been
extensively studied by pharmacologists and toxicologists. Only in recent
years has the technique been employed by those conducting bioassays
(Sanborn and Mai ins, 1977).
Another important characteristic of BaP is its ability to adsorb on
surfaces. BaP has been shown to concentrate onto activated carbon, calcite
material, silica glass, and plastics (Wolk and Schwab, 1968; Andelman and
Suess, 1971). The presence of minerals or suspended or settled
particulates in the water will greatly influence the distribution and
solubility of BaP.
The importance of BaP in the aqueous environment has led to a number
of laboratory experiments designed to study bioaccumulation, toxicity, or
changes in metabolic patterns in a variety of aquatic organisms (i.e.,
Clark and Diamond, 1971; Lee et a\_., 1972; Couch and Winstead, 1979).
However, many of the bioassays have shown variable results because of the
absence of standard exposure and analytical methods and also because of
difficulties involving solubilization, quantitation, and degradation of BaP
in water. The amount of BaP that can be added to pure or saltwater is also
a function of many factors such as temperature, salinity, and amount and
duration of mixing. Once in the water, BaP is subjected to photodegradation
or adsorption to particulate matter and this will considerably alter the
concentrations and distribution (Andelman and Suess, 1970).
Researchers have used organic carriers such as benzene, methanol,
acetone, or ethanol to help solubilize BaP (Suess, 1972a; 1972b).
149
-------
However, addition of BaP to water with small amounts of solubilizing
carrier does not necessarily increase the amount in solution because BaP
tends to precipitate out on water contact. Fairly large proportions or
organic carriers, such as were used in the study by Suess (1972a), are not
feasible for bioassays since even small amounts of solvent could affect the
metabolic patterns of the organism under study.
Our investigation was carried out to explore two possibilities of
making BaP more available to aquatic test organisms without organic
solvents.
MATERIALS AND METHOD
Coating Benzo(a)pyrene on Sand
BaP was coated onto sand particles (8 to 10 mesh) by the following
procedure. Silica sand (1000 g) was first washed with soap and water to
remove dirt and debris, and then rinsed three times with water and acetone.
Washed sand was spread evenly on a flat tray and dried in an oven (110° C)
overnight. BaP was prepared by weighing approximately 500 mg and
dissolving it in 150-200 m£ of methylene chloride in a 1-& Erlenmeyer
flask. The dried sand (cooled to room temperature) was then added to the
methylene chloride/BaP solution and stirred. The flask was placed in a
shallow water, and the methylene chloride was evaporated to dryness by a
nitrogen stream. Subsequently the flask was stoppered and placed in the
dark. BaP-coated sand was placed either in the bottom of the test aquarium
or in a piece of glass tubing incorporated in the water delivery system in
a high flow rate to maximize the opportunity for BaP to dissolve in the
water.
Analysis of BaP in Water
A measured volume of saltwater (usually 5 to 50 mi] was taken from the
aquarium with a volumetric pipette and filtered through a methylene
chloride washed, 0.45-y Gelman filter. The filtered water was then placed
in a 250-nu separatory funnel and extracted three times with one-half the
sample volume of methylene chloride. The organic solvent portions were
then combined, made up to a known volume in a volumetric flask (50 to 100
mi) t adequately mixed, and a portion of this solution was analyzed on a
Perkin-Elmer MPF-3 fluorescence spectrophotometer. The excitation wave
length was at 365 nm (10 nm silt), and the emission at 405 nm was used to
quantitate the amount of BaP in methylene chloride. Appropriate dilutions
of pure BaP dissolved in methylene chloride were used as standards and
analyzed under the same conditions.
Aldrich Chemical Company, Milwaukee, WI
150
-------
BaP Binding to Albumin
The procedure of Bothorel and Demazes (1974) was modified and used in
our studies to bind BaP to Bovine Serum Albumin (BSA): 2-chloroethanol
(75 nu) was slowly added to 50 nu of an aqueous solution containing 250 mg
BSA and allowed to equilibrate for 10 min. A solution of BaP in toluene
(9.7 mg/2.5 nu) spiked with 10 yCi of 14C BaP was then added dropwise
to the BSA-chloroethanol solution with vigorous stirring and this mixture
was immersed in a liquid nitrogen for 3-4 min.
The frozen solution (about 127 ma) was thawed and then placed on a
rotary evaporator to remove chloroethanol and water. The volume of this
solution was reduced to 25 mi and the contents were placed on a gel
filtration column (4.5 cm x 75 cm) packed with sephadex G-25, which was
thoroughly washed with distilled water prior to use.
The albumin-BaP complex was eluted at the void column separating it
from the unbound BaP. The BaP-BSA complex eluted from the gel filtration
column was lyophilized, reconstituted in 25 nu water, and relyophilized
three times. According to Bothorel and Desmazes (1974), repeated freezing
and drying restores the protein conformation. The BaP-BSA complex was
stored in the dark in a dry form before use in the experiments. BaP was
absorbed onto BSA, a peptide chain and whale myoglobin without the aid of
2-chloroethanol treatment by a simple absorption technique as described in
the BaP-BSA complex for the solubility study below.
Solubility Studies
The discrepancy in the literature between the various investigator
claims on seawater solubility of BaP necessitated a carefully controlled
and optimized solubiljty test. The first experiment was performed with
"Marine Environment," an artificial seawater previously adjusted to a
salinity of 1.022 and filtered through a 0.45-y filter and carbon. The
water contained no organics. It was then purged with ultra-pure nitrogen
for 15 min prior to addition of 20 yg of BaP. The flask was protected from
light and slowly stirred for 65 hr. The seawater was filtered through a
0.45 v filter and the flask allowed to dry. The BaP concentration in the
water was measured as previously discussed. Deposits on the flask wall
were extracted into methylene chloride and measured. The amount adhering
to the stirring bar was also measured, since Teflon has a high affinity for
BaP. Care was taken to achieve the best solubility of BaP under these
conditions.
The next set of experiments was done in beakers to test BaP solubility
under three different conditions. In the first experiment, 20 yg of BaP
Import Associates, Inc., San Francisco, CA 94116
151
-------
in 100 y£ ethanol was rapidly stirred into a liter of artificial seawater
and allowed to stir slowly for periodic measurements during seven days.
Additional injections of 20 \ii each were added during the seven-day period
to study photodegradation.
The purpose of the second experiment was to determine the absorption
capabilities and resultant solubility of BaP bound to BSA without the aid
of 2-chloroethanol. The BaP-albumin complex was made by adding 252 yg BaP
dissolved in 100 yi ethanol to a solution of 120 mg BSA in 30 mi water.
This stock solution was allowed to stir for 30 min and then added to H of
artificial seawater and stirred for seven days. The concentration of BaP
was measured in filtered water samples on the first, fourth, sixth, and
seventh days. The third experiment used sand coated by BaP as previously
described (40 g BaP-coated sand was placed in a 1-A beaker of artificial
seawater and allowed to stir at room temperature for seven days). Filtered
samples were measured periodically during the seven days.
The solubility of the BaP-BSA complex in seawater was also determined
by the method of Bothorel and Demazes (1974). A known concentration of the
complex was placed in an artificial seawater aquarium and circulated 9 days
by a submersible. The aquarium contained no fish. A filtered and unfil-
tered sample was assayed periodically for the 9-day period as previously
described.
UPTAKE OF BaP-BSA COMPLEX BY TEST FISH
Lyophilized BaP-BSA Uptake—Two adult English sole were forced-fed
capsules containing 6.0 mg dry lyophilized BaP-BSA complex labeled with
C-benzo(a)pyrene. This work was performed before the water-borne
study in order to ascertain whether the fish would assimilate the BaP after
proteolytic digestion of the BaP-BSA complex. Fish were^sacrificed, after
24 hr and selected tissue was solubilized with protosol. Aquasol was
then added to the solubilized tissue and the amount of radioactivity was
measured in a liquid scintillation counter.
Water-borne Uptake of BaP-BSA Complex from Seawater—Two 24-fc aquaria
containing artificial seawater were used for this experiment. The BaP-BSA
complex was added to one tank to give an initial concentration of 1.7
of BaP (2-4 mg/i BSA; this concentration of BaP also contained 1.07 x
10""5 yCi 1 C BaP/yg unlabeled BaP). The second tank contained only
BSA which had undergone a similar 2-chloroethanol treatment minus the BaP
A submerged circulating pump was placed in each tank and the system was
allowed to equilibrate overnight. Aliquots (50 mi) of water were then
removed, extracted with methylene chloride, and analyzed for BaP.
Three adult English sole were placed in the control and the
experimental tanks. After three days, the fish were sacrificed and the
New England Nuclear
152
-------
liver, kidney, brain, and portions of the skin and muscle tissues were
taken, solubilized with protosol, and the amount of radioactivity present
in these organs was measured by liquid scintillation counting.
RESULTS
Solubility Studies
The experiments with BaP-coated sand showed that very low levels
(about 1 ppb) were present in the water either in a sand column or in sand
placed at the bottom of the aquarium. In both cases, however, small
crystalline particles of BaP, which had sheared off from the sand, were
observed suspended in the water. These suspended particles were filtered
off before any concentration measurements were conducted.
Although levels of BaP in the water were very low, test fish resting
on or burrowing into the sand accumulated substantial amounts of BaP on
their integumental surfaces. Concentrations of BaP found adsorbed onto
integumental surfaces ranged from 14 ng to 4 yg. A discussion of the
adsorption of BaP on the skin and the uptake of BaP by flatfish tissue is
presented in another paper in this symposium (Landolt et ^K, 1978).
In the solubility study with crystalline material, BaP was dissolved
in ethanol and added to saltwater to obtain an initial concentration of
20 ppb. This level of BaP decreased to less than 1 ppb after 24 hr. Even
with repeated additions of 20 ppb BaP to the water, there was no increase
in the overall concentration, and levels in the water remained below 1 ppb
(Figure 1).
In the BaP-albumin solubility study, the complex dissolved completely
in saltwater, and an initial concentration of 16 ppb was obtained.
Analysis of the BaP concentration in the water after 2, 4, 6, and 7 days
showed that the concentration was 16, 10, 6, and 5 ppb, respectively
(Figure 1).
The complexing ability of BaP to unmodified BSA and 2-chloroethanol
treated BSA is shown in Table 1 and illustrates the difference between
absorption of BaP on BSA to that of a smaller protein such as whale
myoglobin and oxidized B-chain of insulin.
TABLE 1. COMPARISON OF THE COMPLEXING CAPACITY OF ALBUMIN FOR BaP BY
SIMPLE ABSORPTION AND BY 2-CHLOROETHANOL TREATMENT
Complexing capacity
Substance Treatment mg BaP/mg protein
Albumin + BaP 2-chloroethanol 3 x 10 .
Albumin + BaP Absorption 4 x 10
Insulin + BaP Absorption 3 x 10~
Myoglobin + BaP Absorption 8.2 x 10~
153
-------
Add20/xg
in Sw
•—•BaP
a—a BaP via sand
BaP-BSA
Figure 1. Solubility comparison of BaP in seawater via three
methods of introduction.
The stability of the BaP-BSA complex in saltwater (Figure 2) followed
a decay curve similar to that obtained by simple absorption and stirred in
a flask (Figure 1). However, the initial concentrations are considerably
different. The rate of decay of the solubilized BaP-BSA complex is much
slower that that of the unfiltered complex. This can be accounted for in
the circulating aquarium system by the action of the sand filter on the
unbound BaP or crystalline BaP.
Figure 3 diagrammatically summarizes the distribution of BaP injected
into a marine system without the aid of a solubilizer. The 10% soluble
fraction is a liberal figure based upon the solution of BaP in saltwater
154
-------
under optimum conditions. This figure illustrates what occurs when BaP
dissolved in an organic solvent such as ethanol is injected into
saltwater.
120-
100-
80-
c
V
J>
60-
40-
20-
Unfiltered BaP assay
D—o Filtered BaP assay
-I r
234 567 89 10
Days
Figure 2. Stability of BaP-BSA complex in a static aquarium system.
155
-------
Figure 3. Schematic distribution of BaP in an aquarium system using an
alcoholic concentrate of BaP.
Uptake of BaP-BSA Complex by Flatfish
The results of the forced feedings with encapsulated BaP-BSA were
quite dramatic. Within 24 hr after feeding the BaP was released from the
albumin binding and was found widely distributed in the various organs and
tissue. This clearly demonstrated that the fish digested the albumin, that
the released BaP could be absorbed by body tissues, and that the complex
could be used in uptake experiments.
A waterborne BaP-BSA experiment had an initial concentration of
1.7 yg/je, (assayed as BaP) added to the 98-* aquarium. Table 2 shows the
results of the uptake of -^-labeled BaP in three fish. High concent-
rations of BaP or its metabolites were found in the kidneys, liver, and
skin. Histopathological examination of the tissues from the fish were also
undertaken.
Tissues from the control fish were essentially normal, but tissues
from the experimental fish revealed a number of lesions. Liver sections
contained large numbers of hepatocytes with clear cytoplasmic vacuoles.
156
-------
These vacuoles were anatomically distinct from those associated with stored
lipid or glycogen and were most frequently clustered in peri pancreatic
zones. Vascular sinusoids were indistinct, and there were focal, although
much smaller, areas of necrosis, as in control specimens. The livers of
these fish were also heavily infested by a sporozoan parasite, Myxidium sp.
The kidney tissue contained multiple foci of pyknosis and necrosis within
hematopoietic tissue. The renal tubules and glomeruli were as expected.
The gills contained a limited number of hyperplastic lamellae in which
there was epithelial proliferation at the distal tips. Several aneurysms
and encysted dinoflagellate parasites belonging to the genus Oodinium were
also present. Focal dermatitis was of limited extent, and the heart
contained an encysted helminth. The other organs were unremarkable.
TABLE 2. UPTAKE OF 14C-LABELED BaP-BSA COMPLEX (1.7 yg/£ OF ASSAY
BaP) BY ENGLISH SOLE FOLLOWING 7-DAY WATER-BORNE EXPOSURE
Fish 1
Tissue
Liver
Brain
Muscle
Skin
Kidney
Fish 2
Tissue
Liver
Brain
Muscle
Skin
Kidney
Fish 3
Tissue
Liver
Brain
Muscle
Skin
Kidney
Wet wt .
used
0.1008
0.0497
0.1113
0.0923
0.1047
0.1025
0. 1013
0.1C25
0.1119
0.1012
0.1078
0.1034
0.0974
0.0943
0.1021
CPM
369
90
91
324
928
898
103
116
484
1526
969
91
80
425
790
uCi in
sample
2.3x10"'
4.3xlO~5
5.3xlO~5
2.2xlO~u
5.8x10"'-*
5. 6x10 "'*
6.2xlO~5
6.7xlO~3
3. 3x10""
9.5x10""
6.1xlO~u
5.5xlO"5
4.6xlO"s
2.9xlO"4
4.5xlO~u
»g BaP per
gm of wet wt .
2.16
0 81
0.44
2.24
5.16 "
5.10
0. 57
0.61
2.75
8.34
5.24
0.50
0.44
2.38
4.53
157
-------
DISCUSSION
Solubility
Our results show that even with a large quantity of BaP coated on
sand, the concentration of water-soluble BaP remained very low. Although
the' solubility of BaP in water has been reported to be from 4 to 12 ppb
(Davis et_ji]_., 1942), studies in our laboratory have shown that 16 to 20
yg/Jt of crystalline BaP can be dissolved in seawater with vigorous shaking.
However, the concentration in the water decreased to about 1 ppb after
several hours.
Even when BaP was added with an organic solvent such as ethanol,
levels of BaP in solution decreased to near zero after 24 hr. Repeated
additions of the same concentrations of BaP in ethanol into the water
showed no increase in BaP concentrations. Andelman and Suess (1971)
reported a loss of BaP from the water to boro silicate glass surfaces at a
rate of about 8 ng per cnr per 3-hr period. For Teflon and plexiglass,
losses were reported to be 38 and 50 ng per crrr/3 hr, respectively.
These results show that it is. very difficult to maintain any concent-
ration of BaP in solution for any length of time, particularly in systems
with glass, Teflon, or PVC surfaces.
On the other hand, the binding of BaP to large molecules, such as BSA
(Bothorel and Desmazes, 1974) or DNA (Liquori et. ^1_., 1962), greatly
increases stability and the solubility of BaP in water. Although the
concentration of the BSA-BaP complex in our test system decreased from 16
ppb to 5 ppb over a 7-day period (70% decrease), this latter concentration
is still five times higher than the concentration that can be obtained by
dissolving crystalline BaP in water.
This experiment also shows clearly that the BSA-BaP complex will
persist for a fairly long time as compared to the addition to water of
crystalline BaP.
The decrease in BaP concentration, due either to adsorption to
surfaces or to bacterial degradation of the protein, could be easily
increased to original levels by the dropwise addition of the BaP-BSA
complex dissolved in water. Concentrations of the BaP-BSA complex in water
can still be easily measured by fluorometric analysis since the BaP
adsorption onto proteins does not significantly alter the fluorescence
spectrum of BaP between 250 and 450 nm.
The technique of utilizing a carrier for a compound, such as BaP, and
maintaining the concentration of dissolved BaP in aqueous systems seems to
be an excellent method for conducting water-borne experiments with low
levels of hydrophobic compounds.
158
-------
Uptake Studies
Bap Coated on Sand—Although very little BaP dissolved in the
saltwater from the coated sand, this coating technique did provide a fairly
large source of crystalline BaP that can be taken up by the test animals,
especially benthic organisms.
Our results (Table 3) clearly indicate that test organisms accumulated
some of the BaP adsorbed onto the sand or sediment on their integumental
surfaces. Presumably, any small particles of crystalline BaP in the sand
will adhere to the mucous layer on the skin surface as the flatfish burrows
or buries itself in the sand. The amounts of BaP found in the methylene
chloride wastes vary from 15 ng up to 4.4 yg per fish. This large
variability in BaP level suggests that there is continual uptake and
release of the compound from the mucous surface.
TABLE 3. CONCENTRATIONS OF BaP FOUND IN METHYLENE CHLORIDE WASHES OF
INTEGUMENTAL SURFACES OF ENGLISH SOLE FOLLOWING 30-DAY
EXPOSURE TO BaP-COATED SAND
Fish
Number
Total
Weight of
Fish Analyzed(g)'
Total BaP
Found in Methylene Chloride
Washes (ng)1
6 (control)
7 (exptl)
8 (exptl)
9 (exptl)
10 (exptl)
11 (exptl)
12 (exptl)
13 (exptl)
14 (exptl)
20.1
10.2
5.6
31.3
46.3
6.9
11.2
4.9
69.6
N.D.2
285
414
230
395
993
4,450
15
1,695
^•Determined by gas chromatography analysis (Landolt et al.
2Not detectable
1978)
159
-------
The results from the analysis of the tissues of flatfish exposed to
BaP-coated sand indicate that there was BaP uptake by the organisms and
measurable amounts of the parent compound were present in the whole body
tissues after 30 days (Landolt et a\_., 1978). There has been no study to
show whether BaP or other hydrophobic compounds can be transported from the
external surfaces through the scale and skin and be either stored or
metabolized by marine flatfish. Since marine fish must drink seawater to
osmoregulate, particles of BaP could be introduced into the gastrointesti-
nal tract by this process and be absorbed. Another path of uptake of BaP
is through the gills. In a study of BaP uptake and metabolism by marine
fish, Lee et ll- (1972) found high levels of BaP accumulation in gill
tissues. This finding led to the postulation that the path of BaP uptake
would be through the gills followed by accumulation of the hydrocarbon and
its metabolites by the organs and tissues.
BSA-BaP Complex—These studies were conducted to determine if BaP
bound to BSA could be absorbed and metabolized. Other studies (Ceas, 1974;
Sanborn and Mai ins, 1977) have shown that certain hydrophobic compounds
bound to BSA or horse albumin can be absorbed and metabolized. Sanborn and
Malins (1971) have shown that larval spot shrimp (Pandalus platycergs) and
larval Dungeness crabs (Cancer magister) absorbed and metabolized C
naphthalene bound to BSA. Ceas (1974) found that a water-soluble horse
albumin-BaP complex adversely affected the development of sea urchin eggs.
Our data clearly show that radioactivity from the BaP was distributed
throughout selected tissues and organs, indicating that the test organisms
were able to absorb BSA-bound BaP.
The higher concentrations of radioactivity found in livers and kidneys
indicate that these organs are the more active of the tissues in
metabolizing and storing BaP. Unpublished results form our laboratory
(Landolt et aj_., 1978) show that livers of English sole exposed to BaP via
intraperitoneal injection rapidly accumulated BaP during the first day, but
no further increase in radioactivity was noted during the next seven days.
This is similar to the results obtained by Lee et _al_. (1972) who also noted
a steady-state level in livers of Gillichthys mirabilis exposed to
3H-labeled BaP.
These studies show that BaP can be administered to test organisms by
coating the sediment or by complexing the compound to a large water-soluble
molecule such as BSA. The coated sediment technique is an excellent method
for benthic test organisms since there is intimate contact between the
organism and the compound. On the other hand, the BaP-BSA complex method
allows increased water solubility of the test compound and provides a
higher exposure concentration to epibenthic and pelagic organisms.
160
-------
REFERENCES
Andelman, J.B., and M.J. Suess. 1970. Polycyclic aromatic hydrocarbons in
the water environment. Bull. WHO 43:479-508.
Andelman, J.B., and M.J. Suess. 1971. The photodecomposition of 3,4
benzpyrene sorbed on calcium carbonate. In: Organic compounds in the
aquatic environments. Marcel Dekker, New York. pp. 439-468.
Bohon, R.L., and W.F. Claussen. 1951. The solubility of aromatic
hydrocarbons in water. J. Am. Chem. Soc. 73:1571-1578.
Bothorel, P., and J.P. Desmazes. 1974. Trapping of benzo(a)pyrene by
bovine serum albumin. Biochem. Biophys. Acta 365:181-192.
Ceas, M.P. 1974. Effects of 3-4 benzopyrene on sea urchin egg development.
Acta Embryo Exp. 3:267-272.
Clark, H.G., and L. Diamond. 1971. Comparative studies on the
interactions of benzopyrene with cells derived from poikilothermic and
homeothermic vertebrates. II. Effect of temperature on benzopyrene
metabolism and cell multiplication: J. Cell Physiol. 77:385-392.
Couch, J., L. Courtney, J. Winstead, and S.S. Foss. 1979. The American
oyster (Crassostrea virginjca) as an indicator of carcinogens in the
aquatic environment. In: Animals as monitors of environmental
pollutants. National Academy of Sciences, Washington, DC.
pp. 65-84.
Davis, W.W., M.E. Krahl, and G.H.A. Clowes. 1942. Solubility of
carcinogenic and related hydrocarbons in water. J. Am. Chem. Soc.
64:108
Hose, Jo Ellen. 1979. Uptake and metabolism of benz(a)pyrene by adult
English sole and by early life history stages of flathead sole and
rainbow trout. M.S. Thesis, Univ. Washington, Seattle, WA.
Landolt, M.L., S.P. Felton, W.T. Iwaoka, and B.S. Miller. 1982.
Bioaccumulations and toxicity in English sole (Parophrys vetulus)
following waterborne exposure to benzo(a)pyrene. In: Symposium:
carcinogenic polycyclic aromatic hydrocarbons in the marine
environment. U.S. EPA, Cincinnati, OH. pp. 268-281.
Lee., R.F., R. Sauerheber, and G.H. Dobbs. 1972. Uptake, metabolism and
discharge of polycyclic, aromatic hydrocarbons by marine fish. Mar.
Biol. 17:201-208.
Liquori, A.M., B. DeLerma, F. Ascoli, C. Botre, and M. Trasciatti. 1962.
Interaction between DNA and polycyclic aromatic hydrocarbons. J. Mol.
Biol. 5:521-526.
161
-------
Sahyun, M.R.V. 1964. Enthalpy of binding aromatic amines to bovine serum
albumin. Nature. 203:1045-1046.
Sanborn, H.R., and D.C. Malins. 1977. Toxicity and metabolism of
naphthalene, a study with marine larval invertebrates. Proc. Soc.
Exp. Biol. Med. 154:151-155.
Suess, M.J. 1972. Aqueous solutions of 3,4 benzpyrene. Water Res.
6:981-985.
Suess, M.J. 1972. Laboratory experimentation with 3,4 benzpyrene in
aqueous systems and the environmental consequences. Zbl. Bakt. Hyg.
155:541-546.
Wolk, M., and H. Schwab. 1968. Zum transportphanomen und wirkungs-
mechanismus des 3,4-benzpyrens in der zelle. Z. Naturforsch.
Z3B:431.
162
-------
ACTIVATION AND UPTAKE OF POLYNUCLEAR AROMATIC HYDROCARBONS
BY THE MARINE CILIATE, PARAURONEMA ACUTUM
by
Donald G. Lindmark
The Rockefeller University
New York, NY
ABSTRACT
The marine ciliate, Parauronema acuturn, can convert
2-aminofluorene and 2-acetylaminofluorene to compounds with
mutagenic activity in the Ames/Salmonella test. The ciliate,
however, does not activate benzo(a)pyrene or benzanthracene or
destroy the mutagenic properties of nitrosoguanidine. Homog-
enates when substituted for a liver microsome fraction (S-9)
in the Salmonella/microsome test activate 2-aminofluorene and
2-acetylaminofluorene to mutagens. Benzo(a)pyrene (BaP) and
benzanthracene are not activated nor is nitrosoguanidine
inactivated. Phenobarbitol does not induce or increase the
amount of activating activity. The activating activity shows
no requirement for the NADPH regenerating system required by
liver microsomes and hence may not be due to a typical mixed
function oxidase. Upon differential sedimentation of cell
homogenates, the majority of the activity sediments along with
a small particulate fraction that has sedimentation properties
of microsomes parallel to those of higher eukaryotes. BaP
though not metabolized is accumulated by cultures of P_. acutum
at a linear rate but not released after removal of BaP from the
incubation medium to a great degree (10%). Hence, this ciliate
can convert certain polynuclear aromatic hydrocarbons to
mutagens and accumulate others, such as BaP, which it cannot
metabolize. The importance of these facts is dependent on the
population of this ciliate in marine environments and the likeli-
hood of contact between ciliate and environmental contaminant.
INTRODUCTION
The importance of polynuclear aromatic hydrocarbons as contaminants of
the marine environment has become more evident in recent years. The
effects of marine organisms on these compounds has only recently become an
*
Present address: The Department of Preventive Medicine
Cornell University, Ithaca NY 14850
163
-------
area of investigation. The abilities of higher eukaryotes to alter and
accumulate compounds, such as benzo(a)pyrene and aminofluorene, have
been studied in some detail. No study on the alteration of compounds or
their uptake by marine single cell eukaryotes has been udertaken. Their
importance, possibly because of their high population in marine
environments and their primary position in marine food webs, cannot be
underestimated. We have been able in this study to investigate the uptake
and alteration of polynuclear aromatic hydrocarbons which may be at
significant levels in the marine environment.
MATERIAL AND METHODS
Organism and growth conditions—The marine ciliate, Parauronema
acutum, was obtained from Dr. A.T. Soldo, Veterans Administration Hospital,
Miami, FL. The ciliate was grown in the dark at 24° C in seawater M medium
(Soldo and Merlin, 1972). Cells (72-hr-old) for experimentation were
collected by centrifugation at room temperature (1000 rpm for 4 min in the
HNS-II centrifuge, Daman/IEC) and either washed (2 times) and resuspended
in 0.25 M sucrose (cell fractionation studies).
Mutagenesis—The mutagenic activity of compounds [benzo(a)pyrene,
2-aminofluorene, 2-acetylaminofluorene, nitrosoguanidine, benzanthracene,
and the solvent in which the compounds were dissolved—dimethylsulfoxide]
and the ability of growing cells and cell homogenates to convert these
compounds to mutagens or eliminate their mutagenic properties was as
described by Ames et aj_. (1975). Histidine revertants as a measure of
mutagenicity were enumerated using Salmonella typhimurium tester strains TA
98, TA 100, TA 1337, TA 1535, TA 1537, and TA 1538 obtained from Dr. Bruce
N. Ames, Biochemistry Dept., University of California, Berkeley, CA.
Strains Ta 100, TA 1535 are used to detect mutagens causing base-pair
substitutions and TA 98, TA 1538, TA 1537 for the detection of various
kinds of frameshift mutations (Ames£t^l_., 1973a).
Conversion of compounds by growing cells—The compounds (listed above)
were added to 48-hr cultures of £. acutum at a final concentration of 500
ng/nu. After an additional 24-hr growth, the cells were removed by
centrifugation and the culture medium tested for mutagenic activity in the
Ames test (Ames et a\_., 1973a). Control cultures had the tested compound
added after 72-hr growth, just prior to removal of cells by centrifugat-
ion.
Conversions of compounds by cell homogenates and fractions—Homogenates
and fractions obtained by differential centrifugation of homogenates were
tested for their ability to convert compounds into mutagens by incorporat-
ion into Salmonella/microsome test of Ames (Ames eit jj]_., 1973a). The
homogenates or fractions were substituted for the S-9 fraction used in the
Salmonel1a/microsome test as a source of activating (microsomal) enzymes
(Ames et"a"l., 1975; Durston and Ames, 1974). The homogenates or fractions
were incorporated into the top agar with the compound to be tested and the
Salmonella tester strain (Ames et al_., 1973a, 1975). Revertants were
enumerated after 2 days incubation at 37° C.
164
-------
Enzyme assays—Mai ate dehydrogenase activity was measured by the
oxidation of NADH (No. O.D. 340 nm) with oxalacetate as a substrate
(Lindmark and Muller, 1974). Acid phosphatase and g-N-acetylglucosaminidase
were measured by release of p-nitrophenol (No. O.D. 410 nm) (Lindmark and
Muller, 1974). Pyruvate kinase was measured by phenylhydrazone production
from phosphoenol-pyruvate (Prichard and Scofield, 1968). NADPH-cytochrome c
reductase was measured by reduction of cytochrome c (Masters j2t jH_., 1967)
Activation of 2-aminofluorene or 2-acetylaminofluorene was measured by
enumerating the number of histidine revertants produced in tester strain TA
1538. The 2-aminofluorene (10 yg) or 2-acetylaminofluorene (10 ug) were
incorporated with the enzyme sample into the top agar with the tester
strain as described by Ames (1973a, b; and Durston and Ames, 1974). Results
presented were obtained from experiments containing optimal concentrations
of enzyme and test compound. Catalase was measured by disappearance of
H202 (Muller, 1973). Fumarase is measured by No. O.D. 240 with malate as a
substrate (Hill and Bradshaw, 1969).
Gel 1 fractionation—Homogenates of £. acutum were prepared by five
successive passages of cells suspended in cold 250 mM sucrose through a
20 ym average pore size stainless steel filter. Differential centrifuga-
tion of the homogenate in a refrigerated centrifuge (SS-34 rotor) (Dupont-
Sorvall, Norwalk, CT) resulted in four fractions: 3 particle fractions
sedimenting at 500 rpm for 4 min, 2500 rpm for 10 min, 19,000 rpm for 60
min, and a nonsedimentable fraction not sedimenting at 19,000 rpm. This
sedimentation procedure resulted in the separation of particles sedimenting
at 2500 rpm containing mitochondrial (malate dehydrogenase, fumarase) and
peroxisomal (catalase) enzymes from those particles sedimenting at 19,000
rpm containing hydrolytic (acid phosphatase, e-N-acetyl glucosaminidase)
and microsomal (NADPH-cytochrome C reductase) enzymese. All particle
associated enzymes are separated from the non-sedimentable cytoplasm
(pyruvate kinase).
Uptake and release of benzo(a)pyrene—The 72-hr cells were harvested
by centrifugation and resuspended in half the original volume of artificial
seawater. C benzo(a)pyrene (0.04 yCi/nu) final concentration was
added, and the sample shaken at 24° C at 200 rpm in a shaking water bath
(Aquatherm, New Brunswick Scientific, New Brunswick, NJ) by centrifugation
and washed twice in seawater. The pellet was resuspended in 1.2 mt of 0.1
NaOH containing 0.4% deoxycholate. A 100 vX, sample was added to 10 mi
Bray's solution and radioactivity determined by scintillation counting.
After 4 hr in the presence of C benzo(a)pyrene, the remaining
cells were washed free of benzo(a)pyrene (2 times) and resuspended in the
original volume.of seawater. Aliquots were taken at hourly intervals, and
the remaining C label in the cells was determined as described above.
Chemicals--Artificial seawater was obtained from Aquarium Systems,
Eastlake, OH. Benzo(a)pyrene and 9,10-dimetyl-l,2-benzanthracene were
obtained from Sigma Chemicals, St. Louis, MO. [1,10- C] Benzo(a)pyrene
165
-------
was obtained from Amersham, Arlington Heights, IL. N-methyl-N'-nitro-n-
nitrosoguanidine, 2-aminofluorene, and 2-acetylaminofluorene were obtained
from Aldrich Chemicals, Milwaukee, WI.
RESULTS
Conversions of compounds by growing cultures—As shown in Table 1,
growing cultures of P_. acutum can convert both 2-aminofluorene and
2-acetylaminofluorene to mutagenic compounds. Benzo(a)pyrene and
benzanthracene are not converted nor is nitrosoguanidine converted into a
non-mutagen. DimethylsuIfoxide (solvent for all compounds) was not
converted into a mutagen. Tester strains TA 1537, 1538, and 98 are
reverted, whereas TA 100, 1535 are not. This specific reversion suggests
the occurrence of frameshift mutations for which these tester strains show
a high sensitivity (Ames et _§]_., 1973a; Durston and Ames, 1974).
Conversions of compounds by homogenates--If homogenates of £. acutum
are incorporated with teest compound and tester strain TA 1538 into the
Salmonella/microsome test, the results obtained are found in Table 2.
2-aminofluorene and 2-acetylaminofluorene are converted to mutagens. As
with growing cells, benzo(a)pyrene and benzanthracene are not converted to
mutagens nor is nitrosoguanidine converted into a non-mutagen. Tester
strains TA 98, 1537 give similar results. TA 100 and 1535 exhibited no
reversion above control. Homogenates prepared from cells grown in the
presence of compounds (phenobarbitol, 1 mg/m£; 2-aminofluorene or
2-acetylaminofluorene, 500 ng/nu) known to induce and increase the cellular
level of microsomal enzymes show no increase in activating activity and
suggest the absence of induction (Ames et jH_., 1973a). The rate of
activation on a per mg protein is 20 times that of Arochlor-induced S-9
fraction obtained from rat liver.
In order for full activity of the mammalian microsomal activating
system of £. acutum, a NADP regeneration system is required. As shown on
Table 3, the activating system of P_. acutum does not have a requirement for
the regeneration system. NADP, glucose-S-PO/j, Mg^ are not required,
suggesting the absence of a typical mixed function oxidase found in higher
eukaryotes.
Differential sedimentation of homogenates—As shown in Table 4,
aminofluorene activating activity shows the highest amount of activity in a
small particle fraction, though 40% of the total activity is in the
non-sedimentable cytoplasm. Further experimentation must be done to define
the particle population with which the activating activity is associated,
although there is a clear separation from the large particle fraction,
sedimenting at a low speed, containing mitochondria! and peroxisomal
enzymes.
Uptake and release of benzo(a)pyrene—Since growing cells and cell
homogenates could not convert benzo(a)pyrene or benzanthracene into a
mutagen, it became of interest to investigate the uptake and release of
benzo(a)pyrene. £. acutum accumulates the label in benzo(a)pyrene at a
166
-------
TABLE 1. CONVERSIONS OF COMPOUNDS BY GROWING CELLS OF P. ACUTUM*
Histidine revertants per plate
Compound
2-aminofluorene
control
2-acety 1 ami hofl uorene
control
benzo(a)pyrene
control
benzanthracene
control
nitrosoguanidine
control
dimethyl sulfoxide
control
TA98
1250
49
1000
31
24
20
36
30
N.D.
N.D.
18
19
TA100
199
152
150
160
140
160
150
110
1350
1200
140
162
TA1535
32
36
30
25
23
18
24
27
890
810
29
36
TA1537
1220
21
1100
14
18
19
20
25
N.D.
N.D.
16
15
TA1538
1095
24
1300
34
22
31
40
35
N.D.
N.D.
30
28
Cells grown for 48 hr, compound added (500 ng/m£), cultures grown addit-
ional 24 hr, harvested, and supernatant solution tested for converted
compound. Control contains an equivalent amount of compound added to
culture after 72 hr growth in the absence of tested compound. N.D.=not
determined.
167
-------
TABLE 2. CONVERSION OF COMPOUNDS BY HOMOGENATES OF P_. ACUTUM
Compound* yg Histidine revertants (TA1538?*per plate
2-anrinofluorene 10 1240
control 10 25
2-acetylaminofluorene 10 1000
control 10 18
benzo(
-------
TABLE 4. DISTRIBUTION OF ENZYME ACTIVITIES AFTER DIFFERENTIAL SEDIMENTATION OF AN HOMOGENATE OF P. ACUTUM
Fraction Catalase
Nuclear 0.5
Large particle 2. 1
Small particle 1.8
Non-sedimentable 1.6
Acid
phosphate
0.2
1.0
3.5
0.3
R.S.A.*
BNAG Fumarase Ma late
dehydrogenase
0.1 0.8 0.7
0.4 x 3.2 3.6
3.0 0.8 0.6
0.8 0.7 0.5
Pyruvate Aminofluo
kinase activati
0.1 0.2
0.3 1.0
0.3 2.1
2.1 1.0
*Percent enzyme/percent protein
-------
linear rate for a period of 4 hr. After removing the benzopyrene from the
environment, the cells lose 10% of the benzo(a)pyrene taken up in the first
hour, but no more up to 48 hr. These results suggest that JP. acutum can
accumulate compounds, such as benzo(a)pyrene, which it cannot metabolize.
DISCUSSION
The marine ciliate, £. acutum, can convert ami no polynuclear aromatic
hydrocarbons, such as 2-aminofluorene and 2-acetylaminofluorene, to
mutagens at a relatively high rate. Other compounds, such as
benzo(a)pyrene and benzanthracene, are not altered. The activating
activity does not require Mg , glucose-6-P04 or NADP, suggesting
that it is not due to a typical mixed function oxidase. The activity
cannot be induced or increased by growing P. acutum in the presence of
phenobarbitol, 2-aminofluorene, or 2-acetylaminofluorene.
The enzyme activity responsible for these processes may be due to a
specific enzyme or, as suggested by the fact the £. acutum can use ami no
acids as its sole carbon source (Soldo and Merlin, 1972), may be due to
enzymes that perform other physiological functions. The activity is
associated with a small particle fraction which corresponds to the
sedimentation properties of microsomes as isolated from other organisms.
This activity can certainly add to the mutagen load in the marine
environment if this ciliate occurs in a high population in an area in
contact with aminofluorenes.
P_. acutum can accumulate benzo(a)pyrene and possibly other compounds,
but it cannot metabolize; this may be of importance in the marine
environment proportional to the ciliate population and the amount of
polynuclear aromatic hydrocabons.
ACKNOWLEDGEMENTS
The author thanks Nancy Dick for her excellent technical assistance;
Miklos Muller and Norman Richards for constructive criticism; Karrie
Polowetzky, Sandra Buck, and Martha Morse for manuscript preparation. This
work was supported by Grant R805364010 from the Environmental Protection
Agency.
170
-------
REFERENCES
Ames, B.N., W.E. Durston, E. Yamasaki, and F.D. Lee. 1973a. Carcinogens
are mutagens: a simple test system combining liver homogenates for
activation and bacteria for detection. Proc. Natl. Acad. Sci.
70:2281-2285.
Ames, B.N., F.D. Lee, and W.E. Durston. 1973b. An improved bacterial test
system for the detection and classification of mutagens and carcino-
gens. Proc. Natl. Sci. 70:782-786.
Ames, B.M., J. McCann, and E. Yamaski. 1975. Methods for detecting
carcinogens and mutagens with the Salmonella/mammalian-microsome
mutagenicty test. Mutat. Res. 31:347-364.
Durston, W.E., and B.N. Ames. 1974. A simple method for the detection of
mutagens in urine: studies with the carcinogen 2-acetyl aminofluor-
ene. Proc. Natl. Acad. Sci. 71:737-741.
Hill, R.L., and R.A. Bradshaw. 1969. Fumarase. In: Methods in
enzymology XIII. J.M. Lowenstein, Ed., Academic Press, New York. pp.
91-99.
Lindmark, D.G., and M. Muller. 1974. Biochemical cytology of trichomonad
flagellates. II. Subcellular distribution of oxidoreductases and
hydrolases in Monocercomonas sp. J. Protozool. 21:374-378.
Master, B.S.S., C.H. Williams, and H. Kamin. 1967. The preparation and
properties of microsomal TPNH-cytochrome c reductase from pig liver.
In: Methods in enzymology X. R.W. Estabrook, Ed., Academic Press,
New York. pp. 565-573.
Muller, M. 1973. Biochemical cytology of Trichomonad flagellates. I.
Subcellular localization of hydrolases, dehydrogenases, and catalase
in Tritrichomonas foetus. J. Cell. Biol. 57:453-474.
Prichard, R., and P. Schofield. 1968. The metabolism of phosphoenol
pyruvate and pyruvate in the adult liver fluke Fasciola hepatica.
Biochem. Biophys. Acta 170:63-76.
Soldo, A.T., and E.J. Merlin. 1972. The cultivation of symbiote-free
marine ciliates in axenic medium. J. Protozool. 19:519-524.
171
-------
EFFECT OF POLYNUCLEAR AROMATIC HYDROCARBONS AND POLYHALOGENATED BIPHENYLS
ON HEPATIC MIXED-FUNCTION OXIDASE ACTIVITY IN MARINE FISH
by
Margaret 0. James*
C. V. Whitney Marine Laboratory, University of Florida
St. Augustine, FL 32084
and
John R. Bend
Laboratory of Pharmacology
National Institute of Environmental Health Sciences
Research Triangle Park, NC 27709
ABSTRACT
The polynuclear aromatic hydrocarbons, 3-methylcholanthrene
and dibenz(a,h)anthracene, induced hepatic microsomal benzo(a)-
pyrene hydroxylase and 7-ethoxycoumarin 0-deethylase activities
in several marine fish, including sheepshead, Archosargus
probatocephalus, little skate, Raja erinacea, southern flounder,
;n\
Paralichthyes lethostigma, and stingray, Dasyatis sabina. The
cytochrome P-450 content of hepatic microsomes from fish treated
with dibenz(a,h)anthracene or 3-methylcholanthrene was usually
not significantly higher than in control fish. 3-Methylchol-
anthrene and dibenz(a,h)anthracene had no effect on hepatic
benzphetamine N-demethylase activity nor on epoxide hydrase or
gluthathione S-transferase activity in these fish. The dose-
response and time course of benzo(a)pyrene hydroxylase
induction was studied in winter and in summer in sheepshead;
the rate of induction was faster in summer. Commercial
mixtures of polychlorinated biphenyls and polybrominated
biphenyls also induced benzo(a)pyrene hydroxylase and
7-ethoxycoumarin 0-deethylase activities and in most experi-
ments caused a statistically significant increase in cytochrome
P-450 content of hepatic microsomes. High individual variation
between fish made it difficult to demonstrate significant
differences with small increases or decreases in enzyme
activities (2-fold less than control). These studies suggest
that the control of cytochrome P-448-dependent mixed-function
oxidation in marine fish is similar to that in mammals.
However, the time course of induction is slower in fish,
especially those acclimated to cold water.
Present address: College of Pharmacy, University of Florida
Gainesville, FL 32610
172
-------
INTRODUCTION
As a result of oil spills and industrial pollution, marine species in
coastal and estuarine environments are frequently exposed to polycyclic
aromatic hydrocarbons (PAHs) and other chemicals, including polychlorinated
biphenyls (PCBs) and even to polybrominated biphenyls (PBBs) in some areas.
These chemicals induce their own metabolism in mammalian species (Conney,
1967; Dent et_ a]_., 1976; Litterst et^ a]_., 1972) and many are carcinogenic
or mutagenic to mammals (Heidelberger, 1975). The effects of PAHs, PCBs,
and PBBs on some hepatic xenobiotic-metabolizing enzymes in the rat are
summarized in Table 1. In this paper we describe the effect of known doses
of selected PAHs, and mixtures of PCBs or PBBs on hepatic
xenobiotic-metabolizing enzymes in marine fish common to the Atlantic Ocean
near Maine (studies carried out at the Mount Desert Island Biological
Laboratory, Salsbury Cove, ME) or Northeast Florida (studies carried out at
the C. V. Whitney Laboratory for Experimental Marine Biology and Medicine
of the University of Florida, St. Augustine, FL). In addition we report
the effects of a pure PCB isomer, 3,3'4,4',5,5'-hexachlorobiphenyl
(3,3',4,4',5,5'-HCB), which is known to cause cytochrome P-448-dependent
induction in mammals (Goldstein et al., 1977) and of a pure PBB isomer,
2,2' ^^'S.S'-hexabromobiphenyl 72",2"r,4,4' ,5,5'-HBB), which causes
cytochrome P-450-dependent induction in rats (Moore ejt al_., 1978).
MATERIALS AND METHODS
Animals—At both laboratories fish were caught locally and maintained
in flowing seawater for the duration of each experiment. Species studied
in Maine were the little skate, Raja erinacea, and the winter flounder,
Pseudopleuronectes americanus. Species studied in Florida were the
Atlantic stingray, DasyatisTabina, the sheepshead, Archosargus
probatocephalus, and the southern flounder, Paralichthyes lethostigma. In
most experiments, five fish were injected intraperitoneally (i.p.) with a
suspension of solution of the chemical under investigation and three fish
were injected with an equivalent volume of the vehicle. In a few
experiments the chemical was administered orally or by intramuscular (i.m.)
injection. The day on which the compound was administered was designated
"Day 1" of the experiment and time (days) to sacrifice varied with the
compound, dosage, and time of year, as indicated in Results. All Florida
fish were fed regularly throughout the experiments. At the designated time
after dosage, animals were sacrificed and livers removed promptly and
placed in ice-cold 1.15% KC1. Washed microsomes and cytosol fractions were
prepared as previously described (James ^t^l_., 1979; Pohl e_t jj]_., 1974).
Chemicals—3-Methylcholanthrene (3-MC), 7,12-dimethylbenzanthracene
(DMBA), 1,2,3,4-dibenzanthracene (DBA), and hexadecane were obtained from
Sigma Chemical Company. Aroclor 1254, Firemaster FF1, 1,3,3'4,4'5,5'-HCB,
and Z.Z'AjVS.S'-HBB were generously supplied by Dr. J. McKinney,
Laboratory of Chemistry, National Institute of Environmental Health
Sciences.
173
-------
TABLE 1. EFFECT OF CHEMICAL PR.ETREATMENT ON HEPATIC XENOBIOTICMETABOLISM IN THE _RAT_
CHEMICAL
Polynuclear Aromatic Hydrocarbons
Polychlorinated and Polybromlnated
Biphenyls
3,3',4,4'-Tetrachlorobipheny1
3,3',4,4',5,5'-Hexachlorobipheny]
2,2',4,4'-Tetrachlorobiphenyl
2,2',4,4',5,5'-Hexachlorobiphenyl
2,2',4,4',6,61-Hexachlorobiphenyl
MIXED-FUNCTION OXIDATION
Induce cytochrome P-448-dependent but not
cytochronte P-450-dependent oxidations
(Conney, 1967; Ryan et ^1_., 1977)
Induce cytochrome P-448 and cytochrome P-450
dependent oxidations (Litterst et _aK, 1972;
Dent et al_., 1976, Ryan e_t aj., 1977)
Induce cytochrome P-448 dependent oxidations
(Goldstein et al.., 1977)
Induce cytochrome P-450 dependent oxidations
(Goldstein et al_., 1977; Moore et^ aJL, 1978)
EPOXIDE METABOLISM
Induce epoxide hydrase and
GSH S-transferase activities
(Oesch et a_l., 1973; Mukhtar
and Bresnick; 1976; Bresnick
et al., 1977; Hales and Neims,
17777
PCBs and PBBs induce epoxide
hydrase (Dent et^l., 1976)
and GSH S-transferase activity
Induce GSH S-transferase
activity (Kohli et^ al_. ,1978)
Induce eposide hydrase and GSH
S-transferase activities (Kohli
et al., 1978)
-------
Assay—Mixed-function oxidase activities measured in hepatic
microsomes were benzo(a)pyrene hydroxylase (AHH), benzphetamine
N-demethylase (BND), 7-ethoxycoumarin 0-deethylase (7-EC), and
7-ethoxyresourfin 0-deethylase (ERF) (James et_ al_., 1979; Pohl et al.,
1974). Cytochrome P-450 content and NADPH-cytochrome £ reductase activity
were quantitated as described previously (James et_ aj_., 1979).
Epoxide-metabolizing enzyme activities measured were microsomal epoxide
hydrase (EH), and cytosol fraction glutathione S-transferase (GSH-T); these
activities were assayed as described by James and coworkers (1979).
Epoxide substrates used were styrene 7,8-oxide, octene 1,2-oxide, and
benzo(a)pyrene 4,5-oxide, although in most experiments styrene oxide was
the substrate. In all cases, assay conditions used were those that gave
maximum in vitro activity in the species under investigation.
RESULTS
Hydrocarbons
Effect on microsomal mixed-function oxidation—Admini strati on of 3-MC
to sheepshead, flounder, stingray, and skate resulted in induction of
hepatic microsomal AHH and 7-EC activities. Table 2 shows results from
four experiments with 3-MC. In other experiments with sheepshead and
little skate, ERF activities were also elevated by 3-MC or DBA treatment.
Mean cytochrome P-450 content of hepatic microsomes from 3-MC-treated fish
was usually somewhat higher than control-values, but the difference was
seldom statistically significant due to wide variation between individual
fish. Hepatic microsomes from only one species, the little skate, had
elevated BND activities after 3-MC treatment. In other species, there was
no difference in BND activities between treated and control fish.
Treatment of sheepshead with DMBA (2x10 mg/kg, i.p. on Days 1 and 3)
caused induction of AHH and 7-EC activities by Day 7, but did not affect
cytochrome P-450 content or BND activity. Similarly, little skates treated
with DBA (10 mg/kg, i.p. on Day 1) had induced AHH and 7-EC activities by
Day 13, but cytochrome P-450 content was unaffected. However, cytochrome
P-448, as well as cytochrome P-450, were isolated from hepatic microsomes
of DBA-treated little skates (Elmamlouk £t al_., 1977).
AHH activity in hepatic microsomes fram control fishwas stimulated by
in vitro addition of 7,8-benzoflavone (10- M)> whereas AHH activity in
hepatic microsomes from 3-MC- or DBA-treated fish was inhibited by in vitro
addition of 7,8-benzoflavone (10~4 M) (Table 3), as previously shown
for rats (Wiebel and Gelboin, 1975).
Aliphatic hydrocarbons are an important quantitative constituent of
crude oils. Consequently, we tested the effect of hexadecane admini-
stration on xenobiotic-metabolizing enzymes. Treatment of sheepshead with
hexadecane (2 x 20 mg/kg, i.p. on Days 1 and 3) did not affect their
hepatic microsomal mixed-function oxidase activities, when assayed on Day 8.
175
-------
TABLE 2. EFFECT OF 3-METHYLCHOLANTHRENE ADMINISTRATION ON HEPATIC MICROSOMAL ENZYMES IN SOME MARINE FISH
o>
SPECIES
Treatment
Sheepshead ^ e
Archosargus probatocephalus '
Corn oil (3)
10 mg/kg, i.p. Day 1
sacrificed Day 8 (4)
Stingray d _
Dasyatis sabina >a
Emulphor (4)
2 x 15 mg/kg, i.m. Days 1 and 5
sacrificed Day 14 (4)
Southern flounder d
Paralichthyes lethostigma '
Corn oil (4)
2 x 10 mg/kg, i.p. Days 1 and 4
sacrificed Day 8 (4)
Little skate • •
Raja erinacea >J
Dimethyl sulfoxide (3)
2 x 50 mg/kg, p.o. Days 1 and 3,
sacrificed Day 11 (3)
CYTOCHROME P-4503 BENZO(a)PYREKED
CONTENT HYDROXYLASE
0.35 1 0.05f 5.3 ± 1.1
0.51 ± 0.21 23.9 1 4.8
0.48 1 0.05 0.59 1 0.12
0.58 1 0.08 6.12 1 1.43
0.11 1 0.05 0.18 1 0.06
0.18 1 0.10 2.74 + 2.07
not assayed 0.84 + 0.28
not assayed 6.50 + 0.24
7-ETHOXYCOUMARINc BENZPHETAMINEC
0-DEETHYLASE N-DEMETHYLASE
0.40 1 0.12 1.15 1 0.25
1.10 1 0.25 1.25 1 0.40
0.08 1 0.08 0.32 + 0.17
4.46 1 0.12 0.30 ± 0.02
not detected not assayed.
0.12 1 0.01 not assayed
0.36 1 0.16 0.96 1 0.16
0.56 + 0.16 1.99 1 0.20
aNmole/mg protein.
Fluorescence um'ts/min/mg protein.
cNmole product/min/mg protein.
dFlorida species.
63-MC was suspended in corn oil at 20 ing/mfc. Controls received an equivalent volume of corn oil.
fMean 1 S.D.
93-MC was suspended in Emulphor:acetone:water (2:2:6, by volume) at 15 mg/mi. Controls received an equivalent
volume of vehicle.
^Less than 0.001 nmol/min/mg protein.
'Maine species.
J-3-MC was dissolved in dimethyl sulfoxide at 50 mg/nt. Controls received an equivalent volume of dimethyl
sulfoxide.
-------
TABLE 3. EFFECT OF Hi VITRO ADDITION OF 7,8-BENZOFLAVONE ON HEPATIC MICROSOMAL AHH ACTIVITY IN SHEEPSHEAD
AND LITTLE'SKATE
7,8-BENZOFLAVONE ADDED
(H)
0
10'7
10'6
ID'5
lO'4
10'3
BENZO(a)PYRENE HYOROXYLASE
CONTROL SKATE DBA SKATE3
0.23C
0.23
0.24
0.21
0.73
0.64
5.23
5.08
3.88
3.30
1.16
0.79
ACTIVITY (F.U./MIN/MG PROTEIN)
CONTROL SHEEPSHEAD 3-MC SHEEPSHEADb
1.37C
1.68
1.72
2.66
5.05
6.05
22.6
22.1
24.1
19.8
8.6
8.3
aSkates were injected i.p. with DBA (10 rag/kg) on Days 1, 2, and 3 and killed on Day 10.
bSheepshead were injected i.p. with 3-MC (20 mg/kg) on Day 1 and killed on Day 9.
cResults shown are from a single experiment, which was repeated twice with similar results.
-------
Having obtained significant increases in hepatic AHH activity with all
species tested following polycyclic hydrocarbon treatment, we further
characterized 3-MC induction in sheepshead. The dose-response of hepatic
AHH induction in sheepshead (sacrificed on Day 8), after various i.p.
doses of 3-MC on Day 1, is shown in Figure" 1. Doses as low as 1 rag/kg
caused hepatic AHH activity to double, but this activity was virtually
unaffected (neither stimulated nor inhibited) by in vitro addition of
7,8-benzpflayone. At all higher doses, hepatic AHH activity of treated
fish was inhibited by in vitro addition of 7,8-benzoflavone. 3-MC doses
above 10 mg/kg did not increase the extent of induction at the time the
experiment was performed (December and January). Higher treated/control
AHH activity ratios were obtained with groups of sheepshead dosed in March
(2 x 20 mg/kg) or June, July, and August (10 mg/kg), although the actual
activities (F.U./min/mg of protein) in hepatic microsomes from control fish
were similar in summer (4.5 +_ 3.2, mean _+ S.D., n=20) and winter (4.2 _+
1.5, n=19). Control values varied up to 5-fold between different groups o1
fish, although AHH activity in almost all of the control fish was
stimulated by in vitro addition of 7,8-benzoflavone.
10 15 20
3-MC, mg/kg
Figure 1. Dose response of AHH induction in sheepshead hepatic micro-
somes.
178
-------
SUMMER
10 15 20
Days After Dose
25
30
8
3
14
WINTER
10 15 20
Days After Dose
25
30
Figure 2. Effect of season on the time course of induction of hepatic
microsomal AHH activity in sheepshead.
179
-------
Since the ocean temperature in St. Augustine, FL, varies from 11° in
January to 28° in August, we studied the time course of induction at both
seasons. Mixed-function oxidase activities were assayed in hepatic
microsomes from groups of sheepshead sacrificed at different times after
single doses of 20 mg/kg 3-MC (winter) and 10 mg/kg 3-MC (summer). In
preliminary experiments, 20 mg/kg 3-MC was found to be relatively toxic
during summer; consequently, the lower dose (10 mg/kg) was used. The
results for AHH activity are shown in Figure 2. In summer, maximum
induction (8.5-fold) was found 72 hr after administration of 3-MC (Day 4).
Hepatic microsomal AHH activities in fish sacrificed only 24 hr after
injection of 3-MC (Day 2) were 3.5-fold higher than in controls, and
7,8-benzoflavone inhibited AHH activities in these treated fish. In
winter, fish sacrificed 96 hr after the dose (Day 5) were induced only
2-fold, and their AHH activities were stimulated by addition of
7,8-benzoflavone. AHH activities in fish sacrificed on Days 8 to 40, after
a dose of 3-MC on Day 1, were inhibited by 7,8-benzoflavone and were induc-
ed to similar extents (4- to 5-fold) throughout this period. By Day 63,
AHH activities were down to 1.8 times control values; 7,8-benzoflavone
stimulated activity in only 1 of 3 treated fish assayed at this time.
AROCHLOR 1254
cytochrome benro(a)pyrene 7-ethoxycoumarin benzphetamine
P 450 hydroxylase 0-deethylase N-demethylase
vehicle-injected (n-5) | 50mg/kg (n=5) g 100 mg/kg (n=4)
Figure 3. Effect of arochlor 1254 on sheepshead hepatic microsomal
mixed-function oxidase activities.
180
-------
Effects on hepatic epoxide hydrase and glutathione S-transferase
activities - None of the aromatic hydrocarbons induced hepatic EH or GSH-T
activities with any epoxide substrate in any of the species studied.
However, in several experiments with sheepshead, usually those in which
high doses were used, EH activity was depressed in treated fish to about
75% of the activity in control fish. For any one experiment, the
depression was seldom statistically significant, due to the high variation
between individuals; but since the depression occurred repeatedly, it can
probably be attributed to the PAH treatment. A similar depression of EH
activity was found in hepatic microsomes from stringrays treated with 3-MC
(2 x 15 mg/kg, i.m.). EH activity in 3-MC-treated rays was 4.8 +.1.2
nmol/min/mg protein (mean=S.D., n=4) and activity in controls was 7.2 +_ 2.1
(mean _+ S.D., n=4). Hepatic GSH-T activity was not depressed or elevated
in any of the hydrocarbon-treated fish; however, a small depression of
activity occurred in 3-MC-treated sting ray. GSH-T activity in hepatic
cytosol from 3-MC-treated rays was 5.22 _+ 0.99 nmol/min/mg protein (mean +_
S.D., n=4); and activity in controls was 6.75 _+ 0.24 (mean _+ S.D., n=4).
Polyhalogenated Hydrocarbons
Effects on mixed-function oxidation -- Mixtures of PCBs (Aroclor 1254)
and PBBs (Firemaster FF1) both caused induction of AHH and 7-EC activities
in sheepshead after i.p. injection (Figures 3 and 4). These mixtures also
slightly induced cytochrome P-450 content and BND activities in livers of
treated fish, but again the increase was not always statistically
significant. Induction with Aroclor 1254 was achieved with doses of 50 and
100 mg/kg, i.p. The effects of lower doses were not studied. In an
initial experiment with Firemaster FF1 at a dose of 50 mg/kg, two of five
sheepshead died within five days, at which time the remaining fish had
induced AHH and 7-EC activities and cytochrome P-450 content. A more
thorough investigation of the effects of Firemaster FF1 on hepatic
xenobiotic-metabolizing enzymes was carried out with a dose of 15 mg/kg.
The results of experiments conducted in winter are summarized in Figure 4.
Peak induction of hepatic AHH and 7-EC activities was observed at Day 20
for fish dosed on Day 1; however, at Days 28 and 56, enzyme actiity with
these two substrates was still induced. Experiments in summer in which
fish sacrificed on Days 28 and 56 after an identical dose of Firemaster FF1
on Day 1 produced different results. In summer, Firemaster FFl-treated
sheepshead sacrificed on Day 28 showed 16-fold induction of AHH, 5-fold
induction of 7-EC, and 2-fold induction of BND in hepatic microsomes
compared with vehicle-injection controls, whereas sheepshead sacrificed
on Day 56 had only 1.7-fold induction of AHH, 2-fold induction of 7-EC, and
no induction of BND compared with controls.
181
-------
cytochrome
P 450
I vehicle-injected (n=!8)
I socnf iced Day 20 (n=5)
benzo(a)pyrene
hydroxylase
7-ethoxycoumarin
0-deethylase
benzphetamine
N-demethylase
sacrificed Day 8 (n=5)
sacrificed Day 28 (n=!0)
! sacrificed Day I5(n=5)
! sacrificed Day 56(n=8)
Figure 4.
Effect of Firemaster FF1 on sheepsheed hepatic microsomal
mixed-function oxidase activities.
The effects of a single dose of S.S'A.A'B.S'-HCB (10 mg/kg) on hepatic
mixed-function oxidation in sheepshead are shown in Figure 5. AHH
activities in treated fish were double those of control values by Day 4
and remained elevated at least until Day 24. However, hepatic microsomal
AHH activity in sheepshead treated with 3,3'4,4'5,5'-HCB was stimulated by
the in vitro addition of 10 ^ M 7,8-benzoflavone, whereas hepatic AHH
activity from sheepshead treated with PAH or mixtures of PCBs and PBBs was
inhibited by 10"/ M 7,8-benzoflavone. 7-EC activity was induced to
about the same extent as AHH activity in 3,3'4,4'5,5'-HCB-treated fish. In
addition, cytochrome P-450 content, BND activity, and NADPH-cytochrome £
reductase activity were increased 2-fold over control values in treated
sheepshead sacrificed on Day 15.
182
-------
TABLE 4. EFFECT OF POLYHALOGENATED BIPHENYL
CO
OJ
ADMINISTRATION ON HEPATIC MICROSOMAL EPOXIDE HYDRASE ACTIVITY
SHEEPSHEAD
EPOXIDE HYDRASE ACTIVITY0
Date sacrificed Dose Time Sacrificed Treated Control Ratio
(month/year) rag/kg (dose given on Day 1) Treated/Control
AROCHLOR 1254
3/77
4/77
3, 3', 4, 4', 5, 5'
6/78
4/78
5/78
5/78
FIREMASTER FF1
10/77
1/78
1/78
1/78
8/77
2/77
9/78
2, 2' ,4, 4' ,5, 5'
3/78
3/78
4/78
100
50
-HEXACHLOROBIPHENYL
10
10
10
10
15
15
15
15
15
15
15
-HEXABROMOBIPHENYL
20
3 x 20
3 x 20
+ 1 x 40
7
7
4
9
15
24
8
15
20
28
28
56
56
17
28
40
4.8 ± 1.5 (4)c
5.1 ± 0.7 (4)
5.9 ± 1.5 (5)
7.4 ± 1.1 (5)
7.9 ± 1.6 (4)
7.9 ± 3.0 (3)
6.7 ± 1.9 (5)
5.0 ± 1.3 (5)
4.3 i 1.3 (5)
6.1 ± 0.8 (5)
5.1 ± 0.6 (5)
4.7 ± 1.0 (5)
7.2 ± 1.5 (5)
4.5 ± 1.6 (5)
5.6 ± 2.4 (5)
5.5 ± 0.5 (5)
5.2 ± 1.3 (3)
2.4 ± 0.6 (3)
4.9 ± 1.1 (3)
4.5 ± 2.1 (3)
4.6 ± 0.6 (3)
5.2 ± 0.1 (3)
6.9 ± 1.4 (3)
4,6 ± 1.7 (3)
2.7 ± 0.4 (3)
5.0 ± 3.6 (3)
3.4 ± 0.9 (3)
5.8 ± 0.2 (3)
6.1+2.1 (4)
4,9 ± 1.3 (3)
5.6 ± 1.3 (3)
5.0 ± 2.0 (3)
0.92
2.13
1.20
1.64
1.72
1.52
0.97
1.09
1.59
1.22
1.50
0.81
1.18
0.92
1.00
1.10
aAll compounds were administered by i.p. injection of a corn oil solution or suspension. Controls received
corn oil.
Nmoles product formed/min/mg of protein.
cMean ± S.D. for (n) individuals.
-------
6
3,3',4,4', 5,5' - HEXACHLOROBIPHENYL
cytochrome
P 450
benzo(a)pyrene
hydroxylase
7-ethoxycoumarin
0-deethylase
benzphelamine
N-demethylase
cytochrome c
reductase
vehicle- injected | sacrificed Day 4 ^sacrificed Day 9
sacrificed Day 15 sacrificed Day 24
Figure 5.
Effects of 3,3'4,4' ,5 ,'5-hexachlorobiphenyl on sheepshead
hepatic enzymes.
Multiple doses of 2,4'4,4' ,5,5'-HBB had no effect on any of the
mixed-function oxidase enzymes assayed (Figure 6). This isomer was
relatively nontoxic, in contrast to the Firemaster FF1 mixture, and no
injected sheepshead died. 2,2' ,4,4' ,5,5'-HBB accounts for about 60% of
Firemaster FF1 (Moore et afK , 1978).
Effects on hepatic epoxide hydrase and glutathione S-transferase
activities—Aroclor 1254 appeared to have an effect on EH activity in
sheepshead treated with 50 rug/kg but no effect was observed with 100 mg/kg
(Table 4). However, this appeared to be due only to the lower EH
activities in the control fish of the 50 mg/kg group. For example, the
mean specific EH activity of all sheepshead assayed from March 1976 to
August 1978 was 5.48 +_ 2.14 (144) (mean +_ S.D., [n]). The effect of
3,3'4,4'5,5'-HCB administration was more consistent. Mean EH activity of
treated sheepshead was increased at each time period studied, compared with
mean control activity, and the difference was statistically significant
(p < 0.05) for sheepshead sacrificed on Day 15.
184
-------
2f2lt4l4ll5,5l-HEXABROMOBIPHENYL
o
<
o>
0)
a:
cytochrome benzo(a)pyrene 7-ethoxycoumarin benzpnetamine cytochrome £
P 450 hydroxylase 0-deethylase N-deemethylase reductase
vehicle-injected(n=9)
sacrificed Day 28(n=5)
sacrificed Day 17 (n=5)
sacrificed Day 40 (n=5)
Figure 6. Effect of 2,2',4,4',5',5,-hexabromobiphenyl on sheepshead
hepatic enzymes.
Firemaster FFl-treated fish sometimes had higher EH activities than
controls, but the difference was significant (_p_ < 0.05) only in one group
of fish. In our experiment, mean EH activity in control sheepshead was
also lower than usual. 2,2'4§4'5,5'-HBB administration had no effect on EH
activities. None of the treatments with halogenated biphenyls caused a
significant change in hepatic cytosolic fraction GSH-T activity with
styrene oxide, except Aroclor 1254 at 50 mg/kg, which again appeared to be
related to lower-than-normal specific activities in the control fish.
DISCUSSION
Polynuclear aromatic hydrocarbons—Polynuclear aromatic hydrocarbons
such as 3-MC, DBA, and DMBA have an effect on hepatic microsomal
mixed-function osidase activity in several species of marine fish, which is
similar to the effect PAHs have on hepatic microsomal mixed-function
oxidase activity in the rat, guinea pig, and several strains of mice. he
lack of induction of EH and GSH-T, and occasional inhibition of EH
are not inconsistent with the results found in mammals. Induction of both
of these enzymes by 3-MC in mammals was usually less than 2-fold; some dose
regimes caused inhibition of EH activity in rats (Bresnick et al_., 1977;
Oesch et.al_., 1973). Inhibition of xenobiotic-metabolizing enzymes soon
after Treatment with inducing agents is a commonly observed phenomenon
(Fouts, 1970). In pretreatment studies with wild species, such as fish,
is often difficult to obtain statistically significant changes of low
magnitude, such as found for EH and GSH-T activities in livers of
PAH-treated inbred rats, due to the variation in enzyme activities.
185
-------
Differences observed in PAH induction of hepatic microsomal AHH
activity in fish vs. mammals relate mainly to the effect of season on the
duration of the induction by 3-MC and the sensitivity to 3-MC induction.
In summer, induction of sheepshead hepatic microsomal AHH activity is
fairly rapid, and peak enzyme activities are attained 72 hr after
administration of the dose, after which time a gradual decline in AHH
activity is observed. A similar pattern is observed in rats treated i.p.
with 80 mg/kg 3-MC (Boobis ejt al_., 1977). In winter, hepatic AHH induction
in the sheepshead is not maximal until at least Day 8 after injection of
3-MC, and activities are still elevated to about the same extent on Day
40. Hepatic microsomal AHH activity in the sheepshead was induced by doses
as low as 1 mg/kg, i.p., whereas rats are not induced by doses of 10 mg/kg
(Boobis et jil_., 1977). We have found that sheepshead held in a floating
dock in tHe intracoastal waterway had induced hepatic AHH activities after
3 to 4 weeks and that AHH activities are inhibited by in vitro addition of
7,8-benzoflavone. This increase in AHH activity may be due to low
concentrations of PAH in the water since barnacles attached to the floating
cages contained phenanthrene, methylphenanthrene, fluoranthrene, and pyrene
in the parts per billion (ppb) range.
Induction of AHH, 7-EC, and ERF activities with no effect or
inhibition of EH and GSH-T activities raises the possibility that
PAH-exposed fish may metabolize PAH to toxic intermediates more rapidly
than they are detoxified. It is possible that the pattern of metabolites
formed in induced fish will be different from those formed in controls and
also that the tissue distribution, rates, and routes of excretion will be
altered in induction. Other factors such as water temperature may also
affect the disposition of xenobiotics in marine species. For species used
for human food, it is important to know the tissue distribution of
xenobiotics and whether they are stored as metabolites. These factors may
also affect the toxicity of a chemical to fish.
A report by Statham and coworkers (1978) showed that benzathracene-
induced trout excreted more 2-methylnaphthalene (as metabolites) into bile
than did control fish, and that livers of induced fish initially took up
more methyl naphthalene than livers of control fish. However, there was
little difference between control and induced fish with respect to blood
and muscle levels of methyl naphthalene. Further studies are needed in this
area to clarify the effect of induction on overall disposition of
xenobiotics and toxicity in marine species.
Polyhalogenated biphenyls--The effects of mixtures of PCBs and PBBs on
hepatic microsomal mixed-function oxidase activities in sheepshead are
generally similar to those found in mammals. Statistically significant
increases in EH activities were found only in one group of Aroclor 1254-
and one group of Firemaster FFl-treated sheepshead, and enzyme activities
in control fish for both of these experiments were unusually low. Analysis
of EH data obtained in 18 months (from March 1976 to December 1977) showed
that EH tended to be lower between March and August, especially in female
sheepshead. However, data obtained so far in 1978 are not consistent with
earlier observations and provide no explanation for the wide variation of
186
-------
EH activity in control sheepshead (activities measured in control
sheepshead ranged from 1.85 to 18.85 nmol/min/mg of protein), other than
genetic heterogeneity.
The two single polyhalogenated biphenyl isomers studied did not appear
to have the same effects in sheepshead as those found in mammals.
3,3',4,4'5,5'-HCB was found to be a polycyclic hydrocarbon-type inducer in
rats (Goldstein et al_., 1977), but the effect of this isomer in sheepshead
hepatic ensymes was not identical to the effect of 3-MC. 3,3'4,4'5,5'-HCB-
treated sheepshead had elevated AHH, 7-EC, BND, and cytochrome £ reductase
activities and cytochrome P-450 content in hepatic microsomes, especially
on Days 9 and 15 after a dose of 10 mg/kg. However, 7-8,benzoflavone did
not depress the AHH activities of treated fish. EH activity was also
significantly higher in 3,3'4,4'5,5'-HCB-treated fish on Day 15.
2,2'4,4'5,5'-HBB was without effect on sheepshead hepatic enzymes,
even at the highest doses tested, although it is a phenobarbital-type
inducer in rats (Moore ^t^l_., 1978). We wanted to study the effect of
this PBB isomer on sheepshead since it is the major compound present in
Firemaster FF1, which induced xenobiotic metabolism in sheepshead. These
experiments suggest that the induction caused by Firemaster FF1 may be due
to a minor component, or minor components, of the mixture. This lack of a
phenobarbital-like induction has been previously reported in fish (Bend and
James, 1978).
The doses of both PAH and polyhalogenated biphenyls which induced AHH
and 7-EC activities in sheepshead were lower than those needed to effect
induction in the rat (Brensnick jit aK, 1977; Dent^t^l_., 1976; Goldstein
et_ _§]_•, 1977). Since sheepshead and other fish exposed environmentally to
very low levels of PAH pollution are often found to have induced AHH
(Payne, 1976), 7-EC, and ERF activities, and sometimes other hepatic enzyme
activities, it appears possible that some species of fish are very
sensitive to induction by PAH. The results obtained with PCBs and PBBs
suggest that the same is true for these chemicals. Thus, it is important
to find out how induction of hepatic enzymes affects the rate of metabolism
of pollutant chemicals in fish and to determine if induction of the hepatic
mixed-function oxidase systems of fish can be used as a sensitive indicator
for the presence of toxic pollutants that behave as polycyclic hydrocarbon-
like inducers in the aquatic environment.
ACKNOWLEDGMENTS
We are grateful to Ms. E. R. Bowen for her excellent technical
assistance and to Ms. D. Ritter who assisted in the preparation of the
manuscript. We also appreciate the support of the U.S. Environmental
Protection Agency.
187
-------
REFERENCES
Bend, J. R., and M. 0. James. 1978. Xenobiotic metabolism in freshwater
and marine species. In. Biochemical and biophysical perspectives in
marine biology. D. Malins and J. R. Sargent, Eds., Academic Press, New
York. pp. 125-188. Vol. 4.
Boobis, A. R., D. W. Nebert, and J. S. Felton. 1977. Comparison of
p-naphthoflavone and 3-methylcholanthrene and inducers of cytochrome
p-448 and aryl hydrocarbon (benzo(a)pyrene) hydroxylase activity.
Mol. Pharmacol. 13:259-268.
Bresnick, E., H. Mukhtar, T. A. Stoming, P. M. Dansette, and D. M. Jerina.
1977. Effect of phenobarbital and 3-methylcholanthrene administration
on epoxide hydrase levels in liver microsomes. Biochem. Pharmacol.
26:891-892.
Conney, A. H. 1967. Pharmacological implications of microsomal enzyme
induction. Phramacol. Rev. 19:317-366.
Dent, J.E., K. J. Netter, and J. E. Gibson. 1976. The induction of
hepatic microsomal metablism in rats following acute administration of
a mixture of polybrominated biphenyls. Toxlcol. Appl. Pharmacol.
38:257-249.
Elmamlouk, I.E., R. M. Philpot, and J. R. Bend. 1977. Separation of 2
forms of cytochrome P-450 from hepatic microsomes of
1,2,3,4-dibenzanthracene (DBA)-pretreted little skates.
Pharmacologist 19: 160.
Fouts, J.R. 1970 The stimulation and inhibition of hepatic microsomal
drug-metablizing enzymes with special reference to effect of
environmental contaminants. Toxicol. Appl. Pharmacol. 17:804-809.
Goldstein, J.A., P. Hickman, H. Bergman, J.D. McKinney, and M. P. Walker.
1977. Separation of pure polychlorinated biphenyl isomers into 2
types of inducers on the basis of induction of cytochrome P-450 or
P-448. Chem. Biol. Interact. 17:69-87.
Hales, B. F., and A. H. Neims. 1977. Induction of rat hepatic glutathione
S-transferase B by phenobarbital and 3-methylcholanthrene. Biochem.
Pharmacol. 26:555-556.
Heidelberger, C. 1975. Chemical carcinogenesis. Annu. Rev. Biochem.
44:79-121.
James, M. 0., E. R., Bowen, P. M. Dansette, and J. R. Bend. 1979. Epoxide
hydrase and glutathione S-transferase activities with selected alkene
and arene oxides in several marine species. Chem. Biol. Interact.
25:321-344.
188
-------
James, M. 0., M. A. Q. Khan, and J. R. Bend. 1979. Hepatic microsomal
mixed-function oxidase activities in several marine species common
coastal Florida. Comp. Biochem. Physiol. 620:155-164.
Kohli, K K., H. Mukhtar, J. R. Bend, P. W. Albro, and J. D. McKinney.
1978. Biochemical effects of pure isomers of hexachlorobiphenyl
(HCB): Hepatic microsomal epoxide hydrase and cytosolic glutathion
S-transferase activities in the rat. Biochem. Pharmacol.
28:144-1446.
Litterst, C. L., T. M. Farber, A. M. Baker, and E. J. Van Loon. 1972.
Effect of polychlorinated biphenyl on hepatic microsomal enzymes in
the rat. Toxicol. Appl. Pharmacol. 23:112-123.
Lu, A. Y. H., W. Levin, S. West, M. Jacobson, D. Ryan, R. Kuntzman, and A.
H. Conney. 1973. The role of cytochrome P-450 and P-448 in drug and
steriod hydroxylations. Annu. N. Y. Acad. Sci. 212:156-174.
Moore, R. W., S. D. Sleight, and S. D. Aust. 1978. Induction of liver
microsomal drug-metabolizing enzymes by 2,2',4,4',5,5'-hexabromo-
biphenyl. Toxicol. Appl. Pharmacol. 44:309-321.
Mukhtar, H., and E. Bresnick. 1976. Effects of phenobarbital and
3-methylcholanthrene administration on glutathione
S-epoxidetransferase activity in rat liver. Biochem. Pharmacol.
25:1081-1084.
Oesch, R., D. M. Jerina, J. W. Daly, and J. M. Rice. 1973. Induction
activation, and inhibition of epoxide hydrase: An anomalous
prevention of chlorobenzene-induced hepatotocicity by an inhibitor of
epoxide hydrase. Chem. Biol. Interact. 6:189-202.
Payne, J. F. 1976. Field evaluation of benzopyrene hydroxylase induction
as a monitor for marine petroleum pollution. Science 191:945-946.
Pohl, R. J., J. R. Bend, A. M. Guarino, and J. R. Fouts. 1974. Hepatic
microsomal mixedfunction oxidase activities of several marine species
from coastal Maine. Drug Metab. Dispos. 2:545-555.
Ryan, D., A. Y. H. Lu, J. Kawalek, S. B. West, and W. Levin. 1975. Highly
purified cytochrome P-448-and P-450 from rat liver microsomes.
Biochem. Biophys. Res. Commun. 64:1134-1141.
Ryan D. E., P. E. Thomas, D. Korzeniowski, and W. Levin. 1977. Separation
of multiple forms of highly purified liver microsomal cytochrome P-450
from rats treated with Arochlor 1254. Fed. Proc. 37:766.
Statham, C. N., C. R. Elcombe, S. P. Szyjka, and J. J. Lech. 1978. Effect
of polycyclic aromatic hydrocarbons on hepatic microsomal enzymes and
disposition of methyl naphthalene in rainbow trout in vivo.
Xenobiotica 8:65-71.
189
-------
Weibel, F. J., and H. V. Gelboin. 1975. Aryl hydrocarbon [benzo(a)pyrene
hydroxylase] in liver from rats of different age, sex, and nutritional
status: distinction of two types by 7,8-benzoflavone. Biochem.
Pharmacol. 24:1511-1515.
190
-------
METABOLISM OF BENZO(a)PYRENE BY CIONA INTESTINALIS
by
*
WILLIAM M. BAIRD, RUJ£ A. CHEMERYS, LEILA DIAMOND,
THOMAS H. MEEDEL, AND J.Richard WHITTAKER
The Wistar Institute of Anatomy and Biology
Philadelphia, PA 19104
ABSTRACT
Some ascidians (sea squirts) such as Ciona intestinal is
thrive in ports and dock regions where they are exposed to
polycyclic aromatic hydrocarbons. To determine if Ciona
intestinal is can metabolize such hydrocarbons, specimens were
exposed to [G- H]benzo(ajpyrene (BaP) (0.5 nmole/mJi seawater)
for 24 hr, and the seawater and several tissues were analyzed
for BaP and BaP metabolites. The BaP'was concentrated by the
organisms; large animals contained as much as 60% of the BaP
after 24 hr. Samples of all ascidian tissues and of seawater
from the ascidian-containing beakers contained slightly higher
proportions of water-soluble radioactive derivatives than did
samples from control beakers. An unidentified BaP derivative
with chromatographic properties similar to those of
BaP-dihydrodiols was detected in the seawater from the
ascidian-containing beaker. Protein-associated radioactivity
was found in several ascidian tissues and was highest in the
intestine. These results suggest that Ciona intestinal is is
able to concentrate a polycyclic hydrocarbon present in the
water, and that the hydrocarbon is slowly lost from the
organism as a water-soluble derivative. Whether this is due to
spontaneous breakdown of the hydrocarbon or to a low level of
hydrocarbon-metabolizing activity of the organism has not been
established, but the very slow rate of conversion of the
hydrocarbon to water-soluble forms probably serves to protect
the organism from biological effects induced by hydrocarbons in
organisms with high hydrocarbon-metabolizing activity.
Present address: Department of Medical Chemistry and Pharmacognosy,
School of Pharmacy, Purdue University, West Lafayette,
** IN 47907
Present address: Boston University Marine Program,
Marine Biological Laboratory, Woods Hole, MA 02543
191
-------
INTRODUCTION
Several polycyclic aromatic hydrocarbons found in petroleum products
or combustion products induce cytotoxicity, mutation, and transformation of
cells in culture and cancer in a number of mammalian species (NAS, 1972;
Dipple, 1976; Freudenthal and Jones, 1976, 1978). Metabolism of the inert
hydrocarbon molecule is required for the induction of such biological
effects. Only a few metabolic pathways lead to the formation of those
metabolites that are more active than the parent hydrocarbons in inducing
biological effects, and most pathways lead to detoxification of the
hydrocarbons through the formation of oxidized derivatives and their
conjugates (NAS, 1972; Heidelberger, 1973; Sims and Grover, 1974; Dipple,
1976; Freudenthal and Jones, 1976, 1978; Jerina and Daly, 1976; Diamond and
Baird, 1977).
The Ascidiacea (subphylum: Tunicata), or sea squirts, are a class of
filter-feeding marine organisms (Millar, 1953, 1971); some species are able
to thrive in areas containing various pollutants (Papadopoulou and Kanias,
1977; Riggio and Mazzola, 1976). We have observed that Ciona intestinal is,
a species frequently used for studies in embryology because of the
determinate cleavage pattern of its eggs, thrives in dock regions where oil
pollution exists. Polycyclic aromatic hydrocarbons are frequently found in
such areas (Kraybill, 1976). Some large adult Ciona collected in the area
of Sandwich, MA; relativley large amounts of petroleum residues on their
surfaces observed showed no apparent toxicity from this exposure.
To find out if Ciona intestinal is can thrive in areas containing
polycyclic aromatic hydrocarbons by metabolizing them through pathways that
do not form biologically active metabolites, the metabolism of
benzo(a)pyrene (BaP), a widespread environmental contaminant, was examined
with techniques developed for the analysis of hydrocarbon metabolism in
tissue culture. These include treatment with an isotopically labeled
hydrocarbon, enzymatic cleavage of conjugated metabolites, isolation of
metabolites by organic solvent extraction, and analysis by chromatography
(Diamond et al., 1967; 1968; Duncan jit ^1_., 1969; Sims, 1970; Huberman et
£]_., 1971; Baird "and Brookes, 1973; Cohen et al_., 1976; Baird et jil_., 1977,
1978; Selkirk, 1978; Philips and Sims, 1979J.
MATERIALS AND METHODS
Animals—Specimens of Ciona intestinal is (L.) were collected off Cape
Cod, MA, shipped in seawater to the Wistar Institute, and maintained 7 to
14 days in an Instant Ocean tank (Aquarium Systems, Inc., East Lake, OH) of
artificial seawater at 15° C. At least 1 hr before experimental use,
animals were transferred to glass beakers (one animal per beaker)
containing *50 mi seawater and 20 jig rifampicin/mA and maintained at 18° C
throughout the experiment. Animals approximately 4 cm in length were
classed as large; those approximately 1.5 cm in length as small.
o
Benzo(a)pyrene metabolism—(G- H)benzo(a)pyrene (Sp. Act. 3-13
Ci/nmole) was purchased from Amersham Corp., IL, and dissolved in dimethyl
192
-------
sulfoxide immediately before use. All studies were carried out in the
presence of yellow light or in the absence of light to prevent photodecomp-
osition of the BaP. The BaP solution was added to the seawater to give
final concentrations of 0.5 nmole BaP/m«, and 0.2% dimethyl sulfoxide.
Immediately after the BaP was added to the beaker, a 2 mi sample of water
was removed for incubation as the control. After 24 hr of incubation,
samples of water were removed for analysis from both the ascidian-contain-
ing and control beakers, the animals dissected, and the tissues frozen for
metabolite analysis.
BaP metabolites and unchanged BaP were extracted from water and tissue
samples by a two-step chloroform:methanol:water procedure and the amount of
radioactivity in each phase determined by liquid scintillation counting of
0.1 ml aliquots (Baird and Diamond, 1976). The material extracted by the
organic solvent was analyzed by high-pressure liquid chromatography (HPLC)
(Selkirk, 1978) or thin-layer chromatography (TLC) (Baird et al_., 1978;
1979). HPLC analyses were carried out on an Altex Model 312 using a 4.6 mm
x 25 cm Spherisorb 5 ym ODS reversed-phase column at 30° C. Samples were
eluted at a flow rate of 1 m£/min with a 60 min concave gradient (Altex
exponent 3) from 3:2 to 4:1 methanol-water followed by 10 min at 4:1; 140
fractions (0.5 min) were collected and radioactivity was measured by liquid
scintillation counting (Baird et _§]_., 1979). Elution positions of markers
of BaP metabolites were determined by UV-absorbance. TLC analyses were
carried out on Eastman 13179 silica gel chrbmogram sheets without
fluorescent indicator and developed in benzene:ethanol (19:1) or toluene:
ethanol (19:1). The samples were then cut into 1 cm squares, eluted with 1
nu methanol, and measured by liquid scintillation counting (Baird and
Diamond, 1976). Markers of BaP metabolites were chromatographed on each
sheet and visualized by short-wave UV light.
Protein-associated radioactivity--Tissues to be analyzed were first
homogenized in distilled water. BaP and its metabolites were extracted
with chloroform:methanol:water as described. The protein interface was
removed, extracted three times with methanol and three times with ethyl
acetate, and dried with nitrogen. The interface was then dissolved in
0.5 N sodium hydroxide solution, neutralized with 1 N hydrochloric acid,
and extracted with two volumes ethyl acetate. The interface from this
extraction was removed, extracted with methanol (three times) and ethyl
acetate (three times), and dried under nitrogen. The sample was dissolved
in 0.5 N sodium hydroxide solution and neutralized with hydrochloric acid.
Aliquots were removed for analysis of radioactivity by liquid scintillation
counting and of protein by a modification of the Lowry procedure (Ross and
Schatz, 1973).
RESULTS
A large ascidian was exposed to (3H)BaP (0.5 nmole/m£ water) for 24
hr, then an aliquot of the water was extracted with chloroform and
methanol, and the chloroform-extractable material was analyzed by HPLC.
The HPLC elution profiles of the chloroform extracts of water from a
control beaker (Figure la) and the ascidian-containing beaker (Figure Ib)
193
-------
show that extracts contained mainly unchanged BaP. The sample from the
ascidian water, however, contained an unidentified peak, designated
"unknown No. 1," that eluted slightly later than the BaP-9,10-diol markers,
With TLC analysis, a similar amount of radioactivity showed an Rp
comparable to markers of BaP-diols. Both ascidian and control water
samples contained similar small amounts of BaP-quinones.
100,000-
30,000
o.
o
4.000-
2,000-
a
o
g
a>
2 9
155
10 30 5O 70 90
FRACTION NUMBER
110
130
Figure 1. HPLC elution profiles of BaP and BaP metabolites in seawater
from a control beaker with no ascidian (a) and a beaker
containing a large Ciona intestinal is (b). ( H)BaP was added
to a beaker containing an ascidian and a sample of water was
immediately removed and incubated separately as a control.
After 24 hr, 0.2 mi samples from each beaker were extracted with
chloroform:methanol:water and the chloroform extracts analyzed
by HPLC, as described in Materials and Methods. The elution
positions of BaP metabolite markers are shown at the top of each
sample. • DPM/0.5 mi fraction.
194
-------
TABLE 1. RADIOACTIVITY REMAINING IN SEAWATER AFTER EXPOSURE OF ASCIDIANS
TO (JH)BaP
Sample
% of radioactivity
not extractable by
organic solvent
Peaks as % of radioactivity
extractable by organic solvent
Unknown #1 Quinones BaP
Large Ascidian
Control
Small Ascidian
Control
26
1
1.5
0.6
15 5
0.2 3
4
4
80
96
93
95
After exposure of ascidians to ( H)BaP (0.5 nmole/nu water) for 24 hr,
water samples from beakers containing either large (4 cm) or small (1.5 cm)
ascidians and from control beakers were extracted with choloroform:methanol:
water and the chloroform phases analyzed by HPLC. Results are averages of
two or three experiments.
TABLE 2. RADIOACTIVITY RECOVERED FROM ASCIDIAN TISSUES AFTER EXPOSURE TO
(3H)BaP
Peaks as Z of radioactivity
Z of radioactivity extractable by organic solvent
aoc extractable by
Procein-associacad
radioactivity
Sample
Control
Tunic
Intestine
Branchial basket
Ovary
organic solvent
1
6
2
2
»>
Unknown 4 1
0.2
0.2
0.2
0,3
ND
Quinone
3
4
4
2
ND
BaP
96
95
95
97
ND
pmoles [JH]BaP/nig
protain
—
5
10
5
2
o
aND = not determined. Large ascidians were exposed to (^HjBaP for 24
hr and dissected. The BaP metabolites in each tissue were extracted and
analyzed by HPLC as described in Table 1. Water from a beaker containing
no ascidian was used as a control. All results are the average of two
experiments.
195
-------
The most striking difference between ascidian water and control water
was in the amount of radioactivity remaining after 24 hr. Control water
contained nearly 70% of the radioactivity added, in contrast to the less
than 7% in the water samples from large ascidians. Thus, most of the BaP
had been removed by the ascidian. This BaP removal was dependent upon the
size of the ascidian, for only a small amount (5 to 10%) was removed by
small ascidians. The water from the small ascidian beakers contained
mainly unchanged BaP, with small amounts of BaP quinones (Table 1). The
amount of BaP that precipitated in the water was similar in all cases: an
acetone rinse of beakers after removal of the water resulted in the
recovery of 20 to 25% of the radioactivity from beakers both with animals
and without.
Tissues were dissected from two large ascidians and analyzed by the
HPLC procedure to determine if BaP metabolites had been formed but not
released into the water (Table 2). No identifiable BaP metabolites were
detected in the three tissues examined. Quinones were present in all
tissues, but the amounts were not significantly greater than those found in
water from the control beaker. All tissue samples contained more
water-soluble radioactivity than the control, but the nature of this
water-soluble material could not be determined. In the large ascidians,
several times as much radioactivity was recovered from the intestine as
from the branchial basket. In the small ascidians, the material recover-
ed from each tissue was similar to that of the large ascidian except that
more was recovered in the branchial basket than in the intestine (data not
shown).
To see if any hydrocarbon metabolites that formed interacted with
cellular components, we measured the amount of protein-associated radioact-
ivity (Table 2). After the tissues were exhaustively extracted by the
procedure described in Materials and Methods to remove any unbound BaP or
BaP metabolites, the protein-associated radioactivity ratios were
calculated (Table 2). Although the amounts were low, BaP appeared to be-
come bound to cellular proteins, especially in the intestine.
DISCUSSION
We have shown that the ascidian, Ciona intestinal is, can remove the
hydrocarbon BaP from seawater. The amount removed by large ascidians was
greater than that removed by small ascidians, but the concentration of BaP
within the organism was greater than in the surrounding water.
No identifiable BaP metabolites were detected in either the
surrounding water or in any of the tissues examined. Some BaP quinones
were found, especially in the intestine where a large portion of the BaP
had accumulated, but the ratio of quinones to BaP did not differ from that
in the control seawater. These quinones may represent a spontaneous
breakdown products of BaP. However, three lines of evidence suggest that .
ascidians possess a low level of BaP metabolizing activity: (1) An
unidentified BaP derivative with chromatographic properties similar to
those of a BaP-diol was recovered in the water from the beaker that
196
-------
contained the large ascidian, but this material was not present in the
control water; (2) A higher percentage of the radioactivity was recovered
in the water phases of chloroform- methanol extractions of tissue samples
and water from the beaker containing the ascidian than from the control
water sample; (3) Protein-associated radioactivity was recovered in
several tissues and was greatest in the tissue with the highest BaP
concentration.
These findings suggest that Ciona intestinal is is capable of
concentrating hydrocarbons such as BaP from seawater. BaP is then slowly
converted to a water-soluble form, either by breakdown, by bacterial
metabolism, or by a low metabolizing activity of the organism. A small
portion of the hydrocarbon also becomes bound to tissue components.
The very low level of hydrocarbon metabolism probably protects Ciona
from the induction of toxicity by such compounds; although most of the
hydrocarbon added to the water was concentrated in the organism, the
ascidian metabolized only a small amount in 24 hr. In contrast, cultures
of rodent cells exposed to similar concentrations of hydrocarbons would
have metabolized almost all of the hydrocarbon within 24 hr (Baird et al.,
1977). Thus, Ciona may be able to gradually remove the hydrocarbons it
accumulates without generating toxic levels of reactive metabolites. The
finding that some BaP becomes bound to cellular components warrants further
investigation. Although the major DNA-binding derivative in several
mammalian systems has been identified as a dihydrodiol-epoxide of BaP
(Weinstein et jil_., 1976; Koreeda et a]_., 1978; Phillips and Sims, 1979),
there is evidence that BaP-quinones may be involved in BaP-DNA interactions
in microsomal systems (Pelkonen et _al_., 1978). BaP-quinones are present in
the tissue of BaP-treated Ciona and therefore may be involved in
BaP-macromolecule interactions in this organism.
ACKNOWLEDGMENTS
This work was supported, in part, by Public Health Service Grants CA
19948, CA 08936, and CA 23394 from the National Cancer Institute.
REFERENCES
Baird, W.M., and P. Brookes. 1973. Isolation of the hydrocarbon-deoxyri-
bonucleoside products from the DNA of mouse embyro cells treated in
culture with 7-methylbenz(a)anthracene- H. Cancer Res.
33:2378-2385.
Baird, W.M., and L. Diamond. 1976. Effect of 7,8-benzoflavone on the
formation of benzo(a)pyrene-DNA-bound products in hamster embryo cells,
Chem. Biol. Interact. 13:67-75.
Baird, W.M., C.J. Chern, and L. Diamond. 1977. Formation of benzo(a)-
pyrene glucuronic acid conjugates in hamster embryo cell cultures.
Cancer Res. 37:3190-3197.
197
-------
Baird, W.M., R.A. Chemerys, C.J. Chern, and L. Diamond. 1978. Formation
of glucuronic acid conjugates of 7,12-dimethylbenz(a)anthracene
phenols in 7,12-dimethylbenz(a)anthracene-treated hamster embryo cell
cultures. Cancer Res. 38:3432-3437.
Baird, W.M., R. Chemerys, A.A. Erickson, C.J. Chern, and L. Diamond. 1979.
Differences in pathways of polycyclic aromatic hydrocarbon metabolism
as detected by analysis of the conjugates formed. In: Polynuclear
aromatic hydrocarbons, proceedings of the third international
symposium. P.M. Jones and P. Leber, Eds. Ann Arbor Science
Publishers, Inc., Ann Arbor, MI. pp. 507-516.
Cohen, G.M., S.M. Haws, B.P. Moore, and J.W. Bridges. 1976.
Benzo(a)pyrene-3-yl hydrogen sulphate, a major ethyl acetate-extract-
able metabolite of benzo(a)pyrene in human, hamster, and rat lung
cultures. Biochem. Pharmacol. 25:2561-2570.
Diamond, L., V. Defendi, and P. Brookes. 1967. The interaction of
7,12-dimethylbenz(a)anthracene with cells sensitive and resistant to
toxicity induced by this carcinogen. Cancer Res. 27:890-897.
Diamond, L., C. Sardet, and G.H. Rothblat. 1968. The metabolism of
7,12-dimethylbenz(a)anthracene in cell cultures. Int. J. Cancer
3:838-849.
Diamond, L., and W.M. Baird. 1977. Chemical carcinogenesis in vitro. In:
Growth, nutrition and metabolism of cells in culture, Vol. III. G.H.
Rothblat and V.J. Cristofalo, Eds. Academic Press, New York.
pp. 421-470.
Dipple, A. 1976. Polynuclear aromatic hydrocarbons. In: Chemical
carcinogens. C.E. Searle, Ed., American Chemical Society, Washington,
DC. pp. 245-314.
Duncan, M., P. Brookes, and A. Dipple. 1969. Metabolism and binding to
cellular macromolecules of a series of hydrocarbons by mouse embryo
cells in culture. Int. J. Cancer 4:813-819.
Freudenthal, R., and P.W. Jones, Eds. 1976. Carcinogenesis: a
comprehensive survey. Vol. I: Polynuclear aromatic hydrocarbons:
chemistry, metabolism, and carcinogenesis. Raven Press, New York.
Heidelberger, C. 1973. Chemical oncogenesis in culture. Adv. Cancer Res.
18:317-366.
Huberman, E., J.K. Selkirk, and C. Heidelberger. 1971. Metabolism of
polycyclic aromatic hydrocarbons in cell cultures. Cancer Res.
31:2161-2167.
Jerina, D.M., and J.W. Daly. 1976. Oxidation at carbon. In: Drug
metabolism—from microbe to man. D.V. Parke and R.L. Smith, Eds.,
Taylor and Francis, Ltd., London, pp. 13-32.
198
-------
Jones, P.W., and R.I. Freudenthal, Eds. 1978. Carcinogenesis: a
comprehensive survey, Vol. Ill: Polynuclear aromatic hydrocarbons:
second international symposium on analysis, chemistry, and biology.
Raven Press, New York.
Koreeda, M., P.O. Moore, P.G. Wislocki, W. Levin, A.H. Conney, H. Yagi, and
D.M. Jerina. 1978. Binding of benzo(a)pyrene 7,8-diol-9,10-epoxides
to DNA, RNA, and protein of mouse skin occurs with high stereoselect-
ivity. Science 199:778-781.
Kraybill, H.F. 1976. Distribution of chemical carcinogens in aquatic
environments. Prog. Exp. Tumor Res. 20:3-34.
Millar, R.H. 1953. Ciona. Liverpool Mar. Biol. Comm. Mem. Typ. Brit.
Plants Anim. 35:1-123.
Millar, R.H. 1971. The biology of ascidians. Adv. Mar. Biol. 9:1-100.
National Academy of Sciences (NAS). 1972. Particulate polycyclic organic
matter. Washington, DC.
Papadopoulou, C., and G.D. Kanias. 1977. Tunicate species as marine
pollution indicators. Mar. Poll. Bull., 8:229-231.
Pelkonen, 0., A R. Boobis, H. Yagi, D.M. Jerina, and D.W. Nebert. 1978.
Tentative identificatio of benzo(a)pyrene metabolite-nucleoside
complexes produced in vitro by mouse liver microsomes. Molec.
Pharmacol. 14:306-322.
Philips, D.H., and P. Sims. 1979. Polycyclic aromatic hydrocarbon
metabolites: their reactions with nucleic acids. In: Chemical
carcinogens and DNA, Volume II. P.L. Grover, Ed., CRC Press, Inc.,
Boca Raton, FL. pp. 29-57.
Riggio, S., and A. Massola. 1976. Preliminary data on the fouling
communities of the harbor of Palermo (Sicily). Arch. Oceanogr.
Limnol. 18 suppl. 3:141-151.
Ross, E., and G. Schatz. 1973. Assay of protein in the presence of high
concentrations of sulfhydryl compounds. Anal. Biochem. 54:304-306.
Selkirk, J.K. 1978. Analysis of benzo(a)pyrene metabolism by
high-pressure liquid chromatography. Adv. Chromatog. 16:1-36.
Sims, P. 1970. The metabolism of some aromatic hydrocarbons by mouse
embyro cell culture. Biochem. Pharmacol. 19:285-297.
Sims, P., and P.L. Grover. 1974. Epoxides in polycyclic aromatic
hydrocarbon metabolism and carcinogenesis. Adv. Cancer Res.
20:165-274.
199
-------
Weinstein, I.B., A.M. Jeffrey, K.W. Jeanette, S.H. Blobstein, R.G. Harvey,
C. Harris, H. Autrup, H. Kasai, and K. Nakanishi. 1976.
Benzo(a)pyrene diol epoxides as intermediates in nucleic acid binding
in vitro and in vivo. Science 193:592-595.
2QQ
-------
BIOACTIVATION OF POLYNUCLEAR AROMATIC HYDROCARBONS
TO CYTOTOXIC AND MUTAGENIC PRODUCTS
BY MARINE FISH
by
John J. Stegeman
Biology Department, Woods Hole Oceanographic Institution
Woods Hole, MA 02543
and
Thomas R. Skopek and William G. Thilly
Department of Nutrition and Food Science,
Massachusetts Institute of Technology, Cambridge, MA 02139
ABSTRACT
Levels of hepatic cytochrome P-450 and mixed-function
oxygenase activity differed markedly between marine fish species
scup, Stenotomus versicolor, and winter flounder,
Pseudopleuronectes americanus, and between male and female winter
flounder. Hepatic preparations from all these fishes, however,
were capable of efficiently activating carcinogenic polynuclear
aromatic hydrocarbons to mutagenic derivatives. The results
indicate that coastal marine fishes may be at a risk to carcino-
genic aromatic hydrocarbons in marine sediments.
INTRODUCTION
Polynuclear aromatic hydrocarbons are metabolized or biotransformed by
microsomal cytochrome P-450 dependent mixed-function oxygenases in tissues
of diverse species (Walker, 1978). The metabolism of carcingogenic poly-
nuclear aromatic hydrocarbons by some species is known to result in
formation of mutagenic derivatives (Wislocki et _§!_., 1976). Studies have
indicated that certain metabolites are responsible for mediating the car-
cinogenic activity of these compounds (Levin et^l_., 1977, 1978; Slaga et
al., 1977; Hecht et_ ^1_., 1978). However, microsomal cytochromes P-450 from
different mammalian tissues, different species, or animals subjected to
different treatment appear to vary in their ability to metabolize and
activate polynuclear aromatic hydrocarbons to mutagens in vitro (Ames
et al_., 1975; Levin et j»l_., 1976). Such variation may indicate differences
201
-------
in susceptibility to mutagenic or carcinogenic potential of polynuclear
aromatic hydrocarbons.
Microsomal electron transport systems in fish tissues are qualita-
tively similar to those in mammals (Pohl je_t aj_., 1974; Chevion £t a\_., 1977;
Stegernan and Binder, 1979). The levels of components of these systems and
the mixed-function oxygenase reactions carried out, however, are generally
lower in fish (Pohl et^ aj_., 1974; Bend et al., 1977). Yet, in some species
the activity of hepatic aryl hydrocarbon [benzo(a)pyrene] hydroxylase is
normally found to be higher than seen in mammals (Ahokas et_ aj_., 1975;
Stegeman and Binder, 1979). Data suggest these fish may have cytochrome(s)
P-450 that catalytically resemble(s) 3-methylcholanthrene-induced
cytochrome P-448 in mammals, whereas others may not (Bend et^ al_., 1977).
We do not know whether such apparent functional differences in cytochromes
P-450 within or between fish species might be associated with varied
capacity to activate polynuclear aromatic hydrocarbons. The present
contribution describes aspects of hepatic cytochrome P-450 systems and the
in vitro activation of selected polynuclear aromatic hydrocarbons by two
species of marine fish, scup (porgy), and winter flounder.
MATERIAL AND METHODS
Chemicals—Benzo(a)pyrene used in enzyme assays was obtained from
Aldrich Chemical Company (Milwaukee, WI). Benzo(a)pyrene, 7,12-dimethyl-
benzanthracene and 1,2,3,4-dibenzanthracene used in mutation assays were
obtained from Sigma Chemical Co. (St. Louis, MO). NADP, NADPH, glucose-6-
phosphate, glucose-6-phosphate dehydrogenase, aminopyrine, Tris, and HEPES
were obtained from Sigma.
Animals—Adult male and female scup, Stejnptomus versicolor, about 100
to 200 g, were collected by angling in Great Harbor, Woods Hole, MA, in
August 1977; winter flounder, Pseudopleuronectes americanus, were obtained
in outer Narragansett Bay, RI, by otter trawl in December 1977. Males were
270 to 350 g and females 480 to 520 g. Scup were maintained for four
months in 800 gallon tanks at 19° _+ 1° C in flowing water filtered through
gravel and sand at the National Marine Fisheries Service, Woods Hole, MA.
Fish were fed a diet of chopped smelt and clams ad libitum every two days.
Flounder were maintained at ambient temperatures at the U.S.
Environmental Protection Agency, Environmental Research Laboratory,
Narragansett, RI, prior to transport to Woods Hole. At the time of use,
the scup were sexually quiescent, and the winter flounder were fully
hydrated, ready to spawn. No fish used in these studies received any
experimental treatment.
Tissue preparations—Animals were killed by decapitation. Excised
livers were placed immediately in ice-cold 0.1 M phosphate buffer, pH 7.3.
Tissues were minced and homogenized in 4 volumes of 0.1 M P04 buffer
pH 7.3, containing 1.15% KC1 and 3 mM MgCl2 by a Potter-Elvehjem tissue
grinder with 4 passes of the pestle at 1350 and 4 at 1900 rpm. Post-mito-
chondrial supernatant (PMS) preparations for use in mutation assays were
202
-------
collected after centrifuging at 9000 xg for 10 min. Microsomal fractions
for enzyme assays were isolated from the 9000 xg supernatant as previously
described (Stegeman and Binder, 1979). PMS preparations were frozen in
liquid N2 and then held at -80° C until use (up to two months). Enzyme
assays on either PMS or microsomes were done immediately.
Enzyme assays--Benzo(a)p.yrene hydroxylase and aminopyrine demethylase
in hepatic preparations were assayed as previously described (Stegeman and
Binder, 1979) by measuring fluorescent product formation and formaldehyde
generation, respectively. NADPH-cytochrome c (P-450) reduction and
cytochrome P-450 were measured as previously described (Stegeman and
Binder, 1979). Protein was determined according to Lowry jjt al_. (1951).
Bacterial mutation assays—Reverse mutation assays to histidine
prototrophy were carried out with Salmonella typhimurium strain TA-98;
forward mutation assays to 8-azaguanine resistance (Skopek et a]., 1978a)
employed _S_. typhimurium strain TM-677. The sources and storage conditions
for these strains have been indicated (Skopek ^t a]_., 1978b).
Basic protocols for both the reverse and forward mutation assays have
been described (Skopek £t jjl_., 1978a; 1978b). Exposure to promutagen was
in liquid culture in 25 m plastic tissue flasks and 5.0 m volumes that
contained an appropriate concentration of bacterial cells, 0.5 nu sterile
PMS, and 6.5 ymoles NADPH. Hydrocarbons were added to duplicate flasks in
50 y£ of dimethyl sulfoxide. Flasks were incubated without shaking for 2
hr at 29° C when scup PMS was employed and 25° C when winter flounder PMS
was employed. The temperatures selected were near optimal temperatures for
benzo(a)pyrene hydroxylase in these two species when assayed over a 2-hr
period. Harvest of cells and plating procedures for estimating both
bacterial survival and mutation are described elsewhere (Skopek, 1978b).
Mutant fractions in both reverse and forward assays are presented as number
of mutant clones (x factor)/number of survivor clones plated.
RESULTS
Levels of microsomal electron transport components and mixed-function
oxygenase activities in hepatic microsomes were compared in scup, winter
flounder, and mice. Results are presented in Table 1. The levels of
cytochrome b$, NADPH- and NADH-cytochrome c reductases in scup were about
2Q% of those measured in mice. Cytochrome P-450 present in scup was of
comparatively greater amounts (about 50% more than observed in mice).
However, the Soret absorption maximum of reduced, CO-treated microsomes was
about 450 nm in both species. Aminopyrine demethylase activity was also
much lower in scup than in mice, but benzo(a)pyrene hydroxylase activity
was almost ten-fold greater in the scup.
203
-------
TABLE 1. HEPATIC MICROSOMAL ELECTRON TRANSPORT COMPONENTS AND
MIXED-FUNCTION OXYGENASES IN SCUP, WINTER FLOUNDER, AND MICE
Component
Scupa
(N > 10)
Winter Flounder
male
(N > 3)
female
(N > 3)
Mouse3
(N > 3)
Liver wt./bndy wt. %
mg mic. protein/g liver
Cytochrome P-450
nmoles «mg
Cytochrome be
nmoles-mg prot."-'-
NADPH-cytochrome c reductase
units*mg ~l c
NAPH-cytochrome c reductase
units-ing prot. ~1
Aminopyrine demethylase
units -rag prot.~l
Benzo (a)pyrene hydroxylase
units*mg prot."-'-
1.01±0.10b 1.0210.05
12.4 ±0.5 17.5 ±0.7
0.62+0.08 0.90+0.21
0.06±0.02
107± 5
183± 6
206+34
693±40
48+ 5
84+13
213+15
2.43±1.7 5.24±0.29
23.0 ±2.9 20.0 ±0.32
0.19+0.2 1.14+0.12
0.33±0.03
38± 5
51+19
77± 5
510± 9
913± 59
800+151
72± 14
f*
Data from Stegeman and Binder (1979). Mice were adult Charles River CD-I females.
b
All values are ± S.E.M.
c _i
Units are nanomoles cytochrome c reduced-min" (reductases), nanomoles HCHO produced
normalized to 1 hour (aminopyrine demethylase) and picomoles 3-OH-benzo(a)pyrene
equivalents produced•min""^ [benzo (a)pyrene hydroxylase].
During spawning, pronounced sex differences in winter flounder were
observed in nricrosomal cytochromes P-450. The Soret absorption maximum of
reduced, CO-treated microsomes from male fish was quite clearly at 448 nm
rather than 450 nm seen in females, or in scup or mice. The levels of
cytochrome P-450 seen in males were more than five times those of females
and were greater than those observed in scup. Unlike cytochrome P-450,
NADPH-cytochrome c reductase activity was quite similar in male and female
flounder, and the levels were lower than in scup. The levels of both
204
-------
aminopyrine demethylase and benzo(a)pyrene hydroxylase activity were also
lower in male and female winter flounder than in scup. However, a sexual
difference was apparent in the levels of these activities (levels in
females were lower than those in males). Benzo(a)pyrene hydroxylase
activity in female flounder was almost as low as that seen in mice, while
this activity in male flounder was several times greater.
The metabolic activation of benzo(a)pyrene to toxic and mutagenic
derivatives by fish liver preparations like those described in Table 1 was
initially determined by a reverse mutation assay. Results indicate that
untreated scup liver PMS when incubated with NADPH and benzo(a)pyrene was
capable of stimulating almost an 85-fold increase in the his revertant
fraction in j>. typhimurium strain TA-98. At the same time, there was a
20-fold reduction in survival of S. typhimurium in the complete incubation
with 50 yM benzo(a)pyrene. The activation indicated in Table 2 was
dependent on the presence of PMS, as well as benzo(a)pyrene and NADPH; a
linear dose-dependent increase was observed in both the mutant fraction and
the toxicity up to a peak at benzo(a)pyrene concentrations between 40 to
60 yM.
TABLE 2. REQUIREMENTS FOR ACTIVATION OF BENZO(a)PYRENE TO TOXIC AND
MUTAGENIC^DERIVATIVES IN S.. TYPHIMURIUM STRAIN TA-98 BY SCUP
LIVER PMS
Incubation Relative Observed Mutant"
Conditions Survival Fraction x 10^
Complete
[50 yM B(a)P] 0.05 ^7.2
Minus B(a)P 1.00 0.55
Minus NADPH 0.91 0.21
*From Stegeman (1977)
^Mutant Fraction refers to the number of his* revertant clones x 10'°
per number of survivor clones. Incubation conditions are as de-
scribed in Materials and Methods.
205
-------
TABLE 3. ACTIVATION OF POLYNUCLEAR AROMATIC HYDROCARBONS TO MUTAGENIC
DERIVATIVES IN 1- TYPHIMURIUM STRAIN TM-677 BY SCUP AND WINTER
FLOUNDER LIVER PMS
PMS Source
(N)a
Scup (9)
Compound^3
B(a)P
DBA
DMBA
Concentration0
(VM)
40
36
20
Relative
Survival
0.48
0.70
0.85
Induced
Mutant d
Fraction
x 105
85.5
32.6
14.6
Winter Flounder
Male (2)
Female (2)
B(a)P
B(a)P
40
40
0.73
0.84
144.5
79.0
Livers from N fish pooled for PMS preparation.
Abbreviations are B(a)P, benzo(a)pyrene; DBA, 1,2,3,4-dibenzanthracene;
DMBA, 7,12-dimethylbenzanthracene.
r»
""Data presented have been selected from dose-response curves and repre-
sent concentrations at which maximal response was detected, except for
female winter flounder. Here the data at 40 yM were selected for com-
parison with both male v/inter flounder and scup.
Mutant fraction refers to the number of 8-azaguanine resistant clones
x 10"^ per number of survivor clones. Background mutant fractions
(0 compound) were 4-6(xlO^) within the range previously reported
(Shopek £t^ al., 1978b) and have been subtracted.
206
-------
Activation of polynuclear aromatic hydrocarbons by scup and winter
flounder liver PMS was compared, using the forward mutation assay to
8-azaguanine resistance in _S. typhimurium strain TM-677. The levels of
benzo(a)pyrene hydroxylase activity per mi of hepatic PMS in the specific
preparations used were 4000 pmol 3-OH-benzo(a)pyrene equivalents/min/m£ for
scup, which was twice that in the male winter flounder, 2018 pmol/min/nu,
and more than six times that in the female, 632 pmol/min/mJi. The results
are presented in Table 3. The mutant fraction resulting from activation of
benzo(a)pyi[;ene by scup liver PMS in this assaay was high, as was observed
in the his reversion assay. Scup liver PMS also readily activated
1,2,3,4-dibenzanthracene and 7,12-dimethylbenzanthracene, but the mutant
fractions observed were lower than with benzo(a)pyrene.
The mutant fraction induced by activation of 40 \M benzo(a)pyrene by
female winter flounder PMS was comparable to that seen with scup. Male
winter flounder PMS, on the other hand, was able to induce a somewhat
higher mutant fraction than scup with 40 yM benzo(a)pyrene. It is
noteworthy that these results are quite different from that seen in levels
of benzo(a)pyrene hydroxylase activity in PMS preparations, or found in
hepatic microsomal preparations of scup and winter flounder. This suggests
that the level of benzo(a)pyrene hydroxylase activity in fish is not a
suitable indicator of the capacity to form mutagenic derivatives from a
polynuclear aromatic hydrocarbon such as benzo(a)pyrene.
DISCUSSION
Both fish species examined possessed marked ability to metabolically
activate polynuclear aromatic hydrocarbons to mutagenic products in vitro.
The mutant fractions induced with 40 yM benzo(a)pyrene in fact exceeded
that [72 (x 10 )] obtained with 40 yM benzo(a)pyrene and hepatic
preparations from Aroclor 1254-induced rats in the forward mutation assay
with S^ typhimurium TM-677 (Skopek et jj]_., 1978b). This is particularly
interesting because fish used in these studies had received no experimental
treatment.
Hepatic preparations from rats uninduced or induced with phenobarbital
usually have much lower capacity for activation of benzo(a)pyrene than
those from animals treated with 3-methylcholanthrene or mixed inducers such
as Aroclor 1254 (Ames et^ ^1_., 1975). This can be attributed to differences
in the ability of cytochromes P-450 in the variously treated animals to
produce certain mutagenic derivatives of benzo(a)pyrene (Levin et al.,
1976). By analogy, the ready formation of mutagenic derivates of
benzo(a)pyrene by the fish here suggests that cytochromes P-450 in these
animals may in some way be catalytically similar to cytochromes P-448 in
some mammals. This is consistent with the observation that hepatic
benzo(a)pyrene hydroxylase in untreated scup at least is strongly inhibited
by 10 M 7,8-benzoflavone (Stegeman and Binder, 1979), a characteristic
of 3-methylcholanthrene-induced cytochrome P-448 in some mammals
(Weibel et al^., 1971). However, it is questionable whether these are
features of constitutive cytochromes P-450 in scup (Stegeman and Binder,
1979). Possibly, the fish used in our study had been exposed incidentally
to aromatic hydrocarbons in the environment, and thus the extent of
207
-------
benzo(a)pyrene activation, as with the reported inhibition by 7,8-benzo-
flavone, may not be characteristic of an uninduced state.
A variety of primary and secondary metabolites of benzo(a)pyrene are
formed by mammalian liver enzymes. Among the primary metabolites,
benzo(a)pyrene-4,5-epoxide is the most potent mutagen, but this arene oxide
is readily inactivated by epoxide hydrase, yielding benzo(a)-
pyrene-4,5,-dihydrodiol (Levin et jil_., 1976). Benzo(a)pyrene-7,8-dihy-
drodiol is formed in a similar manner but, after further metabolism by
cytochrome P-450, can yield benzo(a)pyrene-7,8-diol-9,10-epoxides--highly
mutagenic secondary metabolites (Wislocki et^ jil_., 1976) that are not good
substrates for, and thus not readily inactivated by, epoxide hydrase (Wood
jejt a\_., 1976). An isomer of benzo(a)pyrene-7,8-diol-9,10-epoxide, more
readily formed by 3-methylcholanthrene-induced cytochrome P-448 (Huberman
et ^1_., 1976; Yang et a]_., 1976), is believed to be the ultimate
carcinogenic form of benzo(a)pyrene (Levin et jiK, 1977).
The patterns of metabolites of benzo(a)pyrene produced by scup and
winter flounder have not been established. However, preliminary studies
(Stegeman and Tjessem, unpublished) have indicated that little or no
benzo(a)pyrene-4,5-dihydrodiol is formed j_n vitro by scup liver microsomes,
whereas substantial amounts of 7,8-dihydrodiol and 9,10-dihydrodiol are
formed. At the same time these studies confirm that epoxide hydrase is
present in fish liver (Bend et j*l_., 1977) and indicate that benzo(a)pyrene-
4,5-epoxide is not responsible for mutation induced at least by the scup
preparations. Further, formation of isomeric benzo(a)pyrene-7,8-diol-9,10-
epoxide would be possible, although it is not known if any isomers of this
metabolite would be preferentially formed.
The discrepancy between levels of benzo(a)pyrene hydroxylase activity
and the activation of benzo(a)pyrene by the fish we studied indicates that
metabolite patterns formed by these animals probably differ, although it is
recognized that factors other than catalytic function of cytochrome P-450
may influence mutagenic activity detected using a PMS preparation
(Ames et jjl_., 1975). Yet the formation of mutagenic and carcinogenic
diol-epoxides of polynuclear aromatic hydrocarbons by both species and
sexes is a distinct possibility. It is known that patterns of metabolites
formed by microsomes and intact cells can differ (Selkirk, 1977), but it is
likely that toxic and mutagenic derivatives similar to those formed j_n
vitro can result from metabolism in vivo in these and other (Ahokas et a!.,
1977) fish. Thus, the results clearly suggest that marine fish may be at
risk to carcinogenic activity of polynuclear aromatic hydrocarbons known to
be present in recent coastal marine sediments (Laflamme et a\_., 1978) and
presumably coastal waters.
ACKNOWLEDGEMENTS
This research was supported by NSF Grant OCE 77-24517 (IDOE) and Sea
Grant No. 04-6-158-44106. Technical assistance was provided by J. Seixas,
A. Sherman, and B. Penman. T. Skopek is a predoctoral trainee of the
National Institute of Environmental Health Sciences.
208
-------
REFERENCES
Ahokas, H.T., 0. Pelkonen, and N.T. Karki. 1975. Metabolism of polycyclic
hydrocarbons by a highly active aryl hydrocarbon hydroxylase system in
the liver of a trout species. Biochem. Biophys. Res. Commun.
63:635-641.
Ahokas, J.T., 0. Pelkonen, and N.T. Karki. 1977. The possible role of
trout liver aryl hydrocarbon hydroxylase in activating aromatic
polycyclic carcinogens. In: Biological reactive intermediates:
Formation, toxicity and inactivation. D.J. Jollow, J.J. Kocsis, R.L.
Snyder, and H. Vainio, Eds., Plenum Press, New York. pp. 162-166.
Ames, B.N., J. McCann, and E. Yamasaki. 1975. Methods for detecting
carcinogens and mutagens with the salmonella-mammalian-microsome
mutagenicity test. Mutat. Res. 31:347-364.
Bend, J.R., M.O. James, and P.M. Dansette. 1977. In vitro metabolism of
xenobiotics in some marine animals. Ann. New York. Acad. Sci.
298:505.
Chevion, M., J.J. Stegeman, J. Peisach, and W.E. Blumberg. 1977. Electron
paramagnetic resonance studies on hepatic microsomal cytochrome P-450
from a marine teleost fish. Life Sci. 20:895-900.
Hecht, S.S., E. LaVoie, R. Mazzarese, S. Amin, V. Bedenko, and D. Hoffman.
1978. l,2-dihydro-l,2-dihydroxy-5-methylchrysene, a major activated
metabolite of the environmental carcinogen 5-methylchrysene. Cancer
Res. 38:2191-2194.
Huberman, E., L. Sachs, S.K. Yang, and H.V. Gelboin. 1976. Identification
of mutagenic metabolites of benzo(a)pyrene in mammalian cells. Proc.
Nat. Acad. Sci. U.S. 73:607-612.
Laflamme, R.E. and R.A. Hites. 1978. The global distribution of
polycyclic aromatic hydrocarbons in recent sediments. Geochim.
Cosmochim. Acta 42:289-303.
Levin, W., D.R. Thakker, A. W. Wood, R.L. Chang, R.E. Lehr, D.M. Jerina,
and A.M. Conney. 1978. Evidence that benzo(a)anthracene
3,4-diol-l,2-epoxide is an ultimate carcinogen on mouse skin. Cancer
Res. 38:1705-1710.
Levin, W., A.W. Wood, R.L.Chang, T.J. Slaga, H. Yagi, D.M. Jerina, and A.M.
Conney. 1977. Marked difference in the tumor-initiating activity of
optically pure (+)- and (-)-trans-7,8-dihydroxy-7,8-dihydrobenzo(a)-
pyrene on mouse skin. Cancer Res. 37:2721-2725.
209
-------
Levin, W., A.W. Uood, A.Y.H. Lu, D. Ryan, S. West, A.M. Conney, D.R.
Thakker, H. Yagi, and D.M. Jerina. 1976. Role of purified cytochrome
P-448 and epoxide hydrase in the activation and detoxification of
benzo(a)pyrene. In: Drug metabolism concepts. D. M. Jerina, Ed.,
American Chemical Society Symposium Series 44, Washington, DC.
Lowry, O.H., N.J. Rosenbrough, A.L. Farr, and R.J. Randall. 1951. Protein
measurement with the folin phenol reagent. J. Biol. Chem.
193:265-275.
Pohl, R.J., J.R. Bend, A.M. Guarino, and J.R. Fouts. 1974. Hepatic
microsomal mixed-function oxidase activity of several marine species
from coastal Maine. Drug Metab. Dispos. 2:545-555.
Selkirk, J.K. 1977. Divergence of metabolic activation systems for
short-term mutagenesis assays. Nature 270:604-607.
Skopek, T.R., H.L. Liber, J.J. Krolewski, and W.G. Thilly. 1978a.
Quantitative forward mutation assay in Salmonella typhimurium using
8-azaguanine resistance as a genetic marker. Proc. Natl. Acad. Sci.
75:410-414.
Skopek, T.R., H.L. Liber, D.A. Kaden, and W.G. Thilly. 1978b. Relative
sensitivities of forward and reverse mutation assays in Salmonella
typhimurium. Proc. Natl. Acad. Sci. 75:4465-4469.
Slaga, T.J., W.M. Bracken, A. Viaje, W. Levin, H. Yagi, D.M. Jerina, and
A.H. Conney. 1977. Comparison of the tumor-initiating activities of
benzo(a)pyrene arene oxides and diol-epoxides. Cancer Res.
37:4130-4133.
Stegeman, John J. 1977. Fate and effects of oil in marine animals.
Oceanus 20:59-66.
Stegeman, J.J., and R.L. Binder. 1979. High benzo(a)pyrene hydroxylase
activity in the marine fish Stenotomus versicolor. Biochem.
Pharmacol. 28:1686-1689.
Walker, C.H. 1978. Species differences in microsomal mono-oxygenase
activity and their relationship to biological half-lives. Drug.
Metab. Rev. 7:295-323.
Weibel, F.J., J.C. Leutz, L. Diamond, and H.V. Gelboin. 1971. Aryl
hydrocarbon [benzo(a)pyrene] hydroxylase in microsomes from rat
tissues: differential inhibition and stimulation by benzoflavones and
organic solvents. Arch. Biochem. Biophys. 144:78-86.
Wislocki, P.G., A.W. Wood, R.L. Chang, W. Levin, H. Yagi, 0. Hernandez,
D.M. Jerina, and A.H. Conney. 1976. High mutagenicity and toxicity
of a diol epoxide derived from benzo(a)pyrene. Biochem. Biophys. Res.
Commun. 68:1006-1012.
210
-------
Wislocki, P.G., A.W. Wood, R.L. Chang, W. Levin, H. Yagi, 0. Hernandez,
P.M. Dansette, D.M. Jerina, and A.M. Conney. 1976. Mutagenicity and
cytotoxicity of benzo(a)pyrene arene oxides, phenols, quinones, and
dihydrodiols in bacterial and mammalian cells. Cancer Res.
36:3350-3357.
Wood, A.W., P.G. Wislocki, R.L. Chang, W. Levin, A.Y.H. Lu, H. Yagi, 0.
Hernandez, D.M. Jerina, and A.H. Conney. 1976. Mutagenicity and
cytotoxicity of benzo(a)pyrene benzo-ring epoxides. Cancer Res.
36:3358-3366.
Yang, S.K., D.W. McCourt, P.P. Roller, and H.V. Gelboin. 1976. Enzymatic
conversion of benzo(a)pyrene leading predominantly to the diol-epoxide
r-7,t-8-dihydroxy-t-9,10-oxy-7,8,9,10-tetrahydrobenzo(a)pyrene through
a single enantiomer of r-7,6-8-dihydroxy-7,8-dihydrobenzo(a)pyrene.
Proc. Natl. Acad. Sci. U.S. 73:2594-2598.
211
-------
HYPERSENSITIVITY FOR CARCINOGENESIS RESULTING FROM SPECIES HYBRIDIZATION:
IMPAIRING CONTROL OF CELLULAR ONCOGENES AS TOOL TOWARDS TAILORING TEST
ANIMALS SUITABLE FOR MONITORING CARCINOGENS
by
Manfred Schwab , Safia S. Abdo, Gerhard Kollinger
Genetisches Institut der Justus-Liebig-Universitaet Giessen
Heinrich-Buff-Ring 58-62, D-6300 Giessen, FRG
ABSTRACT
Experimental strategies of selective breeding and
mutagenesis/carcinogenesis have unveiled genes transmitted
through the germ line in development of tumors of apparently
non-viral etiology in the freshwater fish, Xiphophorus. It
seems practicable to construct genotypes in which control of
expression of cancer genes is impaired, although not
abolished—a condition that renders them hypersensitive for
carcinogens.
INTRODUCTION
The degree of genomic contributions to oncogenesis has been debated
for some time. The most general genetic concept was proposed by Comings
(1973), who postulated the existence of two classes of genes relevant to
oncogenesis: the transforming gene (Tr) and the regulating gene (Rj
influencing expression of Jr. Oncogenesis in his model is thought to
result from misguided expression of Tr, due to aberrant function of R. The
recent molecular unveiling of cellular oncogenes in the genome of animal
species (Bishop, 1978), mainly using retroviruses as experimental tools
(Bishop, 1982; Varmus, 1982), and the finding of their likely involvement
in oncogenesis in animals (Hayward «rt afU, 1981; Payne et ^1_., 1982), and
possibly in humans (Cooper et al_., 1980; Shih^t^U, 1981; Perucho et al..
1982), suggests that aberrant expression of cellular oncogenes transmitted
through the germ line may at least be one cause of malignancy.
Present address: Department of Microbiology and Immunology
School of Medicine,
University of California, San Francisco, CA 94143
212
-------
Besides these recent molecular approaches to the role of what we call
today oncogenes, the more classical experimental strategies of formal
genetics have unveiled much earlier the existence of genes with the
potential to elicit cancer ("cancer genes") in the genome of apparently
normal organisms. Among others, hybrid tumors in plants [e.g., Nicotiana
(Kostoff, 1930; Datura (Satina et _§].., 1950; Sorghum (Lin, 1969)] and in
animals [e.g. carp (Sonstegard, 1977); duck (Crew and Koller, 1936);
Sinclair swine (Millikan et jil_., 1974); xiphophorine fish (Gordon, 1959);
Drosophila (Gateff, 1978)"]~as well as apparently genetic tumors in humans
[e.g., neurobastoma (Knudson ert _§]_., 1971); Sparkes ^t a]_., 1979),
polyposis of the colon (Lynch, 1967); melanoma (Anderson, 1971)] are most
pertinent examples. Of these, the xiphophorine fish system may be regarded
as one of the genetically best characterized models (for recent discussion,
see Schwab, 1982a). With the knowledge that the presence of cellular genes
elicits cancer in these fish, henceforth referred to as "oncogenes" or
"cancer genes," our paper aims to evaluate the possibility of constructing
hypersensitive genotypes particularly suitable for the detection of
carcinogens in the water. We shall first give a brief overview on genetics
of hybrid malignancy in the freshwater fish, Xiphophorus, review some of
the results obtained with the primary'carcingoen N-methyl-N-nitrosourea
(MNU) as model agent, and eventually discuss some considerations for
setting up a suitable test system.
GENES INVOLVED IN MALIGNANCY
Genes involved in malignancy were identified essentially by two
experimental strategies: interspecies hybridization and carcinogenesis
studies.
!_. Interspecies hybridization
Gordon (1927), Kosswig (1927), and Haeussler (1927) observed after
crossing procedures (Figure 1) that malignant pigment cell tumors
("melanoma") developed in backcross hybrids. Through subsequent studies
today, at least four kinds of genetic loci involved in determining
malignancy in Xiphophorus can be recognized and shall be subsequently
discussed. It should be kept in mind, though, that their identification is
soley based on formal genetics because nothing is known about their
function, their products, and their architecture; the terminology used here
is regarded only as operational. At least four loci contributing to the
scenario of oncogenesis were identified: macromelanophore determinator
(Mel). differentiated state (Diff), golden (gj and albino (a).
Mel
The simplest dermal pigment cell phenotype of wildtype Xiphophorus
consists of melanophores, pterinophores, and guanophores distributed more
or less uniformly. Within the melanophores, basically two types may be
recognized: a small melanophore usually not exceeding 100 um in diameter
("micromelanophore"), and a large giant melanophore usually approximating
1 mm ("macromelanophore"; Figure 2a). Little hard data exist on the
213
-------
X. macu/atus
Sd/Sd
Figure 1. Experimental strategy of interspecies hybridization to identify
cancer genes in the germ line; 25% of the backcross hybrids
develop malignant melanoma resulting from introduction of the
mel-locus Sd^ of X. maculatus into the genetic environment of X.-
helleri.
214
-------
w& >% ^,
igure 2. Phenotypes encoded by macromelanophore loci.
a. Colonies of macromelanophores determined by the 1
(spots in dorsal fin) and Sp_ (spots on body side).
b. Melanoma in ]3C-hybrid determined by mel-locus Sd.
c. Section through Sd-melanoma.
oci
215
-------
differentiation of the two melanophores and on their interrelation despite
recent speculation (Anders jrt a]_., 1978). It seems established, though,
that they are derived from the neural crest (Humm and Young, 1956). A
general idea for their development would be that at some stage during
differentiation two cell lineages are being determined of which one
eventually yields micromelanophores, the other macromelanophores. The
genetic determinator inducing macromelanophore development might actually
be represented by what is referred to as mel.
Our report deals mainly with five mel -loci that have been described
earlier in detail (Gordon, 1959; Anders jit j»l_., 1973a; Kallman, 1975): the
X-chromosomal spotted dorsal (Sd) and spotted (Sp; Figure 2a) and the
Y-linked stripe sided (Sr; Figure 6), all derived from Xiphophorus
maculatus, and the X-linked lineatus (Li) and the Y-linked punctatus (Pu)
derived from X_. variatus (not shown).
Diff
Diff acts on the expressivity of mel and modulates the pigment cell
phenotype. Genetic analyses unveiled this autosomal locus in_X. maculatus.
Homozygosity for Diff is associated with the development of terminally
differentiated macromelanophores (Figure 2a). A single dose of Diff tends
to inhibit the process of terminal differentiation leading to an increase
of the ratio of incompletely versus terminally differentiated pigment
cells. Lack of Diff causes the majority of the pigment cells to remain in
an incompletely differentiated state, predominantly as melanocytes and
melanoblasts, a condition that eventually leads to a massive increase in
the number of pigment cells and to formation of a melanoma (Figure 2b,c).
Diff and mel are, as far as the loci j>d and $£ are concerned, not
chromosomally linked. A marker gene esterase-1, coding for an
electrophoretically identifiable product (Est-1), was identified to be
chromosomally linked to Diff, thus allowing identification of the presence
of Diff in any given genotype (Siciliano and Wright, 1976; Ahuja et al..
1980; Schwab and Scholl, 1981; Schwab, 1982b; Figure 3). For the mel -loci
Sr, Li, and £u a Diff -locus has not been identified. Crossing procedures
that lead to melanoma in case of the mel -loci Sd[ and S£ generally fail to
produce melanoma in case of Sr, Li, and Pu, except for rare cases, in which
melanoma is produced in L[ -genotypes. Mutagenesis may lead to a malignant
phenotype, however, and it has been concluded that a mel -linked cis acting
controlling element has been impaired by somatic mutation, releasing me!
from negative control, perhaps exerted by a gene functionally homologous to
Diff (Anders £t a]_., 1973a,b). It may not be excluded, however, that
positive control mechanisms also may operate in these cases, such as
chromosome rearrangements leading to a configuration, in which mel is
positioned in the vicinity of a transcriptionally active region ("insertion
mutagenesis" in its broadest sense). The mechanism of Diff action, as that
of mel_, and their interaction in leading to the malignant phenotype, have
not been defined. It may not be excluded that Diff represents a locus
involved directly in differentiation of mel -determined macromelanophores,
as hypothesized earlier (Viel kind, 1976), but it might also be envisioned
216
-------
that it could affect the pigment cell phenotype indirectly, e.g., by
interfering with other genes conferring inhibition of pigment cell
differentiation, or as ,a imitator gene. As it is, Diff seems to represent
an important locus in regulating mel expressivity. Its property of not
being linked to mel renders hybridization of the corresponding mel -carrying
genotypes with others lacking Diff a powerful experimental tool of
producing melanoma in a Mendelian fashion.
The effect of the autosomal recessive locus ^ can be best studied in
%_• maculatus. Animals homozygous for £ virtually lack dermal melanophores,
except for a few (Figure 4). A feasible hypothesis for the lack of
melanophores is that differentiation of the dermal X_. maculatus
melanophores is blocked at an early stage, possibly at the stage of the
chromatoblast, which is the common precursor for all dermal pigment cells.
Guanophores and pterinophores are produced at normal level. Consequently,
in jjg_ individuals carrying mel, elimination of Diff fails to produce
melanoma. Recent studies indicate that their inhibition of differentiation
can be overcome by treatment with tumor promoters, such as
12-0-tetradecanoylphorbol-13-acetate (TPA; Schwab, 1982b), indicating
their potential as test organisms for tumor promoting agents. Genetic
studies have furthermore shown that macromelanophores determined by the
various mel -loci differ: while £ usually completely abolishes
differentiation of macromelanophores determined by j>d_, macromelanophores
determined by other mel -loci, e.g., Lj_, differentiate to a normal extent,
although micromelanophores are still lacking.
The autosomal recessive locus a^ is pleiotropic in the sense that it
both confers the albinotic phenotype and exerts a stimulating effect on the
degree of malignancy: pigment cells in albinotic melanomas show a lower
degree of differentiation and a higher rate of division than in melanotic
melanoma (Vielkind, 1976). Furthermore, carcinogen-induced neruoblastoma
(Schwab et a]_., 1979) is considerably more malignant in albinotic than in
wildtype animals. The mechanism for albinism is unclear, but aa_ -animals
are tyrosinase positive (Vielkind, 1976; Schwab, 1982b).
2. Carcinogenesis studies
a. Hypersensitivity of interspecies hybrids
Studies with various carcinogens formerly used the majority of the
non-hybrids of Xiphophorus available. No tumors were detected. However,
recently, the nitrosamide N-methyl-N-nitrosourea (MNU) was employed in
these studies for the following considerations. In general two types of
chemical carcinogens can be recognized with respect to their mode of action:
indirect acting and direct acting carcinogens. Indirect acting carcinogens,
in order to become active, require metabolism by host enzymes for conversion
to their active forms (Miller, 1970; Lijinsky, 1976; Bridges, 1976).
217
-------
-X
-X
- Est-1
- Esl-1
• Est-1
-X
- EsH
-EsH
Figure 3. Association of terminally differentiated macromelanophores with
Est-1 marked chromosome of X. maculatus. Presence of Est-1
chromosome favors terminal differentiation, lack yields melanoma
(according to Schwab and Scholl, 1980).
l-'igure 4. Effect of locus _g_ on pigment cell phenotype. In jjg_ animals
carrying the me!-locus Sd, differentiation of dermal
melanophores is blocked at an early stage. Few
macromelanophores may develop, and retinal pigment cells are
formed to normal extent. Melanoma does not occur in
BC-hybrids due to differentiation block.
218
-------
Consequently, in experiments using indirect acting carcinogens to analyze
malignancy at least two genotype dependent variables must be taken into
account: (1) enzymatic activation that may vary considerably in the
different genotypes (Neubert, 1974), and (2) the susceptibility of the
genotype to develop a tumor following reaction of the activated carcinogen
with the target molecule in the cell. In addition, the enzymes involved in
activation of the carcinogen may be present only in certain tissues and not
in others due to differential gene activity, which may result in
organotropic activity of the carcinogen. This kind of organotropy is due
to organ-specific activation of the carcinogen, and not to the
susceptibility of the particular genotype to respond to the ultimate
carcinogen with development of cancer.
Figure 5. Effect of locus a_ on pigment cell phenotype. Pigment synthesis
is abolished in aa_ animals, but amelanonic pigment cells
develop. Degree of malignancy of melanoma in ^-hybrids is
increased over that in wildtype animals.
In contrast, direct acting carcinogens undergo conversion to the
ultimate carcinogen spontaneously. They may react instantly with the
target molecules. Thus, by exposure to a direct acting carcinogen, the
susceptibility of a genotype to a carcinogen should be tested directly. It
is quite obvious that although indirect acting carcinogens are more common
in the environment, direct acting carcinogens are more suitable for
analyzing the genetic basis of susceptibility in an experimental system.
The nitrosamide N-methyl-N-nitrosourea (MNU) appears to be
particularly suitable. First, MNU is a very potent direct acting
carcinogen (Druckery et .aj_., 1965; Lijinsky, 1976; Narisawa et _al_., 1976).
Second, MNU data suggest that mutation is the primary event that induces
transformation of normal into malignant cells (Bouck and Mayorca, 1976).
Third, MNU penetrates all tissues of the animal (Magee _et jfl_., 1975).
The recent carcinogenesis study was extended beyond non-hybrids to Fl
and backcross hybrids (BC) that have a spontaneous rate of tumor incidence
below level of detection. The experimental strategy, aiming to identify
chromosomes conferring hypersensitivity, was based on the following
considerations.
-------
X. better/ -/-
X. maculatus Sr I Sr
v"
X. /7e//er/ -/-
BC
Srl-
BC -I-
Figure 6. Experimental strategy of combined selective breeding and
mutagenesis/carcinogenesis to identify cancer genes in the germ
line. A genetically defined chromosome (in this case the
Y-chromosome marked by the mel-locus Sr) of one species
U. maculatus) is introduced by selective breeding into the
genetic milieu of another species (X^. helleri). BC-segregants
differ with their gene pool only by the Sr-chromosome.
Segregants are treated with mutagens/carcinogens and their
susceptibility is compared. The ^-chromosome in this case
confers hypersensitivity for melanoma, -/- segregants are
largely resistant.
220
-------
The various species of Xiphophorus are phenotypically characterized by
the presence of spot patterns consisting of melanophores, guanophores, and
pterinophores. The corresponding pigment cell loci are expressed
codominantly, and the patterns can be used therefore for identifying
chromosomes in any hybrid genotype and, furthermore, as convenient
phenotypic markers for selective breedings (for details on the spot
patterns see Gordon, 1959; Anders ^t al_., 1973a,b; Kallman, 1975). The
basic strategy, displayed in the crossing scheme in Figure 6, was to
introduce defined chromosomes from one species into the genetic environment
of another by crossing and backcrossing. The BC_ segregates with respect to
the marker chromosome, and the gene pool of the ]3C-segregants differs only
with respect to this chromosome (for higher ^-generations the statistics
is better than for lower ones).
The result of the previous study confirmed that nonhybrids proved to
be completely resistant (Schwab and Anders, 1981). In contrast, Fl hybrids
showed slight sensitivity and developed melanoma (0.6% incidence), while
backcross hybrids responded to MNU-treatment with the development of a
large spectrum of tumors, including melanoma, neuroblastoma, carcinomas,
fibrosarcoma, rhabdomyosarcoma, and lymphosarcoma (Schwab and Anders,
1981). By far most of the tumors were represented by melanoma (in 254 of
about 5100 fish treated; Figure 7), followed by neuroblastoma (66 cases;
Figure 8), fibrosarcoma (50 cases; Figure 9), and carcinoma (12 cases;
Figure 10). Other tumors, as the rhabdomyosarcoma (Figure 11) were rare (1
safe case).
Figure 7. Melanoma induced by MNU.
Sensitivity in the j3C-hybrids was not distributed at random. Instead,
certain hybrids showed hypersensitivity for certain tumors with incidence
up to 15% under the experimental conditions. Appearence of some tumors
could be assigned to genetically defined chromosomes, particularly the
neuroblastoma in hybrids derived from _X. variatus - helleri associated with
the J_i_ -chromosome of X^. variatus (Schwab jrt ^1_., 1979), while the
fibrbTarcoma in the same hybrids seems to be associated with an autosome
(Schwab €rt _§]_., 1978). Melanoma in hybrids derived for _X. maculatus -
helleri is associated predominantly with the me! -locus Sr, and the
X-chromosome defined by alleged deletion of Sd~TSchwab and Scholl, 1981).
221
-------
ffc .-•...*•. 'A- --;*
• X
* -
. V i
.
.-
• -
kv
•
Figure 8. Neuroblastoma induced by MNU (from Schwab et^ ^T_., 1979).
a. Fish; b,c. Section; d. EM-section. Insert shows typical
cilium with 9=0 pattern of microtubles.
222
-------
Figure 9. Fibrosarcoma induced by MNU (from Schwab et ^1_., 1978).
a. Fish
b. Section showing invasion into muscle tissue.
223
-------
Figure 10. Carcinoma induced by MNU
a. Fish
b. Section
-------
Figure 11. Rhabdomyosarcoma induced by MNU (from Schwab et al_., 1978)
a. Fish
b,c. Section
Z25
-------
In summary the results of our study indicate that interspecies
hybridization obviously leads to impairment of control of cellular
transformation, a condition that confers sensitivity for induction of
malignancy by carcinogens. Hypersensitivity is associated with genetically
defined marker chromosomes.
b. Genetic modification of MNU-induced melanoma
As in spontaneous hybrid melanoma, two phenotypes could be recognized
in the MNU-induced melanoma: one benign and the other malignant (Schwab
and Scholl, 1981). We attempted to examine whether the degree of was
malignancy associated with Diff. An easy handle for testing this
possibility was to monitor for the Diff -linked Est-1. We found that an
association between the benign phenotype and presence of Est-1 and the
malignant phenotype and absence of Est-1 could be established (as in Figure 3),
thus making an involvement of Diff in the control of the MNU-induced
pigment cell phenotype likely (Schwab and Scholl, 1981; Schwab, 1982b).
SIGNIFICANCE - POSSIBILITIES - POTENTIALS
Experimental carcinogenesis in small aquarium fishes has been carried
out with several species and a variety of chemical carcinogens (Matsushima
and Sugimura, 1976). In recent years, it has been recognized that small
aquarium fishes offer some advantages for carcinogenesis studies: mainly,
fish are more sensitive to carcinogens and, at the same time, are more
resistant to toxic effects than are rodents (Matsushima and Sugimura,
1976). In addition, large numbers of animals can be raised and maintained
easily. Our investigation revealed that a large variety of tumors can be
induced by treatment with MNU in Xiphophorus. Most have been observed to
develop spontaneously in fishes (Ashley _e_t j»]_., 1979; Mawdesley-Thomas,
1975; Harshbarger, 1973). However, they apparently have not been observed
thus far in tests with carcinogens (Matsushima and Sugimura, 1976).
Species differences in the sensitivity to respond with tumor
development to carcinogen treatment have been reported. For instance,
different inbred strains of mice were found to vary in their response to
skin carcinogenesis (Berenblum, 1974). Boxer dogs are particularly
susceptible to tumors following treatment with MNU (Denlinger et a]_., 1978)
and show also a much higher spontaneous tumor incidence than other breeds
(Cohen et al., 1974). Further, in experiments with the guppy (Lebistes
reticulatusT. hepatic tumors were induced by nitrosamines (Sato et al.,
1973; Pliss and Khudoley, 1975), whereas Scherf (1976) showed that the same
species, although most likely another strain, was resistant. All these
results, including our investigation, clearly show the involvement of
genetic factors in susceptibility and urge the use of highly defined
genotypes in carcinogenesis studies.
Differences in susceptibility are often attributed to variations in
the rate of generation of the ultimate carcinogen. For our study this
explanation is rather unlikely: several genotypes found to be sensitive or
resistant to MNU proved to be also sensitive or resistant to X-rays,
226
-------
particularly in the case of the melanoma and the neuroblastoma (Schwab jrt
al., 1979; Schwab and Anders, 1981). Such parallels in the sensitivity
spectrum make it likely that, in short, genetic factors operate directly in
determining sensitivity and resistance.
The fact that hypersensitive genotypes can be constructed from
genotypes resistant to cancer by selective matings gives further support to
the idea that one or several genetic changes are involved in the etiology
of cancer. In the present case of hypersensitivity, apparently new
combination of chromosomes, as well as possibly mutational events, could
resu-lt in, using the terminology of modern molecular genetics, enhancement
of expression of a gene with the potential of triggering malignancy.
Basically two molecular mechanisms could be envisioned. One is impairment
of negative control as proposed earlier (Schwab £t _a]_., 1979; Schwab and
Anders, 1981); the other is chromosomal rearrangement setting up a positive
control mechanism, e.g. insertion mutagenesis in its broadest sense,
resulting in placement of the corresponding oncogene upstream or
downstream in the vicinity of a promoter sequence, as found in chickens
using retroviruses as mutagens (Varmus, 1982; Bishop, 1982).
The molecular nature of the oncogenes identified in Xiphophorus by
genetic strategies remains to be elucidated. It is unlikely from the
present data that any one of the homologues of retrovirus transforming
genes identified by molecular studies in the germ line of Xiphophorus
(manuscript in preparation) is involved in oncogenesis. As it is,
construction of genotypes by selective breeding carrying several oncogenes
conferring hypersensitivity should be a powerful tool for creating suitable
test organisms and at the same time may provide further clues towards an
understanding of cancer genes in malignancy.
ACKNOWLEDGEMENTS
This investigation was supported by Deutsche Forschungsgemeinschaft
(Schw 251/1), Sonderforschungsbereich 103, Marburg, and Land Hessen.
Manfred Schwab is Heisenberg-Fellow of Deutsche Forschungsgemeinschaft.
Safia S. Abdo was on leave from University of Alexandria, supported by
Ministry of Education in Eygpt. The paper is dedicated to the memory of
Curp Kosswig.
227
-------
REFERENCES
Ahuja, M.R., M. Schwab, and F. Anders. 1980. Linkage between a regulatory
locus for melanoma cell differentiation and on esterase locus in
Xiphophorus. J. Heredity 71:403-407.
Anders, A., F. Anders, and K. Klinke. 1973a. Regulation of gene
expression in the Gordon-Kosswig melanoma system. I. The
distribution of the controlling genes in the genome of the
xiphophorine fish, Platypoecilus maculatus and Platypoecilus variatus.
In: Genetics and mutagenesis of fish. J.H. Schroeder, Ed.,
Springer-Verlag, Berlin, Heidelberg, New York. pp. 33-52.
Anders, A., F. Anders, and K. Klinke. 1973b. Regulation of gene
expression in the Gordon-Kosswig melanoma system. II. The
arrangement of chromatophore determining loci and regulation elements
in the sex chromosomes of xiphophorine fish, Platypoecilus maculatus
and Platypoecilus variatus. In: Genetics and mutagenesis of fish.
J.H. Schroeder, ED., Springer-Verlag, Berlin, Heidelberg, New York.
pp. 53-63.
Anders, F., H. Diehl, M. Schwab, and A. Anders. 1979. Contributions to an
understanding of the cellular origin of melanomas in the
GORDON-KOSSWIG xiphophorine fish tumor system. In: Pigment cell.
S. Klaus, Ed., Karger, Basel, pp. 142-149.
Anders, F. 1981. Erb- und Umweltfaktoren im Ursachengefuege des
neoplastischen Wachstums nach Studien an Xiphophorus. Klin.
Wochenschr. 59:943-956.
Anderson, D.E. 1971. Clinical characteristics of the genetic variety of
cutaneous melanoma in man. Cancer 28:721-725.
Aronowitz, 0., R.F. Nigrelli, and M. Gordon. 1951. A spontaneous
epithelioma in the platyfish, Xiphophorus, Platypoecilus variatus.
Zoologica 36:239-241.
Ashley, L.M. 1969. Experimental fish neoplasia. In: Fish in research.
O.W. Neuhaus and J.E. Halver, Eds., Academic Press, New York, London.
pp. 23-43.
Ashley, L.M., J.E. Halver, and S.R. Wellings. 1969. Case reports of three
teleost neoplasms. Natl. Cancer Inst. Monogr. 31:157-165.
Bishop, J.M. 1978. Retroviruses. Ann. Rev. Biochem. 47:35-88.
Bishop, J.M. 1981. Enemies within: the genesis of retrovirus oncogenes.
Cell 23:5-6.
Bishop, J.M. 1982. Oncogenes. Scientific American 246:80-92.
228
-------
Bouck, N., and G. Mayorca. 1976. Somatic mutation as the basis for
malignant transformation of BHK cells by chemical carcinogens.
264:722-727.
Bridges, B.A. 1976. Short-term screening tests for carcinogens. Nature
261:195-200.
Cohen, D., J.S. Reif, R.S. Brodey, and H. Keiser. 1974. Epidemiological
analysis of the most prevalent sites and types of canine neoplasia
observed in a veterinary hospital. Cancer Res. 34:2859-2862.
Comings, D.E. 1973. A general theory of carcinogenesis. Proc. Natl.
Acad. Sci. USA 70:3324-3328.
Cooper, G.M., S. Okenquist, and L. Silverman. 1980. Transforming
activity of DNA of chemically transformed and normal cells. Nature
284:418-421.
Crew, F.A.E., and P. Koller. 1936. Genetical and cytological studies on
the inter-generic hybrid of Cairina moschata and Anas platyrhynchos.
Proc. R. Soc. Edinb. 56:210-220.
Denlinger, R., A. Koestner and J.A. Swenberg. 1978. Neoplasms in purebred
boxer dogs following long term administration of N-methyl-N-nitrosourea.
Cancer Res. 38:1711-1717.
Druckrey, H., S. Ivankovic, and R. Preussmann. 1965. Selektive Erzeugung
maligner Tumoren im Gehirn und Ruckenmark von Ratten durch
N-Methyl-N-Nitrosoharnstoff. Z. Krebsforsch. 66:389-408.
Gateff, E. 1978. Malignant neoplasms of genetic origin in Drosophila
melanogaster. Science 200:1448-1459.
Gordon, M. 1927. The genetics of a viviparous top-m\nnow,
Platypoecilus. The inheritance of two kinds of melanophores.
Genetics 12:253-283.
Gordon, M. 1948. Effects of five primary genes on the site of the
melanoma in fishes and the influence of two color genes on their
pigmentation. In: The biology of melanomas. M. Gordon, Ed., New
York Academy of Science 4:216-268.
Gordon, M. 1959. The melanoma cell as an incompletely differentiated
pigment cell. In: Pigment cell biology. M. Gordon, Ed., Academic
Press, New York, pp. 215-239.
Haussler, G. 1928. Uber Melanombildung bei Bastarden von Xiphophorus
helleri und Platypoecilus maculatus var. Rubra. Klin. Wschr.
7:1561-1562.
229
-------
Harshbarger, J.C. 1973. Activity report. Registry of tumors in lower
animals, 1965-1973. National Museum of Natural History, Washington,
DC.
Hayward, U.S., B.G. Neel, and S.M. Astrin. 1981. Activation of cellular
•oncogene by promoter insertion in ALV-induced lymphoid leucosis.
Nature 290:475-480.
Humm, D.C., and R.S. Young. 1956. The embryological origin of pigment
cells in platyfish-swordtail hybrids. Zoologica 41:1-10.
Kallman, K.D. 1975. The platyfish, Xiphophorus maculatus. In: Handbook
of genetics. R.C. King, Ed., Plenum Press, New York, London.
4:81-132.
Knudson, A.G. 1973. Mutation and cancer. Adv. Cancer Res. 17:317-352.
Kosswig, C. 1927. Ueber bastarde der telostier Platypoecilus and
Xiphophorus. Z. indukt. Abstamm. u. Vereb.-Lehre 44:253.
Kostoff, D. 1930. Tumors and other malformation on Nicotiana hybrids.
Zbl. Bakt. Parasitenk. Abt. II 81:244-260.
Lijinsky, W. 1976. Interaction with nucleic acids of carcinogenic and
mutagenic N-nitroso compounds. Progr. Nucleic Acid Res. Mol. Biol.
17:247-269.
Lin, P.S., and J.G. Ross. 1969. Ovular tumors in a trisomic Sorghum
plant. J. Heredity 60:183-185.
Lynch, H.T. 1967. Hereditary factors in carcinoma. Springer-Verlag,
Berlin, Heidelberg.
Magee, P.N., A.E. Pegg, and P.P. Swann. 1975. Molecular mechanisms of
chemical carcinogenesis. In: Handbuch der Allgemeinen pathologic.
E. Grundmann, Ed., Springer-Verlag, Berlin, Heidelberg, New York.
6:328-419.
Matsushima, T., and T. Sugimura. 1976. Experimental carcinogenesis in
small aquarium fishes. Prog. Exp. Tumor Res. 20:367-379.
Mawdesly-Thomas, L.E. 1975. Some diseases of muscle. In: The pathology
of fishes. W.E. Ribelin and G. Migaki, Eds., University of Wisconsin
Press, Madison, WI. pp. 343-363.
Miller, J.A. 1970. Carcinogenesis by chemicals: an overview. Cancer
Res. 30:559-576.
230
-------
Millikan, I.E., J.L. Boylon, R.R. Hock, and P.J. Manning. 1974. Miniature
Swine = Sinclair Swine. Melanoma in Sinclair Swine: a new animal
model. J. Invest. Dermatol. 62:20-30.
Narisawa, T., C. Wong, R.R. Maronpot, and J.H. Weisburger. 1976. Large
bowel carcinogensis in mice and rats by several intrarectal doses of
methylnitrosourea and negative effects of nitrite plus methyl urea.
Cancer Res. 36:505-510.
Neubert, D. 1974. The toxicological evaluation of mutagenic events.
Mutat. Res. 25:145-157.
Nigrelli, R.F. and M. Gordon. 1951. Spontaneous neoplasms in fishes. V.
Acinar adenocarcinoma of the pancreas in hybrid platyfish. Zoologica
36:121-126.
Payne, G., S.A. Courtneidge, L.B. Crittenden, A.M. Fadly, J.M. Bishop, and
H.E. Varmus. 1981. Analysis of avian leukosis virus DNA and RNA in
bursal tumors: viral expression is not required for maintenance of
the tumor state. Cell 23:311-322.
Perucho, M., M. Goldfarb, K. Shimizu, C. Lama, J. Fogh, and M. Wigler.
1981. Human-tumor-derived cell lines contain common and different
transforming genes. Cell 27:467-476.
Pliss, G.B., and V.V. Kudoley. 1975. Tumor induction by carcinogenic
agents in aquarium fish. J. Natl. Cancer Inst. 55:129-136.
Satina, S.J., J. Rapoport, A.F. Blakeslee. 1950. Ovular tumors connected
with incompatible crosses in Datura. Am. J. Bot. 37:576-586.
Sato, S., T. Matsushima, N. Tanaka, T. Sugimura, and F. Takashima. 1973.
Hepatic tumors in the guppy, Lebistes reticulatus, induced by
aflatoxin B}, dimethylnitrosamine, and 2-acetylaminofluorene. J.
Natl. Cancer Inst. 50:767-778.
Scarpelli, D.G. 1969. Comparative aspects of neoplasia in fish and other
laboratory animals. In: Fish in research. O.W. Neuhaus, and J.E.
Halver, Eds., Academic Press, New York, London, pp. 45-85.
Scherf, R.H. 1976. Toxikologische Wirkungen von Diathylnitrosamin (DANA)
beim Guppy, Lebistes reticulatus. Z. Krebsforsch. 86:155-163.
Schwab, M. How can altered differentiation induced by 12-0-tetradeca-
noylphorbol-13-acetate (TPA) be related to tumor promotion?
In: Carcinogenesis and biological effects of tumor promoters.
E. Hecker, Ed., Raven Ress, New York, pp. 417-426.
Schwab, M. 1982. Biology and genetics of neoplasia in Xiphophorus.
Advances Cancer Res. 58 (in press).
231
-------
Schwab, M., S. Abdo, M.R. Ahuja, G. Kollinger, A. Anders, F. Anders, and
K. Frese. 1978. Genetics of susceptibility in the platyfish/
swordtail tumor system to develop fibrosarcoma and rhabdomyosarcoma
following treatment with N-methyl-N-nitrosourea (MNU). Z.
Krebsforsch. 91:301-315.
Schwab, M., G. Kollinger, J. Haas, M.R. Ahuja, S. Abdo, A. Anders, and
F. Anders. 1979. Neuroblastoma induced in the xiphophorine fish by
N-methyl-N-nitrosourea (MNU) and X-rays: genetic basis for the
susceptibility. Cancer Res. 39:519-526.
Schwab, M., A. Anders. 1981. Carcinogenesis in Xiphophorus and the role
of the genotype in tumor susceptibility. In: Neoplasms. Comparative
pathology of growth in animals, plants and man. H.E. Kaiser, Ed.,
Williams and Williams, Baltimore, pp. 451-459.
Schwab, M., and E. Scholl. 1981. Neoplastic pigment cells induced by
N-methyl-N-nitrosourea (MNU) in Xiphophorus, and genetic control of
their terminal differentiation. Differentiation 19:77-83.
Shih, C., L.C. Padhy, M. Murray, and R. Weinberg. 1981. Transforming
genes of carcinomas and neuroblastomas introduced into mouse
fibroblasts. Nature 290:261-264.
Siciliano, M.J., and A. Wright. 1976. Biochemical genetics of the
platyfish/swordtail hybrid melanoma system. Prog. Exp. Tumor Res.
20:398-411.
Sonstegard, R. 1977. Environmental carcinogenesis studies of fishes of
the great lakes of North America. Ann. N.Y. Acad. Sci. 298:261-269.
Sparkes, R.S., H. Muller, and I. Klisak. 1979. Retinoblastoma with
13q-chromosomal deletion associated with maternal paracentric
inversion of 13q. Science 203:1027-1029.
Varmus, H.E. 1982. Form and function of retroviral proviruses. Science
216:812-820.
Vielkind, U. 1976. Genetic control of cell differentiation in
platyfish-swordtail melanomas. J. Exp. Zool. 196:197-204.
Wellings, S.R. 1969. Environmental aspects of neoplasia in fishes. In:
Fish in research. O.W. Neuhaus, and J.E. Halver, Eds., Academic
Press, New York, London, pp. 3-22.
232
-------
THE USE OF GENETICALLY MODIFIED
FISH IN THE DETECTION AND
MEASUREMENT OF CARCINOGENS IN WATER
by
Linda S. She!ton,
Mary Louise Bellamy, and
Douglas G. Humm
Zoology Department, University of North Carolina
Chapel Hill, NC 27514
ABSTRACT
The platyfish, Xiphophorus maculatus. was challenged with
small concentrations of four carcinogenic polycyclic
hydrocarbons. Using the ratio of melanized fin area to total
fin area, we observed that the fin spot increased in direct
proportion to dosage and duration of exposure. Our tests
demonstrate that the response, which involved increased
cellular activity, can be used as an in vivo monitoring system
for the presence of carcinogens in water.
INTRODUCTION
As carcinogenic substances become more common in the environment, the
need for a sentinel biosystem capable of monitoring chemicals in our waters
becomes more urgent. It is important to search for a in vivo system that
can rapidly and inexpensively test not only a single compound, but the
entire enviromental load of contaminants.
A system has been developed for airborne pollutants by using
Tradescantia puldise and color mutations of the flowering parts (Sparrow
e£ j»l_., 1974). This assay monitors the total environment rapidly and
inexpensively. A bioassay with similar advantages is greatly needed to
monitor the aquatic environment.
Stich and Acton (1976), in a study of "early warning" systems, found
the flatfish, Pleuromectes sp., to be a promising research animal for
carcinogen detection. A great deal of the work involving the relationship
between neoplasias in fish and pollutants has used the commercially
significant flatfish (Mearns and Sherwood, 1976; Bucke, 1976), which,
however, are neither genetically controlled nor readily adaptable to
laboratory conditions.
233
-------
Tumor induction has been studied in guppies and zebra fish (Pliss and
Khudoley, 1975), and in the medaka, Oryzias latipes, (Ishikawa et al..
1975) by organic carcinogens. Matsushima and Sugimura (1976) have reported
experimental carcinogenesis in small aquarium fish. These experiments
involved liver tumors which, unfortunately, cannot be examined in the
living fish.
Ames (Ames j;t a]_., 1973) has developed an excellent experimental
design that employs revertant frequency analysis to draw a correlation
between known carcinogens and the production of mutations in Salmonella
typhimurium. This rapid and convenient bioassay has gained wide acceptance
in testing laboratories and bids well to become accepted an criterion of
mutagenesis. However, mutagenesis and carcinogenesis are not always
synonymous. Therefore, an additional system is needed to measure
carcinogenecity.
In their review, Stichjet^l_. (1975) have stressed the necessity for a
practical screening assay that would actually involve a carcinogenic
process, i.e., morphological transformation. Such an assay, ideally, would
involve a vertebrate with a thoroughly known genetic background, preferably
isogenic.
The most commonly used test animal for detection and dosage studies
involving carcinogens is the mouse. However, the use of small mammals in
laboratory tests is not only time-consuming, but also enormously expensive.
It would be preferable to have a backup sentinel system in which a
reasonably short exposure to the suspected carcinogen results in a visible,
or measurable, cellular change that could be recognized as atypical growth.
One such animal might be the small poeciliid fish, Xiphophorus maculatus
[the platyfish (Ps )]. The platyfish possesses a melanized spot on the
dorsal fin which, in and of itself, is not particularly significant.
However, it can be, and has been, genetically altered in various ways and
therefore is particularly suitable for detection of carcinogens dissolved
in water.
One mechanism of alteration involved crossing Xiphophorus maculatus to
a closely related species, Xiphophorus helleri [the swordtail (S)].The
resultant hybrid (P xS) exhibited a dorsal fin melanophore hyperplasia
known to be pretumorous. Experiments of Gordon (1951), Anders (1973), and
Kail man (1973) have demonstrated that the production of atypical growths in
the poeciliid fish involved the derepressive action of the swordtail genome
upon genes occurring on the X chromosome of the platyfish. As a result of
their research, it is known that the susceptibility of the hybrid to tumor
production may be increased or decreased, almost at will. When the hybrid
is backcrossed to the swordtail, fish developed with dorsal fin pigment
cells which become true malignant melanomas. On the other hand, it was
possible to prepare crosses in which the fish were genetically poised on
the brink of tumor production and a small chemical "push" would be suffi-
cient to start the process of overgrowth and loss of contact inhibition,
the first detectable step toward malignancy. This can be achieved by back-
crossing the hybrid to a platyfish.
234
-------
Platyfish pigment cells also can be altered by X-irradiation to obtain
a viable method of testing carcinogens in water using fish with impaired
control of pigment celi production. Such a group, the sd'sr' platyfish,
was developed by Dr. Fritz Anders of the Genetisches Institute der
Justus-Liebis-Universitat, Geisen, West Germany. These fish differed from
their predecessors in that the gene for striped-side was translocated from
the X chromosome and become attached to the Y chromosome. This
translocation is associated with a loss of represser activity which appears
to be a significant factor in the sensitivity of dorsal fin melanophores to
carcinogens (Anders, personal communication).
TEST METHODS AND MATERIALS
Experiments performed in our laboratory were designed to test both
groups of fish for their susceptibility to carcinogen-induced melanoma
production and to determine which fish would provide the most useful
organism for future carcinogenic testing. The stocks of sd'sr1 platyfish
in our experiments were an inbred stock generously provided by Dr. Fritz
Anders. They were subdivided into male and female groups because the
involvement of the X and Y chromosomes in the pigmentation coding indicated
that response of the sexes would differ:
Platyfish with spotted dorsal fins (P ), Xiphophorus maculatus,
and green swordtails (S), Xiphophorus helleri, were obtained from the
Genetics Laboratory of the New York Zoological Society. The swordtail
females were artificially inseminated with the sperm of the platyfish
males. The hybrids, being sexually compatible with the parent species,
were easily mated to platyfish. The offspring are predominatly female;
only females were used.
All fish were raised to the age of 6 to 8 months and then sexed,
weighed, and photographed. Benzo(a)pyrene (BaP), 3,4,9,10-dibenzpyrene
(DBP), 3-methylcholanthrene (MC), and benzanthracene (BA) were dissolved
independently in 600 mi of aged tank water. Fish were exposed individually
to a single carcinogen for 4 days, and then transferred to fresh water and
fed brine shrimp.
The carcinogens (concentration, 25 yg/nu water/gm of fish) were
suspended in 2 ma of Carbowax 200 (Atlas Powder Co.) as an inert vehicle.
The test compounded BaP, DBP, MC, and BA were obtained from Sigma Chemical
Co. and used without further purification.
Each group was controlled by a fish of its own type and comparable
spot size. The control fish were kept in aged tank water which contained
2 ml Carbowax 200, but no known carcinogens. Additionally, a group of
similar fish were exposed to chrysene as a noncarcinogen with the same
overall chemical structure as the carcinogens. There was no measurable
increase or growth in the spot area of the control of chrysene-treated
fish. During other experiments, diphenylnitrosamine was added as a
noncarcinogenic compound. However, its toxicity was so great that its use
had to be discontinued.
235
-------
Photographs of the dorsal fin were taken every few days before and
after exposure to carcinogen for 1^ months. A 55-mm macro-lens was used
for all pictures. The fins and spots were measured by a pianimeter, and
the data recorded as a ratio of the area of the fin covered by melanized
spot vs. the total fin area.
RESULTS
The measurements of spot vs. fin area from the three experiments are
collated in Tables 1, 2, and 3. The omissions in the tables are due to the
difficulty in taking pictures of live fish. The fish had to be correctly
positioned, the fin perfectly erect, and the melanophores completely
expanded. Many of our photographs were unusable for various reasons. Much
of the scatter seen in our data is due to such complications.
Figures 1, 2, 3, and 4 represent the growth of the pigmented spots
based on the most reliable data. The slopes of the lines drawn determined
by linear regression analysis are presented at the bottom of the columns on
the corresponding table. Figures 1, 2, and 3 indicate that the response of
the three types of fish varied widely. The response of the various fish to
methylcholanthrene is shown in Figure 1.
The response of t-he,psd Sr male was very marked when compared with
the control. The Psd sr female, and the (PsdxS)psd changed only slightly
as compared with the control. The phenomenon was repeated (Figures 2 and 3)
with both .benzo(a)pyrene and 3,4,9,10-dibenzpyre In each case, the response
of the,P sr males was more rapid than the other two. Often, the
psd sr feniaies an(j ^e (psdxs) Psd females did not respond at all; spon-
taneous pigment cell spot growth was observed in the (P xS)P controls
which rendered them less valuable as a comparison standard. In contrast,
the P sr males gave a marked response to all the polycyclic hydrocarbons
used, demonstrating that the Ps s male was the most useful for the testing
of carcinogens in aqueous solutions.
CONCLUSIONS
We believe that there is a real need for a rapid detection system for
carcinogens as distinct from mutagens. This system should employ an
eucariot and, preferably, a vertebrate. Tests for mutagenicity on bacteria
are useful screening devices, but shoud not be considered synonymous with
tests for carcinogenicity. We agree that a single mutation is probably
not sufficient to produce carcinogenesis in higher organisms.
Therefore, we feel that a test system using genetic manipulation to
produce an organism which is already "primed" genetically has a very good
chance of satisfying the need for a rapid and relatively inexpensive test
system for carcinogens. To date, our preliminary data indicate that the
pigment cells under study exhibit at least one criterion of atypical
growth—namely, loss of contact inhibition.
236
-------
TABLE 1. RESPONSE OF THE SWORDTAIL-PLATYFISH HYBRID BACKCROSS
[(PsdXS)Psd] TO POLYCYCLIC HYDROCARBONS IN WATER
DAY CWa
3
4 22.76
7
Exposure
CW MCb BaP
-
-
11.32
19.11<
20.65
_
Substance
BaP DBF
*
-
23.97
_f
-
16.42
25
23
18
BA
.92
.91
.09
BAC
12.30
-
8.77
10 - 20.25 - - 27.57 - 23.28 9.63
12 30.24 24.13e 13.73 19.19 29.21 - 26.69 18.95
13 35.80 21.87 - 21.18 24.72 12.35
17 36.15 - 13.82 25.03 29.94 - - 18.12
18 - 22.12 - - - 22.88 23.12 17.81
20 - - 13.54 - 27.54 18.13
21 - - - 24.99
25 38.78 26.29 14.65 - - 26.09 25.61 19.80
27 ----- 23.20 29.50
.912 .776 .853 .715 .896 .868 .560 .701
m .772 .301 .152 .298 .567 .408 .227 .435
a. Control fish (CW) exposed to 0.3% Carbowax 200 in acclimatized tank
water.
b. The carcinogens, 3-Methylcholanthrene (MC), Benzo(a)pyrene (BaP),
3,4,9,10-Dibenzpyrene (DBP), and 1,2-Benzanthracene (BA) suspended in
Carbowax 200 to a final concentration of 25 ug/mji of water/gm of fish;
duration of exposure to carcinogens: 4 days at 23° C.
c. Each column represents the data obtained from a single fish.
d. Data expressed as percent of total dorsal fin area covered by pigment
cells.
e. Some daily variation in measurements is encountered due to the fact
that it was not always possible to photograph the fin in a fully
extended position.
f. Not measured.
237
-------
TABLE 2. RESPONSE OF THE FEMALE SD'SR1 PLATYFISH (psd'sr'j
TO POLYCYCLIC HYDROCARBONS IN WATER
Exposure Substance
DAY CW MC MC BaP BaP DBP BA
1
10
15
22
24
29
38
43
rT2
m
-
9.60
-
9.00
8.50
-
8.40
9.30
.384
-.015
28.10
19.30 57.00 28.30
58.20 27.00
24.60 60.80
21.60 58.40
59.60
_
69.40
.744 .896 .675
.261 .351 -.067
_
6.99 - 19.95
8.00 - 26.70
8.60 8.90
12.10 25.80
8.50 14.60 26.20
10.30 12.90
10.40
.964 .580 .679
.099 .194 .250
TABLE 3. RESPONSE OF THE MALE SD'SR1 PLATYFISH (psd'sr1) T0
POLYCYCLIC HYDROCARBONS
Exposure Substance
DAY CW MC MC BaP DBP BA
1 38.90 25.80 - 20.10
6 32.70 - 18.70
7 - 23.60 22.30 20.70 25.80 27.40
9 31.20 29.20 20.60 22.50 33.80
13 - 29.30 19.40 23.80 37.00 22.30
14 34.30 - 36.80 25.20
24 - 38.10 24.90 26.60
.590 .911 .850 .932 .904 .024
m -.360 .588 .265 .315 1.430 .008
238
-------
sdsr'o*
LU
< 20
LU
o:
O
a.
CO
10
d'sr'?
rrr-'--- (sxp)x p
^ ^ ^^r*^ ^ ^ ^ ^
^ ^ ^^~ ^ ^ ~
10
20
DAYS
Figure 1.
_,CONTROLS
30
of
The difference is shown in the response to meth> choU ..nrene
three fish. It is derived from data in Tables 1, 2, and 3 by
setting the appropriate controls as 0 growth and plotting the
rate as (spot area/fin area K 100) against time. The equation
for the lines drawn were determined by subtracting the slopes of
the controls from slopes of the experimental s.
mfinal ~ mexper. " mcontrol
239
-------
LU
ce
20
u_
X
<
UJ
cc
o
QL
CO
10
sd'srV
sr
10
20
DAYS
-CONTROL
30(Sxp) x P
Figure 2. Response of the three fish to benzo(a)pyrene derived in the same
manner as Figure 1.
sd'sr'o*
LU
ce
< 20
LU
DC
O
Q.
CO
10
o
.sd'sr'?
..CONTROLS
10
20
DAYS
30
Figure 3. Response of the three fish to 3,4,9,10-dibenzpyrene derived in
the same manner as Figure 1.
240
-------
The possibility exists that these genetically metastable cells have
responded by growth and not by malignant transformation. An increased time
of exposure to the hydrocarbons might well result in the transformation of
the pigment cell, in addition to the stimulation we have reported.
LU
520
< 10
I-
o
Q.
CO
DBF
/
/ X .---BA
X ,...-----:;;:-'.'-'--'-- BQP
S&sS's: . , .CONTROL
is==— ft 20 30
DAYS
Figure 4. Graph showing the response of the sd1sr'or platyfish to the
4 polycyclic hydrocarbons. According to our data, MC and DBP
are are more effective in eliciting a response in the platyfish
pigment cells.
241
-------
ACKNOWLEDGMENTS
This research was supported in part by the Environmental Protection
Agency Grant No. R804650.
REFERENCES
Ames, B.N., F.D. Lee, and W.E. Durston. 1973. An improved bacterial test
system for the detection and classification of mutagens and
carcinogens. Proc. Nat. Acad. Sci. USA 70:782-786.
Anders, A., F. Anders, and K. Klinke. 1973. Regulation of gene expression
in the Gordon-Kosswig melanoma system. In: Genetics and Mutagenesis
of Fish. J.J. Schroder, Ed., Springer-Verlag, New York. pp. 33-63.
Anders, F. 1967. Tumor formation in platyfish-swordtail hybrids as a
problem of gene regulation. Experientia 23:1-10.
Bucke, D. 1976. Neoplasia in roach Rutilus rutilush. from a polluted
environment. Prog. Experi. Tumor Res. 20:205-211.
Gordon, M. 1951. Genetic and correlated studies of normal and atypical
pigment cell growth. Growth, symposium X pp. 153-219.
Ishikawa T., T. Shimamine, and S. Takayama. 1975. Histologic and electron
microscopy observations on biethylnitrosamine-induced hepatomas in
small aquarium fish, Oryzias latipes. J. Nat. Cancer Inst.
55:909-911.
Kail man, K. D. 1973. The sex-determining mechanism of platyfish,
Xiphophorous. In: Genetics and Mutagenesis of Fish. J.H. Schroder,
Ed., Springer-Verlag, New York. pp. 19-28.
Matsushima, R., and T. Sugimura. 1976. Experimental carcinogenesis in
small aquarium fishes. Prog. Exp. Tumor Res. 20:367-379.
Maugh, T. II. 1978. Chemical carcinogens: the scientific basis for
regulation. Science 201.
Mearns, A. J., and M. J. Sherwood. 1976. Ocean wastewater discharge and
tumors in a southern California flatfish. Prog. Exp. Tumor Res. 20:
75-85.
Pliss, C.B. and V.V. Khudoley. 1975. Tumor induction by carcinogenic
agents in aquarium fish. J. Nat. Cancer Inst. 55:129-134.
Sparrow, A.M., L.A. Schairer, and R. Villalobos-Pietrini. 1974.
Comparison of somatic mutation rates induced in Tradescantia by
chemical and physical mutagens. Mutat. Res. 26:265-276.
242
-------
Stich, H.F. and A.B. Acton. 1976. The possible use of fish tumors in
monitoring for carcinogens in marine environment. Prog. Exp. Tumor
Res. 26:265-276.
Stich, H.F., P. Lam, L. W. Lo, D.J. Koroipatnich, and R.H.C. San. 1975.
The search for relevant short term bioassays for chemical carcinogens:
The tribulation of a modern Sisyphus. Can. J. Genet. Cytol. 17:
471-492.
243
-------
THE DISTRIBUTION OF BENZO(a)PYRENE IN BOTTOM SEDIMENTS
AND OF NEOPLASMS IN BOTTOM-DWELLING FLATFISH SPECIES OF
THE PACIFIC AND ATLANTIC OCEANS, NORTH, CHINA, BERING AND
BEAUFORT SEAS, AND SEA OF OKHOTSK
by
H.F. Stich, B.P. Dunn, and A.B. Acton
Environmental Carcinogenesis Unit,
British Columbia Cancer Research Centre,
601 West 10th Avenue, Vancouver B.C., Canada V5Z 1L3
and
F. Yamasaki, K. Oishi, and T. Harada
Hokkaido University, Hakodate, Japan
ABSTRACT
Human and animal populations are continuously exposed to
hundreds of carcinogenic, mutagenic, clastogenic and recombino-
genic agents. These chemicals can interact, leading to
extremely large numbers of possible permutations and combinations
that can either enhance or reduce their genotoxic or carcino-
genic activity. It is impossible to examine all these inter-
actions for economic and logistic reasons. Thus, other approaches
to identify high risk areas must be sought. In this paper, we
explore the feasibility of using naturally occurring animal
populations as an early warning system to detect carcinogen-
mutagen contamination of a particular environment. The basic
idea is simple. In animals, tumors may appear within months,
whereas in man the latency period may well exceed 16 or more
years. Thus, by screening indigenous animal populations for
benign and malignant neoplasms, we should be able to recognize
carcinogen-contaminated areas long before they adversely affect
man. There is no shortage of examples of making use of "built-in"
organisms to detect agents with a general toxic action.
However, the practicability of a naturally occurring indicator
organism for chemical carcinogens or mutagens is still unproven.
In this paper, we critically review the difficulties encountered
in the attempt to use neoplasms of bottom-dwelling fish popu-
lations as a possible indicator for man-made or naturally ocurr-
ing contamination of shallow estuarine nursery grounds that
might at some time be developed for intensive aquacultural usage.
244
-------
INTRODUCTION
Epidemiological studies of human populations coupled with chemical
analysis have made it possible to identify numerous carcinogenic and
mutagenic agents in man's environment and to trace their origins. Based on
this experience, it has been proposed that naturally occurring animal
populations be used in an early warning system to detect carcinogen-mutagen
contamination of a particular environment. The basic idea is simple. In
animals, tumors may appear within months, whereas in man the latency period
may well exceed 16 or more years. Thus, by screening indigenous animal
populations for benign and malignant neoplasms, we should be able to
recognize carcinogen-contaminated areas long before they adversely affect
man. There is no shortage of examples of making use of "built-in"
organisms to detect agents with a general toxic action. However, the
practicability of a naturally occurring indicator organism for chemical
carcinogens or mutagens is still unproven. In this paper, we critically
review the difficulties encountered in the attempt to use neoplasms of
bottom-dwelling fish populations as a possible indicator for man-made or
naturally occurring contamination of shallow estuarine nursery grounds that
might at some time be developed for intensive aquacultural usage.
Advantages of Flatfish as an Indicator Organism
In choosing a suitable animal test species, we applied several
criteria (Stich and Acton, 1976; Stich ^t al_., 1976a,b): (1) Availability.
Flatfish samples can be easily collected by bottom trawling in shallow
waters, requiring only a small boat and a trawl net (available at all
marine stations). Moreover, several flatfish species are of commercial
importance and thus can be obtained on regular fishing boats or at
processing plants. (2) Global distribution. Flatfish have also the
advantage of a wide geographic distribution. Some species are common along
the shores of the entire northern Pacific Ocean or northern Atlantic Ocean.
(3) Migratory behavior. Several flatfish populations have nursery grounds
in the shallow waters of estuaries or within deltas of rivers that are
areas of greatest concern because of their potential for becoming polluted.
The movements of juvenile populations into deeper waters and their
migratory pattern as adults are fairly well-known. (4) Tumor prevalence.
Papillomas have been reported to occur among flatfish at such high
frequencies that their precise quantitation would require the screening of
relatively only small numbers (Nigrelli j!t j|l_., 1965; Wei lings et al.,
1965, 1977; Cooper and Keller, 1969; Wellings, 1969; Stich et aj_., 1976b,
1977a; Yamazaki eit a\_., 1978). (5) Age-adjusted tumor prevalence. Skin
papillomas develop on young flatfish. Skin nodules which precede
papillomas start to appear on lemon soles, Parophrys vetulus. about 4 to 5
months after metamorphosis and reach a peak within the first year post-
metamorphosis. A valid comparison can only be achieved by comparing the
peak frequencies at each sampling station. (6) Diagnosis. The skin
papillomas are easily detectable by macroscopic examination. At the
microscopic level, papillomas are characterized by rounded enlarged cells
with vacualization and degeneration of cytoplasmic organelles (Wellings
et^l-, 1965, 1977; Brooks et aj_., 1969; Well ings, 1969; Stich jt al.,
1976b, 1977a; Yamazaki et aK, 1978).
245
-------
It was suggested that these cytoplasmic organelles, the so-called "X" cells
(Brooks et a\_., 1969) could be protozoa (Wellings j^t ail_., 1977), a
hypothesis which we have previously criticized and which seems to lack
factual support (Peters et jj]_., 1978).
Restriction of the Flatfish-Tumor System: Geographic Distribution
If the flatfish-papilloma system has the many advantages mentioned
above, the questions must be asked, why then do we not have a "fish tumor
watch", or why have regulatory agencies not yet introduced a fish-tumor
monitoring program in areas of aquaculture and recreational activity. One
of the reasons is the unique global distribution pattern of the skin
papillomas (Figure 1). Flatfish with skin nodules and papillomas
characterized by typical X-cells and envelope cells seem to be restricted
to the northern Pacific Ocean and the adjacent Bering Sea and Sea of
Okhotsk. To the best of our knowledge, papillomas composed of X-cells have
not been seen outside of what we previously called areas of potential skin
papilloma risk (Stich et_ ^K, 1977a). Of four skin papillomas of flatfish,
Platichthys flesus. caught in the North Sea, all lacked the X-cells and
their histopathology was more comparable to the skin papillomas found among
European eels. Obviously, any correlation between tumor frequency and
level of carcinogens in the marine environment can be sought only in those
regions in which flatfish can develop skin papillomas.
SKIN PAPILLOMAS OF '-LATFiSM SPECIES
Figure 1. Distribution of flatfish species, Pleuronectidae, with skin
papillomas (^). The examined species are listed by
Stich et a]_. ;i977b). The marked areas are free of skin
papillomas ( £\ ).
246
-------
In spite of considerable effort, the cause(s) of the global
distribution pattern has still eluded us. Previously, we have speculated
about the involvement of a virus and the possibility that the papilloma
cells result from an abortive lymphocystic infection in flatfish of the
northern Pacific Ocean (Stich et al_., 1977a). However, there are other
equally plausible explanations. For example, the global distribution
pattern of skin papillomas could reflect the distribution of flatfish
populations that are sensitive or resistant to the induction of skin
neoplasms. Such an assumption would in turn raise the question as to the
reasons why resistant fish species can evolve in one location and not at
another.
Whatever explanation may finally prove to be correct, there is one
important lesson to be learned from our global study on flatfish
papillomas. An attempt to find a correlation between fish neoplasms and
environmental agents should be preceded by a careful investigation of
whether or not the selected fish species at the location under study is
capable of developing a particular neoplasm. Such a preliminary
exploratory study may avoid a lot of frustration. If feasible, we should
even try to obtain quantitative information about the variations in
sensitivity of fish populations from different geographic locations.
Otherwise, there will always be the nagging suspicion that observed
differences in tumor prevalences may reflect differences in sensitivity
towards carcinogenic and mutagenic agents rather than differences in
environmental contaminants.
Consistency of Tumor Prevalences
The prevalence of skin papillomas among flatfish populations varies
markedly according to geographic location, season, and age group. At
nursery grounds, a sampling period of about one year is required to gain
accurate information on the peak of tumor prevalence. Considering the
large geographic and age-dependent variations, it was somewhat unexpected
that the prevalence figures of skin papillomas in flatfish populations did
not change from year to year (Table 1). Even during longer periods, the
peak prevalence was surprisingly constant, as exemplified by rock sole
populations sampled off the Queen Charlotte Islands (Canada).
If tumor frequencies are to be used successfully in a monitoring
program, this observed consistency in papilloma frequencies cannot be
simply discarded, but should be explained. We could assume that a
contaminated estuary does not rapidly change, and thus juvenile fish are
exposed to similar carcinogenic levels for long periods. One of the more
intriguing explanations would be to assume a genetic heterogeneity within
the fish population, consisting of sensitive and resistant individuals and
further, to postulate that their ratios differ among populations in
different locations. The latter idea is by no means farfetched. For
example, there appears to be a particular distribution of aryl hydrocarbon
hydroxylase inducibility among the normal human population that apparently
increases the risk of lung and laryngeal carcinomas (Kellermann et al_.,
1973; Kellermann, 1977; Brandenburg and Kellerman, 1978). Again, the
247
-------
different sensitivities of various mouse strains to cancer induction are to
a great extent based on the genetic control of metabolism, activation, and
detoxification of carcinogens (Nebert and Felton, 1976; Nebert et jiK, 1976;
Kouri and Nebert, 1977). But, even among fish (platyfish-swordtail system),
the different responses to carcinogenic agents, including X-rays and
N-methyl-N-nitrosourea (MNU), have been traced to various combinations of
tumor and regulatory genes (Anders jet _al_., 1971; Anders and Anders, 1978;
Schwab et ^1_., 1978a,b). If differing sensitivities to environmental
carcinogens are really genetically based, then a balanced polymorphism may
exist and subgroups with different ratios of sensitive and resistant fish
could arise by selection. This selection could be by the action of
carcinogens, and one could imagine a situation where increased death rate
of sensitive individuals could produce a population with a lower tumor
prevalence. Such an assumption could also explain the observation that the
prevalence figures are relatively constant over a time period and that
"constant" prevalence figures vary between populations of the species
inhabiting different geographic locations.
TABLE 1. CONSTANCY OF SKIN PAPILLOMA PREVALENCES AMONG FLATFISH
POPULATIONS SAMPLED AT DIFFERENT TIME PERIODS
Location
Queen Charlotte
Islands, B.C.
Bay of San
Francisco
Vancouver, B.C.
Everett,
Washington
Species
Psettichtys Melanosticus
(sand sole)
Parophrys vetulus
(English sole)
Parophrys vetulus
(English sole)
Parophrys vetulus
(English sole)
Year of
Sampling
1953
1975
1968
1975
1973
1974
1976
1975
1976
(%)
Prevalence
Skin Papilloma
33%
31%
29%
28%
56%
58%
60%
44%
41%
TUMOR PREVALENCES AND B(a)P IN BOTTOM SEDIMENTS
Since flatfish stay for prolonged periods in their life span within
layers of mud, it was expected that carcinogens in the bottom sediments
could directly affect the surface fish, leading to the observed skin
papillomas. It was further argued that, in such a case, a correlation
between prevalence of skin papillomas and the level of carcinogen in the
sediments of nursery grounds should become easily evident. Actually no
simple link between neoplasms of flatfish and levels of benzo(a)pyrene
(BaP) in bottom sediments was observed (Table 2). For example, relatively
high levels of BaP occur in the Jade (Northern Germany) or near Hong Kong,
but not a single skin papilloma was seen among six local flatfish species.
But even within the northern Pacific, which we previously called an area of
potential risk regarding skin papilloma risk (Stich et a\_., 1977a), no
simple quantitative relationship between BaP and skin papillomas could be
243
-------
established (Table 2). The high prevalence of skin papillomas found among
fish populations in the vicinity of urban and industrial centers (Stich and
Acton, 1976; Stich et^U, 1976a,b; 1977a,b) cannot be due only to PAH.
Moreover, it would "Be misleading to assume that all PAHs originate from
man-made activity. The fairly high levels of BaP at Furen Lake in Hokkaido
and at the MacKenzie delta and surrounding coast are likely to be due to
naturally occurring phenomena, such as forest fires, which in the northern
regions may contribute a greater amount of PAH to the environment than
generally assumed (Beumer and Youngblood, 1975; Lunde and Bjorseth, 1977;
McMahon and Tsoukalas, 1978). PAHs can also be transported over wide areas
(Lunde and Bjorseth, 1977) and may have, in cold arctic or subarctic
waters, different rates of biodegradation (Atlas, 1977) that could lead to
their accumulation over longer time periods.
Numerous reasons can be cited in an attempt to explain the lack of an
obvious correlation between skin papilloma prevalences and BaP levels in
bottom sediments: (a) PAHs are potent skin carcinogens for mammals, but may
not have a similar effect on fish skin; (b) the concentration of PAHs is
too low; (c) PAHs are bound to particulate matter within the bottom
sediment and thus do not enter the fish skin; (d) BaP and other PAHs do not
represent a major part of carcinogenic compounds in marine bottom
sediments.
Before the question of involvement in induction of tumors of marine
organisms is discarded or downgraded, the measurements of BaP and other
carcinogenic compounds in the marine environment should be critically
reviewed. Is it valid to assume that bottom sediments are the source of
PAHs that enter bottom-dwell ing or free-swimming organisms? If PAHs pass
into invertebrate and vertebrate via the food chain, then the BaP content
of the biotope of the food chain organisms is more crucial than that of the
indicator organisms. Thus PAH measurements of the bottom sediment could
easily be misleading. At present it is difficult to assess whether
sediments properly reflect the PAH levels at the water surface or PAH in
suspension. The use of several accummulator organisms that inhabit a
variety of habitats will lead to a more complete evaluation of the PAH
actually available to living organisms (Mix et jf[., 1977).
BaP in Bottom Sediments and AHH Activity in Flatfish
Aryl hydrocarbon hydroxylase (AHH) is involved in the activation and
detoxification of a wide array of carcinogens. In rodents and man, the
inducibility of this enzyme appears to be genetically controlled, thus
dividing many populations into responsive and nonresponsive subgroups with
a high and low sensitivity to tumor induction, respectively (Kellermann
et£l_., 1973; Nebert and Felton, 1976; Nebert et al_., 1976; Kellermann,
1977; Kouri and Nebert, 1977; Poland and Kende, 1977; Boulos, 1978). AHH
activity is also found in fish (Lee £t aj_., 1972; Payne and Penrose, 1975;
Payne, 1976; Bend jit a]_., 1977). If PAHs comprise a major component of
carcinogens in the marine sediments, we should expect to find a correlation
between BaP levels and AHH activity of bottom-dwelling fish. Such a corre-
lation was actually observed by plotting BaP levels against AHH levels in
livers of lemon soles sampled at different locations off the coast of
249
-------
TABLE 2. THE CONTENT OF BaP IN BOTTOM SEDIMENTS AND THE PREVALENCE OF SKIN
PAPILLOMAS AMONG FLATFISH AT VARIOUS LOCATIONS
Location
Wi Ihelmshaven
sand
si It
Hong Kong
sand
Kotzeb ue
Shesbal Fk
Baldwin Peninsula
Prince of Wales
Nome
Seward
Kittigazuit
Kings Point
Kay Point
Shingle Point
MacKenzie Delta
Fu ren Lake
Shore of Okhotsk
Eve rett
Port Susan
Utsaladdy
Vancouver
Port Renfrew
B(a)P yg/kg
dry weight
35-1660
4- 51
2- 38
0.4-0.5
0.1-2.2
0.2-1.9
<0. 1-1.1
<0.1-0.8
0.4
1.8/24.3
0.6-17.3
117
0.3/6.3
2-33
55"
Sea 0.2"
0.3
0.7
0.6
11.1
Prevalence(i)
Skin papi 1 lomas
0
0
C-23
0
0
0
33
0.3
42-44
20-26
1- 3
56-60
0
0
0
0
Species and
References
11 species-''
Lepidopsetta
bill neata*""
5 species-*"
Liopsetta qracialis
Platichthys ste 1 1 at us
Limanda schrenki
Limanda schrenki
Parophrys vetulus
Parophrys vetulus
Parophrys vetulus
Parophrys vetulus
Platichthys stelatus
Microstomus pacificus
Parophrys vetu 1 us
Platichthys stelatus
"B(a)P yg/kg wet weight
»«Stich e± aj_. , 1977a
Dwellings et al., 1977
250
-------
a
in
_ 2O-
10-
•s
• 11
• 6
100 1000
B(a)P ug/kg sediment org. concent
10000
Figure 2. Correlation between BaP levels in bottom sediments and AHH
activity (p mole/min/mg/liver) in the liver of lemon sole
sampled at the following locations in British Columbia and the
State of Washington: 1. Vancouver, 2. Gibsons, 3. Crescent
Beach, 4. Coal Harbor, 5. Port Moody, 6. Denman Island,
7. Tofino, 8. Port Renfrew, 9. Everett, 10. Port Susan,
11. Henderson Inlet, 12. Utsaladdy, 13. Bellingham, and
14. Olympia. Per point between 2 and 12 bottom samples were
analyzed according to the method of Dunn (1976), Dunn and Stich
(1975), and Dunn and Young (1976).. AHH was assayed following
the procedure of DePierre et_ aK (1975).
251
-------
British Columbia and Washington (Figure 2). As expected, some fish popu-
lations did not fit into the overall picture. An exceptionally high AHH
activity was found in the liver of flatfish collected near a paper mill in
an area with a relatively low BaP concentration. The observed correlation
seems to be in good agreement with reports on elevated AHH activity in
livers and gills of the cunner, Tautogolabrus adspersus, in oil-contaminated
bays of Newfoundland (Payne and Penrose, 1975; Payne, 1976). However, a
correlation can only be seen when relatively large numbers of flatfish are
examined and the fish populations are sampled at the same period of the
season. Even when age, size, and weight are adjusted, and the fish are
collected in one small territory from a uniform type of bottom sediments,
the AHH activity can vary considerably (Figures 3,4,5). This variation in
AHH activity becomes particularly evident if flatfish are collected from
offshore areas contaminated by complex mixtures of urban and industrial
effluents (Figure 3). On the other hand, the AHH levels can be quite
similar, this pattern is frequently encountered in areas without any
obvious pollution by industrial, urban, or agricultural discharges. Again,
we would like to point out that a simple correlation may either not exist
or be found only in exceptional cases with large differences between
pristine and contaminated areas. As seen in Figure 4, even within the same
habitat, different species of comparable age can greatly differ in their
liver AHH activities. Of particular interest is that the variations in AHH
activities of 1-, 2-, and 3-year-old lemon soles inhabiting the same
contaminated environment did not significantly differ (Figure 3.)
I
I
< 10
15O
LENGTH mr
Figure 3. AHH activities in the liver of lemon sole of various lengths
sampled at Port Moody. A point represents the AHH activity in
the liver of our fish.
252
-------
AHH ACTIVITY (LIVER)
1O 20
3O
ROCK SOLE
LEMON SOLE
SAND DAB
Figure 4. AHH activity in the liver of two sole species and sand sole
collected off Denman Island.
PAPILLOIVIA
LOCATION PREVALENCE o
"/. |—
AHH ACTIVITY ( LIVER )
10 ao 30
*ir n iv *
RENFREW O.D
BELLllMGHAN" 33 8
••••»•: •*•!• i
EVERETT
Figure 5. AHH activity of the liver of lemon sole collected at several
locations which differed in their skin papilloma prevalences,
253
-------
AHH Activity in Flatfish and Tumor Prevalence
The possibility of a positive or negative correlation between AHH
levels in the livers of flatfish and tumor prevalence was examined. First,
we measured the enzyme activity of lemon sole populations that differed
greatly in their prevalences of skin papillomas. As seen in Figure 5,
populations with high prevalences of skin tumors can have low or high AHH
levels. Then, we compared the AHH levels of tumor-bearing lemon soles with
those in tumor-free fish of the same age group and from the same location.
No significant difference between these two fish groups was observed.
Outlook
Based on our experience with indicator organisms, we would like to
comment on possible future developments of this attractive, but as yet
unproven approach to assessing environmental hazards to man or to animal
populations. Foremost on a list of priorities should be the introduction
of genetically well-defined organisms as subjects in examining the
carcinogenic or mutagenic load of the marine environment. Although this
suggestion may smack of heresy, the fact that our current progress in the
use of lower vertebrates or invertebrates in detecting carcinogens lags far
behind the considerable advances made with inbred strains of rodents cannot
be ignored.
The use of wild, randomly mating fish populations for examining the
carcinogenicity or mutagenicity of compounds would be comparable to
performing tests on field mice collected from different geographic areas—a
situation which would be discarded as totally useless. The selection of
highly sensitive strains of bacteria, yeast, and molds for the numerous
short-term tests to detect the mutagenic properties of compounds or complex
mixtures is another example which must be emulated if we are to become
successful in introducing marine or fresh water organisms as test subjects.
The selection of genetically well characterized subjects with known
sensitivity to one or another group of obnoxious compounds appears to be a
prerequisite to establishing the use of a marine organism at a sound
scientific level.
The opinion has been repeatedly expressed that the ultimate judgment
about the carcinogenic hazard of an environment can only be obtained by
establishing tumor prevalences of the population under discussion and its
subgroups. This would mean that only a human population could provide any
indication as to whether or not a particular environment is carcinogenic
for man. Although one cannot deny the validity of such an approach, the
objective of monitoring programs, early-warning systems, and analysis of
contaminations is to prevent such an occurrence. The question is not
whether predictive tests should be introduced into the repertoire of
preventive medicine, but rather which of the numerous biological and
chemical procedures can best reflect the carcinogenic or mutagenic hazard
in a marine environment. At present, chemical analysis cannot reveal the
carcinogenic/mutagenic property of a compound. Microbial short-term
bioassays can detect the mutagenic capacity of a compound, but whether a
254
-------
positive result indicates that a compound is carcinogenic to, for example,
man, rodents, fish, or oysters, is a debatable issue. Thus, tumor
frequencies of indigenous indicator organisms could represent a highly
promising approach when the sensitivities and responses of these organisms
to particular carcinogens or complex carcinogenic mixtures become known.
Such information could be readily obtained by properly designed experiments
under controlled laboratory conditions.
We would be remiss if we listed only the difficulties encountered in
the use of indicator organisms and omitted the encouraging observations.
For example, a study of tumor prevalences along the shores of Hokkaido has
revealed an exceptionally high value (33%) among flatfish in one shallow,
poorly flushed bay, which is in a region devoid of industry, urban
activity, or agricultural usage. Subsequent analysis of BaP in the bay has
shown concentrations in the order of 55 yg/kg, whereas areas open to the
sea had low BaP concentrations, and their flatfish population also had a
low frequency of neoplasms. This pattern seems to exemplify how even
preliminary data on tumor frequencies can help to detect hotspots of
contamination. But much more impressive in its predictive value was the
discovery that tumor-like growth anomalies appeared among cultured algae,
Porphyra tenera, Nori, at the mouth of particular rivers in Kyushu, Japan
(Ishio et al., 1970, 1972a,b). The distribution pattern of algae "tumors"
combined with tumor induction experiments and chemical analyses led to the
source of contamination: PAH-releasing industrial complexes. But probably
the most convincing evidence in favor of the indicator organism was the
gradual disappearance of the tumor-like growth anomalies following the
introduction of preventive measures by the factories responsible. This
example seems to rank in importance equal to the discovery of the
carcinogenic potential of aflatoxins by investigating outbreaks of
hepatomas among trout kept at particular hatcheries and fed a particular
contaminated diet. Other fascinating, but less well-documented results
have been reported from China where chickens developed esophageal
tumors in regions in which a high frequency of esophageal carcinomas also
occurred in man. Pigs also have been found with nasopharyngeal tumors in
districts with comparable carcinomas among the human population (Lawrence,
1977).
REFERENCES
Anders, A., F. Anders, and D.L. Pursglove. 1971. X-ray-induced mutations
of the genetically-determined melanoma system of Xiphorin fish.
Experientia 27:931-932.
Anders, A., and F. Anders. 1978. Etiology of cancer as studied in the
platyfish-swordtail system. Biochim. Biophys. Acta 516:61-95.
Atlas, R.M. 1977. Studies on petroleum biodegradation in the Arctic. In:
Fate and effects of petroleum hydrocarbons in marine organisms and
ecosystems. D.A. Wolfe, Ed., Pergamon Press, New York. pp. 261-269.
255
-------
Bend, J.R., M.O. James, and P.M. Dansette. 1977. In vitro metabolism of
xenobiotics in some marine animals. Ann. N.Y. Acad. Sci.
298:505-521.
Beumer, M., and W.W. Youngblood. 1975. Polycyclic aromatic hydrocarbons
.in soils and recent sediments. Science 188:53-55.
Boulos, B.M. 1978. High risks in exposure to polycyclic aromatic
hydrocarbons. In: Carcinogenesis. P.W. Jones and R.I. Freudenthal,
Eds., Raven Press, NY. 3:439-449.
Brandenburg, J.H. and G. Kellermann. 1978. Aryl hydrocarbon hydroxylase
inducibility in laryngeal carcinoma. Arch. Otolaryngol.
104:151-152.
Brooks, R.F., G.E. McAru, and S.R. Wellings. 1969. Ultrastructural
observations on an unidentified cell type found in epidermal tumors of
flounders. J. Nat. Cancer Inst. 43:97-109.
Cooper, R.C., and C.A. Keller. 1969. Epizootiology of papillomas in
English sole, Parophrys vetulus. Natl. Cancer Inst. Monogr.
31:173-185.
DePierre, J.W., M.S. Morgan, K.A.M. Johannesen, and L. Ernster. 1975. A
reliable, sensitive, and convenient radioactive assay for benzpyrene
monooxygenase. Anal. Biochem. 63:470-484.
Dunn, B.P., and H.P. Stich. 1975. The use of mussels in estimating
benzo(a)pyrene contamination of the marine environment. Proc. Soc.
Exp. Biol. Med. 150:49-51.
Dunn, B.P. 1976. Techniques for determination of benzo(a)pyrene in
marine organisms and sediments. Environ. Sci. Technol. 10:1018-1021.
Dunn, B.P., and D.R. Young. 1976. Baseline levels of benzo(a)pyrene in
Southern California mussels. Mar. Pollut. Bull. 7:231-234.
Ishio, S., T. Yano, and H. Nakagawa. 1970. Algal cancer and causal
substances in wastes from the coal chemical industry. Presented
at Fifth International Water Pollution Research Conference, San
Francisco, CA.
Ishio, S., K. Kawabe, and T. Tomiyama. 1972a. Algal cancer and its
causes. Carcinogenic potencies of waste and suspended solids
discharged to the river Ohmuta. Bull. Jpn. Soc. Sci. Fish.
38:17-24.
Ishio, S., H. Nakagawa, and T. Tomiyama. 1972b. Algal cancer and its
cause. Separation of carcinogenic compounds from sea bottom mud
polluted by waste of the coal chemical industry. Bull. Jpn. Soc. Sci.
Fish. 38:571-576.
256
-------
Kellermann, G., C.R. Shaw, and M. Luyten-Kellermann. 1973. Aryl
hydrocarbon hydroxylase inducibility and bronchogenic carcinoma.
N. Engl. J. Med. 289:934-937.
Kellermann, G. 1977. Hereditary factors in human cancers. In: Origins
of human cancer. H.H. Hiatt, J.D. Watson, and J.A. Winsten, Eds.,
Cold Spring Harbor Laboratory, Cold Spring Harbor, NY. pp. 837-845.
Kouri, R.E., and D.W. Nebert. 1977. Genetic regulation of susceptibility
to polycyclic-hydrocarbon-induced tumors in the mouse. In: Origins
of human cancer. H.H. Hiatt, J.D. Watson, and J.A. Winsten, Eds.,
Cold Spring Harbor Laboratory, Cold Spring Harbor, NY. pp. 811-855.
Lawrence, E. 1977. Urban surveys to detect cancer. Nature 270:464-465.
Lee, R.F., R. Sauerheber, and G.H. Dobbs. 1972. Uptake, metabolism and
discharge of polycyclic aromatic hydrocarbons by marine fish. Marine
Biol. 17:201-208.
Lunde, G., and A. Bjorseth. 1977. Analysis of polycyclic aromatic
hydrocarbons in long-range transported aerosols. Nature
268:518-519.
McMahon, C.K., and S.N. Tsoukalas. 1978. Polynuclear aromatic
hydrocarbons in forest fire smoke. In: Carcinogenesis. P.W. Jones
and R.I. Freudenthal, Eds., Raven Press, New York. 3:61.
Mix, M.C., R.T. Riley, K.I. King, S.R. Trenholn, and R.L. Schaffer. 1977.
Chemical carcinogens in the marine environment. Benzo(a)pyrene in
economically important bivalve mollusks from Oregon estuaries. In:
Fate and effects of petroleum hydrocarbons in marine ecosystems and
organisms. D.A. Wolfe, Ed., Pergamon Press, NY. pp. 421-431.
Nebert, D.W., and J.S. Pel ton. 1976. Importance of genetic factors
influencing the metabolism of foreign compounds. Fed. Proc.
35:1133-1141.
Nebert, D.W., S.S. Thorgeirsson, and G.H. Lambert. 1976. Genetic aspects
of toxicity during development. Environ. Health Perspect. 18:35-45.
Nigrelli, R.F., K.S. Ketchen, and G.D. Ruggieri. 1965. Studies on virus
diseases of fishes. Zoologica 50:115-122.
Payne, J.F. 1976. Field evaluation of benzopyrene hydroxylase induction
as a monitor for marine petroleum pollution. Science 191:945-946.
Payne, J.F., and W.R. Penrose. 1975. Induction of aryl hydrocarbon
[benzo(a)pyrene] hydroxylase in fish by petroleum. Bull. Environ.
Contam. toxicol. 14:112-116.
-------
Peters, N., G. Peters, H.F. Stich, A.B. Acton, and G. Bresching. 1978. On
differences in skin tumors of Pacific and Atlantic platyfish. J. Fish
Dis. 1:3-25.
Poland, A., and A. Kende. 1977. The genetic expression of aryl
hydrocarbon hydroxylase activity: evidence for a receptor mutation in
nonresponsive mice. In: Origins of human cancer. H.H. Hiatt, J.D.
Watson and J.A. Winsten, Eds., Cold Spring Harbor Laboratory, Cold
Spring Harbor, NY. pp. 847-867.
Schwab, M., S. Abdo, M.R. Ahuja, G. Kollinger, A. Anders, and F. Anders,
1978a. Genetics of susceptibility in the platyfish/swordtail tumor
system to develop fibrosarcoma and rhabdomyosarcoma following
treatment with N-methyl-N-nitrosourea (MNU). Z. Krebsforsch.
91:301-315.
Schwab, M., J. Haas, S. Abdo, M.R. Ahuja, G. Kollinger, A. Anders, and F.
Anders. 1978b. Genetic basis of the susceptibility for the induction
of neoplasms by N-methyl-N-nitrosourea (MNU) and X-rays in the
platyfish/swordtail tumor system. Experientia 34:780-782.
Stich, H.F., and A.B. Acton. 1976. The possible use of fish tumors in
monitoring for carcinogens in the marine environment. Progr. Exp.
Tumor Res. 20:44-45.
Stich, H.F., A.B. Acton, and B.P. Dunn. 1976a. Carcinogens in estuaries,
their monitoring and possible hazard to man. INSERM 52:83-94.
Stich, H.F.,A.B. Acton, and C.R. Forrester. 1976b. Fish tumors and
sub-lethal effects of pollutants. J. Fish. Res. Board Canada
33:1993-2001.
Stich, H.F., A.B. Acton, K. Oishi, F. Yamazaki, T. Harada, and H.G. Moser.
1977a. Systematic collaborative studies on neoplasms in marine animals
as related to the environment. Ann. N.Y. Acad. Sci. 298:374-388.
Stich, H.F., A.B. Acton, B.P. Dunn, K. Oishi, F. Yamazaki, T. Harada, G.
Peters, and N. Peters. 1977b. Geographic variations in tumor
prevalence among marine fish populations. Int. J. Cancer
20:780-791.
Wellings, S.R., R.G. Chuinard, and M. Bens. 1965. A comparative study of
skin neoplasms in four species of pleuroneitid fishes. Ann. N.Y.
Acad. Sci. 126:479-501.
Wellings, S.R. 1969. Neoplasms and primitive vertebrate phylogeny:
echinodenns. prevertebrates and fish—a review. Natl. Cancer Inst.
Monogr. 31:59-128.
Wellings, S.R., C.E. Alpers, B.B. McCain, and M.S. Myers. 1977. Fish
disease in the Bering Sea. Ann. N.Y. Acad. Sci. 298:290-304.
258
-------
Yamazaki, F., T. Hibino, K. Oishi. T. Harada, H.F. Stich, and A.B. Acton.
1978. X-cell morphology in the epidermal papillomas of flatfish
collected from coastal waters of Hokkaido, Japan. Bull. Jpn. Soc.
Sci. Fish. 44:407-413.
259
-------
POLYNUCLEAR AROMATIC HYDROCARBONS
IN ESTONIAN UATER, SEDIMENTS, AND AQUATIC ORGANISMS
by
Pavel Bogovski, Ingeborg Veldre, and Aino Itra
Institute of Experimental and Clinical
Medicine, Ministry of Health,
Tallinn, the Estonian SSR,
and
Lia Paalme
Institute of Chemistry, Academy of Science,
Tallinn, the Estonian SSR
ABSTRACT
Benzo(a)pyrene (BaP) was selected as a model polynuclear
aromatic hydrocarbon in a survey of Estonian marine and
freshwater water column, sediments, algae, aquatic plants, and
fish. Uater samples had low BaP content; sediments were much
higher. Water samples taken in the summer season had the lowest
BaP content. Fish livers had higher BaP residues than gills and
gonads contents. Weight class of fish was not correlated with
BaP residues in the case of herring. However, large predatory
fish had lov/er residues than medium-sized predatory fish.
Species of fish with high fat content had higher residues of BaP
than low-fat fish.
INTRODUCTION
Estonia has over 1500 lakes, 7000 rivers, and 3780 kilometers of
coastline that provide important water resources for freshwater and Baltic
Sea fisheries, agar harvests for use in ice cream and other foods, recrea-
tion, and a source of drinking water from lakes and rivers.
Many research papers have shown that polycyclic aromatic hydrocarbons
(PAHs) are ubiquitous in the aquatic environment and that some are
carcinogenic. Benzo(a)pyrene (BaP) has been used as an indicator of the
diverse group of PAHs because of its known carcinogenic activity,
stability, frequency of occurrence, and ease of chemical analysis. There
are no previously reported studies on the occurrance of BaP in the aquatic
environment of Estonia.
260
-------
MATERIAL and METHODS
Our preliminary survey was undertaken to determine the distribution of
BaP in water, sediments, and tissue residues of various aquatic organisms.
Sampling sites selected were of economic significance and represented
different types of water bodies (Maemets and Raitviir, 1977). Most of the
sampling was undertaken at least once per season, but some lakes were
sampled only during the summer. Over 2000 samples were analyzed from 64
lakes, 21 rivers, and 8 bays of the Baltic Sea. Fish samples were
collected from 22 species; more than 400 samples were analyzed.
BaP was determined by the following method: 3 a water was repeatedly
extracted with ethyl ether; 100 m£ aliquots of the solvent were
successively used to extract each water sample. The extractions were
pooled, steam distilled to dryness, and extracted with n-octane. BaP was
determined by quasi-linear luminescent spectra using a modification of the
method of Khesina (1968). Samples were read at - 196° C in solid
parafins.
Fish residues were analyzed for BaP as follows: organs or fillets
were homogenized, 100 g homogenized tissues were immersed in 100 mi ethyl
alcohol, and 15 to 25 g KOH added. Then the mixture was boiled for 2 hr.
The resultant mass was immersed in water and extracted five times with
ethyl ether. The extract was washed initially with acidized water (5%
^$04) and finally with distilled water. NA2S04 was added to the
extract, and the ether was removed by steam distillation. The residue was
dissolved in n-octane and read as in the water analyses procedure.
Water plants were extracted in benzene with a Soxleth apparatus. The
benzene was removed by steam distillation and the residue dissolved in
n-octane. The BaP in n-octane was separated by thin layer chromatography,
using several solvent systems. The fluorescent front was scraped off the
plate, eluted with benzene, and then measured by luminescence as in water
samples. The fish used in this study are described in Table 1.
RESULTS
Table 2 summarizes the BaP content of water, sediments, and aquatic
organisms. All water samples had BaP concentrations below the USSR
sanitary limit (0.005 yg/£ water). Seawater samples contained slightly
more BaP than freshwater samples; the lowest concentrations were found in
lakes and rivers of non-industrialized areas (Veldre j2t _al_., 1977; 1979).
During summer, BaP content was lower than in winter when lakes in Estonia
are covered by ice. Although water concentrations were low, BaP
concentrations were high in saltwater and freshwater sediment (Bogovski
.et _a]_., 1977). Freshwater aquatic plants had exceptionally high
concentrations of BaP in tissue residues.
261
-------
TABLE 1. CHARACTERISTICS OF FISH STUDIED
Fish
Carp
Seatrout
Tench
Salmon
Pike
Eel
Perch
Roach
Bream
Pike-perch
Baltic
sprat
Baltic
herring
Silver
bream
White bream
Brown trout
Minnow
Species
Cyprinius
carpio
Salmo trutta
trutta
Tinea tinea
Salmo salar
Esox lucius
Anguilla an-
guilla
Perca fluvia-
tilis
Rutilus ru-
tilus
Abramis
brama
Lucioperca
lucioperca
Sprattus
sprattus
balticus
Clupea haren-
gus membras
Vimba vimba
Blicca
bjoerkna
Salmo trutta
Phoxinus
Water
Freshwater
Saltwater
Freshwater
Freshwater
Salt & Freshwater
Freshwater
Fresh & Saltwater
Fresh & Saltwater
Freshwater
Freshwater
Fresh &
Saltwater
Saltwater
Saltwater
Salt & Freshwater
Freshwater
Freshwater
Freshwater
Fat Content
high
high
high
high
low
high
low
low
high
low
high
low
low
low
high
low
phoxinus
262
-------
TABLE 2. LEVEL OF BaP IN WATER, SEDIMENTS, AND AQUATIC ORGANISMS
Source
Saltwater
Freshwater
Sample
water
sediments
algae ^
zooplankton
baltic sprat
baltic herring
pike-perch
water
sediments
Number of
Samples
29
16
2
13
23
12
9
32
2
Concentration (yg/kg, and
yg/£, respectively)
mm
<0.0000
0.24
2.40
0.26
0.22
0.08
0.06
<0.0000
2.10
max
0.0123
11.50
3.00
5.10
4.48
3.10
8.50
0.0047
4.30
arithmetic
mean
0.0007
2.57
2.70
1.11
0.99
1.22
2.25
0.0006
3.20
waterplants
reed(Phragmites)
duckweed
10
1.32 129.0
36.90
(Lemna)
roach
perch
3
6
8
19.10
0.03
0.02
112.4
3.04
1.90
52.40
0.73
0.83
Collected by hand with aqualung
Collected by plankton tongs using a silk net No.
39
Table 3 summarizes the BaP content in fresh and saltwater fish
fillets. Most fish contained 1 to 2 yg/kg BaP (10 to 20 yg/kg on a dry
weight basis). Some fish (salmon, eel) contained 5 to 10 yg/kg BaP. Among
the freshwater fish tested, the pike, minnow, and pike-perch had low tissue
residues as compared to eel, salmon, and some saltwater species. Table 4
lists the BaP content in muscular tissue of predatory and non-predatory
fish. The results showed no marked difference in the BaP content of the
fish. Table 5 demonstrates that fish livers have higher BaP residues than
gills or roe and, in some cases, fish fillet.
Figure 1 summarizes the BaP content in low- and high-fat fish.
Species of fish with high-fat content have higher residues of BaP than
low-fat fish. The relationship of BaP tissue residues to the weight of
fish is illustrated in Figure 2. Weight class of fish is not correlated
with BaP residues in the case of Baltic herring. Large (700 to 1100 g)
predatory fish (pike-perch and tench) have lower residues than medium-sized
(400 to 600 g) fish.
263
-------
TABLE 3. CONTENT OF BaP IN THE MUSCULAR TISSUE OF FISH
Source
divers of
North
Estonia
Various
lakes
Bays
and
gulfs
Species
trout
minnow
salmon
tench
pike
eel
perch
roach
bream
pike -perch
baltic
sprat
baltic
herring
pike -perch
perch
silver-breetfn
white bream
Number
Samples
4
2
2
8
11
7
8
6
32
4
23
12
9
2
2
2
of Concentration yg/kg
min
< 0.01
0.17
0.11
0.09
0.05
0.00
0.02
0.03
<0.01
0.23
0.22
0.08
0.06
0.10
0.08
0.20
max
1.72
0.19
11 .82
2.94
0.41
24.00
1.90
3.04
10.00
0.76
4.48
3.10
8.50
0.52
0.13
0.48
arithmetic
mean
0.685
0.180
5-960
0.962
0.187
4.370
0.835
0.733
1.694
0.417
0.99
1 .22
2.25
0.31
0.11
0.34
264
-------
TABLE 4. THE LEVEL OF BaP IN PREDATORY AND NON-PREDATORY FISH
Fish Number of
Samples
Predatory
p ike 1 1
pike -perch 4
t ro ut 4
eel 7
Non predatory
bream 32
tench 8
roach 6
Concentration p_g_/kg
mm
0
0
*• 0
-------
4
5
I 1 high fat
I
n 1 2 3 4 BaP,ug/kg
Figure 1. Relationship of BaP tissue residues in high and low fat fish,
1 - eel
2 - tench
3 - bream
4 - perch
5 - roach
6 - pike
400-600
<300
>100
60-99
40-49
30-39
20-29
<20
[§^§j pike- perch
I ^^
|
1 I baltic herring
0
3 BaP, yg/kg
Figure 2. Relationship of BaP tissue residues to weight of fish.
266
-------
REFERENCES
Bogovski, P.A., I.A. Veldre, A.R. Itra, and L.P. Paalme. 1978.
Benzo(a)pyrene in fish of Estonian water bodies. Gigiena and
Sanitaria, Nr. 4:111-113.
Bogovski, P.A., I.A. Veldre, A.R. Itra, and L.P. Paalme. 1977. Some data
on the migration of benzo(a)pyrene in water environment. In: Proc.
of the third symposium of epidemiologists, microbiologists, and
hygienists of Estonia, November 22-23, 1977. Tallinn, pp. 244-245.
Fedoseeva, G., and A. Khesina. 1968. Determination of polycyclic aromatic
hydrocarbons by using quasi-linear luminescence spectra. J.
Prikladnoy Spektroskopii 9(2):282-288.
Maemets, A., and A. Raitviir. 1977. The classification of Estonian lakes
based on the analyses of principal components and coordinates. Proc.
Acad. Sci. Estonian SSR. Biol. Ser. 26(2):138-148.
Veldre, I.A., M.A. Rahu, A.P. Ilnitski, L.G. Lohova, and N.I. Sherenesheva.
1977. A study of carcinogenic hydrocarbons in special benzo(a)pyrene
in the water bodies of Estonian SSR. "Water Resour. 3:147-152.
Veldre, I., A. Itra, and L. Paalme. 1979. Levels of Benzo(a)pyrene in oil
shale industry wastes in some bodies of water in the Estonian S.S.R.
and in water organisms. Environ. Health Perspect. 30:211-216.
267
-------
BIOACCUMULATION AND TOXICITY IN ENGLISH SOLE PAROPHRYS VETULUS
FOLLOWING WATERBORNE EXPOSURE TO BENZO(A)PYRENE
by
M.L. Landolt, S.P. Felton, W.T. Iwaoka,
B.S. Miller, D. DiJulio, and B. Miller
University of Washington, School of Fisheries,
Seattle, Washington, 98195
ABSTRACT
Juvenile English sole, Parophrys vetuliis, measuring 10 to
30 cm in total length were held in an all glass/teflon aquarium
system at 11° C and exposed continuously for a 30-day period to
artificial seawater containing low levels of benzo(a)pyrene
(BaP). The BaP was present both in solution [(1.0 parts per
billion (ppb)] and as crystals coated onto the sand substrate.
At the termination of the test, the fish were examined to
determine the level of BaP in the tissues, the level of hepatic
aryl hydrocarbon hydroxylase (AHH) activity, and the extent of
tissue damage.
Gas chromatographic analysis revealed detectable levels of
BaP in tissue extracts as well as markedly significant
quantities adsorbed onto integumental surfaces. Awareness of
this adsorption phenomenon is critical to an understanding and
an accurate determination of whole body uptake. Enzyme analysis
by fluorescence spectroscopy using pooled data indicated that
mixed function oxygenase activity in the experimental fish was
increased by over 100%. However, closer inspection of
individual enzyme levels showed wide variability and a bimodal
pattern of AHH activity. Histopathological examination
revealed the presence of free blood in either the pericardial or
abdominal cavities in 55% of the test fish. This change was not
noted in control animals. In addition, the exposed fish had
increased numbers of immature white blood cells in peripheral
circulation.
268
-------
INTRODUCTION
Benzo(a)pyrene (BaP) is a polycyclic aromatic hydrocarbon which has
frequently been selected as a model compounmd for studies on the fate and
effects of this ubiquitous group of chemicals. BaP occurs in such media as
cigarette smoke, automobile exhaust, coal tar, and soot (IARC, 1973), and
resultant environmental contamination by this compound has been noted in
air, vegetation, fresh and salt water, food, and soil (IARC, 1973). It is
not surprising therefore that significant bioaccumulation of BaP has been
noted in several aquatic organisms (Neff and Anderson, 1975; Dunn and
Stich, 1976; Varanasi and Mai ins, 1977).
BaP is known to be carcinogenic in a variety of homeothermic animals
(IARC, 1973), but little is known of its toxic effects or mutagenic
potential in marine poikilotherms (Hodgins et^^l_., 1977). The experiment
herein reported represents an interdisciplinary approach to assessing such
effects and uses the English sole, Parophrys vetulus, as a model species.
English sole are members of the family Pleuronectidae. They begin
life as pelagic eggs, live for a time as littoral larvae, and, after
metamorphosis, become benthic organisms living in intimate contact with
sediments; thus, in the course of their development these fish are exposed
to all levels of the water column. They have world-wide distribution, are
not migratory, and their diet includes, among other things, bivalve
molluscs which are capable of concentrating and storing environmental
contaminants such as BaP (Lee et^ a]_., 1972a; Fossata and Canzonier, 1976).
In our study juvenile English sole were exposed for 30 days to chronic
low levels of waterborne BaP. At test termination, the fish were analyzed
to determine whether bioaccumulation of BaP had occurred, whether hepatic
BaP hydroxylase activity was altered, and whether tissue lesions were
present.
MATERIALS AND METHODS
Collection and Housing of Fish
Fifty juvenile English sole 10 to 30 cm in total length were collected
by beach seine at Eagle Cove, San Juan Island, WA. The fish were
transported to the University of Washington, School of Fisheries, where
they were placed in five 230-£ aquaria. These holding tanks contained
artificial seawater and were part of a closed aquarium system constructed
entirely of glass and teflon as developed for toxiological studies. The
fish were held at 11° C and allowed to acclimate for one month prior to
testing. The actual exposure period lasted for 30 days.
Addition to and Quantitation of BaP in Seawater
Fine sand was washed with soap and water, then rinsed with water and
acetone, and baked dry. Purified BaP was dissolved in methylene chloride,
coated onto the sand (500 mg BaP/kg sand), and the methylene chloride was
269
-------
subsequently evaporated under a stream of nitrogen. To each of four
aquaria was added 750 g of the BaP-coated sand. In addition, 2 kg of
cleaned coarse sand, similarly coated, was packed into glass tubing and
placed in the water delivery system of the four experimental aquaria. Only
clean sand was placed in the control tank.
The concentration of dissolved BaP was monitored throughout the test
period by the following procedure. A known volume of water (50 mi) was
filtered through methylene-chloride-washed filter paper, then placed in a
250 mi separatory funnel, and extracted three times with 25 mi of reagent
grade methylene chloride. The organic solvent portions were subsequently
combined, made up to a known volume (100 nu), and adequately mixed. An
aliquot of the solution was analyzed on a Perkin-Elmer MPF fluorescence
spectrophotometer (excitation wavelength 365 nm, slit 10 nm, spectrum
scanned 340 to 500 nm). Maximum emission was obtained at 405 nm, and the
height of the peak was used to quantitate the amount present. Large
quantities of crystalline BaP were in circulation in the system; however,
due to its non-polar nature only about 1.0 ppb BaP was actually in solution
at any given time. Filtration of the water prior to analysis allowed for
removal of the crystals and for accurate determination of water-borne
concentrations.
Quantitation of BaP Tissue Levels by Gas/Liquid Chromatography
Fish were sacrificed by a blow to the head and then frozen for
analysis. Before analysis, each fish was washed with methylene chloride to
remove any crystalline BaP adsorbed onto integumental surfaces. The entire
animal was then homogenized. A portion of the tissue was accurately
weighed and placed in a flask containing 100 mi 0.5 M KOH in methyl
alcohol, 100 mi contaminant-free distilled water, and boiling stones. The
mixture was refluxed for 18 to 24 hr. After saponification, the mixture
was transferred to a l-£ separatory funnel containing 100 mi nanograde
hexane, 35 mi nanograde toluene, and 100 mi solution of saturated
N32S04 in distilled water. The funnel was shaken for 2 min, the layers
allowed to separate, and the hexane-toluene layer decanted into a clean
100-m£ flask. The remaining aqueous-methanol phase was subsequently
partitioned with two additional aliquots of 100 mi hexane plus 35 m£
toluene, and all three hexane-toluene aliquots were combined. The mixture
was then reduced to a small volume (3 to 5 mi) on a rotary evaporator and
quantitatively transferred to a 25-m£ mixing cylinder with hexane toluene
rinses.
Between 0.2 and 0.5 g of non-saponifiable fraction was dissolved in no
more than 5 mi hexane and transferred to a chromatograph column (22 mm
I.D.) containing 16.5 cm Florisil and an upper 1.2-cm layer of anhydrous
^$64; 300 mi hexane was passed through the column at a flow rate of
10 nu/min followed by 300 mi of a 30% (v/v) mixture of methylene chloride
in hexane at the same flow rate. The volume of this fraction was reduced
on a rotary evaporator, transferred quantitatively to a glass vial, and
reduced to dryness with a stream of N£ gas.
270
-------
Gas liquid chromatographic analysis was performed on the Hewlett-
Packard 402 high performance gas chromatograph under the following
conditions: Column: 6 ft ,x 1/4 inches O.D. all glass column
Packing: Supelco SP 2350 on acid-washed 100/120 chromosorb W
Carrier Gas: N£, 50 nu/min, Rotameter setting 4.0
Hydrogen: 35 nu/min, Rotameter setting 3.5
Air: 20 nu/min, Rotameter setting 3.5
Temperature Column: 255° C, detection 290° C, injector 280° C
Range 1, attenuation 16 or 30 X
Histopathological and Hematological Analysis
Approximately 1 ma of blood was withdrawn into a heparinized syringe
by caudal vein puncture of unanesthetized animals. The fish were then
sacrificed by immersion in Bouins solution, and the abdominal cavity was
injected with more fixative. After 24 hr, tissues were washed and
transferred to 70% ethanol.
Serial whole body cross sections were dehydrated in a graded series of
ethanols, cleared in xylene, and embedded in paraffin; 6-y sections were
prepared using a rotary microtome and were .stained with hematoxylin and
eosin. Smears were prepared from the heparinized blood and stained with
Leishman and Giemsa stains. Capillary tubes were filled with a portion of
the remaining blood for hematocrit determination.
Hepatic Aryl Hydrocarbon Hydroxylase Analysis
The fish were sacrificed by a blow to the head and the liver was
immediately removed in toto, weighed, diluted 1:23 (wt:vol) in cold 0.15%
KC1, and homogenized. The homogenate was placed in a refrigerated (4° C)
centrifuge and spun for 20 min at 100,000 rpm (9000 x g), and the clear
supernatant fluid was collected.
The supernatant fluid was next incubated for 20 min at 19.5° C in a
mixture of Tris buffer (0.1 M pH 7.2), MgCl (33 mM), distilled water, BaP
(0.4 mg/nu in methanol), and NADPH (8 mg/nu). The reaction was stopped by
the addition of 1 ml cold acetone, and the mixture immediately extracted by
constant shaking (25° C) for 10 min in 3.25 nu reagent grade hexane. A
portion of the hexane layer (1 mi) was decanted, placed in 3 ma cold NaOH,
and extracted for 5 min at 25° C with constant shaking (Nebert and Gelboin,
1968).
The NaOH layer was then analyzed on a Perkin-Elmer MPF spectro-
fluorimeter (396 nm excitation, 522 nm emittance) and the results expressed
in fluorescence units (Fu)/mg protein. The protein determinations were
performed according to the method of Lowry (Lowry et^^l_., 1957).
271
-------
RESULTS
At the termination of the 30-day exposure, all of the fish appeared to
be in generally good health and were feeding regularly. Of 40 experimental
fish, 32 survived. No control animals died during the test.
Levels of BaP in Seawater
BaP is a non-polar compound which is sparingly soluble in seawater
(20 ng/mz saturation) and which, due to its hydrophobic nature, is
difficult to maintain in solution. Sufficient quantities of BaP were added
to our system to achieve supersaturation; however, upon letting the water
stand for a few hours, the amount of BaP in solution declined to less than
1 ppb and persisted at that level throughout the experiment. Large
quantities of crystalline BaP were present in the aquaria at all times as
evidenced by filtration prior to water analysis.
Uptake of BaP into Fish Tissue
No BaP was found in the three control fish analyzed (Table 1);
however, measurable levels were found in 10 of 11 experimental animals.
Fish 1 through 5 were analyzed without benefit of prior washing and yielded
whole body accumulations of from 137 to 282 ppb in test animals (Table 1).
BaP has a strong tendency to adsorb onto surfaces, such as glass
aquarium walls and sand substrate. Because of this phenomenon, the
integuemental surfaces of fish 6 through 14 were washed repeatedly with
methylene chloride, without scraping, to remove loosely bound BaP. Table 2
shows that significant amounts of BaP were removed in the course of these
washings. This material represents superficial adsorption rather than true
uptake.
Table 1 shows the uptake of BaP by fish 6 through 14 following removal
of the compound from integumental surfaces. High individual variability
was noted with values ranging from undetectable levels to a high of
499 ppb. Such discrepancies in the amount of parent compound present in
tissues may have related to differential capacity to metabolize BaP (see
AHH activity section).
HEPATIC AHH Activity
At test termination, 11 experimental and four control fish were
sacrificed for analysis of hepatic AHH activity. On the basis of pooled
data, the mean level for the experimental fish was 57.9 Fu/mg protein as
compared with 22.3 Fu/mg protein for control animals (Table 3). Thus, an
increase of over 100% in enzyme activity appeared to occur following
continuous exposure to BaP.
272
-------
TABLE 1. UPTAKE OF BaP BY ENGLISH SOLE FOLLOWING 30-DAY WATERBORNE
EXPOSURE
Total
Fish Weight of
Number Fish (g)
Unwashed
1 (exptl)
2 (control)
3 (exptl)
4 (control)
5 (exptl)
Washed
5 (control)
7 (exptl)
8 (exptl)
9 (exptl)
10 (exptl)
11 (exptl)
12 (exptl)
13 (exptl)
14 (exptl)
77.9
15.0
40.6
18.5
50.0
20.1
10.2
5.6
31.3
43.3
6.9
11.2
4.9
69.6
Wet
Weight
Analyzed (g)
40.1
11.0
34.8
16.1
38.2
15.3
7.2
5.3
21.4
33.1
5.4
8.7
4.9
60.0
Total
Amount of BaP
in Tissue
Analyzed (ng)
5,500
N.D.
7,150
N.D.
10,770
N.D.
270
690
810
16,530
30
3,230
2,240
1,710
Concentration (ppb)
137
N.D.
205
N.D.
282
N.D.
37
130
38
499
N.D.
371
457
28
TABLE 2. CONCENTRATION OF BaP FOLLOWING WASHING WITH METHYLENE CHLORIDE
OF INTEGUMENTAL SURFACE OF ENGLISH SOLE FOLLOWING 30-DAY
WATERBORNE EXPOSURE
Fish Number
Total Weight of Fish (g) Total BaP found in Extract (ng)
6 (control)
7 (exptl)
8 (exptl)
9 (exptl)
10 (exptl)
11 (exptl)
12 (exptl)
13 (exptl)
14 (exptl)
20.1
10.2
5.6
31.3
43.3
6.9
11.2
4.9
69.6
N.D.
285
414
230
395
993
4,450
15
1,695
273
-------
TABLE 3. HEPATIC AHH ACTIVITY IN ENGLISH SOLE FOLLOWING 30-DAY EXPOSURE
TO WATERBORNE BaP
Control Fish Activity FU/mg Prot. Experimental Fish Activity FU/mg Prot.
1 18.2 1
2 42.9 2
3 7.9 3
4 20.3 4
5
6
7
8
9
10
11
17.3
16.7
2.3
209.1
5.9
2.4
42.3
146.5
10.3
179.3
5.0
MEAN 22.3 57.9
Closer inspection of the data revealed that wide individual variation
occurred among the test animals, the level of enzymes varying from 2.3 to
209.1 Fu/mg protein. If the three fish with highest activities (Nos. 4, 8,
and 10) were eliminated, the mean level of AHH activity among the
experimental fish (12.8 Fu/mg protein) would be significantly decreased to
a point even lower than that of the controls.
Since the compound was administered via a waterborne route rather than
by feeding, all of the experimental animals should have been exposed to
comparable toxicant levels. Thus, it appeared that the population of fish
tested fell into two categories on the basis of BaP hydroxylase activity:
those that were capable of responding to induction and those that were
not.
Histopathology
In general, the animals were in excellent health. It was not possible
to reliably distinguish treated from the untreated fish, because most of
the lesions noted were common to both groups. None of the fish contained
abdominal fat, although many had food in the lumen of the gastrointestinal
tract.
Parasitism was noted in all the fish and consisted primarily of
external infestations by the monogenetic trematode, Gyrodactylus sp., and
of internal infestation by an unidentified microsporidan found both inter-
274
-------
and intracellularly in the skeletal muscle of the body wall and intestine.
Little inflammatory response was generated toward the organisms, which are
common parasites of Puget Sound flatfish in their natural environment.
Limited numbers of fish in both groups evidenced subacute dermatitis
apparently of bacterial origin. These lesions were characterized by
epidermal erosions, peripheral epithelial hyperplasia, dermal inflammatory
infiltrates, hemorrhage, edema, and limited involvement of underlying
muscular tissue. Bacterial colonies were prominent and located primarily
in dermal connective tissue.
Hepatic vacuolation was found in both groups and, depending upon the
individual, had either a diffuse or multifocal distribution. Similarly,
there was also variability in the relative abundance of melanin macrophage
centers in the liver.
One change was noted solely in the experimental animals. Five of the
nine exposed fish were found to have extravasated blood either in the
peritoneal or pericardial cavity, the extent of which varied from nild to
severe (Figure 1).
M
,
B
Figure 1. Extravasation of blood (B) into the peritoneal cavity. Liver
(L), body wall musculature (M). Hematoxylin and eosin, 80 x.
275
-------
Hematology
At the termination of the test, blood was drawn from 15 fish (11
experimental, 4 control) for determination of hematocrit levels and for
differential blood cell counts.
The hematocrit values were quite low compared with those published for
other fish species, but were fairly consistent for both exposed and
unexposed fish (Table 4). The experimental fish did have a slightly lower
mean value of 12.7% compared to 16.2% for the controls.
TABLE 4. PACKED RED BLOOD CELL VOLUMES (HEMATOCRIT) OF JUVENILE ENGLISH
SOLE EXPOSED AND UNEXPOSED TO WATERBORNE BaP FOR 30 DAYS
Group Specimen Number Hematocrit (*)
I. Control 24
22
19
20
MEAN
II. Experimental 13
15
16
18
17
27
21
14
25
26
28
MEAN
18
13
16
JjJ
16.2
7
n
12
10
20
13
12
13
3
16
23
12.7
Leucocyte differentials were classified according to Ellis (1976) and
revealed a disparity between the experimental and control fish (Table 5).
Because of poor fixation, it was not possible to reliably distinguish
lymphocytes from thrombocytes, which were therefore scored together.
276
-------
In the control fish, the predominant cell type was the small lymphocyte/
thrombocyte. These cells accounted for an average of 66% of the white
cells. In experimental fish, these cells were also the most predominant
but represented an average of only 40%. In the control fish, granulocytes
accounted for 26% of the white blood cells and in the experimental fish,
38%. The most striking difference between the two groups was in the number
of very large cells which appeared to represent immature blast cells and,
to a lesser extent, monocytes. In the control fish, the large cells made
up an average of 8% of the cells, but in the experimental animals this
value almost tripled to an average of 22%. Without special stains the
identity of the blast forms could not be determined; however, they were
most compatible with cell types from either the erythrocytic or granu-
locytic series.
TABLE 5. LEUCOCYTE DIFFERENTIAL COUNTS FOR JUVENILE ENGLISH SOLE EXPOSED
AND UNEXPOSED TO WATERBORNE BaP FOR 30 DAYS
Group
Control
Specimen Lymphocytes/
Number thrombocytes (%)
24
22
19
20
MEAN
66
82
68
52
66
Granulocytes (%'.
32
U
18
38
26
Monocytes/
) blast cells (%
2
4
14
10
8
Experimental
13
15
16
18
27
14
25
26
28
MEAN
24
16
14
36
48
60
20
58
72
40
28
52
40
36
48
38
46
22
26
38
48
32
46
28
4
2
28
10
_6
22
DISCUSSION
Toxicity tests involving heterogeneous populations of wild fish
present a formidable challenge which should not be undertaken lightly.
Unlike the laboratory mouse, fish collected in the field represent a
diverse group of individuals with varied genetic, nutritional, and health
histories. In light of this inherent diversity, test conditions should be
as meticulous as possible so that data can be interpreted reasonably. This
care is particularly critical when low levels of toxicants are used.
277
-------
To this end, the fish used in our study were collected from a site
removed from industrial activity. The fish were maintained in artificial
seawater free of petroleum hydrocarbons and were housed in an aquarium
system constructed entirely of glass and teflon to avoid leaching of
extraneous compounds associated with certain plastics, PCV tubing, and
metals. The use of control animals housed under similar conditions does
not lessen the need for such precautions since a synergistic effect may be
utterly different from that caused by the compound under study. One
further protective measure was purification of the test compound. Such
processing revealed to us the presence of a polar contaminant which behaved
as a direct mutagen (unpublished data) in the Ames Salmonella assay (Ames
et_ _al_., 1975), and which could have altered the results of our experiment.
Quantitation of BaP in Seawater
Studies have been conducted exposing fish and shellfish to waterborne
BaP (Lee et al_., 1972a, 1972b; Lee, 1975; Meff and Anderson, 1975; Lee et
al., 197677 In some of these studies, the actual concentration of the
compound in solution was not distinguished from that in crystalline form.
Tests in our laboratory indicate that BaP is rapidly removed from solution
and that large portions of the compound are not readily available to fish.
Investigators should note this fact when reporting exposure conditions.
Uptake of BaP
Studies of whole-body accumulation of waterborne BaP (Lee et al.,
1972b) have frequently either failed to consider surface adsorption of the
chemical or have attempted to remove adhering particles by aqueous
washings. Due to the non-polar nature of BaP, such procedures are
unsatisfactory. The substitution of organic solvents, such as methylene
chloride, provides more complete removal. Distinction between adherent
particles and tissue concentrations are critical to accurate determinations
of body burdens.
Our study did not attempt to distinguish metabolites; however, further
tests in our laboratory (unpublished data), in which metabolites were
identified, verified that substantial quantities of BaP are stored as
parent compound. The fish used in the current experiment exhibited marked
variability in bioaccumulation of BaP. This may have reflected
differential ability to metabolize the substance.
Hepatic AHH Activity
Measurement of piscine microsomal enzyme activities has been suggested
as a means of monitoring aquatic pollution (Payne, 1976). Results from our
study suggest that AHH activity may be a poor index of chemical exposure.
278
-------
In the fish examined following 30-day exposure, two distinct groups of
organisms emerged. Enzyme levels in one group were equal to or lower than
those of control fish; but dramatically increased in the other. This
disparity might be explained by genetic factors that render the animal
either responsive or nonresponsive to induction for a given enzyme system.
Such a phenomenon is known in man for AHH (Conney and Burns, 1972). If
piscine familial differences exist in terms of enzyme activity, they might
help to identify sub-populations of high-risk animals.
Histopathology and Hematology
The presence of extravasated blood in 55% of the test animals is
suggestive either of endothelial damage or of alterations in membrane
permeability. The possibility cannot be excluded, however, that bleeding
may have resulted from trauma, such as that induced by migrating parasites.
Further light and electron microscopy will be needed to resolve the
question; however, similar results have been noted in other fish species
exposed to petroleum hydrocarbons (Vishnevetskii, 1961).
The shift toward more immature white blood cells in peripheral blood
may have resulted from alterations in, hematopoietic tissue. Even though no
lesions were noted in the anterior kidney, results from other waterborne
and injection experiments that we conducted corroborate these findings.
The implication of such damage is obvious. If leucopoiesis were impaired,
the animal would have decreased resistance to infectious agents and might
be more susceptible to neoplasia due to diminished cell mediated immunity
(Rubin, 1964).
ACKNOWLEDGEMENTS
This study was supported by National Institute of Environmental Health
Sciences Contract number N01-ES-7-2101.
REFERENCES
Ames, B.N., J. McCann, and E. Yamasaki. 1975. Methods for detecting
carcinogens and mutagens with the Salmonel1a-mammalian-mlcrosome
mutagenicity test. Mutation Research 31:347-364.
Conney, A.H., and J.J. Burns. 1972. Metabolic interactions among
environmental chemicals and drugs. Science 178:576-586.
Dunn, B.P., and H.F. Stich. 1976. Release of carcinogen benzo(a)pyrene
from environmentally contaminated mussels. Bull. Environ. Contam.
Toxlcol. 15:398-401.
Ellis, A.E. 1976. Leucocytes and related cells in the plaice,
Pleuronectes platessa. J. Fish. Biol. 8:143-156.
Fossato, V.U., and W.J. Canzonier. 1976. Hydrocarbon uptake and loss by
the mussel Mytilus edulis. Marine Biol. 36:243-250.
279
-------
Hodgins, H.O., B.B. McCain, and J.W. Hawkes. 1977. Marine fish and
invertebrate diseases, host disease resistance and pathological
effects of petroleum. In: Effects of petroleum on arctic and
subarctic marine environments and organisms. Vol. II Biological
effects. D.C. Malins, Ed., Academic Press, New York. pp. 95-128.
International Agency for Research and Cancer (IARC). 1973. Monographs on
the evaluation of carcinogenic risk of chemicals to man, Vol. III.
World Health Organization (WHO), Lyon, France, pp. 91-136.
Lee, R.F. 1975. Fate of petroleum hydrocarbons in marine zooplankton.
In: Proceedings of 1975 Conference on Prevention and Control of Oil
Pollution. American Petroleum Institute, Washington, DC.
pp. 549-553.
Lee, R.F., R. Sauerheber, and A.A. Benson. 1972a. Petroleum hydrocarbons:
uptake and discharge by the marine mussel, Mytilus edulis. Science
77:344-346.
Lee, R.F., R. Sauerheber, and G.H. Dobbs. 1972b. Uptake, metabolism and
discharge of polycyclic aromatic hydrocarbons by marine fish. Mar.
Biol. (Berl.) 17:201-208.
Lee, R.F., C. Ryan, and M.L. Neuhauser. 1976. Fate of petroleum
hydrocarbons taken up from food and water by the blue crab,
Callinectes sapidus. Marine Biol. 37:363-370.
Lowry, O.H., N.R. Robert, and J.J. Kapphaha. 1957. The fluorometric
measurement of pyridine nucleotides. J. Biol. Chem. 224:1047-1064.
Nebert, D.W., and H.V. Gelboin. 1968. Substrate inducible microsomal aryl
hydroxylase in mammalian cell culture. J. Biol. Chem.
243:6242-6249.
Neff, J.M., and J.W. Anderson. 1975. Accumulation, release and
distribution of benzo(a)pyrene -C in the clam, Rangia cuneata.
In: Proceedings of 1975 Conference on the Prevention and Control of
Oil Pollution. American Petroleum Institute, Washington, DC. pp.
469-471.
Payne, J.F. 1976. Field evaluation of benzopyrene hydroxylase induction
as a monitor for marine petroleum pollution. Science 191:945-946.
Rubin, B.A. 1964. Carcinogen-induced tolerance to homotransplatantion.
Prog. Exp. Tumor. Res. 5:217-292.
Varanasi, U. and D.C. Malins. 1977. Metabolism of petroleum hydrocarbons:
accumulation and biotransformation in marine organisms. In: Effects
of petroleum on arctic and subarctic marine environments and
organisms. Vol. II. Biological effects. D.C. Malins, Ed., Academic
Press, New York. pp. 175-270.
280
-------
Vishnevetskii, F.E. 1961. Pathomorphology of fishes poisoned with phenol
and water-soluble components of crude oil, coal tar, and fuel oil (an
experimental study). Tr. Astrkh. Gos. Zopov. 5:350-352.
281
-------
ACCUMULATION AND RELEASE OF POLYCYCLIC AROMATIC
HYDROCARBONS FROM WATER, FOOD, AND SEDIMENT BY
MARINE ANIMALS
by
Jerry M. Neff,
Battelle New England Marine Research Laboratory
397 Washington, Street, Duxbury, MA 02332
ABSTRACT
All species of marine organisms studied to date rapidly
accumulated polycyclic aromatic hydrocarbons (PAH) from low
concentrations in the ambient water. Bioaccumulation factors
tend to increase as the molecular weight of the PAH increases.
The patterns of PAH accumulation from dispersed or water-soluble
fractions of oil are complex and variable. PAHs adsorbed to food
are accumulated only to a very limited extent by marine
polychaete worms and fish. However, accumulation of PAH from
food is many times more efficient than accumulation from water
by some marine crustaceans. The bioavailability to benthic
marine animals of sediment-adsorbed PAH is very limited.
Animals collected from PAH-contaminated sediments generally have
lower concentrations of PAH in their tissues than the PAH
concentration in the sediment. When returned to a PAH-free
environment all marine animals studied to date rapidly released
PAH from their tissues to low or undetectable levels. PAH
metabolism is important but not essential for PAH release. Many
endogenous and exogenous factors affect the rates of PAH uptake
and release by marine organisms.
INTRODUCTION
The presence of polycyclic aromatic hydrocarbons (PAHs) in tissues of
a wide variety of marine organisms (Neff, 1978) strongly indicates that
these organisms are able to accumulate PAH present at low concentrations in
the ambient medium, food, or sediments. This is not surprising, since PAHs
are highly hydrophobic and lipophilic. Intrinisic lipid/water partition
coefficients favor their rapid transfer from the aqueous phase into
282
-------
lipophilic compartments such as biological membranes, macromolecules, and
depot lipid stores in organisms (Leo et^ a\_., 1971; Neely et^ aj_., 1974).
Release of PAH from tissues of contaminated organisms may be passive,
reflecting an equilibrium distribution between the aqueous phase and
lipophilic compartments in contract with it, or it may be active and
involve metabolic transformation of PAH to polar water-soluble metabolites
which are more readily excreted. A large amount of research has been
conducted in recent years on accumulation and release of petroleum
hydrocarbons by marine organisms (see reviews of Anderson e^t £I_., 1974a;
Varansi and Mai ins, 1977). Only literature dealing specifically with
accumulation and release of PAH by marine organisms will be reviewed.
Accumulation and Release of PAH from Water
Polychaete worms, Neanthes arenaceodentata, exposed for 24 hr to
seawater containing 0.15 parts per million (ppm) 14C-naphthalene, accumulated
naphthalene to a maximum of 6 pg/g tissue (ppm) in 3 to 24 hr (Rossi, 1977).
On return to isotope-free seawater, 14C-naphthalene was rapidly
released and reduced to nondetectable levels within 300 hr. Approximately
one-third of the radioactivity released by Neanthes during the first 24 hr
of depuration was in the form of unmetabolized naphthalene. The remainder
was in the form of polar metabolites. Significant levels of ^C-naphthalene
metabolites remained in tissue material after 504 hr (21 days) of depuration,
suggesting that some metabolic products were covalently bound to tissue
macromolecules.
Several studies have been performed on accumulation and release of PAH
in solution by marine bivalve molluscs. Lee et,ll- (1972a) demonstrated
accumulation by the mussel, Mytilus edulis. of ^-naphthalene and
3H-benzo(a)pyrene dissolved in seawater. The highest levels of activity
were recorded in the gill; the authors hypothesized that gill tissue of
mussels has a micellar layer which absorbs hydrocarbons and then passes
them to other tissues. Although no evidence of MFO activity could be
detected, mussels rapidly released accumulated PAH when returned to
isotope-free seawater.
Neff et^jLL- (1976a) measured relative rates of accumulation and
release of 4 PAHs from seawater by the estuarine clam, Rangia cuneata,
(Table 1). Phenanthrene was accumulated most rapidly and released most
slowly. The rapid release of naphthalene from clam tissues probably masked
a similarly rapid uptake during exposure, since both influx and efflux of
this compound undoubtedly occurred simultaneously. These results can best
be explained in terms of relative aqueous solubilities and lipid/water
partition coefficients of the four PAHs. Naphthalene is the most
water-soluble of PAHs tested and thus more readily bioavailable. Although
it has a high affinity for lipids, its lipid/water partition coefficient
favors rapid release to water when naphthalene concentrations in the medium
are reduced. Benzo(a)pyrene (BaP), at the other extreme, has a very low
aqueous solubility (1 ppb) so that most BaP in the exposure water was in
colloidal or particulate form, decreasing its bioavailability.
283
-------
Once absorbed, BaP would not readily partition back into the aqueous phase
even when concentrations in the medium were extremely low. In other exper-
iments R^cuneata accumulated up to 7.2 ppm BaP during 24-hr exposure to
0.03 ppm 14C-BaP in seawater (Neff and Anderson, 1975). Nearly 75% of
the activity was located in the viscera (digestive gland, gonad, etc.).
When returned to isotope-free seawater, clams released BaP to undetectable
levels in 58 days. Tissue distribution of BaP remained relatively constant
during the depuration period. Dunn and Stich (1976) measured the rate of
release of BaP from Mytilus edulis naturally contaminated with PAH.
Concentration of BaP in mussel tissues was approximately 45 ug/kg wet
weight at the beginning of the experiment. When placed in clean seawater,
mussels released BaP at an approximately exponential rate over the six-week
depuration period. The overall half-life of BaP in mussel tissues was
approximately 16 days.
TABLE 1. ACCUMULATION FROM SEAWATER AND RELEASE OF PAH BY THE ESTUARINE
CLAM, RANGIA CUNEATA (FROM NEFF et al_., 1976A)
PAH Naphthalene Phenanthrene Chrysene Benzo(a)pyrene
y
Exposure nr.n
Concentration 0.071 0.089 0.066 0.052
(ppm)
Tissue concentration
after 24 hr exposure 0.43±0.1 2.85±1.1 0.54±0.3 0.45±0.1
(ppm) ± S.D.
Bioaccumulation
factor 6.1 32.0 8.2 8.7
[tissue]/[water]
Tissue concentration
after 24 hr depura- 0.15±0.02 2.47±1.2 0.40±0.15 0.38*
tion (ppm) ± S.D.
% Released in 24 hr 6j[ ]! 26_ 1_6
only one sample analyzed.
Lee et _al_. (1978) suspended oysters, Crassostrea virgi'nica. at 7-m
depth in a controlled ecosystem enclosure, which was dosed with Prudoe Bay
crude oil enriched with several PAHs. Some oysters were subsequently
removed to hydrocarbon-free seawater for depuration studies.
284
-------
Naphthalene and alkylnaphthalenes were accumulated most rapidly by the
oysters (Table 2). Rate of accumulation of other PAHs decreased with
increasing molecular weight. When oysters were returned to clean seawater,
naphthalenes were released rapidly and reached undetectable levels in 23
days. Anthracene, fluoranthene, benz(a)anthracene, and BaP were released
much more slowly. Based on the depuration experiments, calculated half-
lives of the naphthalenes, anthracene, fluoranthene, benz(a)anthracene, and
BaP were 2, 3, 5, 9, and 18 days, respectively. Ihus, despite the very
limited ability of molluscs to metabolize PAH, all species so far studied
are able to release the majority of accumulated PAH from their tissues in
periods varying from a few days to several weeks.
Patterns of accumulation and release of PAHs in solution by aquatic
crustaceans are somewhat different, probably reflecting their more active
mode of life and greater PAH-metabolizing abilities. Ihe copepod, Calanus
helgolandicus, was able to accumulate naphthalene during exposure for 24 hr
to seawater solutions containing as little as 0.1 yg naphthalene/A (ppb)
(Corner et jjl_., 1976a). When returned to naphthalene-free seawater,
copepods released naphthalene rapidly (half-life ^1.5 days). Depuration
was slightly more rapid in copepods feeding on algal cells than in starved
individuals. Subsequently, Harris £t al_. (1977b) studied accumulation of
14C-naphthalene by £. helgolandicus and Eurytemora affinis during exposure
to concentrations of 0.2 to 992 yg C-naphthalene/Jl seawater for up to
15 days. Initial uptake was rapid but after exposure for seven to eight
days to seawater containing 50 yg/£ an equilibrium-condition was
approached. At an exposure concentration of 1 yg Onaphthalene/£.,
the quantity of radioactivity accumulated in 10 days was nearly 50 times
greater in the smaller estuarine species, £. affinis, than in the larger
oceanic species, £. helgolandicus, when expressed in terms of body weight.
Ihe difference in uptake rate was not due to differences in lipid content
between the two species. When returned to hydrocarbon-free seawater, £.
hegolandicus released up to 90% of the accumulated radioactivity in 24 hr.
However, 5% of the accumulated radioactivity remained in the copepods after
11 days. Much of the radioactivity in the copepods and depuration water
was identified as naphthalene metabolites.
Lee (1975) obtained similar results when he exposed several species of
marine zooplankton to seawater solutions of radio-labeled BaP, methychol-
anthrene, and naphthalene. The copepod, Calanus plumchrus. accumulated up
to 22 x 10"4 ug BaP/individual during a three-day exposure to 1 yg
BaP/A in seawater (Figure 1). When returned to isotope-free seawater, the
copepod released most of the accumulated radioactivity in 17 days. When
depuration was continued beyond 17 days, no further hydrocarbon loss was
observed. PAH metabolism contributed significantly to release of PAHs by
all crustaceans studied.
Pink shrimp, Penaeus duorarum, exposed to 1 or 5 ug chrysene/*
seawater accumulated the PAH in both the cephalothorax and abdomen
(Mi 11 er et al_*» 1978). At an exposure concentration of 5 ppb, shrimp
accumulated approximately 1.8 yg chrysene/g tissue in the cephalothorax and
0.4 yg chrysene/g tissue in the abdomen in 28 days. When returned to clean
285
-------
ro
cc
• Copepodsexposed to one jUg of H-benz-
pyr«ne (none liter of sea water.
A Copepods exposed to one |ig of H-
benzpyrene In one liter of sea water;
after three days of exposure cope-
pods transferred to radioactive free
sea water.
17
18
19 20
TIME (days)
Figure 1. Accumulation of benzo(a)pyrene by the copepod Calanus plumchrus during exposure to
1 yg BaP/£ seawater and BaP release following return to clean seawater (From Lee, 1975)
-------
TABLE 2. ACCUMULATION AND RELEASE OF PAH BY THE OYSTER, CROSSOSTREA VIRGINICA, EXPOSED TO CRUDE
ro
oo
•-g
OIL TREATED WITH SEVERAL PAHs IN A CONTROLLED ECOSYSTEM ENCLOSURE (FROM LEE ET AL . , 1978)
Duration of
Exposure
(days )
2
8
2
8
8
Duration of
Exposure
(days)
2
8
2
8
8
Duration of
Depuration
(days)
7
7
23
Duration of
Depuartion
(days)
7
7
23
Naohthalene Methylnaphthalenes Dimethyl naphthalenes
Oyster Water Oyster Water Oyster Water
(yg/g) (yg/0 (ug/g) (vs/*) (ns/g) (yg/0
30 5 58 8 84 10.
12 3 36 3 72 2
1 1 8 -
2 2 - 4 -
N.D* - N.D. - N.D.
Anthracene Fluoranthene Benz[a]anthracene Benzo[a]pyrene
Oyster Water Oyster Water Oyster Water Oyster Water
(yg/g) (yg/a) (yg/g) (yg/0 (yg/g) (vg/0 (yg/g) (yg/^)
5.6 13 5.0 7.2 2.8 5.3 0.36 1.9
2.5 1 4.0 0.4 1.8 0.1 0.30 0.1
1.2 - 1.7 - 1.9 - 0.40
0.4 - 1.4 - 1.0 - 0.20
0.1 - 0.4 - 0.3 - 0.12
* N.D.=not detected, less than 0.5 yg/g.
-------
seawater, shrimp released most of the chrysene in 10 days. However, a
small but measurable amount of chrysene remained in their tissues after 28
days of depuration. No attempt was made to measure metabolite formation.
Juvenile blue crabs, Callinectes sapidus. accumulated isotopically
labeled BaP, methylcholanthrene, and fluorene during exposure to seawater
containing 2.5, 1.0, and 30 pg/A of these compounds, respectively
(Lee jit jil_., 1976). Maximum radioactivity in the crabs was reached after
two days, although uptake from the water continued beyond this time. After
a two-day exposure, rapid discharge of PAHs and metabolites from the crabs
balanced uptake from the medium. Initial uptake took place via gill
tissue, from which radioactivity was transferred primarily to the blood and
hepatopancreas. The pattern of accumulation, distribution in the body, and
metabolism and release of H-BaP by crabs is summarized in Table 3. Main
sites of BaP accumulation were the hepatopancreas and gill. During both
the exposure and depuration periods the fraction of total radioactivity
present as unmetabolized BaP decreased with time in all tissues analyzed.
At every sampling period, more than 50% of accumulated radioactivity was in
the hepatopancreas and, in all but the Day 1 samples, more than 50%
hepatopancreatic activity was present as hydrocarbon metabolites. During
depuration, concentrations of BaP and its metabolites in the crab tissues
dropped rapidly, so that after 20 days only small amounts of radioactivity
remained, primarily in the hepatopancreas as polar metabolites. BaP and
metabolites were recovered from depuration water.
Throughout the first four days of depuration, the major excretory
product was unmetabolized BaP. After longer depuration times, major
excretory products were various polar metabolites. More than 50%
accumulated BaP, and its metabolites were excreted in six days. More than
70% radioactivity excreted was in particulate form, suggesting that fecal
material was the main route of PAH excretion. A fraction of the total
radioactivity in crab tissues (varying from 0.3 to 50% of the total)
occurred in a form not extractable with the solvents used. This material
may have consisted of highly polar compounds or BaP metabolites that were
covalently bound to tissue macromolecules. The relative proportion of this
non-extractable radioactivity in the tissues increased during the
depuration period, implying that it was not as readily excreted as BaP or
its major metabolites. Similar conclusions were obtained with fluorene and
methylcholanthrene. All results dramatically demonstrate the importance of
PAH metabolism in elimination of these materials from tissues of PAH-
contaminated crustaceans.
The observations reported in several of the studies discussed above
suggest that some of the products of PAH metabolism may be retained in
animal tissues longer than unmetabolized PAH. A recent study by Sanborn
and Mai ins (1977) supports this view. Stage V spot shrimp larvae, Panda!us
piatyceros,..accumulated high concentrations of radioactivity during exposure
to 8-12 pg C-naphthalene/£ in water. Metabolic products of naphthalene
accounted for up to 21% of the radioactivity in larval tissues.
C-naphthalene was almost completely depurated from the tissues during
24 to 36 hr in isotope-free seawater. However, metabolic products were
strongly resistant to depuration.
288
-------
TABLE 3. FATE OF 3H-BENZO(a)PYRENE ACCUMULATED FROM WATER (2.5 yg/£ BaP) BY
JUVENILE CRABS, CALLINECTES SAPIDUS. AFTER TWO-DAY EXPOSURE, CRABS
WERE TRANSFERRED TO HYDROCARBON-FREE SEAWATER FOR DEPURATION. VALUES
ARE MEANS FOR 3 CRABS SEPARATELY ANALYZED +1 S. D. (LEE et a]_. , 1976)
Product
Total
Benzo(a)pyrene
Hydroxybenzo (a ) pyrene
Polar metabolites
Time
(days)
1
_2
4
8
12
20
1
_2
4
8
12
20
1
_2
4
8
12
20
1
2
4
8
12
20'
Gill
150±70
410±85
100±32
70±55
35±12
1±2
no
290
70
22
10
t
10
15
5
11
7
t
15
3
12
17
11
t
A
Quantity
Blood
90±17
85±20
80±52
65±21
20±13
6±5
50
30
12
8
1
1
27
12
21
9
6
2
10
26
32
36
10
2
in tissue
Hepato-
pancreas
270±90
590±210
300±18
210±16
320±74
40±26
160
280
40
21
32
2
20
84
90
40
24
5
89
210
140
130
224
29
(vg) x 1
Stomach
14±2
11±6
10±4
4±5
6±2
2±1
3
2
1
t*
t
t
2
2
2
t
1
t
8
6
7
2
2
1
04
Muscle
10±3
18±7
8±5
4±3
2±2
2±3
4
2
2
1
t
t
2
2
1
t
t
t
2
8
4
2
1
1
289
-------
Most fish are able to metabolize and excrete PAH accumulated from the
medium even more rapidly than crustaceans. Anderson et _al_. (1974b)
reported that accumulation and release of naphthalene and 1-methylnaphtha-
lene were very rapid in the estuarine sheepshead minnow, Cyprinodon
variegatus. When exposed for 4 hr to 1 ppm of each compound in seawater
the-fish accumulated 60 ppm naphthalene and 210 ppm 1-methylnaphthalene.
Nearly 90% of the accumulated hydrocarbons were released after 29 hr in
hydrocarbon-free seawater.
Lee et_ _a]_. (1972b) studied accumulation and release of 14C-naphthalene
and H-BaP by three species of marine fish. All three species rapidly
accumulated the PAH from diluted solution in seawater. Equilibrium levels
in the tissues were reached in about 1 hr. The main route of uptake
appeared to be through the gills. Both naphthalene and BaP tended to
accumulate primarily in the liver and gallbladder; the latter organ
contained the majority of the activity after several hours depuration.
Following 24 hr in clean seawater, more than 90% of radioactivity
accumulated as C-naphthalene was lost from most of the fish tissues.
BaP was released more slowly with losses of 50, 50, 90, and 20% radioact-
ivity in liver, gut, gill, and flesh, respectively, after 24 hr in clean
seawater. A significant portion of the radioactivity excreted following
exposure of fish to both C-naphthalene and H-BaP was in the form
of polar metabolites. The main routes of excretion of PAHs and their
metabolites were the gallbladder and urine.
Statham et^ jiK (1976) showed that the gallbladder was a principal site
of accumulation of several hydrophobic xenobiotics in trout, Sal mo
gairdneri. However, the highest concentrations of radioactivity detected
jo mummichogs, Fundulus heteroclitus. following exposure for 4 hr to
C-naphthalene was in the spleen (34 to 105 times the exposure
concentration)(DiMichele and Taylor, 1978). Liver, brain, anterior kidney,
and gallbladder also contained high concentrations of radioactivity. No
attempt was made to identify metabolites. On the other hand, mangrove
snapper, Lutjanus griseus. exposed to 1 or 5 yg chrysene/Jt seawater,
accumulated PAH only in the liver and not in other tissues examined
(gallbladder, white muscle, intestine)(Miller_et a^., 1978). These
disparate results suggest either that there are interspecific differences
in distribution in fish tissues of PAH accumulated from water or that
different PAHs have different distribution patterns in the tissues.
The role of the MFO system in excretion of PAH by rainbow trout, Salmo
gairdneri, was investigated by Statham et_ al_. (1978). Trout that were
pretreated (induced) with benzanthracene and exposed to
1 C-2-methylnaphthalene in solution, exhibited significantly elevated
rates of bilary excretion of accumulated C-2-methylnaphthalene and
metabolites. The bile of induced fish contained a higher proportion of
polar metabolites of 2-methylnaphthalene than that of controls. Initial
levels of C were higher in livers of induced trout than in uninduced
controls, and radioactivity appeared to be retained longer in livers of
induced fish during the depuration period. Pretreatment with benzanthra-
cene had little effect on the rate of disappearance of radioactivity
290
-------
from blood and muscle. The greater retention of 14C in livers of
induced fish may be due to the metabolic production of electrophylic
metabolites which became bound to tissue macromolecules.
4 Sharp et.il. (1978) showed that rates of uptake and release of
C-naphthalene changed during the timecourse of embryonic development
of the mummichog, Fundulus heteroclitus. The uptake rate of naphthalene
decreased markedly from 1265 DPM 14C/embryo/hr on day two of
development to 195 DPM i4C/embryo/hr on Day 10 of development. Water
influx rates varied only slightly during the same time period. Rate of
C-naphthalene efflux from the embryos, when returned to isotope-free
seawater, varied from 8.65 to 9.95%/hr between Day 2 and Day 6 of
development and then dropped to 5.85%/hr on Day 12 of development. Embryos
were unable to metabolize either naphthalene of chrysene to polar
metabolites at any stage of embryonic development, possibly accounting for
the relatively slow depuration rate.
Accumulation of PAH from Food
Considerably less research has been done on the ability of aquatic
animals to accumulate PAH from food sources. Such information is, of
course, necessary to assess the potential for biomagnification of PAH in
aquatic food chains.
Rossi (1977) fed young adult polychaete worms, Neanthes arenaceodentata,
powdered alfalfa (tbeir normal diet in laboratory culture) contaminated
with 10 to 15 ppm C-2-methylnaphthalene for 16 days in succession.
Radioactivity was detected in the worms after feeding for 192 and 384 hr on
contaminated food. If worms were given uncontaminated fish food for 24 hr
to allow purging of labeled unassimilated food, no radioactivity was
detected in tissue material at any sampling time. In addition more than
85% of radioactivity recovered from the water of the exposure chambers
occurred as the unmetabolized parent compound. Thus, N^. arenaceodentata
had little if any ability to accumulate 2-methylnaphthalene from its food.
The situation is quite different in marine crustaceans and fish.
Corner et jal_, (1976b) studied dietary uptake of C-naphthalene by the
marine copepod, Calanus helgolandicus. When Calanus were fed
C-naphthalene-contaminated living.or dead copepod nauplii, El mini us
sp., or living algae, Biddulphia, r4C-naphthalene was accumulated.
Uptake from food was much more efficient than uptake from solution.
Approximately 35% of radioactivity taken from the food source was lost
during 24-hr depuration, after which 94% of that remaining in the copepods
could still be identified as unmetabolized hydrocarbon. However, less than
one-third of the radioactivity in depuration water existed as unmetabolized
naphthalene. Similar results were obtained by Harris et^ al_. (1977a), who
compared ^C-naphthalene uptake from solution versus food supply
(Biddulphia sinensis cells). Compared to the quantity of C-naphthalene
present as suspended food, the amount in solution alone required to give
the same increase in hydrocarbon level in copepods was several orders of
2yi
-------
magnitude greater in experiments with female Calanus and two orders of
magnitude greater with males (Table 4). The discrepancy between male and
female responses was attributed to the higher feeding rate of females in
comparison to the smaller males. Approximately 60% of naphthalene ingested
with the food was assimilated by both male and female copepods (Table 5).
Of the assimilated naphthalene, almost half was retained in copepod
tissues; the other half was released as either unmetabolized naphthalene or
its metabolites.
TABLE 4. QUANTITATIVE IMPORTANCE OF THE DIETARY PATHWAY IN THE ACCUMULATION
OF 14C-NAPHTHALENE BY THE MARINE COPEPOD, CALANUS HELGOLANDICUS
(FROM HARRIS et al_. , 1977A)
Hydrocarbon Radioactivity as
concentration pg hydrocarbon/animal
In soln As food Soln Soln &
(ugA) (ng/«.) alone Food
A B_
Hydrocarbon
in soln. alone
equivalent to
(soln & food)
level (yg/2,)
C
Ratio
(C-A:B)
0.96
4.69
25.52
26.00
134.10
0.93
Experiments with females
0.66
2.38
122.40
29.00
61.40
1.10
53
179
759
749
2997
106
386
4538
1830
6774
Experiments with males
75 88
2.40
11.55
230.80
76.54
375.40
1.24
2186
2892
1677
1743
3930
282
TABLE 5. FATE OF 14C-NAPHTHALENE ACCUMULATED BY CALANUS HELGOLANCICUS
FROM INGESTION OF CONTAMINATED BIDDULPHIA CELLS
(FROM HARRIS et al., 1977A)
% Ration
Retained
Feces Assimilated Retained Soluble Assimilated
release %
Soluble release
Assimilated
Femal e
Male
41.9
39.3
58.1
60.7
31.2
26.8
26.9
33.9
53.7
44.1
46.3
55.9
2y2
-------
Lee et al. (1976) showed that juvenile blue crabs, Callinectes sapidus,
were~~abTe to accumulate naphthalene, methyl naphthalene, fluorene, and BaP
from food. When crabs were fed shrimp or oysters containing radio!abeled
PAH, between 7 and 10% of the radioactivity was transferred from the
stomach to other body tissues. Most of the remainder was unabsorbed and
lost in the feces during the first two days after feeding. Radioactivity
appeared in the hepatopancreas after 6 hr and in the blood after 18 hr.
Gills, muscles, and gonads also contained significant levels of radio-
activity. The pattern of PAH release form the tissues following ingestion
of contaminated food was qualitatively similar to that described earlier
for depuration of PAH accumulated from water. A significant fraction of
radioactivity was released as polar metabolites.
Dixit and Anderson (1977) administered 14C-naphthalene to Gulf
killifish, Fundulus similis. by stomach tube. After 2 hr, 34% of the
administered radioactivity was recovered in tissue material, 34% of this
was in stomach tissue. Thus, approximately 12% of the administered
^C-naphthalene was assimilated by Fundulus. The majority of this act-
ivity was localized in the liver and gallbladder after 2 hr. Significant
amount of activity were also present in the heart and lateral body
musculature. After 8 hr, 79% of the radioactivity recovered was present in
the gallbladder. These results strongly suggest that PAHs absorbed from
the gut are transported to the liver where they are rapidly metabolized and
excreted in bile.
Roughly similar results were obtained when 14C-BaP contaminated
squid were fed to young cod, Gadus morrhua, (Corner et^ al_., 1976b) or
juvenile herring, Clupea harengus (Whittle et _§]_., 1977); 48 hr after
feeding, 83.5% of the radioactivity was stiTT present in the stomach and
12% in intestinal contents of the cod (Table 6). The liver, bile fluid,
and gills were the only other tissues containing significant activity.
After 72 hr the bile fluid contained 12.5% of the total activity and after
96 hr the intestinal contents and feces contained 57% of the total
recovered radioactivity; most of the remainder (37.1%)1was associated with
the stomach. In herring, nearly 80% of the recovered ^C remained in
the lipid fraction (unmetabolized BaP) of the stomach 43 hr after ingestion
of l4C-BaP contaminated squid. The largest fraction of the remaining
activity was found in the lipid fraction of the intestine (10.3%) and the
residual fraction (unextractable BaP metabolites) of the stomach, pyloric
caecae, and intestine (4.9%). Most of the activity recoverd in bile was
water-soluble, indicating polar metabolites. Although the digestive tract
of fish represents the major site of both uptake and excretion of orally
administered PAH, the fact that more than 98% of radioactivity recovered
from the fish 43 hr after feeding was in the digestive tract strongly
suggests that there was very little assimilation of ingested BaP, and that
the small amount of BaP that was assimilated was rapidly metabolized and
excreted via the gallbladder into the intestine. The authors concluded
that retention of BaP in the stomach implies strong adsorption or binding
to the stomach wall. This binding prevents subsequent absorption an
assimilation of ingested BaP.
293
-------
TABLE 6. DISTRIBUTION OF 14C ACTIVITY (AS % OF TOTAL ACTIVITY RECOVERED
AT INTERVALS AFTER FEEDING SQUID CONTAINING 14C-BaP AND
14C-HEXADECANE TO YOUNG CODFISH,GADUS MORRHUA (FROM CORNER et j»l_
1976B)
% C activity recovered after
Sample
48 hr 72 hr 96 hr
Stomach
Liver
Bile fluid
Intestinal contents
Urine
83.5
2.2
0.6
12.9
0
32.9
3.6
12.5
8.6*
trace
37.1
3.4
1.9
36.3
0.03
Aquarium residue
(mainly feces)
Aquarium water
Plasma**
Blood**
Gills
Spleen
0.5
0
0.03
0
0.2
0.06
41.4
0.2
0.08
0.06
0.5
0.1
20.7
0.12
0.05
0.04
0.4
0.1
*A loss of at least 50% occurred during dissection.
"Expressed on a per gram basis.
Twenty-four hours after fingerling coho-salmon, Oncorhynchus kisutch,
were fed food containing C-naphthalene or C-anthracene, 0.03 and
0.17%, respectively, of the administered radioactivity was associated with
the liver, brain, and flesh of the fish (Roubal et al.., 1977b). The
specific activity in the liver and brain was higher than in flesh, but the
flesh contained more total activity. Anthracene and its metabolites were
retained longer in fish tissues than were naphthalene and its metabolites.
From the limited data available, it would appear that there are large
interspecific differences in ability to absorb and assimilate PAH from
food. Polychaete worms have a very limited ability to absorb and
assimilate PAH, whereas fish adsorption of PAH from the gut is limited and
variable depending on species of fish, the PAH, and possibly the food
matrix in which PAH is administered. Crustaceans, on the other hand,
apparently readily assimilate PAH from contaminated food. In all cases
where assimilation of ingested PAH was demonstrated, metabolism and
excretion of PAH were rapid. Thus, the potential for food chain
biomagnification of PAH seems to be limited. For such biomagnification to
294
-------
occur, the material must be readily absorbed from food, and once
assimilated, it must be relatively resistant to metabolism or excretion.
Accumulation of PAH from Petroleum
When petroleum is spilled in water, PAHs in the oil may enter the
water column as PAH in solution, in dispersed form (micro- or macro-oil
droplets), or adsorbed to organic or inorganic particulates. The physical
form in which oil-derived PAHs occur in the water column may significantly
affect the rate at which they are accumulated by aquatic organisms.
Several investigators have studied accumulation of petroleum in different
aqueous forms.
As indicated above, when aquatic animals are exposed to a single PAH
or a simple mixture of the PAH in solution, there is usually a good corre-
lation between octanol/water partition coefficient of each PAH and its
bioaccumulation factor (concentration in tissue/concentration in water)
(Neeley et a\_., 1975; Neff et a!., 1976a; Lu et .al_., 1977). Therefore,
biomagniTTcation factors increase as molecular weight of the PAH increases.
Based on these considerations, one might predict that the aromatic fraction
in tissues of oil-contaminated marine animals would be enriched in higher
molecular weight PAH as compared to the aromatic fraction of the contam-
inating oil. This is not the case when short-term exposure to dispersed
oil occurs. Oysters, Crassostre^a virginica, subjected for 8 hr to a
concentrated oil-in-water dispersion of No. 2 fuel oil (302 ppm total
hydrocarbons at 8 hr), accumulated a wide spectrum of hydrocarbons
(Table 7)(Neff et al_., 1976b). By the end of the exposure period, oysters
had accumulated^" total of 311 ppm total hydrocarbons, including 76.7 ppm
PAH. The relative concentrations of different PAHs in the oyster tissues
were similar to those in the oil; there was no evidence of selective
accumulation of higher or lower molecular weight PAH. When returned to
oil-free seawater, the oysters released n-paraffins rapidly; PAHs were
released more slowly. All PAHs showed roughly similar behavior. Concent-
rations of different PAHs remained essentially constant during the first
120-hr post-exposure. However, after 672 hr (28 days), the concentration
of PAH in oyster tissues reached background levels. Clams, Rangia cuneata,
exposed to the same oil dispersion accumulated only 89.6 ppm total
hydrocarbons in 8 hr (28.5% of the amount accumulated by oysters in the
same time period). The pattern of petroleum hydrocarbon accumulation and
release by the clams was qualitatively similar to that of the oysters.
•
Bieri and Stamoudis (1977) obtained similar results when they exposed
oysters, Crassostrea virginica, and hard shell clams, Mercenaria mercenaria
to an experimental spill of No. 2 fuel oil. The oysters rapidly accumu-
lated aliphatic and aromatic hydrocarbons after the spill. Between 6 and
25 hr after the spill, most of the n-alkanes had been released from oyster
tissues. Branched alkanes and olefins were released more slowly. Patterns
of PAH accumulation and release were more complex (Table 8). The
concentrations in oyster tissues of alkyl naphthalens with up to five alkyl
carbons, biphenyls with up to two alkyl carbons, and fluorenes with up to
one methyl group increased until 25 hr after the spill. Maximum tissue
295
-------
TABLE 7. ACCUMULATION OF PETROLEUM HYDROCARBONS BY OYSTERS CRASSOSTREA VIRGINICA DURING"
EXPOSURE TO DISPERSED NO. 2 FUEL OIL IN A FLOW THROUGH SYSTEM AND SUBSEQUENT RELEASE
OF HYDROCARBONS WHEN OYSTERS WERE RETURNED TO OIL-FREE SEAWATER. THE COMPOSITION OF
THE NO. 2 FUEL OIL IS INCLUDED FOR COMPARISON (FROM ANDERSON et al. ,19748; NEFF
et al., 1976B)
ro
Time (hr)
Exposure
0
8
Depuration
3
6
24
120
672
No. 2 fuel oil
(% composition)
Petroleum
n-P
,N
_ ** 0.2
235
156
68
18
10
-
7.38
14.7
12.0
7.3
6.5
4.7
-
0.4
1-MM
0.1
8.7
8.4
5.1
5.7
4.7
-
0.82
2-MN
0.3
15.0
12.0
7.3
7.6
6.8
0.1
1.89
DMN
1.0
21. S
22.7
13.2
14.8
13.4
0.5
3.11
hydrocarbon concentration (ppm, yq/g wet wt)*
TMN
0.8
9.1
10.8
5.7
9.5
4.9
0.9
1.84
B MB
_
0.3 0.5
0.3 0.4
0.1 0.2
0.2 0.2
0.1 0.1
-
0.16
F MF
-
1.0 1.2
0.7 0.7
0.4 0.2
0.5 0.7
0.2 0.1
-
0.36
P MP
-
1.9 1.9
1.3 1.3
0.6 0.6
1.2 1.3
0.4 0.4
-
v s—
0.53
DMP DBT
-
0.3 0.3
0.2 0.3
0.1 0.1
0.3 0.2
0.2 0.1
-
*
0.07
Total
2.4
311.7
227.1
108.9
66.7
46.1
1.5
16.56
*n-P, C12-C2q n-paraffins; N, naphthalene; 1-MN, 1-methylnaphthalene; 2-MN, 2-methylnaphthalene; DMN, dimethyl-
naphthalenes; TMN, trimethylnaphthalenes; B, biphenyl; MB, methylbiphenyls; F, fluorene; MF, methylfluorenes;
P, phenanthrene; MP, methylphenanthrenes; DMP, dimethylphenanthrenes; DBT, dibenzothiophene.
**
less than 0.1 ppm
-------
concentrations for more highly substituted naphthalenes, biphenyls, and
fluorenes as well as for dibenzothiophenes and phenanthrenes were attained
100 hr post-spill. Many of the PAHs which reached peak concentrations at
25 hr were released to low or undetectable levels after 100 hr. Higher
molecular weight PAHs were released to undetectable levels in 242 to 509
hr.
TABLE 8. CONCENTRATIONS OF PAH IN THE TISSUES OF OYSTERS, CRASSOSTREA
VIRGINICA, EXPOSED TO A SPILL OF 85 i OF NO. 2 FUEL OIL IN A 24
x 24 m ENCLOSED SHORELINE AREA NEAR YORKTOWN, VA.
CONCENTRATIONS IN THE OYSTER IN PPM (yg/g wet vrt) AND IN
SEAWATER IN PPB (ug/i) (FROM BIERI AND STAMOUDIS, 1977)
Time after spill
CoinDound +6 hr +25 hr +100 hr +242 hr +509 hr
2-Methyl naphthalene
1 -Me thy! naphthalene
Biphenyl & 2,6-dimethyl-
naphthalene
1,3-Diniethylnaphthalene
1,5- & 2,3-Dimethyl-
naphthalene
water
11
9
8
11
3
oysters
0.13
0.11
0.20
0.25
0.07
water
0.2
0.2
0.2
0.3
_
oysters water oysters
0.15 -*
0.12
0.28
0.35
0.13
oysters oysters
_
-
_ _
-
_ _
3- & 4-Methylbiphenyl &
C,-naphthalene 2 0.10 - 0.16 - -
Cj-naphthalenes 7 0.32 - 0.48 - 0.25 0.03
2,3,5-Trimethylnaphthalene &
C.-naphthalene 2 0.13 - 0.23 - 0.07
C~-Biphenyl, C,-naphthalene &
* C4-naphthalene 1 0.07 - 0.19
Fluorene, C.-naphthalenes,
methylacenaphthenes &
C2-biphenyl 2 0.09 - 0.18
C^-Naphthalene & C2-biphenyl 0.4 0.04 - 0.09 - 0.13 0.05
C.-Naphthalenes 4
4 C5-naphtha1ene 0.5 0.05 - 0.13 - 0.37 0.13
Cr-Naphthalene, C--biphenyl
D & methylfluorehes 2 0.16 - 0.40 - 0.47
Cc-Naphthalene, C,-bipheny1
3 & C3-biphenyl ^ - 0.08 - 0.14 - 0.24 0.10
Cg-Naohthalene, C?-biphenyl,
C,-bi phenyl, methylf1uorene,
&JC4-biphenyl - .0.05 - 0.11 - 0.20 0.10
Di benzothi ophene, C4-bi phenyl,
C2-f1uorene, & phenanthrene 2 0.20 - 0.71 - 1.41 0.56
297
-------
TABLE 8 (CONTINUED)
• Compound
Methyldibenzothiophene
Methyldibenzothiophene,
C3-fU:orene, & 2-methyl-
phenanthrene
1-Methylphenanthrene,
C2-dibenzothiophene 5
+6 hr
Time after spill
+25 hr +10C hr
+242 hr +509 hr
Cp-Di benzothi ophene
3,6-Dimethy!phenanthrene
& C~-phenanthrene
Cp-phenarthrenes &
C3-dibenzothiophene
C-j-Phenanthrene
Total unresolved
envelope
water oysters .'later oysters water oysters oysters oysters
0.3
1 0.07
0.05
0.1 0.02
0.2 0.03
0.4 0.10
0.03
0.18
0.33
0.17
0.06
0.09
0.27
0.04
0.32
0.32
0.34
0.18
0.35
0.77
0.21
24
0.17
0.27
0.17
0.13
0.2:0
0.48
0.15
12
*, Compound either absent, below background or cannot be identified.
1, Semiquantitative estimate from unresolved area.
All PAHs reached maximum concentrations in the water in 6 hr, then
decreased to low or undetectable levels after 25 hr. Bieri and Stamoudis
suggested that continued increase of PAH concentrations in the tissues of
oysters long after water-accomodated PAHs had disappeared indicated that
oysters acquired PAHs from oil-contaminated organic detritus via the
digestive tract. The authors went on to hypothesize that apparent PAH
bioaccumulation factors and changes in PAH concentrations in the tissues of
animals exposed to an oil spill in their natural environment are determined
mainly by residence time of different PAHs in the environment. Hard shell
clams accumulated only about one-tenth as much PAH as the oysters,
indicating a substantial interspecific difference in ability of molluscs to
accumulate petroleum hydrocarbons.
298
-------
Bieri et _al_. (1977) used a similar experimental design to investigate
accumulation of petroleum hydrocarbons from spills of new and weathered
south Louisiana crude oil by the mummichog, Fundulus heteroclitus. The
concentration of all n-alkanes in the fish reached a maximum 76 hr after a
spill of fresh crude. Naphthalene, monomethylnaphthalenes, and 2,3-dimethyl'
naphthalene reached maximum concentrations in fish tissues 31 hr after the
spill, while all other PAHs analyzed reached maximum concentrations after
76 hr (Table 9). All PAHs except naphthalene, 2-methylnaphthalene, and the
dibenzothiophene-C4-biphenyl-C5-naphthalene group were detectable in
fish tissues 216 hr post-spill. Nearly all PAHs reached a maximum concent-
ration in the exposure water 31 hr after the spill. Naphthalene, methyl-
naphthalenes, and some dimethyl naphthalenes disappeared rapidly from the
water between 31- and 76-hr samplings. Mean apparent bioaccumulation
factors (concentration in tissues/concentration in water) were about 1000
at 31 hr for PAHs and 290 for aliphatics. At 76 hr, bioaccumulation
factors were 2700 for PAHs and 220 for aliphatics. Bioaccumulation factors
for alkyl naphthalenes were similar although higher than that for
naphthalene, but there was no relation between degree of alkylation and
bioaccumulation factor.
PAHs were accumulated by the fish much more rapidly from the weathered
crude oil and all experimental fish died within 120 hr. The difference
was partially explained by the observation that PAH were accomodated into
the water column much more rapidly from weathered than fresh crude oil.
Although maximum concentrations of individual PAHs in the water
following spills of crude oil (fish experiment) were much lower than
maximum PAH concentrations in the water following the spill of No. 2 fuel
oil (oyster experiment), the fish accumulated substantially higher
concentrations of PAH than oysters. Mummichogs also tended to retain PAH
in their tissues longer than oysters. Thus, ability to metabolize PAH
(high in fish, low in oysters) may have little effect on the rate of PAH
uptake and release in an acute oil-spill situation. The greater accumu-
lation of PAH by mummichogs is probably attributable primarily to the fact
that they could not escape from the contaminants, while oysters could
remain isolated through valve closure from the contaminated medium for long
periods.
In the experiments described above, animals were exposed to both
dispersed oil droplets and water-soluble fractions of oil. It is likely
that some PAHs measured in the animals were actually present as
unassimilated micro-oil droplets in the gut or adsorbed to the gills or
other body surfaces. This may partly account for the lack of differential
uptake of various PAHs from the oil as opposed to exposure to water-soluble
fractions (WSF) of oil alone where this form of uptake is sometimes
observed. Rossi and Anderson (1977) reported that mature male polychaete
worms, Neanthes arenaceodentata. accumulated higher concentrations of
methyl naphthalenes than naphthalene and dimethyl naphthalenes during 8-hr
exposure to a WSF of No. 2 fuel oil, although the WSF contained a higher
concentration of naphthalene than of methyl naphthalenes. Gravid female
worms accumulated more dimethyl naphthalenes than methyl naphthalenes and
299
-------
naphthalene. All naphthalenes were released at the same rate when the
worms were returned to clean seawater. Neff et jil_. (1976b) reported 24-hr
bioaccumulation factors of 2.3 naphthalene, 8.1 to 8.5 methyl naphthalenes,
and 17.1 and 26.7 dimethyl- and trilmethylnaphthalenes for clams exposed to
a WSF of No. 2 fuel oil. Cox and colleagues spilled No. 2 fuel oil on the
surface of a shrimp mariculture pond and measured concentrations of
naphthalenes in the water and tissues of several species of marine animals
from the pond. Naphthalenes were greatly enriched in the tissues of
crustaceans in comparison to aqueous concentrations (Table 10).
TABLE 9. CONCENTRATIONS OF PAH IN THE TISSUES OF MUIMMICHOGS, FUNDULUS
HETEROCLITUS, EXPOSED TO A SPILL OF 570 A OF SOUTH LOUISIANA
CRUDE OIL IN A 810 n/ ENCLOSED SHORELINE AREA NEAR YORKTOWN, VA.
CONCENTRATIONS ARE IN PPM (ug/g wet wt.)(FROM BIERI et al_., 1977)
Time after spill
uuuipuuiiui,i ;
Naphthalene
2-Pethylnaphthalene
1 -Methyl naphthal ene
Biphenyl & 2,6-dimetnylnaphthalene
1 ,3- Dimethyl naphthal ene
1 ,5-Dinethylnaphthalene
2, 3-Oi methyl naphthalene
3- & 4-MethylbiphenyI & C., naphthalene
C.,-Naphthalene
Methyl biphenyl & C,-Naphthalene
2, 3. 5-Trimethyl naphthal ene
C,-flaphthalene & C^-biphenyl
Fluorene, 04- « Cg-naphthalenes
6 C2-tiphenyl
Cjj-fiaphthalene, C2- 1 C3-biphenyls
Methyl fluorene, C.-naphthalene
& C3-biphenyl
Oibenzothiophene, C^-biphenyl
6 C5-raphthalene
Phenanthrene
+6 hr
0.25
0.34
0.34
0.25
0.36
0.13
0.04
0.10
0.09
0.07
0.07
0.03
0.04
0.03
0.11
0.01
0.08
+31 hr
0.68
1.56
1.26
0.81
1.18
0.37
0.05
0.27
0.20
0.17
0.13
0.09
0.11
0.04
0.23
0.02
0.11
+76 hr
0.14
1.12
0.98
0.99
1.54
0.47
0.03
0.56
0.42
0.44
0.25
0.22
0.45
0.11
0.54
-
0.26
+216 hr
_*
-
0.13
0.25
0.40
0.16
0.05
0.14
0.12
0.09
0.08
0.03
0.05
0.03
0.13
-
0.08
*Peak rot detectable on the gas chromatogram.
300
-------
Alkyl naphthalenes were more enriched than naphthalenes. There were
quantitative differences in the amounts of naphthalenes accumulated by the
three species, but the qualitative pattern of naphthalene and alkyl
naphthalene uptake was similar. Coho salmon, Oncorhynchus kisutch, exposed
to a dilute WSF of Prudoe Bay crude oil for five weeks, accumulated the
more highly alkylated benzenes and naphthalenes faster than the less
substituted aromatics (Table ll)(Roubal et ail_., 1977a). €4- and 65-
benzenes and 2-methylnaphthalene had the highest bioaccumulation factors.
Therefore, there is differential bioaccumulation of PAHs from water-soluble
fractions of oil but not from dispersed oil.
TABLE 10. CONCENTRATIONS OF NAPHTHALENES IN THE WATER AND SELECTED
CRUSTACEANS FROM AN EXPERIMENTAL SHRIMP POND TO WHICH NO. 2
FUEL OIL WAS ADDED. THE WATER SAMPLES WERE COLLECTED 24 HR
AFTER THE EXPERIMENTAL SPILL AND THE TISSUE SAMPLES WERE
COLLECTED 72 HR AFTER THE SPILL (FROM COX et al_., 1975)
Mean Concentration (ppb)
Sample
Naphthalene I'ethylnaphthalenes Dimethylnaphthalenes
Water
Brown shrimp
(Penaeus aztecus)
Fiddler crab
(Uca minax)
Warf crab
(Sesarma cinereum)
2.3
450
750
930
15.4
3,610
4,520
6,050
15.2
14,700
16,800
24,700
TABLE 11. CONCENTRATIONS OF AROMATIC HYDROCARBONS IN THE WSF OF PRUDOE
BAY CRUDE OIL IN THE MUSCLE TISSUE OF YOUNG COHO SALMON,
ONCORHYNCHUS KISUTCH. THE RESULTING BIOACCUMULATION FACTORS
ARE ALSO INCLUDED. THE FISH WERE EXPOSED TO THE WSF AT 10° C
FOR FIVE WEEKS IN A FLOW-THROUGH SYSTEM (FROM ROUBAL et al_.,
1977A)
Hydrocarbon
C2-Benzenes
C, -Benzenes
C.- & C5-Benzenes
Naphthalene
1 -Methyl naphthal ene
2-Methylnaphthalene
C2-Naphthalenes
C0-Naphthalenes
Concentration
Water Muscle
352
40 1,
12 5,
4
4
4
10
6
(ppb)
Tissue
490
500
500
240
400
560
850
680
Bioaccumulation
factor
1.39
37.5
458.3
60
TOO
140
85.0
113.3
301
-------
Distribution of PAH in Tissues of Aquatic Animals
Patterns of accumulation and release of PAH by different body regions
of oil-exposed aquatic animals also vary. When shrimp, Penaeus aztecus,
were exposed to a dilute WSF of No. 2 fuel oil for 20 hr, maximum concent-
trations of naphthalenes were reached in the head region, abdomen, gill,
and exoskeleton within the first hour of exposure (Figure 2)(Neff et al.,
1976b). The digestive gland continued to accumulate naphthalenes for the
full 20-hr exposure period and contained more than ten times the
concentration of naphthalenes than other tissues analyzed at the end of the
exposure period. When shrimp were returned to clean seawater, the
abdominal muscle and exoskeleton released naphthalenes very rapidly to
undetectable levels after a 25-hr depuration. The head region released
naphthalenes more slowly. Nearly 250 hr were required for complete
depuration of naphthalenes from the gill and hepatopancreas, suggesting
that in shrimp the hepatopancreas is an important site of PAH storage and
metabolism and the gills are an important route of PAH excretion.
1000,-
2
Ul
O
§
u
0 I
• HEAD REGION
0 ABDOMEN
• GILL
D EXOSKELETON
A DIGESTIVE GLAND
A WATER
3 5
EXPOSURE
10 20 30
DEPURATION
SO
100
TIME (HOURS)
300
1975
Figure 2.
Accumulation and retention of naphthalenes (naphthalene,
methyl naphthalenes, and dimethyl naphthalenes) by different body
regions of juvenile brown shrimp, Penaeus aztecus, exposed to a
20% dilution of the water soluble fraction of No. 2 fuel oil
(From Neff et al., 1976b).
302
-------
3000
2000
1000
800
600
400
300
5 goo
0.
100
5 eo
<
K
CO
60
40
30
20
2 3 4 6 8 I
EXPOSURE
20 I
3 4 6 8 10 20 30
DEPURATION
60 100 200 366
TIME (HOURS)
Figure 3. Distribution of total naphthalenes in the tissues of Gulf kill-
fish, Fundulus similis, during exposure to the water soluble-
fraction of No. 2 fuel oil and at different times following
exposure (From Neff jrt aK, 1976b).
Naphthalenes also accumulated very rapidly in Gulf killifish, Fundulus
similis. during 2-hr exposure to the WSF of No. 2 fuel oil (Figure 3)(Neff
e£ jiK, 1976b). Maximum naphthalene concentrations were reached in most
tissues after 1-hr exposure. The gallbladder and brain contained the
highest concentrations of total naphthalenes (2300 and 620 ppm, respect-
ively, after 1 hr). When fish were returned to clean seawater, all organ
systems inmediatly began to release naphthalenes. The somatic muscles
released naphthalenes most rapidly, whereas the gallbladder and brain
released naphthalenes much more slowly. Complete depuration of
naphthalenes from all tissues required 366 hr. Slightly different results
were obtained when pink salmon fry, Oncorhynchus gorbuscha, were exposed to
a WSF of Cook Inlet crude oil for four days (Rice et aj_., 1977). Maximum
concentrations of naphthalenes were reached in gut, gill, and skeletal
muscle (the only tissues analyzed) after 10-hr exposure; the highest
concentration of naphthalene was found in the gut at all sampling times,
the next highest in skeletal muscle tissue. Methyl naphthalenes were
released more rapidly than dimethyl naphthalenes from the gut during the
latter part of the exposure period and after fish were returned to clean
seawater.
Mallard ducks, Anas piatyrhynchos. fed 5 mi of south Louisiana crude
oil per day for 14 days demonstrated high concentrations of aromatic hydro-
carabons in their tissues (Lawler et j*l_., 1978). The skin and underlying
adipose tissue accumulated greater concentrations of aromatic hydrocarbons
than did other tissues examined including liver, breast muscle, heart
303
-------
muscle, brain, uropygial gland, and blood. Two- and three-ring aromatics
were accumulated to a greater extent than were benzenes. Ringed seals
Phoca hispida. rapidly accumulated petroleum hydrocarbons, including
aromatics, when exposed to Normal Wells crude oil by immersion or ingestion
(Engelhardt et a]_., 1977). Relatively low levels of hydrocarbons were
detected in such tissues as blubber, brain, liver, kidney, muscle, and
lung. Somewhat higher levels were found in whole blood. Hydrocarbon con-
centrations in the bile and urine were high, indicating that these were
major routes of hydrocarbon excretion by the seals.
Effect of Dissolved Organic Matter on Uptake of PAHs
Marine waters nearly always contain significant amounts of dissolved
and colloidal organic matter and suspended inorganic and organic
particulate matter. These materials may interact with petroleum or PAH
entering the water column in such a way as to change bioavailability of
these^ompounds to marine organisms. Sanborn and Mai ins (1977) reported
that 14C-naphthalene was accumulated nearly four times more rapidly
than C-naphthalene complexed to bovine serum albumen by stage V spot
shrimp, Panda!us platyceros. Dunn and Young (1976) inferred from their
data on BaP contamination in mussels, Mytilus edulis. from California that
BaP associated with particulate matter of pyrolytic origin was not readily
available to marine molluscs. Studies on bioavailability of PAH from food
and sediments also seem to indicate that, in most cases, PAH complexed to
colloidal organic materials or adsorbed to organic or inorganic
particulates are less bioavailable than PAH in solution or fine dispersion
in water. Several types of dissolved organic compounds may actually
increase the solubility or accomodation of PAHs and other petroleum
hydrocarbons in water. Humic substances are complex organic macromolecules
of plant origin commonly found in fresh and coastal marine waters in
soluble or colloidal form. Because of their detergency, they tend to
increase accomodation of poorly soluble organic materials in water. Boehm
and Quinn (1976) studied the effect of natural dissolved organic matter
(DOM), primarily humic substances, from Narragansett Bay, RI, on the
bioavailability of selected petroleum hydrocarbons and No. 2 fuel oil to
the hard shell clam, Mercenaria mercenaria. Removal of DOM from seawater
significantly increased uptake of n-hexadecane from water. No significant
difference was evident in uptake of phenanthrene from seawater containing
DOM compared to seawater from which DOM had been removed. The clams
accumulated, on the average, seven times more total hydrocarbons from
dispersions of No. 2 fuel oil in DOM-free seawater than from dispersions in
seawater containing natural levels of DOM. In the absence of DOM, uptake
of saturated hydrocarbons from fuel oil was increased 17-fold whereas
uptake of aromatic hydrocarbons was increased 5-fold. Because of extremely
low aqueous solubilities, alkanes above about C$ are present in water
primarily as finely dispersed droplets (McAuliffe, 1966). Addition of DOM
either increases their solubility or decreases droplet size so that alkarces
can pass through the gill filter of Mercenaria. Dispersed oil droplets
would behave similarly to alkanes. However, phenanthrene is sufficiently
soluble at the concentrations used (1 mg/z) and therefore its physical
state in the water would not be affected by presence of DOM.
304
-------
Accumulation of PAH from Sediment
Bottom sediments of lakes, rivers, and coastal marine waters are the
ultimate repository for much of the PAH entering the aquatic environment
from various sources, large populations of microbes, plants, and animals
live on the surface of or within bottom sediments, particularly in
estuarine and coastal marine environments. Some benthic animals actually
ingest bottom sediments and remove organic materials from them as a source
of nutrition (deposit feeder). The question of bioavailability of
sediment-adsorbed PAH to these benthic organisms has received relatively
little attention. This is unfortunate since assimilation of PAH from
sediments by benthic organisms would provide a mechanism by which PAH could
be cycled from sediments into the aquatic food chain. In addition,
bioaccumulation of PAH from sediments by benthic organisms might pose a
health hazard to the predators (including man) of benthic organisms.
Rossi (1977) exposed the marine polychaete worm, Neanthes
arenaceodentata, to sediment contaminated with No. 2 fuel oil. Total
concentrations of naphthalenes in the sediment dropped from 9 to 3 ppm
during the 38-day exposure. However, naphthalenes concentrations in water
of the flow-through exposure system never rose above the 0.01 ppm detection
limit. Analysis of more than 20 replicate samples of worm tissues, taken
periodically throughout the exposure period, showed that the polychaetes
contained less than 0.1 ppm total naphthalenes at all sampling times. The
worms were apparently unable to accumulate naphthalenes from the sediment
although they were observed to ingest sediment and pass it through their
digestive tracts.
Anderson jit jil_. (1977) obtained similar results when they exposed the
sipunculid worm, Phascolosoma agassizii, to sediments contaminated with
Prudoe Bay crude oil. Total hydrocarbon concentration in the sediment
ranged from 475 to 765 ppm and total naphthalenes from 2.6 to 3.7 ppm
during the two-week exposure period. The highest mean concentrations of
total naphthalenes in the worms were 3.8 to 4.8 ppm, reached after 40-hr
exposure. Worms exposed to contaminated sediment for two weeks contained a
mean 1.9 ppm total naphthalenes. When worms were returned to clean
sediment, they rapidly released accumulated naphthalenes to undetectable
levels in two weeks or less. Sipunculids never contained naphthalene
concentrations significantly higher than those in the contaminated
sediment, and much of the hydrocarbon present in the worms was probably
associated with sediment in the gut. The authors concluded that the
sipunuclid worms were unable to accumulate naphthalenes from sediment to a
significant extent.
Fucik^t jil_. (1977) transferred the clam, Rangia cuneata, to sediments
at several stations in the vicinity of an oil separator platform in Trinity
Bay, TX. The bottom water near the platform contained 0.19 ppb total
naphthalenes and bottom sediments contained 27.7 ppm total naphthalenes.
Clams accumulated up to 33.6 ppm total naphthalenes in 97 days. There was
a good correlation between rates of uptake of naphthalenes by the clams and
levels of naphthalenes in sediments at the different stations. Gas
305
-------
of clam extracts showed a large unresolved envelope of weathered petroleum
similar to that in sediments. The authors concluded that at least part of
the naphthalenes accumulated by Rangia was derived from sediments. The
majority of tissue samples contained lower concentrations of naphthalenes
than did sediments, indicating that uptake of naphthalenes from sediments
by Rangia was very inefficient.
Accumulation of naphthalenes from oil-contaminated sediment and
detritus by the detritivorous clam, Macoma inquinata, was investigated by
Roesijadi et_ jj]_. (1978b). Sandy sediments artificially contaminated with
Prudoe Bay crude oil contained initial concentrations of 90.2 to 2750.8 ppm
total hydrocarbons and 0.45 to 11.96 ppm total naphthalenes. Total hydro-
carbon and naphthalene concentrations in the sediments decreased by 44.6
and 88.8%, respectively, during the 15-day exposure period. Tissues of
sediment-exposed clams contained 0.01 to 0.15 ppm total naphthalenes com-
pared to 0.01 to 0.07 ppm in controls. Thus, the clams were unable to
accumulate naphthalenes from the heavily contaminated sediment. Clams
exposed to detritus contaminated with Prudoe Bay crude oil spiked with
1 C-2-methylnaphthalene failed to accumulate significant concentrations
of C-2-methylnaphthalene from the detritus. All of the 2-methylnaph-
thalene in clam tissues could be accountecLfor via uptake from seawater,
presumably of solubilized material or of C-methylnaphthalene adsorbed
to very fine suspended matter.
Roesijadi et_ jil_. (1978a) studied accumulation of Prudoe Bay crude oil
and specific PAHs from oil-contaminated sediments by three benthic infaunal
invertebrate species: the sipunculid worm, Phascolosoma agassizii, and
clams, Macoma inquinata and Protothaca staminea. £. agassizii and JJ1.
inquinata are deposit feeders while £. staminea is a suspension feeder.
The animals were exposed on the lower shore of Sequim Bay, WA, in boxes
containing beachsand contaminated with Prudoe Bay crude oil. Total hydro-
carbon concentrations in the exposure sediment were 887.4 yg/g initially
and declined to 443.8 and 420.6 yg/g after 40 and 60 days in the field.
Hydrocarbon accumulation was generally greater in the two deposit feeding
species than in the suspension feeder (Table 12). The amounts of aliphatic
and diaromatic hydrocarbons accumulated by all species were quite low even
after 60 days of exposure to contaminated sediment. M. inquinata accumu-
lated up to 2.24 to 2.68 ppm total naphthalenes (mostly dimethyl naphthal-
enes) in 60 days. These results again illustrate the very limited bioavail-
ability of sediment-adsorbed petroleum hydrocarbons even to sediment-
ingesting benthic invertebrates.
Clams, M_. inquinata, were also exposed in the laboratory to sediments
containing detritus contaminated with Prudoe Bay crude oil and spiked with
four different C-PAH. One group of clams was exposed directly to the
sediment, while another group was suspended in the overlying water column.
This allowed the investigators to estimate the relative efficiency and
magnitude of PAH uptake from detritus and water. After seven days, the
sediment, water, and clams were analyzed for C activity. Efficiency
of PAH uptake from sediments was much lower than from water (Table 13).
Bioaccumulation factors for uptake of the four PAHs from contaminated
306
-------
sediments were 0.2 or less indicating no significant bioaccumulation of PAH
by this route. Bioaccumulation factors for uptake of the four PAHs from
seawater were in the 10.3 to 1349 range indicating a significant potential
for bioaccumulation, particularly of higher molecular weight PAHs, from
seawater. Thus, accumulation of PAH from sediment, when it occurs at all,
may be attributed in large part to uptake of PAH desorbed from sediment
particles into the interstitial water.
TABLE 12. CONCENTRATIONS OF ALIPHATIC AND DIAROMATIC HYDROCARBONS IN THE
TISSUES OF THREE BENTHIC INFAUNAL INVERTEBRATES, PHASCOLOSOMA
AGASSIZII, MACOMA INQUINATA. AND PROTOTHACA STAMINEA. FOLLOWING
EXPOSURE IN THE FIELD FOR 40 OR 60 DAYS TO SEDIMENT
CONTAMINATED WITH PRUDOE BAY CRUDE OIL AT AN INITIAL
CONCENTRATION OF 887.4 yg TOTAL HYDROCARBONS/g SEDIMENT (FROM
ROESIJADI et al., 1978A)
Species
Treatment
Hydrocarbon (uq/g '.vet
C12~C28
MN
DM.N
TN
P. agassizii
fl. inquinati
P. stamirea
Control
Control
Control
<0.10
<0.10
<0.10
'<0.005
<0.005
<0.005
<0.005
<0.005
<0.005
<0.01
<0.01
<0.01
<0.02
<0.02
<0.02
P.. aqassizii
P. agassizii
P. staninea
40 day exp.
40 day exp.
4Q cay exp.
40 day exp.
1.90
0.73
0.59
<0.10
-------
TABLE 13. ACCUMULATION OF 14C-PAH FROM SEDIMENT BY THE DEPOSIT-FEEDING
CLAM, MACOMA INQUINATA. CLAMS WERE EXPOSED TO SEVEN DAYS TO
SEDIMENT CONTAINING AN INITIAL CONCENTRATION OF 2000 yg/g
PRUDOE BAY CRUDE OIL SPIKED WITH 10 y CI OF THE PAH INDICATED
(FROM ROESI.JAOI etal., 1978A)
Parameter
.'Jet uptake froir sediment (ug/g)
UotaKe from seaivater (yg/g)
Concentration in sediment at
seven days Ug/g)
Concentration in seawater at
seven days (tig/ml )
Bioaccurrulation factor for
uptake frcn sedinent**
Bicaccurulation factor for
uotake from seawater***
PAH*
Phe Chry
0.096 0.308
0.038 0.297
0.49 8.37
3.7 x 10"3 4.3 x 10"4
0.20 0.04
10.3 694
DMBA
0.297
.0.856
4.53
6.3 x 10"4
0.06
1349
BaP
0.059
0.037
0.64
4.3 x 10"5
0.09
861
*Phe, phenanthrene; Cliry, chrysene; DMBA, dimethylbenz[a]anthracene; BaP, benzo[a]py-
rene.
*JC3lculated as net upta
-------
of aromatic hydrocarbons and after 51 days the liver contained only 124 ppb
1,2,3,4-tetramethylbenzene and 60 ppb 2-methylnaphthalene. The concentrat-
ions of aromatics found in the tissues of the fish throughout the experi-
ment were of the same order of magnitude as aromatic hydrocarbon concentra-
tions in the oiled sediment. However, trimethylnaphthalene, fluorene, and
phenanthrene, although present in the oiled sediment, were not detected in
the tissues of the fish. Thus, accumulation of aromatic hydrocarbons from
oiled sediment by fish is substantially less efficient than uptake from the
water.
All studies discussed above lead to the general conclusion that
sediment-adsorbed PAHs are not readily assimilated by benthic animals.
However, relatively few species have beeen studied. Nothing is known about
PAH uptake from sediments by deposit-feeding fish (e.g., mullet) or
attached aquatic plants.
Effects of Endogenous .an_d_E.xo_g_enp_u_s_factor;_s__p_n the Accumulation and Release
of PAH by Aquatic Organisms
Several endogenous biological factors such as size, nutritional status
body composition, age, and sex, as well as exogenous physical factors such
as salinity and temperature, may affect patterns of uptake and release of
PAH by aquatic organisms. Relatively few investigations have been
performed in these areas. Studies of this sort might provide valuable
information on mechanisms of PAH accumulation and retention by aquatic
organisms.
Endogenous Factors
Stegeman and Teal (1973) observed that oysters, Crassostrea virginica,
containing high concentrations of tissue lipids accumulated petroleum
hydrocarbons to higher concentrations than did low fat oysters. Several
other investigators have noted a positive correlation between lipid content
of marine organisms and their ability to accumulate petroleum hydrocarbons,
including PAH, from water. Harris et_ a\_. (1977a) studied accumulation of
^C-naphthalene by nine species of marine and estuarine copepods.
Retention of 14C-naphthalene following 24-hr exposure to this PAH in
solution varied nearly 16-fold among the different copepod species. Strong
positive correlations were drawn between body weight, ash-free dry weight
or total lipid content, and naphthalene retention, especially between total
lipid content and naphthalene retention. The authors went on to study the
uptake and retention of 14C-naphthalene by male and female Gal anus
helgolandicus which had either been starved for five days or fed Biddulphia
cells before the uptake experiment began. Starved male and female copepods
showed significantly lower levels of both total lipid and lipid as percent
of ash-free dry weight than did fed copepods. Starved animals also accumu-
lated significantly less 14C-naphthalene than fed individuals. Inter-
estingly, both starved and fed male copepods contained nearly three times
as much total lipid as starved and fed females, yet males and females
accumulated roughly equivalent concentrations of ^C-naphthalene. It
was suggested that lipid composition of males and females may be different
309
-------
and that these compositional differences may affect naphthalene retention
by each sex. Although C-naphthalene uptake and retention from
solution was positively correlated with total lipid content of the
copepods, rate assimilation of naphthalene from food was unrelated to total
lipid content of the animals. Lee (1975) observed that rate of
accumulation of H-BaP was significantly greater in large than small
copepods.
Mature male and gravid female polychaete worms, Neanthes
arenaceodentata, accumulated naphthalenes at a similar rate from the water-
soluble fraction of No. 2 fuel oil (Rossi and Anderson, 1977). However,
males rapidly released naphthalenes when returned to oil-free seawater,
while gravid females retained them until spawning occurred about 300 to 500
hr after the beginning of the depuration period. Newly released zygotes
contained high concentrations of naphthalenes and retained them during
early developmental stages to the trochophore stage several days later.
The trochophore and later juvenile stages rapidly released naphthalenes.
Gravid female Neanthes contain high concentrations of lipids, primarily
associated with the ovaries and developing eggs. Naphthalenes apparently
accumulated in these gonadal lipid stores and were released in eggs at
spawning. Yolk lipid stores are not utilized in this species until larvae
reach the trochophore stage. When the lipids were mobilized, naphthalenes
were released rapidly. Thus, hydrocarbons, which become associated with
stable lipid pools, such as depot lipids and gonadal-lipid stores, may be
retained until the animals mobilize lipids for nutritional purposes.
Hydrocarbons, which become associated with more labile hydrophobic compart-
ments, such as membrane lipids and cellular macromolecules, may be released
rapidly when ambient levels of hydrocarbons decrease. Such a two-compart-
ment model may partially explain the observation of several investigators
that depuration of chronically accumulated hydrocarbons is a two-phase
process characterized by an initial rapid release of hydrocarbons followed
by a second phase of very gradual release of remaining hydrocarbons.
Exogenous Factors
Temperature and salinity of the ambient medium have a profound effect
on may physiological functions in marine organisms. These factors also
affect solubility, adsorption-desorption kinetics, octanol/water partition
coefficients, etc., of PAH in water. Therefore, one would expect that
salinity and temperature would have a significant effect on accumulation
and release of PAH by marine organisms. Nevertheless, little research has
been done in this area.
Young (1977) acclimated groups of estuarine grass shrimp, Palaemonetes
pugio, to salinities of 5, 15, and 35 °/0<> (parts per thousand) or 2,
17, and 32 °/00 S. Each group was then exposed to either 3 ppm naphtha-
lene or 0.3 ppm phenanthrene at the acclimation salinity. Naphthalene up-
take was rapid at all salinities. Tissue naphthalene concentrations reached
a maximum in 2 hr. The highest level (57 ppm) was observed in animals
acclimated and exposed at the intermediate salinity, 15 °/°° S. At 5
310
-------
and 35 °/00 S, maximum naphthalene uptake was much less with 2-hr
values of 27.5 and 41 ppm, respectively. Similar results were obtained in
the phenanthrene exposure. Phenanthrene uptake was maximal (14.7 ppm) at
the intermediate salinity and significantly lower at 2 and 32 °/°° S
(both approximately 9.8 ppm)(Figure 4). When returned to phenanthrene-free
seawater of the acclimation salinity, all three groups of shrimp released
phenanthrene rapidly and at approximately the same rate. £. pugio is an
excellent osmoregulator over its entire normal environmental salinity range.
Body fluids are regulated hyperosmotic to the medium at low salinities and
hypoosmotic to the medium at high salinities (Roesijadi &t _§]_., 1976). At
intermediate salinities in the 15 to 17 °/00 range, body fluids of the
shrimp are isomotic to the medium. Water turnover rate measurements show
that permeability of_P. pugio is greatest at the isosmotic salinity and is
reduced at salinities which are associated with active osmoregulation.
These changes in apparent permeability of the shrimp are reflected in
changes in rate of uptake but not release of PAHs at different salinities.
By comparison, salinity had only a marginally significant effect on
rate of naphthalenes uptake from a WSF of southern Louisiana crude oil by
the estuarine marsh clam, Rangia cuneta (Fucik and Neff, 1977). Naphtha-
lenes uptake was lowest at 30 °/00 S and variable at lower slainities
(Figure 5). J}. cuneata is an osmoconformer throughout most of its normal
salinity regime and osmoregulates only at salinities below about 5 °/00
(Bedford and Anderson, 1972). For another clam, Protothaca staminea, which
is an osmoconformer throughout its entire environmental salinity regime,
salinity did not have a statistically significant effect on uptake of
naphthalenes from the WSF of southern Louisiana crude oil. Therefore,
physiological processes underlying osmoregulation by aquatic animals may
influence bioavailability to the animals of PAH in solution.
Temperature had a highly significant effect on rate of uptake of
naphthalenes from a WSF of southern Louisiana crude oil by both Rangia
cuneta and Protothaca staminea (Figure 5)(Fucik and Neff, 1977). Rate of
naphthalenes uptake was highest at the lowest temperature used and
decreased with increasing temperature. This inverse relationship between
temperature and naphthalene uptake rate was not due to the influence of
temperature on filtration rate of clams or on residence time of naphtha-
lenes in the water. Filtration rate of R. cuneata (the rate at which water
was pumped over the gills) increased in a nearly linear fashion with
increasing temperature, so that the gills presumed to be the major sites of
PAH uptake by bivalve molluscs (Lee _e_t aj_., 1972a) were actually exposed to
larger volumes of the WSF at higher temperatures where naphthalenes uptake
was lowest. Temperatures in the range used in these experiments did not
have a statistically significant effect on initial concentration or
residence time of naphthalenes in the exposure water. Rate of naphthalene
release from oil-contaminated clams returned to clean seawater was not
significantly affected by temperature. The influence of temperature on
rate of accumulation of C-naphthalene and 14C-2-methylnaphthalene
from solution by j*. cuneata was also investigated with similar results
(Fucik and Neff, unpublished observations): There was nearly a linear
inverse relationship between 14C uptake and temperature.
311
-------
O-
Q-
o
o
to
10
2%oS
17%o S
O24681O1224681OT2
EXPOSURE DEPURATION
SAMPLING TIME I Hours)
Figure 4. Accumulation and retention of phenanthrene by grass shrimp,
Palaemonetes pugio, which were acclimated to and exposed to 0.3
ppm phenanthrene at salinities of 2, 17 or 32°/°° (From
Young, 1977).
312
-------
Harris _et_ _al_. (1977a) also demonstrated an inverse relationship
between temperature and amount of -^C-naphthalene accumulated by the
copepod, Calanus helgolandicus, during exposure to 1 yg/£ -- C-naphthalene
for 24 hr (Figure 6). Accumulation of C-naphthalene in each copepod
decreased by about 39 pg (picograms), or by 3.23 pg/ug copepod lipid, per
10° C rise in temperature. The authors suggested that the effect of
temperature on "4C-naphthalene accumulation could be explained if rate
of hydrocarbon metabolism increased more rapidly than rate of uptake as
temperature increased. This would not explain the effect of temperature on
naphthalenes accumulation by molluscs because they have little if any
ability to metabolize naphthalenes. A better explanation of this phenomen
is provided by Herbes (1977). He measured the effect of temperature on the
adsorption of anthracene to non-living yeast cells from solution (0.02 yg/Ji)
The fraction of anthracene adsorbed by the yeast cells decreased signifi-
ficantly with increasing temperature. Calculated heat of adsorption for
this process was 5.2 kcal/mole which is characteristic of simple physical
(Van der Waals) adsorption. As temperature rises, the strength of this
weak chemical bond decreases, favoring desorption of PAH from the particles
Because partitioning of PAH between soluble and adsorbed phases is determi-
ined by relative rates of adsorptive and desorptive reactions, adsorption
will be increasingly favored over desorption as temperature decreases.
Therefore, temperature exerts its effect on PAH uptake primarily at the
initial step of the process--adsorption of PAH from water onto the surface
of a biological membrane.
!•
<
K
z
0
(J
W 3-
LU
2-
UPTAKE
DEPURATION
E
Figure 5. Uptake and release of total naphthalene by the clam, Rangia
cuneata, exposed to a 25% water-soluble fraction of south
Louisiana crude oil following acclimation for 14 days to
different combinations of salinity and temperature (From Fucik
and Neff, 1977).
313
-------
100-1
90-
a
o
0
£ 7a"
a
c 6O-
o>
>• 50-
30-
**-
o
c 20-|
c
• 10-
I
' 10 '
Temperature C
I
15
20
Figure 6
Effect of temperature on the accumulation of
14C-naphthaleneby the copepod, Calanus helgolandicus, during
exposure for 24 hr to 1 jig 14C-naphthalene/i seawater (From
Harris et^ al_., 1977a).
From the very limited research in this area, it is apparent that many
biotic and abiotic factors influence rates and patterns of PAH uptake,
retention, and release by marine organims. Since many of these variables
are not controlled in laboratory hydrocarbon uptake studies, and none are
controlled in field studies, it is not surprising that published rates of
hydrocarbon uptake and release by marine organisms are so variable. For
instance, several investigators have reported that the majority of hydro-
carbons accumulated by marine animals are released rapidly when the animals
are returned to clean seawater (Lee^t^K, 1972a,b; Stegeman and Teal,
1973; Neff et al., 1976b), whereas others have reported that hydrocarbons,
once accumulated", are released very slowly, if at all, even after long
periods of depuration (Blumer et^ al_., 1970; Boehm and Quinn, 1977). These
differences may be attributed to differences in such factors as the dur-
ation of exposure to hydrocarbons, lipid content and nutritional status of
experimental animals, age and sex of animals, and salinity, temperature,
and other physical environmental factors during the exposure and depuration
periods.
314
-------
REFERENCES
Anderson, J.W., R.C. Clark, and J.J. Stegeman. 1974. Petroleum
hydrocarbons. In: Marine bioassays workshop proceedings. Marine
Technology Society, Washington, DC. pp. 36-75.
Anderson, J.W., L.J. Moore, J.W. Blaylock, D.L. Woodruff, and S. L.
Kiesser. 1977. Bioavailability of sediment-sorbed naphthalenes to
the sipunculid worm, Phascolosoma agassizii. In: Fate and effects of
petroleum hydrocarbons in marine ecosystems and organisms. D.A.
Wolfe, Ed., Pergamon Press, New York. pp. 276-285.
Anderson, J.W., J.M. Neff, B.A. Cox, H.E. Tatem, and G.M. Hightower.
1974b. Effects of oil on estuarine animals: toxicity, uptake and
depuration, respiration. In: Pollution and physiology of marine
organisms. F.J. Vernberg and W.B. Vernberg, Eds. Academic Press, New
York. pp. 285-310.
Bedford, W.B., and J.W. Anderson. 1972. The physiological response of the
estuarine clam, Rangia cuneata, (Grey) to salinity. I. osmoregulation
Physiol. Zool. 45:255-260.
Bieri, R.H., and V.C. Stamoudis. 1977. The fate of petroleum hydrocarbons
from a No. 2 fuel oil spill in a semi natural estuarine environment.
In: Fate and effects of petroleum hydrocarbons in marine ecosystems
and organisms. D.A. Wolfe, Ed., Pergamon Press, New York. pp. 511-516
Bieri, R.H., V.C. Stamoudis, and M.K. Cueman. 1977. Chemical
investigations of two experimental oil spills in an estuarine
ecosystem. Proc. oil spill conference (prevention, behavior, control,
cleanup). American Petroleum Institute, Washington, DC. pp.
511-516.
Blumer, M., G. Souza, and J. Sass. 1970. Hydrocarbon pollution of edible
shellfish by an oil spill. Mar. Biol. 5:195-202.
Boehm, P.O., and J.G. Quinn, 1976. The effect of dissolved organic matter
in seawater on the uptake of mixed individual hydrocarbons and No. 2
fuel oil by a marine filter-feeding bivalve Mercenaria mercenaria.
Estuarine Coastal Mar. Sci. 4:93-105.
Boehm, P.O., and J.G. Quinn. 1977. The persistence of chronically
accumulated hydrocarbons in the hard shell clam, Mercenaria
mercenaria. Mari. Biol. 44:227-233.
Corner, E.D.S., R.P. Harris, C.C. Kilvington, and S.C.M. O'Hara. 1976a.
Petroleum compounds in the marine food web: short-term experiments on
the fate of naphthalene in Calanus. J. Mar. Biol. Assoc. U.K.
56:121-123.
315
-------
Corner, E.D.S., R.P. Harris, K.J. Whittle, and P.R. Mackie. 1976b.
Hydrocarbons in marine zooplankton and fish. In: Effects of
pollutants on aquatic organisms. A.P.M. Lockwood, Ed., Cambridge
. University Press, Cambridge, England, pp. 71-106.
Cox, B.A., J.W. Anderson, and J.C. Parker. 1975. An experimental oil
spill: the distribution of aromatic hydrocarbons in the water,
sediment, and animal tissues within a shrimp pond. Proc. conference
on prevention and control of oil pollution. American Petroleum
Institute, Washington, DC. pp. 607-612.
DiMichele, L., and M.H. Taylor. 1978. Histopathological and physiological
responses of Fundulus heteroclitus L. to naphthalene esposure. J.
Fish. Res. Bd. Canada.
Dixit, D., and J.W. Anderson. 1977. Distribution of naphthalenes within
exposed Fundulus similis and correlations with stress behavior. Proc.
oil spill conference (prevention, behavior, control, cleanup).
American Petroleum Institute, Washington, DC. pp. 633-636.
Dunn, B.P., and H.F. Stich. 1976. Release of the carcinogen
benzo(a)pyrene from environmentally contaminated mussels. Bull.
Environ. Contam. Toxicol. 14:398-401.
Dunn, B.P., and D.R. Young. 1976. Baseline levels of benzo(a)pyrene in
southern California mussels. Mar. Pollut. Bull. 7:231-234.
Engelhardt, R.R., J.R. Geraci, and T.G. Smith. 1977. Uptake and clearance
of petroleum hydrocarbons in the ringed seal, Phoca hispida. J. Fish.
Res. Bd. Canada. 34:1143-1147.
Fucik, K.W., H.W. Armstrong, and J,M, Neff. 1977. The uptake of
naphthalenes by the clam, Rangia cuneata, in the vicinity of an
oil-separator platform in Trinity Bay. TX. Proc. oil spill conference
(prevention, behavior, control, cleanup). American Petroleum
Institute, Washington, DC. pp. 637-640.
Fucik, K.W., and J.M. Neff. 1977. Effects of temperature and salinity on
naphthalenes uptake in the temperate clam, Rangia cuneata, and the
boreal clam, Protothaca staminea. In: Fate and effects of petroleum
hydrocarbons in marine ecosystems and organisms. D.A. Wolfe, Ed.,
Pergamon Press, New York. pp. 305-316.
Harris, R.P., V. Berdugo, E.D.S. Corner, C.C. Kilvington, and S.C.M.
O'Hara. 1977. Factors affecting the retention of a petroleum
hydrocarbon by marine planktonic copepods. In: Fate and effects of
petroleum hydrocarbons in marine ecosystems and organisms. D.A.
Wolfe, Ed., Pergamon Press, New York. pp. 2867-304.
316
-------
Harris, R.P., V. Berdugo, S.C.M. O'Hara, and E.D.S.
Accumulation of C-1-naphthalene by an oceani
Corner. 1977b.
oceanic and an estuarine
copepod during long-term exposure to low-level concentrations. Mar.
Biol. 42:187-195.
Herbes, S.E. 1977. Partitioning of polycyclic aromatic hydrocarbons
between dissolved particulate phases in natural waters. Water Res.
11:493-496.
Lawler, G.C., W.-A. Loong, and J.L. Laseter. 1978. Accumulation of
aromatic hydrocarbons in tissues of petroleum-exposed mallard ducks,
Anas platyrhynchos. Environ. Sci . Technol . 12:51-54.
Lee, R.F. 1975. Fate of petroleum hydrocarbons in marine zooplankton.
Proc. conference on prevention and control of oil pollution. American
Petroleum Institute, Washington, DC. pp. 549-554.
Lee, R.F., W.S. Gardner, J.W. Anderson, J.W. Blaylock, and J. Barwell -Clark.
1978. Fate of polycyclic aromatic hydrocarbons in controlled
ecosystem enclosures. Environ. Sci. Technol.
/
Lee, R.F., C. Ryan, and M.L. Neuhauser. 1976. Fate of petroleum
hydrocarbons taken up from food and water by the blue crab,
Callinectes sapidus. Mar. Biol. 37:363-370.
Lee, R.F., R. Sauerheber, and A. A. Benson. 1972a. Petroleum hydrocarbons:
uptake and discharge by the marine mussel, Mytilus edulis. Science
177:344-346.
Lee, R.F., R. Sauerheber, and G.H. Dobbs. 1972b. Uptake, metabolism, and
discharge of polycyclic aromatic hydrocarbons by marine fish. Mar.
Biol. 17:201-208.
Leo, A., C. Hansch, and D. Elkins. 1971. Partition coefficients and their
uses. Chem. Rev. 71:525-616.
Lu, P.-Y., R.L. Metcalf, N. Plummer, and D. Mandel . 1977. The environ-
mental fate of three carcinogens: benzo(a)pyrene, benzidine, and
vinyl chloride evaluated in laboratory model ecosystems. Arch.
Environ. Contam. Toxicol. 6:129-142.
McAuliffe, C. 1966. Solubility in water of paraffin, cycloparaffin,
olefin, acetylene, cycloolefin, and aromatic hydrocarbons, J.
Physical Chem. 70:1267-1275.
McCain, B.B., H.O. Hodgins, W.D. Gronlund, J.W. Hawkes, D.W. Brown, M.S.
Myers, and J.H. Vandermeulen. 1978. Bioavailability of crude oil
from experimentally oiled sediments to English sole, Parophrys
vetulus, and pathological consequences. J. Fish. Res. Bd. Canada.
35:657-664.
317
-------
Miller, D.L., J.P. Corliss, R.N. Farragut, and H.C. Thompson, Jr. 1978.
Accumulation and elimlination of the polynuclear aromatic hydrocarbon
chrysene by mangrove snapper, Lutjanus griseus, and pink shrimp,
Penaeus durarum. Unpublished manuscript, Southeast Fisheries Center,
Miami Lab, NMFS, NOAA, Miami, FL. 18 pp.
Neely, W.B., D.R. Branson, and G.E. Blau. 1974. Partition coefficient to
measure bioconcentration potential of organic chemicals in fish.
Environ. Sci. Technol. 8:1113-1115.
Neff, J.M. 1978. Polycyclic aromatic hydrocarbons in the aquatic
environment: sources, fates, and biological effects. American
Petroleum Institute, Washington, DC. 350 pp.
Neff, J.M., and J.W. Anderson. 1975. Accumulation, release, and
distribution of benzo(a)pyrene-C in the clam, Rangia cuneata.
Proc. conference on prevention and control of oil pollution. American
Petroleum Institute, Washington, DC. pp. 469-472.
Neff, J.M., J.W. Anderson, B.A. Cox, R.B. Laughlin, Jr., S.S. Rossi, and
H.E. Tatem. 1976a. Effects of petroleum on survival, respiration and
growth of marine animals. In: Sources, effects and sinks of
hydrocarbons in the aquatic environment. American Institute of
Biological Sciences, Washington, DC. pp. 515-540.
Neff, J.M., B.A. Cox, D. Dixit, and J.W. Anderson. 1976b. Accumulation
and release of petroleum-derived aromatic hydrocarbons by four species
of marine animals. Mar. Biol. 38:279-289.
Rice, S.D., R.E. Thomas, and J.W. Short. 1977. Effect of petroleum
hydrocarbons on breathing and coughing rates and hydrocarbon
uptake-depuration in pink salmon fry. In: Physiological responses of
marine biota to pollutants. F.J. Vernberg, A. Calalbrese, F.P.
Thurberg and W.B. Vernberg, Eds., Academic Press, New York.
pp. 259-278.
Roesijadi, G., J.W. Anderson, and J.W. Blaylock. 1978a. Uptake of
hydrocarbons from marine sediments contaminated with Prudoe Bay crude
oil: influence of feeding type of test species on availability of
polycyclic aromatic hydrocarbons. J. Fish. Res. Bd. Canada.
35:608-614.
Roesijadi, G., J.W. Anderson, S.R. Petrocelli, and C.S. Giam. 1976.
Osmoregulation of the grass shrimp, Palaemonetes pugio, exposed to
polychlorinated biphenyls (PCB). I. Effect on chloride and osmotic
concentrations and chloride- and water-exchange kinetics. Mar. Biol.
38:343-355.
318
-------
Roesijadi, G., D.L. Woodruff, and J.W. Anderson. 1978b. Bioavailability
of naphthalenes from marine sediments artificially contaminated with
Prudoe Bay crude oil. Environ. Pollut. 15:223-229.
Rossi, S.S. 1977. Bioavailability of petroleum hydrocarbons from water,
sediments and detritus to the marine annelid, Neanthes arenaceodentata.
Proc. oil spill conference (prevention, behavior, control, cleanup).
American Petroleum Institute, Washington, DC. pp. 621-626.
Rossi, S.S., and J. W. Anderson. 1977. Accumulation and release of
fuel-oil derived diaromatic hydrocarbons by the polychaete, Neanthes
arenaceodentata. Mar. Biol. 39:51-55.
Roubal, W.T., K.H. Bovee, T.K. Collier, and S.I. Stranahan. 1977a.
Flowthrough system for chronic exposure of aquatic organisms to
seawater soluble hydrocarbons from crude oil: construction and
applications. Proc. oil spill conference (prevention, behavior,
control, cleanup). American Petroleum Institute, Washington, DC.
pp. 551-556.
Roubal, W.T., T.K. Collier, and D.C.-Malins. 1977b. Accumulation and
metabolism of carbon-14 labeled benzene, naphthalene, and anthracene
by young coho salmon, Oncorhynchus kisutch. Arch. Environ. Contam.
Toxicol. 5:513-529.
Sanborn, H.R., and D.C. Malins. 1977. Toxicity and metabolism of
naphthalene: a study with marine larval invertebrates. Proc. Soc.
Exper. Biol. Med. 154:151-155.
Sharp, O.R., K.W. Fucik, and J.M. Neff. 1978. Physiological basis of
differential sensitivity of fish embryonic stages to oil pollution.
In: Marine pollution: Functional processes. F.J. Vernberg, W.B.
Verngerg, A. Calabrese, Eds., Academic Press, New York.
Statham, C.N., C.R. Elcombe, S.P. Szyjka, and J.J. Lech. 1978. Effect of
polycyclic aromatic hydrocarbons on hepatic microsomal enzymes and
disposition of methyl naphthalene in rainbow trout in vivo.
Xenobiotica 8:65-71.
Statham, C.N., M.J. Melancon, Jr., and J.J. Lech. 1976. Bioconcentration
of xenobiotics in trout bile: a proposed monitoring aid for some
waterborne chemicals. Science 193:680-681.
Stegeman, J.J., and J.M. Teal. 1973. Accumulation, release, and retention
of petroleum hydrocarbons by the oyster, Crassostrea virginica. Mar.
Biol. 22:37-44.
319
-------
Varansi, U., and D.C. Malins. 1977. Metabolism of petroleum
hydrocarbons: accumulation and biotransformation in marine organisms,
In: Effects of petroleum on artic and subartic marine environments
and organisms. Vol. II. Biological effects. D.C. Malins, Ed.,
Academic Press, New York. pp. 175-270.
Whittle, K.J., J. Murray, P.R. Mackie, R. Hardy, and J. Farmer. 1977.
Fate of hydrocarbons in fish. In: Petroleum hydrocarbons in the
marine environment. A.D. Mclntyre and K.J. Whittle, Eds., Cons.
Intern. Explor. Mer. Vol. 171, Charlottenlund Slot, Denmark, pp.
139-142.
Young, G.P. 1977. Effects of naphthalene and phenanthrene on the grass
shrimp, Palaemonetes pugio, (Holthuis). Master's Thesis, Texas A&M
University, College Station, TX. 67 pp.
320
-------
SOME.ASPECTS OF THE UPTAKE AND ELIMINATION OF THE
POLYNUCLEAR AROMATIC HYDROCARBON CHRYSENE
BY MANGROVE SNAPPER, LUTJANUS GRISEUS, AND
PINK SHRIMP, PENAEUS DUORARUM
by
Donald L. Miller, Jane P. Corliss, Robert^N. Farragut, and
Harold C. Thompson, Jr.
U.S. Department of Commerce
National Marine Fisheries Service
National Oceanic and Atmospheric Administration
Southeast Fisheries Center
Miami, FL 33149
ABSTRACT
The accumulation of the polynuclear aromatic hydrocarbon
(PAH), chrysene (found in shale and crude oil), was studied in
the mangrove snapper, Lutjanus griseus, and its accumulation and
elimination were studied in pink shrimp, Penaeus duorarum. When
exposed to 1 and 5 yg/fc chrysene in a closed seawater environ-
ment, snapper were found to concentrate the contaminant in their
livers, but not in other tissue (gallbladder, white muscle,
intestine). The shrimp accumulated chrysene in both the
cephalothorax and the abdomen. After exposure to chrysene for
28 days, shrimp transferred to fresh seawater released most of
the contaminant rapidly, but detectable amounts remained in
their bodies 28 days after the transfer.
INTRODUCTION
Spiral ing energy needs of the United States have put a strain on
presently exploited petroleum resources, requiring the petroleum industry
to continue a search for oil farther offshore in deeper waters. Expanded
oil production from offshore wells and the installation of the offshore
deep-water ports off the coast of Texas and Louisiana have increased the
potential for oil spills in the marine environment. According to NAS
(1975), oil released in offshore production accidents represent about 1.3%
(72.6 million kg/year) of the total 5.5 billion kg/year discharged in oceans,
Oil spillage could increase to 181.4 million kg/year by the early 1980s.
Present address: Department of Health and Human Services, Public Health
Service, Food and Drug Administration, National Center
for Toxicological Research, Jefferson, AR 72079.
321
-------
Moreover, offshore oil platforms attract large communities of marine
species comprising the lower as well as higher trophic levels. The most
productive shrimping grounds in the northern Gulf of Mexico are also major
oil producing areas. Consequently, shrimp and fish attracted to rig
structures are extremely vulnerable to both acute and chronic exposure to
petroleum hydrocarbons. Since petroleum contains carcinogenic aromatic
hydrocarbons (Mamedov, 1959; Gilchrist et jil_., 1972; McKay and Latham,
1973; Hurtubise, 1977), contamination of commercial fish species—such as
shrimp and snapper—may pose a potential health hazard to consumers of
marine foods. Other researchers have proven that upon exposure to certain
polynuclear aromatic hydrocarbons (PAHs), bivalves (Cahnmann and Kuratsune,
1957), other invertebrates (Koe and Zechmeister, 1952; Corner et al., 1973;
Rossi and Anderson, 1977), as well as some fish (Neff et^ jfl_., 197^7,
accumulate these contaminants in certain tissues.
The purposes of our study were two-fold. The first was to determine
if the carcinogen chrysene (Hecht et_ ^1_., 1974) found in shale (Lahe and
Eisen, 1968) and crude oils (Mamedov, 1959) is accumulated by two commer-
cially important marine organisms—pink shrimp, Penaeus duorarum, and
mangrove snapper, Lutjanus griseus—after exposure to 1 and 5 yg/£
concentrations of the contaminant in a closed seawater environment. The
second was to determine which tissues, if any, accumulate chrysene and how
rapidly the contaminant is eliminated after the organisms are transferred
to fresh seawater. Chrysene was chosen for its stability and relatively
low danger to the researchers.
MATERIALS and METHODS
Extraction
All solvents used were UV grade (Burdick and Jackson, Inc. ). The
extraction procedure for both water and tissue samples was a modification
of the technique developed by Bligh and Dyer (1959). In extracting
chrysene from seawater samples, a system of methanol, chloroform, and water
(sample) was established in the proportion 2:1:0.8 (v/v/wt) in a separatory
funnel. Then the solution was shaken vigorously for 2 min, 1 volume of
chlorofrom was added, the solution was shaken for 30 sec, 1 volume of
distilled water was added, and the solution was shaken again for 30 sec.
The phases were allowed to separate; the chloroform phase was removed,
filtered through glass wool, and concentrated via rotoevaporation for
analysis. This technique resulted in 90 to 95% recovery of chrysene.
For recovery of chrysene from tissue samples, whole lipid extracts
were prepared. The wet tissue was weighed and homogenized for 2 min in a
Virtis homogenizer (if the tissue contained tough connective tissue) or a
Potter-Elvejhem glass homogenizer with methanol and chloroform in the ratio
2:1:1 (MeOH:CHCl3:tissue v/v/wt). As in water extractions, the
chloroform phase was recovered by adding 1 volume CHC13, homogenizing for
The use of trade names does not imply endorsement by the National
Marine Fisheries Service or the U.S. Environmental Protection Agency.
322
-------
30 sec, then adding 1 volume distilled water, and homogenizing again for 30
sec. The solution was transferred to polyethlylene centrifuge tubes and
centrifuged for 10 min at 5000 rpm in a Sorvall RC-5 centrifuge equipped
with a SS-34 rotor. The chloroform phase was removed via a pipette,
filtered through glass wool, and concentrated via rotoevaporation. All
procedures were conducted at room temperature (25° C). This technique
resulted in 90 to 95% recovery of chrysene.
Analysis of Extracts
Samples were analyzed using a Hewlett-Packard Model 1084A liquid
chromatograph with either a Hewlett-Packard Model 1030B variable-wavelength
UV detector (set at 268 nm) or a HP fixed-wavelength UV detector (254- nm).
The reversed-phase columns used were either a HP RP-8 or LDC Datasorb ODS. 0
For most analyses, an isocratic solvent mode was effective, using MeOH:H2
in the proportion 82.5:17.5% at a flow rate of 1.5 nu/min. For some
separations, a linear gradient mode was utilized, starting with 50:50%
MeOH:H20, and increasing the MeOH at the rate of 2%/min. Although
chromatograms varied from animal to animal, no pre-column cleanup was
necessary with the lipid extracts. A solution of 1 yg/ms. chrysene
dissolved in chloroform was used as an external standard. All samples were
run in duplicate; values presented in tables and figures are averages of
duplicate determinations.
Snapper
Thirty locally caught mangrove snapper, Lutjanus griseus (mean wet
weight 163.8 g), were held in five circular (2081.9 t) fiberglass tanks
(1.83 m in diameter and 0.91 m high). Each tank contained 1690 a natural
seawater and a closed filtration system. Undergravel filters were
constructed from 2.54-cm PVC pipe; seven airlift pipes were arranged to
create a circular current. The filter bed consisted of crushed oyster
shell. Overhead mercury vapor lamps provided 12 hr of illumination. The
fish were fed frozen shrimp daily until they stopped feeding.
Before exposure, the natural disappearance rate of chrysene in the
tanks was determined to maintain a constant chrysene concentration of
either 1 or 5 yg/4. We determined that one-half of the chrysene had
disappeared within 3.5 hr after the addition of sufficient chrysene to
achieve the desired concentration. After 7 to 8 hr, chrysene had virtually
vanished (Figure 1). Therefore, one-half of chrysene necessary to achieve
the desired initial concentration was added every 3.5 hr following initial
addition of chrysene. Chrysene was applied to a tank as described; water
samples were removed, extracted, and analyzed at various time intervals
after the addition of contaminant. Water samples were periodically
withdrawn, extracted, and analyzed during the exposure period to insure
that a proper chrysene concentration was maintained.
Chrysene dissolved in acetone was added to the tanks via polyethylene
tubing connected to glass syringes. The chrysene concentration of the
standard in acetone delivered to 1 yg/A tanks was 0.184 mg/nu. The 1 yg/£
323
-------
5-1
4-
NJ
Wl
a.
O
I
bl
O
O
O
u
z
u
tt
X
o
2-
Figure 1.
1 23456
TIME (HOURS AFTER CHRYSENE INTRODUCTION)
Decreasing concentration of chrysene as a function of time for
fiberglass snapper tanks (volume = 1690 *). Values are averages of
duplicate samples.
-------
tanks received 4.6 mi of the standard during a 15-min period every 3.5 hr.
The chrysene concentration of the standard in acetone delivered to the 5
vg/i tanks was 0.92 mg/m£. The 5 yg/A tanks also received 4.6 nu of the
standard during 15-min period every 3.5 hr. The plungers of the syringes
driven by a Harvard syringe pump attached to an electric timer provided
desired concent rations of the contaminant.
Chrysene was maintained at 1 and 5 yg/A in two tanks for each
concentration. Only one tank was used as the control. Two fish were
sampled after 4 days, 7 days, and every 7 days thereafter from each tank
that recieved chrysene. Four fish were sampled from the control tank after
4 days and every 14 days thereafter. One fish also was sampled from each
tank before chrysene was added. Immediately after capture, fish were
rinsed with chloroform, and the brain was pithed. The fish were then
weighed, dissected, and filleted, and tissues (liver, gallbladder,
intestine, and muscle) were analyzed.
Shrimp
At the outset of the accumulation-elimination study with pink shrimp,
we determined appropriate parameters concerning the disappearance rate of
chrysene in 75.8-z glass aquaria, as determined for the fiberglass (snapper)
tanks. Again, one-half of the chrysene disappeared after 3.5 hr and
virtually vanished within 6 to 7 hr. Chrysene was added to glass aquaria
in the same manner as for the fiberglass tanks to maintain constant
chrysene concentrations. The chrysene concentration of the standard in
acetone delivered to l-yg/£ aquaria was 5.37 yg/mi; l-yg/£ aquaria received
4.6 ml of the standard during a 15-min period every 3.5 hr. The chrysene
concentration of the standard in acetone delivered to 5-yg/£ aquaria was
55 yg/nu. The 5-yg/£ aquaria received 2.22 nu of the standard over a 15-
min period every 3.5 hr.
Chrysene dissolved in acetone was pumped by a Harvard Metering Device
into 75.8-Ji glass aquaria containing 48.7 a natural seawater and maintained
at 1 and 5 yg/i. Four aquaria received each chrysene concentration, and
two aquaria received acetone only; three aquaria served as control. Twenty
pink shrimp were held in each aquarium. Filtration was under gravel with a
crushed oyster-shell filter bed. The shrimp were fed daily a prepared food
from the University of Arizona. The food was too small to be seen on the
crushed oyster shell; therefore, the amount was increased or decreased
according to the appearance of the water (yellowing of the water indicated
leaching of food). The number of molts was recorded daily.
Shrimp were sampled from each aquarium containing both concentrations
at Day 4,7, and every 7 days thereafter. Water samples were withdrawn
periodically to check contaminant levels. After chrysene contamination was
discontinued, each aquarium was drained, rinsed, and refilled with clean
seawater three times. Shrimp were sampled from the four aquaria that
received chrysene at the same intervals as during accumulation. Accumulation
and elimination were both measured for 28 days. Shrimp were sampled from
each control aquarium at Day 0, 4, 14, and 28 during both accumulation and
325
-------
elimination periods. No shrimp were sampled from acetone control aquaria
(maintained only during accumulation period). Immediately after removal
from the aquaria, the shrimp were washed with seawater and frozen. Since
the exoskeleton was removed, the shrimp were not rinsed in chloroform.
Before analysis, each sample was thawed and dissected. Tissues of the
cephalothorax, including the appendages, abdomen, and intestine (when 0.2 g
was available) were analyzed. The carapace was removed from the
cephalothorax and the exoskeleton, telson, and uropods removed from the
abdomen before extraction.
RESULTS
Snapper
This experiment was terminated prematurely when snapper died in all
tanks, including those in the control on Day 20. Deaths of the fish began
5 days after exposure to 1 and 5 yg/£ chrysene and after the 13th day in
the control tank. No difference in mortality was noted in fish exposed to
1 or 5 yg/z chrysene.
There appeared to be little difference in total food consumption by
fish exposed to the two chrysene concentrations either at the end of the
experiment and before the mass mortalities. On Day 4, food consumption by
fish declined in all tanks except the control, where a higher consumption
level was observed for the remainder of the experiment.
Tables 1 and 2 summarize results of 20-day exposure of Lutjanus
griseus to chrysene. After exposure to both 1 and 5 yg/z chrysene, JL.
griseus concentrated the PAH in the liver (Figure 2), but no consistent
significant accumulation was observed in other tissues studied. After 20
days, fish exposed to 5 yg/£ chrysene accumulated 1300 yg/kg of the PAH
(260 times the exposure level); fish exposed to 1 yg/i accumulated
360 yg/kg chrysene in their liver (360 times the exposure concentration).
Livers of fish that received 1 yg/£ chrysene accumulated 27% of the amount
of chrysene that was accumulated in the livers of fish exposed to the
5 yg/A chrysene.
Shrimp
Shrimp exposed to both 1 and 5 yg/i chrysene accumulated the PAH in
the cephalothorax and abdomen (Tables 3, 4); no chrysene was detected in
the intestines, nor was any significant amount detected in the control
organisms.
More chrysene was concentrated in the cephalothorax than in the
abdomen at both 1 and 5 yg/ji concentrations. Shrimp exposed to 5 yg/£
chrysene concentrated the PAH approximately 360-fold in the cephalothorax,
and approximately 84-fold accumulated in the abdomen by the end of the
exposure period (28 days) (Figure 3). Chrysene accumulated in the
cephalothorax throughout the exposure period, whereas concentrations in the
abdomen remained fairly constant after 14 days. The cephalothorax of
32fi
-------
TABLE 1. ACCUMULATION OF CHRYSENE IN VARIOUS TISSUES AFTER MANGROVE
SNAPPER, LUTJANUSGRISEUS, W£RE EXPOSED TO A lyg/* CONCENTRATION
IN A CLOSED SEAWATER SYSTEM
DAYS EXPOSURE
Tissue 0 4 7 14 20
White Muscle n.d.** n.d. n.d. n.d. 17.4
Liver n.d. 104 105 308 367
Intestine n.d. n.d. n.d. n.d. 190
Gallbladder n.d. n.d. n.d. n.d. n.d.
lues given are in yg/kg. Wet weight basis.
le detected which indicates amounts < 5 x 10 g/injectii
TABLE 2. ACCUMULATION OF CHYRYSENE IN VARIOUS TISSUES AFTER MANGROVE
SNAPPER, LUTJANUS GRISEUS. WERE EXPOSED TQ A 5 yg/JZ,
CONCENTRATION IN A CLOSED SEAWATER SYSTEM
DAYS EXPOSURE
Tissue 0 4 7 14 20_
White Muscle n.d. 15.8 n.d. n.d. n.d.
Liver n.d. 415 406 810 1,294
Intestine n.d. n.d. n.d. n.d. n.d.
Gallbladder n.d. n.d. n.d. n.d. n.d.
^ Values given are in yg/kg. Wet weight basis.
none detected
327
-------
TARI.F 3. ACCUMULATION AND ELIMINATION OF CHRYSENE IN VARIOUS TISSUES
Of PEHAEUS DUORARUM EXPOSED TO A 1 ug/fc CONCENTRATION IN A
CLOSFD SEAWATER SYSTEM
DAYS EXPOSURE
ACCUMULATION
Tissue
Cephal other ax
Abdomen
Intestine
0
n.d
25
n.d
4
.**84
91
. n.d .
7
170
124
^"
14 21
251 140
188 188
n.d. -
28 33
248 229
199 155
n.d.
ELIMINATION
35
196
142
^m
42
151
91
^"
49
89
90
—
56
48
91
"*
^ Values given are in ug/kg. Wet weight basis.
none detected.
TABLE 4. ACCUMULATION AND ELIMINATION OF CHRYSENE IN VARIOUS TISSUES OF
PENAEUS DUORARUfJ EXPOSED TO A 5 ug/4 CONCENTRATION IN A CLOSED
SEAWATER SYSTEM
DAYS EXPOSURE
ACCUMULATION ELIMINATION
Tissue 0 4 7 14 21 28 33 35
**
Cephalothorax n.d. 476 450 634 626 1,809 525 418
Abdomen 25 260 199 423 197 418 218 216
Intestine n.d. n.d. n.d. - - - - -
42 49 56
190 536 105
148 104 112
— _ —
*
AA. Values given are in pg/kg. Wet weight basis.
none detected.
shrimp exposed to 1 jig/A levels accumulated chrysene by approximately
250-fold over exposure levels, while the abdomen concentrated chrysene by
approximately 200-fold {Figure 4).
By the end of the experiment pink shrimp exposed to 5 yg/£ chrysene
had accumulated approximately 4.3 times more chrysene in the cephalothorax
than in the abdomen; pink shrimp exposed to 1 ug/z chrysene accumulated
approximately 1.2 times the amount of chrysene in the cephalothorax than in
the abdomen.
328
-------
12-
2.0-
X
Ml
JC
^
w
o
£
6-
KJ
O
Z
O
o
lit
Z
u
K
4-
o —o \mt/S. expotur*
• *5tif/£. exposure
.a
10
12
14
16
18
20
TIME (DAYS)
Figure 2. Accumulation of chrysene in mangrove snapper, Lutjanus griseus, liver
tissue.
-------
CO
CO
o
16-
2
X 14-
M
"12-
Z
O
8-
6-
O
o
"
U
-• Tail
• -o Cephalothorax
— 2
*""
8 12 16
Accumulation
20
24
28
32
36
TIME (DAYS)
40 44
Elimination
48
52
56
Figure 3.
Accumulation and elimination of chrysene by pink shrimp, Penaeus
duorarum, exposed to the 5 ug/£ level.
-------
CO
CO
O261
'Tail
° o Cephalothorax
8 10 12 14 16 18 20 22 24 26 28 30 32 34 36 38 40 42 44 46 48 50 52 54 56
Accumulation Elimination
TIME (DAYS)
Figure 4. Accumulation and elimination of chrysene by pink shrimp, Penaeus
duorarum, exposed to the 1 yg/£ level.
-------
The cephalothorax of shrimp exposed to 5 yg/£ chrysene contained
approxinatly 7.3 times more of the PAH than the cephalothorax of shrimp
exposed to 1 yg/£ chrysene. Abdomens of shrimp exposed to 5 pg/£
chrysene incorporated approximately 2.1 times as much chrysene as those of
shrimp exposed to 1 yg/i chrysene.
When the shrimp were returned to chrysene-free water, an initial
rapid loss of chrysene occurred in both body sections, followed by a longer
phase of gradual release. Release was more rapid in the cephalothorax than
in the abdomen, and the most rapid in cephalothorax from shrimp exposed to
5 pg/ji chrysene. There was little difference in food consumption between
shrimp receiving either chrysene concentration or acetone without chrysene.
Control shrimp consumed slightly more food than shrimp receiving chrysene.
Frequency of molting was approximately equal among shrimp during the 28-day
exposure to both chrysene concentrations or acetone without chrysene, and
shrimp used as controls. Molting during the following 28 days, however,
was more frequent in shrimp at both chrysene concentrations than in control
shrimp.
During the accumulation period mortality was higher in the aquaria
that received both concentrations of chrysene and acetone only than the
control aquaria. No difference in mortality was noted among shrimp at both
concentrations of chrysene. With the exception of one aquarium, mortality
began occurring on the 13th day of the chrysene exposure. From 1 to 9
shrimp were found dead each day for the remaining 15 days in the aquaria
receiving chrysene. Three deaths occurred in the control aquaria after air
lines had become disconnected. During the elimination period there were
only two deaths in the aquaria that received chrysene and none in the
control aquaria. There was no visible evidence of disease in any of the
aquaria; however, no histological examinations were made.
DISCUSSION
The results of our study indicate that the two organisms studied,
mangrove snapper and pink shrimp, can rapidly accumulate chrysene in their
tissues. In mangrove snapper, chrysene accumulated in significant amounts
only in the liver after 20-day exposure. In contrast to the findings of
other researchers (Neff et a\_.» 1976), snapper continued to concentrate the
aromatic hydrocarbon after 20 days. Other researchers have found that
aromatic hydrocarbons accumulate in the livers of other fish species (Lee
et^ aK, 1972) and therefore are transferred to the gallbladder. Although
no chrysene was detected in the gallbladder in our study, it is possible
that its metabolites, which were not monitored, could have been present.
Pink shrimp accumulated chrysene in the cephalothorax and abdomen with
increasing concentration as a function of time.
Difficulty in preventing disease in fish held in a closed laboratory
system was encountered in tests using the chrysene/acetone mixture. We
observed that when acetone was added to tanks, shrimp suffered a severe
sloughing of the mucous membrane. Sloughing probably also occurs in
similar tests with fish, increasing their susceptibility to disease.
332
-------
The first appearance of disease in the snapper tanks occurred only 5 days
after the chrysene/acetone mixture was added. At that time, foam was
prevalent on the water's surface. On the following day, the first two
mortalities occurred. The water had a slimy appearance that remained
throughout the experiment.
Airstones were frequently clogged and had to be changed. By Day 10,
the majority of fish in one tank had cloudy eyes, and on the following day,
11 died. (The cause was diagnosed as Cryptocaryon.) Also on Day 11, the
copepod, Argulus, was visible in each tank and on the fish, and was more
numerous in the tanks receiving chrysene than in the control tank. The
water was changed in all tanks and 0.25 yg/«, dylox added. No copepods were
visible in the water or on the fish the following day or thereafter. Water
was changed again that day because some fish in every tank had cloudy eyes;
25 yg/£ formaldehyde was also added to the tanks that received chrysene.
Shrimp exposed to either 1 or 5 yg/£ chrysene concentrated more of the
contaminant in the cephalothorax than in the abdomen, especially shrimp
exposed to 5 vg/x, chrysene. Other researchers, using different aromatics,
have found increased retention in the cephalothorax, especially in the
digestive gland of brown shrimp, Penaeus aztecus (Neff et a]_. , 1976). This
organ may act as a storage place for lipids (Vonk, 196977 After chrysene
exposures ceased, an initial rapid decrease in amounts of contaminant
occurred, followed by a prolonged slower release. Perhaps the initial
phase represents active excretion, while the second phase represents
passive diffusion. Whatever the mechanism of release, it should be noted
that detectable levels of the carcinogen were present even 28 days after
termination of exposure to 1 yg/fc concentration of chrysene.
Aquaria that received the chrysene/acetone mixtures and acetone only
had a large buildup of slime in the air tubes and on the inner sides. The
air tubes and stones frequently clogged and had to be cleaned. This
reduction in air was probably responsible for some deaths. During the last
8 days of the tests, water in the aquaria that received the chrysene/acetone
mixtures and acetone without chrysene became so cloudy that three water
changes were required. The rapid and extensive production of foam was also
evident on the water's surface. This excessive production of slime is
probably due to a sloughing of the mucous membranes in the presence of
acetone. There was no slime buildup or foam in the control aquaria.
Our study demonstrates that mangrove snappers and pink shrimp
accumulate chrysene after prolonged exposures. Although mangrove snappers
accumulate chrysene in their livers, liver tissue is not used for human
consumption and does not pose a health hazard to humans at this time. Pink
shrimp, on the other hand, accumulated chrysene in the cephalothorax and
abdomen (tail). Shrimp exposed to 5 yg/i chrysene concentrated the
hydrocarbon by 360-fold in the cephalothorax as compared to approximately
84-fold in the tail. Initially pink shrimp were able to eliminate chrysene
very rapidly from the cephalothorax and tail after exposure to the
contaminant was terminated. After the rapid initial elimination, a rather
333
-------
slow elimination process was observed. The shrimp still contained large
amounts of chrysene in both the cephalothorax and tail after 28 days in
chrysene-free seawater.
In our opinion, the level of chrysene remaining in the shrimp after 28
days of elimination could pose a serious health hazard if these shrimp were
consumed by humans for an extended period of time. Our primary
consideration at this point should be a more intensive study to project the
probable concentrations and duration of exposure of these two organisms to
chrysene during a crude-oil or shale-oil spill in the natural environment.
The data derived from a study of this scope would be extremely valuable in
determining whether an oil spill would actually contaminate commercially
important marine species used for human consumption with carcinogens of
sufficient quantity to pose a serious human health hazard.
ACKNOWLEDGMENTS
This work was supported in part by a contract with the U.S.
Environmental Protection Agency, Gulf Breeze Laboratoy, Gulf Breeze,
Florida, Contract EPA-IAG-D6-0084. We thank Dr. Donald V. Lightner of the
University of Arizona, Tuscon, for supplying the food to maintain the
shrimp for our study.
REFERENCES
Bligh, E.G., and W.J. Dyer. 1959. A rapid method of total lipid
extraction and purification. Can. J. Biochem. Physiol.
37(8):911-917.
Cahnmann, H.J., and M. Kuratsune. 1957. Determination of polycyclic
aromatic hydrocarbons in oysters collected in polluted water. Anal.
Chem. 29(9):1313-1317.
Corner, E.D.S., C.C. Kilvington, and S.C.M. O'Hara. 1973. Qualitative
studies on the metabolism of naphthalene in Mai a squinado (Herbst).
J. Mar. Biol. Assoc. U.K. 53:819-832.
Gilchrist, C.A., A. Lynes, G. Steel, and B.T. Whitham. 1972. The
determination of polycyclic aromatic hydrocarbons in mineral oils by
thin-layer chromatography and mass spectrometry.
Analyst 97:880-888.
Hecht, S.S., W.E. Bondinell, and D. Hoffman. 1974. Chrysene and
methylcnrysenes: presence in tobacco smoke and carcinogenicicty.
J. Nat. Cancer Inst. 53(4):1121-1133.
Hurtubise, R.J., J.F. Schalron, J.D. Feaster, and D.H. Therkildsen. 1977.
Fluorescence characterization and identification of polynuclear
aromatic hydrocarbons in shale oil. Anal. Chem. Acta 89:377-382.
334
-------
Koe, B.K., and L. Zechmeister. 1952. The isolation of carcinogenic and
other polycyclic aromatic hydrocarbons from barnacles. II. The goose
barnacle, Mitelia polymerus. Arch. Biochem. Biophys. 41:396-403.
Lahe, I., and 0. Eisen. 1968. Composition of polynuclear aromatic
hydrocarbons from heavy fractions of shale oil. Eesti NSV Teasd. Akad.
Toim. Khim. Geol. 17(1):30-31.
Lee, R.F., R. Sauerheber, and G.H. Dobbs. 1972. Uptake, metabolism, and
discharge of polycyclic aromatic hydrocarbons by marine fish.
Mar. Biol. 17:201-208.
Mamedov, C.I. 1959. Luminescence spectra of high molecular weight
petroleum hydrocarbons. Izv. Akad. Nauk. SSSR Ser. Fizicheskaya
23:126.
McKay, J.F., and D.R. Latham. 1973. Polyaromatic hydrocarbons in high
boiling petroleum distillates. Isolation by gel permeation
chromatography and identification by fluorescence spectrometry.
Anal. Chem. 45:1050-1055.
National Academy of Sciences (NAS). 1975. Petroleum in the marine
environment. Ocean Affairs Board, National Academy of Sciences,
Washington, DC. 107 p.
Neff, J.M., B.A. Cox, D. Dixit, and J.W. Anderson. 1976. Accumulation and
release of petroleum-derived aromatic hydrocarbons by four species of
marine animals. Mar. Biol. 38:279-289.
Rossi, S.S., and J.W. Anderson. 1977. Accumulation and release of
fuel-oil-derived diaromatic hydrocarbons by the polychaete, Neanthes
arenaceodentata. Mar. Biol. 39:51-55.
Vonk, H.J. 1969. Digestion and metabolism. In: The physiology of
Crustacea Vol. I. T.H. Waterman, Ed., Academic Press, NY.
pp. 291-316.
335
-------
ACCUMULATION, TISSUE DISTRIBUTION, AND DEPURATION OF BENZO(a)PYRENE
AND BENZ(a)ANTHRACENE IN THE GRASS SHRIMP, PALAEMONETES PUGIO
by
F.R. Fox and K. Ranga Rao
Department of Biology
University of West Florida
Pensacola, FL 32504
ABSTRACT
The short-term uptake, tissue distribution, and depuration
of two polvcyclic aromatic hydrocarbons, C-benzo(a)pyrene
(BP) and L C-benz(a)anthracene (BA), were studied utilizing
the grass shrimp, Palaemonetes pugio, at known stages of the
molt cycle. Premolt shrimp accumulated less BP and BA than
intermolt shrimp. The newly molted shrimp accumulated more BA
than intermolt shrimp. At each of the concentrations tested
[1.25, 2.5, 5.0, 10.0 parts per billion(ppb)], intermolt shrimp
accumulated BA to a greater extent than BP. The BA or BP
accumulated by shrimp increased in relation to environmental
levels of these compounds. The accumulation of BP and BA in
tissues examined was in the following order: digestive tract
(stomach + intestine)> hepatopancreas> cephalothorax> abdomen.
All tissues accumulated more BA than BP. When exposed to media
containing 2.5 ppb BP or 2.8 ppb BA, a rapid uptake by shrimp
was noted during the first 6-hr exposure, subsequently uptake
was reduced for BP. However, at termination of 96-hr exposure,
shrimp exhibited a trend of continual accumulation of BA and BP.
When transferred to seawater, shrimp appeared to depurate BA
more rapidly than BP. In the shrimp exposed to BA, the level of
radioactivity declined by 80% after a 7-day depuration; under
similar conditions, the BP level (radioactivity) declined by
only 35%.
INTRODUCTION
Concern with the possible contamination of the aquatic environment by
polycyclic aromatic hydrocarbons (PAHs) has led to an increase in studies
of the effects of these compounds on marine animals. Although PAHs are
derived from airborne particulates produced by forest fires, refuse burning
and the combustion of fossil fuels, petroleum and its products are also
implicated in PAH contamination. The detection of PAH in marine fish
336
-------
(Pancirov and Brown, 1977) and shellfish (Cahnmann and Karatsune, 1957;
Erhardt, 1972; Dunn and Stich, 1976; Bravo et aj_., 1978;) collected from
polluted waters adds relevance to the investigations of PAH contamination
of the marine ecosystem.
The accumulation, distribution, and release of PAH have been
investigated in commercially important shellfish (Lee £t aj_., 1972; Neff
and Anderson, 1975; Dunn and Stich, 1976) that are raised for human
consumption. The clam, Rangia cuneata, took up approximately 200 times the
ambient level of benzopyrene (BP) and retained about twice the ambient
level of BP at the end of thirty hours (Neff and Anderson, 1975).
Studies of PAH in crustaceans have been limited to copepods (Lee,
1975) and the blue crab, Callinectes sapidus (Lee ejt jil_., 1976). In both
studies it was found that crustaceans rapidly accumulate BP, reaching a
maximum in 2 days, whereas the release of BP is slow, occurring over two to
three weeks. Studies using the blue crab examined the distribution of BP
and its metabolites in various tissues.
In studies with crustaceans, the stage of the molt cycle must be
considered. Changes in cuticle permeability occur in relation to cyclic
shedding, secretion, and hardening of the exoskeleton in crustaceans
(Passano, 1960; Conklin and Rao, 1978). Conklin and Rao (1978) showed that
the uptake of pentachlorophenol, a chlorinated aromatic hydrocarbon, varied
with the stage of the molt cycle, the greatest accumulation occurring
immediately after ecdysis. Our paper discusses the accumulation, tissue
distribution, and retention of the polycyclic aromatic hydrocarbons,
benzopyrene and benzanthracene, at various stages of the molt cycle.
Materials and Methods
Animals—Grass shrimp, Palaemonetes pugio. were collected from grass
beds in Santa Rosa Sound, Gulf Breeze, FL, and maintained in large aquaria
containing filtered (5 ym) seawater of 10 °/oo salinity. Shrimp were
used within three weeks of collection and were not fed during experiments.
Experimental design--!ptermolt shrimp were exposed to 1.25, 2.5, 5.0,
10.0 ppb of either [7, 10- C]-benzo(a)pyrene (60.7 mCi/mmole) or
[12- C]-benz(a)anthracene (49 mCi/mmole) in 10 °/oo seawater (200 mA/animal)
animal). At the end of 6-, 12-, and 24-hr exposures, shrimp were
transferred to PAH-free seawater for a 2-min wash. Animals were blotted
dry, weighed, and placed into Protosol tissue solubilizer. Digested
samples were neutralized with acetic acid, and 15m«. of Aquasol II were
added. Samples were counted in a Beckman LS-133 liquid scintillation
counter, and values were corrected for quench and machine efficiency. A
sample (0.5 mi] of the medium was taken for each isotope to determine the
exposure levels of each isotope at the beginning of the experiment.
Animals at various stages of the molt cycle were exposed to a medium
(400 m£/shrimp) of 10 °/oo seawater containing either 2.5 ppb (2.5
337
-------
benzopyrene (BP) or 2.8 ppb benzanthracene (BA) for 3 hr to test the effect
of ecdysis on accumulation. The molt cycle stage was determined by the
method of Conklin and Rao (1978). After 3 hr, shrimp were placed in clean
seawater for two min, blotted dry, weighed, and placed in tissue
solubilizer.
Grass shrimp in stage C (intermolt stage of the molt cycle) were used
to determine the accumulation, retention, and distribution of BP and BA.
In the accumulation experiment, shrimp were exposed from 15 min to 96 hr in
either 2.5 ppb BP or 2.8 ppb BA. Animals were washed in clean seawater,
dried, weighed, and solubilized before scintillant was added; samples were
counted for radioactivity.
Tissue distribution of BP and BA was followed during the first 24 hr
of accumulation. At the appropriate time intervals, shrimp were washed and
dried before dissection to remove the hepatopancreas, the digestive tract
(stomach and intestine), and the abdominal muscle. The remainder of the
shrimp (cephalothorax) as well as the tissues dissected were solubilized
and counted for radioactivity after being weighed. Shrimp were exposed to
either 2.5 ppb BP or 2.8 ppb BA for 12 hr. They were then transferred to
uncontaminated seawater and sacrificed at intervals ranging from 3 to 168
hr for analysis of radioactivity. The tissue distribution of radioactivity
from BP or BA was observed during the depuration process. The four parts
of the shrimp discussed above were analyzed for retention of radioactivity.
Intermolt shrimp and newly molted shrimp were exposed to BP or BA for
3 hr to compare tissue distribution. The four parts of the dissected
shrimp were analyzed for radioactivity.
Chemicals— C-Benzopyrene and C-benzanthracene were purchased from
Amersham/Searle, Arlington Heights, IL. Protosol and Aquasol II were
purchased from New England Nuclear Corp., Boston, MA.
Results
The accumulation of [7, 10- C]-benzo(a)pyrene in grass shrimp
exposed to 1.25, 2.5, 5.0, and 10.0 ppb of benzopyrene (BP) medium is shown
in Figure 1. Uptake in shrimp exposed to 10.0 ppb was approximately 5
times as great as that of shrimp exposed to 1.25 ppb at 12 and 24 hr
exposure. The difference was much less at 6 hr exposure. Shrimp exposed
to 2.5 ppb accumulated about 1.5 times as much as shrimp in 1.25 ppb, while
shrimp in 5.0 ppb medium took up about 2.5 times that of shrimp in the
least concentrated medium.
The accumulation of [12- C]-benz(a)anthracene in shrimp exposed
to the four concentrations is shown in Figure 2. Shrimp exposed to 10.0
ppb benzanthracene (BA) medium had an uptake about 8 times that of shrimp
exposed to 1.25 ppb medium when exposed for 12 and 24 hr; those in 2.5 ppb
took up almost twice the mount of shrimp exposed to 1.25 ppb. The
difference was less after 6 hr exposure. The shrimp in 5.0 ppb BA
accumulated about 4 times as much as those in 1.25 ppb medium.
338
-------
To conserve labeled compounds yet maintain sufficient levels of
radioactivity, we used 2.5 ppb BP or 2.8 ppb BA (equivalent in DPM to 2.5
ppb BP) for the remaining experiments.
Accumulation of BP and BA by the grass shrimp, Palaemonetes pugio, in
different molt stages is shown in Figure 3. Although BP uptake varies
considerably in the stages of the molt cycle, BA was taken up more than BP
at every stage of the molt cycle (the greatest difference was observed
immediately after the molt).
Shrimp were exposed to experimental media for various times to observe
the accumulation rate. As shown in Figure 4, grass shrimp accumulated
benzanthracene to a greater extent than benzopyrene for the time intervals
tested. Both compounds were taken up quickly during the first 6 hr.
During the next 3 days, BA continued to accumulate while BP appeared to
maintain the same level of incorporation.
Accumulation of benzopyrene in the tissues paralleled that observed in
the whole body (Figure 5). The amount increased in all the tissues in the
first 3 hr but changed relatively little during the next 18 hr. The
digestive tract (stomach and intestine) showed the greatest uptake at 6 hr
while hepatopancreas had about one-fourth the amount observed in the
digestive tract. The abdominal muscle and cephalothorax were relatively
low in radioactivity.
Benzanthracene uptake in tissues also paralleled the whole animal
trend. Figure 6 shows that only the abdominal muscle failed to show an
increase in uptake after 1 hr exposure. The digestive tract had the
greatest accumulation over the 24-hr period. The hepatopancreas
accumulated about one-half and the abdominal muscle and the cephalothorax
about one-twentieth of that found in the digestive tract.
The distribution of BP and BA in the four parts of the intermolt and
newly molted shrimp was examined after a 3 hr exposure (Table 1). The
accumulation of BP in the tissue other than the abdominal muscle of newly
molted shrimp was not significantly different from that in intermolt
shrimp. But a nearly three-fold increase in uptake of BP occurred in the
abdominal muscle of newly molted shrimp in comparison with intermolt
shrimp. The accumulation of BA in the abdominal muscle, hepatopancreas,
and digestive tract of newly molted shrimp was significantly (P = 0.05)
different from that in intermolt shrimp. A nearly two-fold increase was
observed for these tissues.
Figure 7 shows the retention of BP and BA radioactivity after an
exposure period of 12 hr. Benzanthracene was lost from grass shrimp faster
and to a greater extent over the seven day depuration period than
benzopyrene; 35% of the accumulated BP radioactivity was lost by the shrimp
after 7 days, while about 80% of the initial BA radioactivity was lost from
shrimp during the same period.
339
-------
4OO
-*-*
I)
o>
Q
o 200
u
u
BP
12
Time(hr)
24
Figure 1. Accumulation of benzopyrene in stage C shrimp which were exposed
to four concentrations (ppb) of the hydrocarbon for three time
intervals. The medium (200 ma/shrimp) contained 17 DPM/Ji.
Values are the mean _+ SEM for 8 shrimp.
12
Time(hr)
Figure 2. Uptake of benzanthracene in stage C shrimp which were exposed to
four concentrations (ppb) of the hydrocarbon for three time
periods. The medium (200 mi/shrimp) contained 17.3
Values are the mean _+ SEM for 8 shrimp.
340
-------
-420-
CD
£300-
Q_
Q
c
o
To
Zi
E
D
U
u
< 60-
BAQ
BPm
i
11
rr
f
rh
i
[
i
^
*
:;;^
I
i:
-
D
Stage
3 12 18
Postmolt time(hr)
Figure 3. Accumulation of benzopyrene and benzanthracene at various stagesp)
of the molt cycle in the grass shrimp. The medium (400 mji/shrim
contained 17 DPM/y£ or 17.3 DPM/y£, respectively. Shrimp
remained in the medium for 3 hr. C = intermolt, D = premolt,
E = ecdysis (molt). Values are the mean +_ SEM for 8 shrimp.
24
Time (h1-)
Figure 4. Accumulation of benzopyrene and benzanthracene over 96-hr
period in stage C shrimp. The medium contained 2.5 ppb BP or
2.8 ppb BA (200 m£/shrimp) which had 17 DPM/u£ or 17.3
respectively. The values are the mean _+ SEM for 8 shrimp.
341
-------
Figure 5. Uptake of benzopyrene (2.5 ppb) in the tissues of the grass
shrimp in 24 hr. The medium (200 mi/shrimp) contained 17
DPM/yfc. A=abdominal muscle, C=cephalothorax, D=digestive tract,
H=hepatopancreas. Values are given as the mean _+ SEM for 7
shrimp.
I
375-"
525
Figure 6. Uptake of benzanthracene (2.8 ppb) in the tissues of the stage C
grass shrimp in 24 hr. The medium (200 mA/shrimp) contained
17.3 DPM/yi. A=abdominal muscle, C=cephalothorax, D=digestive
tract, H=hepatopancreas. Values are the mean _+ SEM for 8 shrimp.
342
-------
TABLE 1. DISTRIBUTION OF PAH IN TH TISSUES OF INTERMOLT AND ECDYSIAL GRASS SHRIMP
Tissues
Digestive tract
(stomach + intestine)
Hepatopancreas
CM
£ Cephalothorax
Abdominal muscle
Benzopyrene Uptake
Intermolt
2604 + 162 (7)
694 +_ 78 (7)
166 + 23 (7)
49 +_ 10 (8)
(DPM/mg wet wt.)
Molt
3028 + 355 (8)
1089 + 182 (7)
207 + 22 (7)
131 + 20 (7)
Benzanthracene Uptake
Intermolt
5409 + 462 (8)
2528 + 363 (8)
387 + 29 (8)
143 + 16 (7)
(DPM/mg wet wt.)
Molt
8190 +_ 620 (7)
4838 ±620 (7)
479 + 53 (6)
327 _+ 41 (7)
Shrimp were exposed to either 2.5 ppb benzopyrene (17 DPM/yji) or 2.8 ppb benzanthracene (17.3
for 3 hr. After transfer to seawater, the shrimp were dissected for the tissues which were weighed,
solubilized and counted in a liquid scintillation counter. The values are expressed as the mean +_ SEM
and the number of animals are given in parenthesis. A t-test was done on the intermolt and molt values
for each compound. BP values for the abdominal muscle were significantly different at the 0.05 level.
The BA values for digestive tract, hepatopancreas and abdominal muscle are significantly different at the
0.05 level. Other values are not significant.
-------
= 360
5
a
E
Q.
Q
c
or
I
2
Time(hr)
72
168
Figure 7. Retention of benzopyrene (2.5 ppb) and benzanthracene (2.8 ppb)
in the stage C shrimp during a 7-day depuration period. The
shrimp were exposed to the PAH (200 nu/shrimp) for 12 hr before
transfer to seawater. The medium contained 17 DPM/y* or 17.3
respectively. Values are the mean +_ SEM for 7 shrimp.
3OOO
168
Time(hr)
Figure 8. Retention of benzopyrene (2.5 ppb) in the tissues of the shrimp
during a 7-day depuration. The shrimp were exposed to the PAH
(200 mi/shrimp) for 12 hr prior to removal to uncontaminated
seawater. The medium contained 17 DPM/uJi. A=abdominal muscle,
C=cephalothorax, D=digestive tract, H=hepatopancreas. Values
are the mean + SEM for 6 shrimp.
344
-------
The retention of benzopyrene radioactivity by the different tissues is
shown in Figure 8. Although the four parts of the shrimp showed little
reduction in radioactivity, the digestive tract decreased appreciably over
the 7-day period. These results agreed with the whole-body counts.
The pattern of depuration of benzanthracene radioactivity in the grass
shrimp tissues (Figure 9) is consistent with that of whole animals. The
abdominal muscle showed the least change, whereas digestive tract again
had the greatest loss of radioactivity. The hepatopancreas and the
cephalothorax lost very little radioactivity until the fourth day.
900O-
BA
SOO-i
1 300-t
100
72
Time Ihr)
168
Figure 9.
Retention of benzanthracene (2.8 ppb) in the tissues of the
shrimp during a 7-day depuration . The shrimp were exposed to
the compound (200 ma/shrimp) for 12 hr prior to removal to clean
seawater. The medium contained 17.3 DPM/pJi. A=abdominal
muscle, C=cephalothorax, D=digestive tract, H=hepatopancreas.
Values are the mean _+ SEM for 7 shrimp.
DISCUSSION
The accumulation of aromatic hydrocarbons by a variety of aquatic
animals has been studied (Lee et^ al_., 1972a,b, 1976; Stegeman and Teal,
1973; Lee, 1975; Neff and Anderson, 1975; Neff et _al_., 1976; Harris et_al_.,
1977; Rossi and Anderson, 1977; Roubal et al., 1977; Melancon and Lech,
1978). Accumulation is rapid in fish (Melancon and Lech, 1978) as well as
in crabs (Lee et_aj_., 1976) and oysters (Neff et a]_., 1976). Most of the
compounds reach a maximum concentration in one to two days; however,
depuration generally occurs over a 3-week period.
345
-------
The uptake of PAH in grass shrimp was rapid as observed in previous
studies of aromatic hydrocarbon uptake in crustaceans. However, the
influence of the molt cycle showed considerable variation in uptake of the
hydrocarbons. Our studies indicate that PAHs may enter the shrimp more
easily at ecdysis when the cuticle is more permeable. Conklin and Rao
(1978) presented evidence that the grass shrimp accumulates pentachloro-
phenol to a greater extent at ecdysis than at any other stage in the molt
cycle.
The incorporation of radiolabeled benzopyrene and benzanthracene into
shrimp within minutes of exposure to these compounds has shown the perme-
abililty of crustaceans to these PAH. Lee and others (1972b) detected
large amounts of benzopyrene in tissues of marine fish within minutes of
exposure. The uptake of naphthalenes occurs within the first 10 hr of
exposure in clams (Neff et al., 1976), within the first 3 hr exposure in
marine polychaete worms "[Rossi and Anderson, 1977), and within 30 min in
brown shrimp (Anderson et^ al_., 1974), indicating that marine invertebrates
are rapid accumulators of aromatic hydrocarbons. The bioaccumulation of BP
in clams was 200 times the ambient level (Neff et^ al_., 1976) in 24 hr,
whereas the grass shrimp accumulated 18 times the ambient level at the end
of 96 hr. However, grass shrimp accumulated 24 times the ambient level of
BA over the same period.
Benzopyrene and benzanthracene distribution in grass shrimp was not
in total agreement with studies on other marine animals. Lee and coworkers
(1972b) showed that fish liver accumulates more benzopyrene than stomach
and that crab hepatopancreas took up considerably more benzopyrene than the
stomach (1976). However, in terms of uptake per unit weight of the tissue,
the stomach (digestive tract) accumulated more PAH than hepatopancreas in
the grass shrimp. The abdominal muscles in the crab (Lee et al., 1976) and
brown shrimp (Neff et^ aK, 1976) appeared to take up as much" "Rydrocarbon as
the stomach; yet the grass shrimp muscle accumulated very low quantities
compared to digestive tract on a tissue weight basis. The digestive tract
accumulated BP 154 times the ambient level, while BA accumulation was 376
times the ambient level. The muscle and the cephalothorax accumulated the
same or less than the whole shrimp: muscle had 8.5 times the BA and 3.5
times the BP; the cephalothorax, 25 times the BA and 10 times the BP. The
hepatopancreas accumulated almost 5 times as much BA (bioaccumulation
factor = 231) than BP (bioaccumulation factor =49).
In studies of the distribution of BP and BA in the tissues of
intermolt and newly molted shrimp, BA accumulated to a higher level than
BP. While the abdominal muscle showed greater accumulation of BP at the
molt, the digestive tract, hepatopancreas, and abdominal muscle showed
greater uptake of BA in newly molted animals.
When comparing the rate of discharge of aromatic hydrocarbons in .
marine invertebrates (Harris jet al., 1977), copepods appeared to lose
naphthalene rapidly in the first~4~8 hr of depuration. Brown shrimp
eliminated a large amount of naphthalene in minutes (Anderson e_t a\_., 1974),
whereas Neff and coworkers (1976) demonstrated that initially brown shrimp
346
-------
rapidly released naphthalene followed by a long period of gradual
depuration. In other studies, the blue crab slowly depurated aromatic
hydrocarbons (Lee et al., 1976); in two days, the blue crab lost about half
of the benzopyrene accumulated over the previous two days. Benzopyrene
loss in grass shrimp was slower than in blue crab and may be equated with
the release from the clam, Rangia cuneata (Neff and Anderson, 1975), which
lost 10% of its PAH burden in 6 days. Lee and others (1972b) found that
fish discharged half of their benzopyrene burden from gut and liver in the
first day of depuration. The blue crab released about half of its burden
from hepatopancreas but showed little change in stomach benzopyrene after
two days of depuration (Lee jst jal_., 1976).
Our investigation as well as previous studies using invertebrates and
fish suggests that aromatic hydrocarbons are rapidly accumulated in marine
animals and that their release is dependent upon the number of benzenoid
rings present. Naphthalenes are rapidly discharged, while PAHs are slowly
released. Benzathracene appears to depurate more easily than its related
carcinogen, benzopyrene. Further investigations of several aromatic
hydrocarbons with differing numbers of benzenoid rings in a variety of
marine animals would help in clarifying PAH contamination of the marine
environment. Contradictions observed in the limited study of these
compounds in marine animals can only be "eliminated by further studies.
ACKNOWLEDGEMENTS
This investigation was supported by Grant R-80454-01 from the U.S
Environmental Protection Agency.
REFERENCES
Anderson, J.W., J.M. Neff, B.A. Cox, H.E. Tatem, and G.M. Hightower. 1974.
The effects of oil on estuarine animals: toxicity, uptake and
depuration, respiration. In: Pollution and physiology of marine
organisms. F.J. Vernberg and W. Vernberg, Eds., Academic Press, Inc.,
New York. pp. 285-310.
Bravo, A.H., S. Salazar, A.V. Botello, and E.F. Mandelli. 1978.
Polyaromatic hydrocarbons in oysters from coastal lagoons along the
eastern coast of the Gulf of Mexico, Mexico. Bull. Environ. Contam.
Toxicol. 19:171-176.
Cahnmann, H.J., and M. Karatsune. 1957. Determination of polycyclic
aromatic hydrocarbons in oysters collected in polluted water. Anal.
Chem. 29:1312-1317.
Conklin, P.J., and K.R. Rao. 1978. Toxicity of sodium pentachlorophenate
to the grass shrimp, Palaemonetes pugio, in relation to the molt
cycle. In: Pentachlorophenol. K.R. Rao, Ed., Plenum Publishing Co.,
New York. pp. 181-192.
347
-------
Dunn, B.P., and H. F. Stich. 1976. Release of the carcinogen
benzo(a)pyrene from environmentally contaminated mussels. Bull.
Environ. Contain. Toxicol. 15:398-401.
Dunn, B.P., and D.R. Young. 1976. Baseline levels of benzo(a)pyrene in
southern California mussels. Mar. Pollut. Bull. 7:231-234.
Erhardt, M. 1972. Petroleum hydrocarbons in oysters from Galveston Bay.
Environ. Pollut. 3:257-271.
Harris, R.P., V. Berdugo, S.C.M. O'Hara, and E.D.S.
Accumulation of ^C-l-naphthalene by an oceani
Corner. 1977.
oceanic and an estuarine
copepod during long-term exposure to low level concentrations. Mar.
Biol. 42:187-195.
Lee, R.F. 1975. Fate of petroleum hydrocarbons in marine zooplankton.
In: Proceedings of 1975 Conference on Prevention and Control of Oil
Pollution. American Petroleum Institute, Washington, DC. pp. 549-553.
Lee, R.F., R. Sauerheber, and A.A. Benson. 1972a. Petroleum hydrocarbon
uptake and discharge by the marine mussel, Mytilus edulis. Science
177:344-346.
Lee, R.F., R. Sauerheber, and G.H. Dobbs. 1972b. Uptake, metabolism, and
discharge of polycyclic aromatic hydrocarbons by marine fish. Mar.
Biol. 17:201-208.
Lee, R.F., C. Ryan, and M.L. Neuhauser. 1976. Fate of petroleum
hydrocarbons taken up from food and water by the blue crab,
Callinectes sapidus. Mar. Biol. 37:363-370.
Melancon, M.J. Jr., and J.J. Lech. 1978. Distribution and elimination of
naphthalene and 2-methylnaphthalene in rainbow trout during short-
and long-term exposure. Arch. Environ. Contam. Toxicol. 7:207-220.
Neff, J.M., and J. W. Anderson. 1975.. Accumulation, release and
distribution of benzo(a)pyrene-C1 in the clam, Rangia cuneata.
In: Proceedings of the 1975 Conference on Prevention and Control of
Oil Pollution. American Petroleum Institute, Washington, DC.
pp. 469-471.
Neff, J.M., B.A. Cox, D. Dixit, and J.W. Anderson. 1976. Accumulation and
release of petroleum-derived aromatic hydrocarabons by four species of
marine animals. Mar. Biol. 38:279-289.
Panicrov, R.J., and R. A. Brown. 1977. Polynuclear aromatic hydrocarbons
in marine tissues. Environ. Sci. Tech. 11:989-992.
Passano, L.M. 1960. Molting and its control. In: The physiology of
crustracea, Vol. I. T.H. Waterman, Ed., Academic Press Inc., New
York. pp. 473-536.
34R
-------
Rossi, S.S., and J.W. Anderson. 1977. Accumulation and release of
fuel-oil-derived diaromatic hydrocarbons by the polychaete, Neanthes
arenaceodentata. Mar. Biol. 39:51-55.
Roubal, W.T., T.K. Collier..and D.C. Mai ins. 1977. Accumulation and
metabolism of carbon- labeled benzene, naphthalene, and
anthracene by young coho salmon, Oncorhynchus kisutch. Arch. Environ.
Contam. Toxicol. 5:513-529.
Stegeman, J.J., and J.M. Teal. 1973. Accumulation, release and retention
of petroleum hydrocarbons by the oyster, Crassostera virgim'ca. Mar.
Biol. 22:37-44.
349
-------
AN ECOLOGICAL PERSPECTIVE ON HUMAN FOOD WEBS
by
Rufus Mori son
Integrated Pest Management Research Program
Office of Processes and Effects Research
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC 20460
ABSTRACT
Summaries in this text illustrate the complexity of human
food webs and indicate the present lack of understanding in the
area of human food web ecology. Increases in the numbers and
rates of introduction of hazardous chemicals in effect may
reduce the ability of organisms essential to food webs to react
in a timely manner. Many facets of this complex problem are
being addressed at present, but no structured program exists.
It is recommended that an ad hoc committee be sponsored by
the U.S. Environmental Protection Agency (EPA) to organize
ecological research needs and priorities on the subject of
xenobiotics in human food webs. Also, a strategy should be
developed to assist in the implementation of an information
management system.
INTRODUCTION
Chemical constituents of food relevant to food safety include a wide
spectrum of substances that are introduced through varied routes. The
purpose of this document is to describe the need for an understanding of
the origin of environmental contaminants in the human diet. The complex
subject might best be approached through a basic knowledge of the movement
of xenobiotics through food chains to man.
Food-chain studies can provide a framework for an integrated approach
to ongoing research. Results from such studies could be applied to develop
a strategy to prevent the exposure of man to xenobiotics through the food
chain.
The development of such a strategy requires a comprehension of the
transport and fate of food contaminants that move through the abiotic
(physical) compartments and biotic (living) compartments of the food chain.
Of particular complexity are the interfaces of physical compartments, i.e.,
air and water, or water and soil.
350
-------
The biotic compartments of food webs are poorly understood. Processes
such as biotransformation of chemicals and the structure of food webs are '
not well-known. Very little is known about diet-related human behavior
such as actual dietary intake, nutritional characteristics and regional'
preferences, and changes in dietary habits. Thus, the prediction of human
exposures becomes a speculative art. Because human dietary intake is
poorly elucidated, there is an unclear relationship between human nutrition
and ecological food-web trophic-dynamics.
Food chains that directly or indirectly involve human consumers are
influenced by many factors. Some of the penultimate problems are
resolvable, but others present interactions which are beyond resolution
with current techniques and have little more than an empirical basis for
investigation. These questions include broad areas of human nutrition,
induction of pathological process by toxicants, and oncogenesis.
The suggested methods of approach include the monitoring of the
pathways of xenobiotics in "human food chains" by using the
trophic-dynamic/systems theoretic as the basis for the program of research
on food webs. This theory incorporates and integrates humans into the
energy flow in the ecosphere rather than isolating them.
The impact of human activities since the industrial revolution has
affected global climates, atmospheric and oceanic nutrient cycling, crop
nutrient cycling, and perhaps even the energetics that drive these cycles.
Natural resource exploitation and the attendant problems associated with
the rapid population increase and elevated living standards in developed
countries have greatly accentuated the introduction of hazardous levels of
xenobiotics into humans. The rates of introduction of compounds and the
enormous scale of industrial activity has exacerbated the "ecological"
problems and greatly reduced the time available to react to these insults.
The environmental legislation which has been enacted in recent years is a
response to the threat posed by gross chemical introductions.
The basis for a food-chain research program presently exists. A
number of individual projects are already underway in the EPA Office of
Research and Development. The organization of these ongoing projects may
not be strictly directed toward food chains. However, with careful
coordination and planning by scientists, the effort may result in
significant accomplishments in relating and applying ecological theory to
problems regarding human diets.
Background
Pollutants can reach man in a variety of ways—in the air he breathes,
the food and water he ingests, through his skin, from his surroundings
(including air, water, and solids), and from a combination of routes.
There can be continuous exposure, intermittent but repeated exposure, or
sporadic exposure; concentrations may remain constant or vary greatly;
351
-------
exposures may result from direct, intentional, or unavoidable applications
(i.e., food, drugs, water); and exposures may be effected through complex
environmental pathways.
A thorough understanding of the pathways of a contaminant in the
environment affords a reasonable basis for estimating human exposure.
Understanding of what happens to a chemical in the environment provides the
only rational basis to prevent or minimize human exposure through food
chains. Prudence dictates that it is much more desirable to limit the
level of exposure to prevent ill effects than to attempt to reduce exposure
to an acceptable level of chemicals after an effect has occured. For
example, if certain physical conditions enhance the growth of toxin-
producing fungi, perhaps such conditions can be avoided. It is known that
cotton grown in irrigated desert areas often habors extensive growths of
Aspergillus flavus, which produces aflatoxin in the cotton seed. (Cotton
seed meal is a component of human food and animal feeds.) Another example
is the fallout of toxic elements (metals) in flyash from power plants and
from emissions of smelters of refineries, which may result in crop uptake
of excessively high levels of undersirable trace substances. Attention to
power plant placement or control of crops planted in the vicinity of such
facilities could eliminate or reduce human exposure to such material. The
knowledge of the transport and alteration of a chemical can be used to
avoid or prevent exposure, i.e., the sequestration of mercury by sediments
and improved methods and processes in handling and using such organics as
pentachlorophenol.
Chemicals may be transported and transformed in complex ways (Figure 1)
Once released (volatilized) into the environment, they may be transported
in currents of air or water, or in association with solid particles. They
may be transformed into hazardous compounds by chemical or biochemical
reactions, diluted by diffusion, or concentrated by physical or biological
processes. Biochemical transformation in the cell/organ system may produce
oncogens from moderately toxic precursors. Human exposure to chemicals
takes place not only at point of use or discharge of the chemical but also
at points distant in the space and time horizon.
To assess the total magnitude of human exposure, we must trace the
movement as well as the transformation of a chemical from the point of its
release in the environment to the site where it may be degraded as a less
harmful substance.
Figure 2 shows pathways by which chemicals reach man. In this simple
scheme, the environment is represented by "boxes" or compartments through
which chemicals move at various rates. Arrows between the boxes denote
transfer between compartments. To estimate the concentration of a chemical
in a compartment, we must know not only the rate of release of the chemical
(a function of its use patterns), but also its retention time within the
compartment. The retention time is determined by the rates of transfer to
352
-------
Biotic
Components
Abiotic
Components
D
Return to Abiotic
Components
D = degradation
Figure 1. Model of chemical transport and degradation in biosphere, (from
Robinson, 1973).
Biomass
Biomagnification
Trophic level
111
1OOOppm
10
II
i^-v—— __
100 ppm
-IflJL
1 ppm
Figure 3. Biomagnification/bioconcentration in a food web (from De Santo,
1978).
353
-------
NATURAL TOXICANTS
Figure 2. General diagram of chemical flow in food web.
354
-------
to and from other compartments, by the rates of alteration of the chemical,
and by diffusion processes within the compartment. Thus, to understand the
behavior of chemicals in food chains, we should study the processes of
dispersion, the transfer between and within compartments, the chemical and
biological transformations in compartments, and bioaccumulation in the
biota of a compartment.
Many environmental processes result in dispersion, dilution, or
degradation of chemicals. A most important reconcentration process in
human food webs is "biconcentration," in which plants and animals may
accumulate certain chemicals to levels much higher than those in their
ambient environment (Figure 3). Other processes include the absorption of
chemicals onto airborne or waterborne particles and the concentration of
certain chemicals from liquid effluents into sewage sludge and subsequent
possible entry into human food crops.
Transport of chemicals takes place on various geographic scales. On
both regional and global scales, an important transport takes place in the
atmosphere. For some substances with low reactivity, passive transport and
diffusion are the factors of importance. The processes that remove
chemicals from the atmosphere are less well-known. Some substances diffuse
upward and are degraded by ultraviolet radiation. Others diffuse downward
to be adsorbed onto the surfaces of suspended particles; others are
dissolved in water droplets and returned to earth in rainfall. In some
cases, global atmospheric processes may include cycles of elimination and
reintroduction of a substance, as the balance and flow are determined by
the air and water or air and terrestrial interface reactions. In any case,
man is the ultimate food-chain receptor of such chemicals (Figure 1).
Thus, as in the atmosphere, the transport and retention time of
chemicals in water may be controlled by processes which take place at phase
boundaries and which, as yet, are poorly understood. On balance, the
direction of movement is toward the oceans, although processes, such as
chemical precipitation or adsorption on solids, may result in deposition in
"sinks" that intercept the movement entirely or "reservoirs" that delay the
movement. These sinks and reservoirs contribute to accumulations in the
local biota, in food organisms, and ultimately in humans.
Pollutants can be actively transported in biological systems. On a
regional basis, migrating birds or fish may carry small amounts of
pollutants for long distances. The amounts involved are for the most part
insignificant in terms of global or regional redistribution. On the other
hand, the quantities of pollutants carried in or on plants or animals used
as food may be of major concern in relation to human exposure.
In contrast to the atmosphere or water, the soil compartment serves
both as a reservoir which receives and disperses pollutants and as a
"chemical reactor" in which transformations take place (Figure 4).
355
-------
Movement between compartments
>
Degradation or transformation
of a chemical compound
V
Figure 4. Chemical transport model between compartments (from Robinson, 1973)
Substantial quantities of chemicals released into the environment
reach the soil either through direct application or transfer from air or
water. It should be recognized that there is a constant interchange
between the compartments of the environment. Consequently, a chemical
applied to soil may transfer in a reversible process to air or to water
through run-off and leaching. Nonetheless, the soil becomes an important
reservoir for many chemicals (Figure 2).
Three major interactions or processes are of concern in soil:
sorption, Teaching-diffusion, and alterations through chemical and
biochemical processes. Obviously these interactions affect the plants and
animals consumed as food. For example, it is reasonable to ask how rapid
and in what quantity a chemical may move from a landfill disposal site as a
result of precipitation percolating through the soil profile or of ground
water moving through the landfill. If leached in considerable quantities,
the chemicals may subsequently reach aquifers used as human water sources
or be carried into streams and bodies of water containing food-chain
biota. Adsorption plays a very important role in leaching and diffusion
behavior so that strongly adsorbed compounds are found to be poorly leached
by water. Substances, such as higher molecular weight halogenated organics,
i.e., polychlorinated biphenyls and DDT may show little movement even after
several years of exposure to percolating water.
Many organic chemicals become solids at ambient temperatures. The
kinetic motion of the molecules causes them to have a finite vapor pressure
356
-------
even at these temperatures. As temperatures increase, the transition from
solid to vapor with or without an intermediate liquid phase tends to result
in increased vapor pressure. The vapor pressure then is related in a
complex fashion to the rate of evaporation of the compound and to its
tendency to exchange across air/water or soil/air interfaces or pass
directly into the atmosphere. The chemical in the vapor state will quickly
establish an equilibrium state of adsorption on particles suspended in the
atmosphere. Photodecomposition may occur in either the sorbed or vapor
state. The contributions to food chains are indirect but nevertheless
significant via the evaporation-volatilization of hazardous substances.
The human habitat plays a role in the transport of polluting
chemicals. The structure of the outdoor environment of towns and cities
provides semienclosed spaces from which chemicals diffuse slowly. Paved
areas, drainage networks, and sewage systems constitute a pathway for
transporting chemicals out of urban areas. Sewage treatment plants
separate the dissolved and solid-phase components of the chemical mixture.
Disposal of the dissolved materials into waterways and discharge of the
sewage sludge on land may lead indirectly to significant human exposures
via food chains. These urban subsystems may be ultimately more important
in determining human exposure to chemicals than the rural environment.
The physical behavior of these substances in the environment must be
understood, as well as their fate and accumulation. Three major processes
are involved in breakdown and metabolism: photochemical, chemical, and
biologically mediated alterations. Photochemical breakdown induced by
absorption of light will induce alterations. For some substances, the
quantum efficiency is high and consequent breakdown is rapid. Chemical
breakdown, on the other hand, is dependent upon molecular structure. If a
chemical is susceptible to nucleophilic attack, oxidation, or hydroxylation,
alterations can occur fairly rapidly. Similarly, a chemical may undergo
hydrolysis, the rate of which will be dependent on pH, temperature, and the
presence of catalytic sites. For example, certain organophosphates in soil
are rapidly hydrolyzed following sorption (presumably the clay affords
catalysis for this reaction). On the other hand, halogenated hydrocarbons
tend to be more refractory toward purely chemical reactions and hence
persist for long periods in food-chain compartments.
Alterations from biologicaly mediated reactions are variable.
Although considerable research has been devoted to metabolic studies and
investigations of alterations occurring in the environment, complete
understanding for all but a few compounds is lacking. In describing the
behavior of organic compounds in the environment, observers have reported
that compounds partition into a particular phase. The partition
coefficient of an organic substance from water into a lipid solvent may be
related (in certain instances) to its biological activity. It has been
observed that the partition coefficient in octanol/water correlated very
closely with the propensity of an organic substance to accumualte in fatty
tissues of organisms. It has been demonstrated that the octanol/water
partition coefficient is related to the adsorbability of a compound by soil
357
-------
organic matter and certain clays. This can be of predictive value in
studying exposure of human food to hazardous chemicals.
An important example of partitioning is the phenomenon of bioconcen-
tration. Many aquatic plants and animals are able to concentrate persis-
tent chemicals, heavy metals and lipophilic organic compounds into their
tissues to levels many times those in the ambient water (Figure 2). In
extreme cases, such as the concentration of cadmium by shellfish, and DDE
and PCBs by fish, the concentration factors may be as high as one hundred
thousand or even one million times the ambient levels. For such chemicals,
consumption of contaminated fish or shellfish is often the principal route
of human exposure. Although many measurements of bioconcentration factors
have been made, there is still an incomplete understanding of the
biological mechanisms that determine the exact degree of concentration
achieved for various chemicals. However, it is now possible to estimate if
some chemicals will bioconcentrate based upon structure activity
relationships. Structure activity correlations can be used to predict
biodegredability under waste-treatment and environmental conditions and to
provide valuable insight into the assimilating capacity of the environment.
Techniques are well-established to predict the toxicity of organic
chemicals on structural properties such as lipid solubility and
electronegativity and an appropriate structure-activity correlation. With
respect to toxic effects to aquatic organisms, the major research need is
an adequate data base of toxicological information from which necessary
quantitative correlations can be established. The establishment of this
predictive capability will permit estimates of toxicity. Detailed
toxicological studies can be made on those chemicals with the greatest
potential for adverse reactions, including contaminants of the human
food-chain.
The problems associated with acid rain are related to food web
dynamics. This is a serious concern because of the increased heavy metal
(Pb, Cd, As, Se) uptake by food plants with altered soil characteristics.
Also, the changed pH of atmospheric moisture results in increased
solubility and consequently increased foliar absorption of heavy metals by
leafy vegetables and animal forages. This is a relevant example of a
disturbance in an ecological compartment felt throughout the bioshpere with
man as the witless target species.
Human exposure estimation is an area of principal concern in
attempting to resolve some of the food-chain dynamics. Human nutrition
data should be collected on a national as well as a regional basis.
Significant regional differences in food intake are apparent. When this
information becomes available and the quantities of ingested foods are
known, food-chain pathways of pollutants can be ascertained. In order to
assess the dietary intake of chemicals present in food, the amount of a
given substance present in different foods can be measured. In addition,
it should be determined if the source of pollutants is environmental or
industrial (food processing). By combining this information with data on
food consumption, estimates can be made. Factors to be taken into account
358
-------
in this approach include: variations in dietary patterns, quantities of
given foods consumed by different age or ethnic groups, and variations in
concentrations of a given chemical in different foods at different times.
A substance that occurs in more than one compartment of the human food
chain (Figure 2) may promote exposure through several routes varying in
form and quantity; hence, total exposure can only be estimated by combining
the separate sources and measuring them in different trophic levels.
Food-Chain Related Programs
The problems associated with the addition of chemicals are not easily
resolved because of the complexity of the systems in terms of the numbers
of interacting organisms and "level of organization" through which these
compounds and their metabolites pass. In related research, we must also
consider how and where these compounds bioaccumulate, whether or not the
various organisms are capable of depurating, and if the compounds are
altered or transformed to produce carcinogens, mutagens, and teratogens.
The techniques for recognition of carcinogens in human food organisms
and those organisms indirectly important in food webs are under development
by EPA researchers. At present, these techniques and their causal
association are being studied by a few investigators. This research has
been pursued by scientists at EPA, the Food and Drug Administration (FDA),
the National Cancer Institute (NCI), and the Smithsonian Institution. The
induction of carcinogenic processes and the description of inducing
mechanisms in food web organisms are important areas of investigation.
There is a limited program of biological tissue monitoring by EPA
scientists (mussel watch) for toxic substances. U.S. coastal areas range
from the heavily contaminated to almost pristine and are monitored for data
on the incidence of neoplasms and tissue degeneration in four species of
bivalve mollusks. This effort appears to be the only investigative program
of its type which deals with anomalies in organisms related to the human
food webs. However, low level funding has hampered its effectiveness.
Other EPA research projects underway in 1978 are listed below.
Atmospheric Research
Airborne contaminants from mobile sources in human food crops
(plants), animals forages and meat animals, dairy products (these
include heavy metals, hydrocarbons, oxides of nitrogen and sulfur,
ozone, and trace metals from emission controls).
Atmospheric source influence on ground water quality from NOx,
S02> 03, HCN, CO, trace metals: Mn, Ni, Ru, Ir, S02; heavy
metals: Pb, As CN, Se, Cd.
Effects of atmospheric contaminants on marine food organisms.
359
-------
Bloaccumulation and concentration, sediment uptake and release,
levels of contaminants in primary producers and consumers.
Effects of muHiroute exposure (ingestion) to man of Cd and Pb.
Atmospheric sources to human food chain from petrochemical
complexes; incidence of tumorogenic disorders in populations of
food-chain organisms.
Dendrochronology - relation of time of depositon to pollutants.
Exposure assessment of pollutants in oil and water exposure
compartments; predictive model of volume and distribution in
environmental compartments.
Response of critical receptor organisms (including man) from
atmospheric toxics.
Atmospheric chemical fluxes relative to fluxes in sediments.
Surveys of atmospheric hazard substances in the Great Lakes.
Fluxes of hazardous substances in Saginaw Bay from atmosphere.
Toxics uptake by phytoplankton in the Great Lakes.
Terrestrial ecosystem pathways and nutrient cycling effects from
airborne contaminants (pesticides, trace elements, metals, and
gaseous air pollutants).
Pollutant pathways in plants, soils, and animals.
Energy-Related
Determination of toxicity and bioaccumulation of polyaromatic
hydrocarbons from energy sources in aquatic and terrestrial
animals.
Effect of shale oil development by-products on aquatic
organisms in the food chain.
Effects of pollutants from coal gasification and liquefaction on
aquatic and terrestrial life (including man) through the food
chain.
Impacts of biocides on food webs through the processes of
bioaccumulation and interaction with sediments.
Effects of compounds on histopathology of marine organisms
(fish), and teratology of organisms from halocarbon exposure.
Effects of carcinogens from shale oil in the marine food web.
360
-------
Effects of SC>2 and other pollutants on grasslands.
Sewage wastes effects and toxic uptake by crop plants.
Fate and transport of energy-related pollutants in biological
systems.
Characterization of ground water geohydrology in relation to
oil-shale processing.
Food Crop Pathways
Models to predict behavior of agricultural chemicals and animals
wastes in major non-irrigated crop regions under various edaphic
and climatic conditions, i.e., bioaccumulation and
bioconcentration of chemicals in food organisms.
Trace element transport and organic biotransformation resulting
from passage through plant and soil systems.
Coal-fired power plant vicinity sampling and analysis of soils
and plants for Hg, Pb, Cd, and As.
Drinking Water Contamination
Effects of trihalomethanes/halorganics (as carcinogens, mutagens,
teratogens) in drinking water.
Alternate disinfection chemical by-products as carcinogens,
mutagens, teratogens.
Water/food supply availability of inorganics and metals: Hg, As, Ba,
Se, Cd and their influence on cardiovacular disease.
Effects of asbestiform fibers in drinking water.
Effects of geochemicals and industrial organics from water
supply.
Persistence and availability of pesticides (from soil to food
chain organisms).
Storage, depuration, and excretion of halogenated and
nonhalogenated pesticides in animals and man.
Aquatic Ecological Effects
Bioaccumulation of organics in primary producers and primary
consumers.
361
-------
Mass balance for toxics in the Great Lakes ecosystem;
bioaccumulation of hazardous substances in fish.
Hazardous substances and heavy metals uptake and release in
sediments and in benthic algal communities (marine and
freshwater).
Analysis and tracking of organic chemicals {Mirex, Kepone, PCPs,
and PCBs) in fish of the Great Lakes.
Accumulation of DDT and PCBs in food fish.
Pathway of asbestos to human water sources.
Land bioaccumualtion and alteration of communities from
wastewater and sludge application.
Distribution and source of PCBs in dredge spoils.
Sediment processes, fate, bioaccumulation, and toxicity of Kepone
in food-chain organisms and sediment processes in the James River
estuary.
Fate and effects of petroleum hydrocarbons, transuranics,
pesticides and heavy metals in organisms in the estuary.
Structure activity studies from which to make predictions on
bioaccumulation and toxicity of certain classes or organics.
Water Quality
Human health effects associated with treatment and disposal of
wastewater with reference to persistent organics (and
particularly PCBs) in food web; Cd, Pb, and trace metals
translocated from sludge to soil, to plants, to man; persistent
organics in human food web organisms from wastewater sources.
Development of and testing for persistence of hazardous
substances in water.
Bioaccumulation and health hazard test protocols for pollutants.
Literature search for documentation of water quality for
shellfish using chemical and biological data.
Ecological processes and effects of land application of municipal
wastewater and sludge and non-point runoff on plant and animal
communities including bioaccumulation, population dynamics, etc.,
in aquatic systems.
362
-------
Land application sewage sludge and pathways of contaminant
through soils, groundwater, surface water, plants, and animals to
humans as evidenced by food chains.
Food-chain effects of heavy metals on public health (epidemiology).
Monitoring
Development of monitoring methods for toxic industrial and
municipal wastes and hazard evaluation applicable to human
populations and aquatic food-chain organisms.
Fate and effects of organics in aquatic systems, particularly
carcinogens in estuarine and marine systems.
Routes and rates of pesticide movement through ecosystems to man.
Bioaccumulation and concentration of chemicals in estuarine food
web organisms.
Development and use of aquatic indicator species of carcinogens,
mutagens, and teratogens in food web.
Transport of substitute chemicals (pesticides) and degradation
products in model systems.
Hazardous chemicals and pesticides bioaccumulated in terrestrial,
estuarine/marine, and freshwater systems and organisms.
Substitute chemical (pesticides) interaction with other chemical
agents, pathogens, environmental conditions relative to
disposition in microcosms (rates and pathways of accumulation).
Other Federal Programs
There are other Federal Programs outside EPA that are relevant to food-
chain research. The Food and Drug Administration has research projects
related to the sources of various contaminants which directly enter human
food webs. These projects include the following:
PCBs in food and food packaging materials
Organochlorine pesticide residue analysis in fish
Herbicides and fungicides in the human diet
Total diet identification of pesticide residues
Dioxins as food contaminants
Mercury, cadmium, lead, and heavy metal contaminant in food
Chlorinated dibenzofurans as food contaminants
Sewage-deri ved chemicals
Food Contaminants
Pesticides and metals in food
363
-------
Radionuclides in food
Heavy metals in processed foods
Total diets in adults, infants, and toddlers
Carbamates in foods
Plant toxins, biogenic toxicants, and marine toxicants in foods.
Atmospheric-Food Chain Inputs—There have been assessments of the problems
associated with a portion of this general area. The National Academy of
Science (NAS) publication, The Tropospheric Transport of Pollutants and
Other Substances to the Oceans (1978), outlines the problems. A brief
synopsis follows:
I. The current data are insufficient to design a comprehensive and
sound monitoring program; therefore, research lines must be
followed until some limited set of quality data make the
definition of monitoring network feasible.
II. There is a great need for atmospheric and oceanic concentration
and composition data on such pollutant groups as:
(1) Metals, in particular, their concentration in surface slicks
and water layers, and a description of their continental
sources.
(2) Halogenated hydrocarbons, in particular, the high molecular
weight hydrocarbons, such as the chlorinated hydrocarbons.
It is important to measure concentration in phytoplankton
and zooplankton to determine biomagnification of these
compounds.
(3) Low molecular weight halocarbons and monohalomethanes. Their
source is probably oceanic. These can be divided into two
groups: short and long residence time.
Short residence time - halocarbons (6 months to 1 year).
Their source is probably from rivers and not the atmosphere.
Long residence time - halocarbons, such as the low molecular
weight CC14, CHCL3, (CHo) CCL3, would remain in
solution when dumped into rivers or oceans. Oceans act as a
reservoir for these compounds from atmospheric sources, for
example.
(4) High molecular weight halocarbons. Removal by attachment to
aerosol particles may be important for high molecular weight
chlorinated hydrocarbons. There is evidence that aerial
fallout on a 50 x 200 km nearshore area in the Southern
California Bight for DDT and PCBs exceeds runoff and
wastewater (sewage outfalls) inputs to a comparable area.
364
-------
CONCLUSIONS
EPA Environmental Research Laboratories have a unique capability to
advance the food-chain research. Their research objectives might emphasize
and elucidate the following areas:
(1) Environmental samples and tissues from human populations and
establish banks for monitoring.
(2) Establishing a bank for the tissues of regional terrestrial and
aquatic species to monitor the population fluctuations.
(3) The fate and transport of xenobiotics through compartments and
bioaccumulation and a transformation of xenobiotics by the
biota.
(4) The global atmospheric processes of elimination and
reintroduction of a substance with the balance and flow
determined by the air and water or air and terrestrial interface
reactions.
(5) A working model of the urban food web--the cycling and pathways
of industrial contamination of biospheric compartments.
(6) Data on food consumption and amount of chemicals present in food
to establish human exposure estimates.
(7) A comprehensive model and a scientific information access system
relative to food chain xenobiotics.
(8) An environmental forecasting and technology assessment of human
food-chain problems.
EPA laboratories should be given adequate funding and personnel to
meet these objectives. In such a comprehensive program, the laboratory
research scientist should be able to set priorities and determine the
validity and nature of such projects.
BIBLIOGRAPHY
Bresler, J.B., Ed. 1968. Environments of man. Addison-Wesley Publishing
Co., Reading, MA. 289 pp.
Colinvaux, Paul. 1973. Introduction to ecology. John Wiley and Sons, New
York. 621 pp.
De Santo, R.S. 1978. Concepts of applied ecology. Springer-Verlag, New
York. 310 pp.
Edwards, C.A. 1973. Environmental pollution by pesticides. Plenum Press,
New York. 542 pp.
365
-------
Gevantman, L.H., Ed. 1976. Program and abstracts. Symposium on
nonbiological transport and transformation of pollutants on land and
water. U.S. Department of Commerce, National Technical Information
Service. Springfield, VA. PB-257 347, 181 pp.
Hynes, H.B.N. 1970. The Ecology of running Water. University of Toronto
Press, Toronto, Canada. 555 pp.
McRae, A., and L. Whelchel, Eds. 1978. Toxic Substances; control
sourcebook. Aspen Systems Corp., Germantown, MD. 609 pp.
Muirhead-Thompson, R.C. 1971. Pesticides and freshwater fauna. Academic
Press, New York. 248 pp.
National Academy of Sciences (NAS). 1975a. Assessing potential ocean
pollutants. NAS, Washington, DC. 438 pp.
National Academy of Sciences. 1975b. Principles for evaluating chemicals
in the environment. NAS, Washington, DC. 434 pp.
National Academy of Sciences. 1978. The tropospheric transport of
pollutants and other substances to the oceans. NAS, Washington, DC.
484 pp.
Odum, E.P. 1971. Fundamentals of ecology. W.B. Saunders Co.,
Philadelphia. 574 pp.
Smith, R.L. 1966. Ecology and field biology. Harper and Row, New York.
686 pp.
Palmisano, J.F., and J.A. Estes. 1977. The environment of Foucjotka
Island, Alaska. M.L. Merritt and R.G. Fuller, Eds., ERDA, Oak Ridge,
TN. 567 pp.
Russell, C.S. 1975. Ecological Modeling in a resource framework: The
proceedings of a symposium. Resources for the Future, Washington, DC.
366
-------
THE CELLULAR FATE OF BENZO(A)PYRENE
by
Vesna Ivanovic and I. Bernard Weinstein
Division of Environmental Sciences and Institute of Cancer Research
Columbia University College of Physicians and Surgeons
New York, New York 10032
ABSTRACT
Recent work in mammalian systems emphasizes covalent
binding of chemical carcinogens to DNA as the initiating event
in the process of cell transformation. In our paper cellular
and molecular aspects of carcinogen action are summarized partic-
ularly with respect to covalent binding of derivatives of
benzo(a)pyrene (BaP) to DNA and RNA in hamster embryo cell
cultures. After addition of BaP, the binding to RNA proceeds
rapidly and follows a linear time course for at least 48 hr,
presumably because of a lack of an RNA repair mechanism. In
contrast, after approximately 18 hr of incubation with
benzo(a)pyrene, the extent of binding to DNA reaches a plateau,
reflecting equilibrium between de novo binding and DNA repair
processes. A detailed analysis provides evidence that both
stereoisomers of benzo(a)pyrene 7,8-dihydrodiol 9,10-oxide
(BaPDE) are involved in covalent binding to hamster embryo
cellular nucleic acids.
The relevance of these studies in mammalian systems to the
marine environment and marine organisms is discussed. The
possible consequences of bioaccumulation of polycyclic aromatic
hydrocarbons in marine organisms and the metabolic processes and
neoplasia induction in these organisms are explored in terms of
potential hazards to humans.
INTRODUCTION
During this meeting, it has become apparent that oceans are the
recipients of innumerable foreign organic compounds that occur as
environmental pollutants. Consequently, it is not surprising that this
contamination affects marine biota. Deposition of these pollutants in
tissues, bioaccumulation, metabolism, degradation or depuration, detoxi-
fication, and neoplasia induction are well-documented in marine organisms.
Of additional concern is the possible impact of this type of marine
367
-------
pollution on human cancer induction. In the complex marine food web,
contaminated organisms may be a potential food source for other organisms
at higher trophic levels and thus eventually appear in human food sources.
These considerations have major implications in terms of both marine
ecology and human health.
Although many studies have focused on the mixed function oxidase (MFO)
enzyme system in marine organisms, there has been considerably less
attention to the cellular and molecular aspects of chemical carcinogenesis
in marine species. Historically, basic discoveries in this area have
primarily involved nonmarine organisms. In the present paper, an attempt
is made to evaluate the status of our understanding of the cellular fate of
certain carcinogens in mammalian systems. Emphasis will be placed on the
compound BaP and on studies analyzing its covalent binding to mammalian
cell DNA.
Celluar Events in Mammalian Systems
Figure 1 summarizes the cellular fate of BaP or related carcinogens in
mammalian systems. The first critical step in the encounter between a cell
and a potential carcinogen is metabolism of the compound. This subject has
been discussed in detail, with respect to fish, during this meeting. Most
of the metabolites that are formed are detoxification products. During
this process, however, highly reactive activated intermediates are formed
which, unless further metabolized, can act as "proximate" and "ultimate"
carcinogens. Within this context, carcinogenesis can be thought of as an
error in drug detoxification. Ultimate carcinogens are highly reactive
electrophiles (Miller, 1970) that can bind covalently to nucleophilic
residues in cellular macromolecules—DNA, RNA and proteins (Brookes and
Lowley, 1964). At present, it is not known with certainty which target is
critical in terms of the process of cell transformation. Current studies
favor DNA as the critical target but its role in transformation is not
understood. A number of laboratories have shown that when DNA is modified
with chemical carcinogens, there is impairment of not only DNA replication
but also RNA transcription (Weinstein, 1976). It is possible, therefore,
that serious distortions in gene transcription and the control of gene
expression occur as a result of the modification of cellular DNA.
How do organisms protect themselves against this attack of foreign
materials on DNA and thereby avoid genetic damage? Nature has evolved a
highly efficient and complex enzyme mechanism, called DNA repair, which can
recognize damaged regions of DNA, excise and then repair them (Friedberg et
al_., 1977). If DNA repair systems operate efficiently and with high
fidelity (error free repair), the host has solved the problem. If, in
contrast, the repair is incomplete, or "error-prone," then mistakes are
likely to occur in the daughter strand when the DNA replicates, and
mutations will occur in the daughter cell. The damage to DNA by
carcinogens and its consequent repair provide a basis for some of the
short-term tests for carcinogens that assay for mutagenicity or induction
of DNA repair.
368
-------
Detoxification
Products
Procarcinogen
Metabolic Activation
Proximate Carcinogen
—Ultimate Carcinogen (Electrophile)
I
Covalent Binding (DNA, RNA, Proteins)
No Repair
Error-Prone Repair
T
Error-Free Repair
Survival
CYTOTOXICITY
MUTAGENESIS
CARCINOGENESIS?
Figure 1. Cellular fate of carcinogens in mammalian systems.
12 I
Benzo[a]pyrene
Smooth endoplasmic
reticulum
(Golgi apparatus)
Mitochondrion
Rough
endoplasmic
reticulum
Ribosomes
Lysosome
Membrane
Figure 2. The encounter between an environmental carcinogen and a target
enkaryotic cell.
369
-------
Our laboratory has focused on the molecular details of the covalent
binding of chemical carcinogens to DNA that appear to reflect the initating
events in cell transformation. The study of these covalent adducts rather
than simple metabolites has several advantages since cellular DNA acts as a
trapping agent for ultimate carcinogenic metabolite(s), and the extent of
binding may provide an index of the potential potency of the compound in
question.
Benzo(a)pyrene in Rodent Cell Cultures
The most extensively studied polycyclic aromatic hydrocarbon (PAH)
carcinogen is BaP. As presented in Figure 2, this carcinogen enters the
cell most probably by passive diffusion through the lipid layer of the
membrane (Burnette and Katz, 1975). It reaches the endoplasmic reticulum
and nucleus and induces the so-called aryl hydrocarbon hydroxylase (AHH) or
MFO enzyme system. This monooxygenase cytochrome P-450 system oxidizes BaP
at a variety of positions to form more than 35 metabolites (Yang and
Gelboin, 1977a) (Figure 3). Most of metabolites are detoxification
products, while some bind covalently to DNA, RNA, and proteins.
S~~\ l_-».[2,3-EPOXIDEJ
[9.10-EPOXIDE]
\ OH
[7.8-EPOX10E]
6.12-QUINONE
CONJUGATES BOUND MACROMOLECULES
DNA
RNA
PROTEIN
PHENOL
-»
CONJUGATES
CONJUGATES
-^ ?
OH
OH
4.5-DIHYDRODIOL
Figure 3. Benzo(a)pyrene metabolism (Selkirk et a]_., 1974),
370
-------
In order to study this covalent binding, we exposed various confluent
rodent cell cultures to 1 yg/mfc of (14C)-labeled BaP. After 24-hr
incubation, cells were harvested and nucleic acids extracted and rigorously
purified to exclude contaminants and noncovalently bound material. The
amount of carcinogen covalently bound was determined by the radioactivity
of the purified nucleic acids (Table 1). Table 1 indicates that the K-22
epithelial rat liver cells bind BaP to DNA at a rather low level, i.e., one
residue of BaP per 150,000 nucleotides. We have found considerable
variations between cell types. Primary hamster embryo cells were chosen
for more detailed studies of BaP binding, since the modification was
considerably higher (1/45,000), and hamster embryo cells are readily
transformed in culture by BaP (Berwald and lachs, 1965; DiPaolo and
Donovan, 1967). These studies have been published (Grover et_ a]_., 1974)
and results are summarized below.
TABLE 1. THE EXTENT OF IN VIVO BaP BINDING TO CELLULAR NUCLEIC ACIDS
Cell Culture Nucleic Acid Extent
K-22 epithelial
rat liver cells DNA+RNA 1/150,000
Hamster embryo (confluent RNA 1/30,000
cells primary DNA 1/45,000
culture)
The time courses of covalent binding of BaP to RNA and DNA in hamster
embryo cell cultures are compared in Figure 4. The binding to RNA
proceeded rapidly and followed an approximately linear time course for at
least 48 hr. This indicates that metabolites are available for covalent
binding during this period. In contrast, after approximately 18 hr
incubation with BaP, DNA binding reached a plateau. This plateau is due to
an equilibrium between de np_vp_ binding and DNA repair processes.
These studies with radioactively labeled carcinogen-bound nucleic
acids, of course, do not provide information on the nature of the
intermediates and nucleoside adducts formed. In order to understand the
relation between the formation of covalently bound adducts to DNA and
induction of cancer, it is necessary to identify the ultimate metabolite
and to determine the base specificity of nucleotides involved in binding.
371
-------
100
Ld
o
o
LJ
_J
u
LJ
_l
O
O
m
o.
m
UJ
_J
O
50
0
20
INCUBATION
40 60
TIME (HOURS)
Figure 4. Time course of bindinn of ( C)BaP to DMA and RNA in ronfluent
HEC cultures. HEC DHA, Q Q ; HEC RNA,M [^. The
results represent mean values obtained from rive independent
studies.
372
-------
The detection by standard analytical techniques of the reactive metabolites
j_n vivo has been hindered for several reasons: (1) the large number of
metabolites produced; (2) the very low levels of reaction with nucleic
acids (1/10,000 - 1/1000,000), and (3) the structural complexity of the
nucleoside adducts.
Identification of an Ultimate Metabolite of Benzo(a)pyrene
Some of the multiple products of BaP metabolism are illustrated in
Figure 3. The AHH enzyme system oxidizes BaP to a variety of quinones,
phenols, dihydrodials and epoxides (Selkirk £t aU, 1974, 1976). Boyland
(1950) was the first to propose that arene oxide derivatives of
carcinogenic PAH molecules are the reactive intermediates responsible for
the in vivo binding of the parent hydrocarbon to nucleic acids. The
properties of synthetic K-region epoxides lent some support to this
hypothesis. They are alkylating agents that react covalently with nucleic
acids (Grover et_ a]_., 1972): they are mutagenic in several systems and can
induce malignant transformation of rodent cells in culture (Heidelberger,
1974; Marquardt et^ ^1_., 1974). Studies during the late 1960s until the
mid-1970s tended to favor, therefore, the K-region of BaP (BaP-4,5-oxide,
see Figure 4) as the important reactive metabolite of this hydrocarbon
(Grover et a±., 1972; Heidelberger, 1974).
The first technique successfully utilized for identification of the
covalently bound form of BaP involved enzymatic degradation, to the
deoxynucleoside level, of the modified DNA obtained from cells exposed to
radioactive BaP. These products were co-chromatographed on a Sephadex
LH-20 column, which separates materials on the basis of hydrophobicity,
with UV markers of DNA-bound model compounds synthesized in vitro. Using
this approach, Baird (1975) established that the digest of DNA reacted jm
vitro with BP-4,5-oxide did not co-chromatograph with in vivo products
(Jeffrey et^al_., 1976). At the same time, it was demonstrated that when
BaP and a series of BaP metabolites were added to a microsomal system in
the presence of DNA, BaP 7,8-dihydrodiol (see Figure 3) was the most active
compound in terms of covalent binding (Borgen jrt al., 1973). The latter
compound bound to DNA to a tenfold greater extentThan BaP, suggesting that
it is a proximate carcinogen which is converted by the microsomes to an
ultimate carcinogen. One year later, Sims and co-workers (1974) extended
this result by providing evidence for a two-stage metabolic activation of
BaP. This initially involves formation of BaP 7,8-dihydrodiol, which then
is further metabolized on the 9,10 ring positions to give BaP
7,8-dihydrodiol 9,10-oxide (BaPDE).
In addition to column chromatography of enzymatic digests,
fluorescence spectroscopy is an extremely useful technique for detecting
and identifying BaP adducts in nucleic acids (Ivanovic ejt al_., 1976). We
have used this procedure at liquid nitrogen temperature which increases the
sensitivity about tenfold (Ivanovic et ^1_., 1976). Fluorescence
measurements were utilized to obtain information regarding the structure of
373
-------
the DNA- and RNA-bound chromophore formed in HEC at different time points.
As illustrated in Figure 5, the fluorescence emission spectra of DNA and
RNA. samples obtained from HEC following an 18-hr incubation with the parent
hydrocarbon closely resembled those obtained with DNA or RNA reacted with
BaPDE in vitro. The fluorescence spectra of the 24- and 42-hr DNA and RNA
samples (Figure 5) and the 48-hr sample (not shown here) were also
qualitatively similar to the 18-hr samples. These studies provided
fluorescence spectral evidence that BaPDE is the ultimate metabolite of BaP
responsible for covalent binding to DMA and RNA.
Two stereoisomers of BaPDE have been systhesized (Figure 6). In
isomer I, sometimes called anti, the 7-hydroxyl and 9,10-oxide groups are
on opposite sides of the plane of the ring system. In isomer II, also
called syn, the 7-hydroxyl and 9,10-oxide groups are on the same side.
Depending on the location of the 7-hydroxyl group (whether above or below
the plane), we can further distinguish two enantiomers for each
stereoisomer. As presented in Figure 7, in the 7R enantiomer of BaPDE I
the 7-hydroxyl group is above the plane, whereas in its enantiomeric pair,
7S BaPDE I, the 7-hydroxyl group is below the plane. Similarly, there are
two enantiomers for isomer II. In 7R BaPDE II, the 7-hydroxyl is above,
and in 7S BaPDE II it is below the plane of the ring system. An additional
complexity is introduced by the nature of the 9,10-oxide ring opening,
which can be cis or trans, in the reaction with base residues in nucleic
acids.
Base Specificity of Benzo(a)pyrene-Nucleic Acid Binding In Vivo
Fluorescence spectroscopy does not distinguish between either the
different BaPDE isomers or different nucleosides involved in nucleic acid
binding. To obtain information on these compounds, it was necessary to use
high pressure liquid chromatography (HPLC) analysis of the nucleoside
adducts.
Benzo(a)pyrene-RNA Adducts in Hamster Embryo Cells—
RNA samples obtained from HEC cultures exposed to (14C)BaP for
either 18 or 42 hr (see Figure 3) were hydrolyzed with KOH (Ivanovic et
al., 1978) and BaP modified nucleosides were analyzed utilizing various
BaPDE nucleoside adducts synthesized in vitro as markers.. Figure 8A
illustrates the elution positions of several of these markers, whose
structures and stereochemistry have been previously elucidated by NMR, mass
spectroscopy, and circular dichroism (Moore et a]_., 1977; Jeffrey et al.,
1976; 1977; Weinstein et al_., 1976). All of the peaks are guanosine
adducts resulting from addition of the 2-amino group of guanosine to the 10
position of BaPDE. The first peak in Figure 8A, GI-1, results from trans
opening of 7S BaPDE I, whereas the last peak, GI-3, represents its
enantiomeric pair derived from 7R BaPDE I. Peak 61-2 (Figure 8A) is
analogous to GI-3 but is formed by cis addition. GII-1 marker in Figure 8A
is a result of trans opening of 7S BaPDE II (Moore et .§]_., 1977; Jeffrey et
al., 1977). Figures 8B and 8C represent RNA digests obtained from cells
incubated for either 18 or 42 hr with BaP. The 18-hr time point HPLC
374
-------
DNA
RNA
100
50
o 0
£ 100
00
50
UJ
O
z o
LU
00 100
UJ
CC
O
50
I8hr
24 hr
42 hr
18 hr
24 hr
42 hr
350 400 450 40O 450
WAVELENGTH (nm)
500
Figure 5. Low temperature fluorescence emission spectra of DNA and RNA
obtained from HEC cultures exposed to (14C)BaP for 18, 24,
or 42 hr, with comparisons to in vitro modified DNA. Jji vivo
modified nucleic acid samples,~T ); control NA, (.....); DNA
of RNA modified in vitro with BaPDE ( ).
375
-------
HO' 8
HO'
I E
Figure 6. Stereochemical structures of two isomers of BaPDE.
0,
HO
OH
7S BaPDE
HO
OH
7R BaPDE
7R BaPDE
0
JCC
7S BaPDE
Figure 7. Structures of enantiomers of BaPDE I. (+)-7B,8a-dihydroxy-9a,
10a-epoxy-7,8,9,10-tetrahydrobenzo(ajpyrene, and BaPDE II,
(j:)-73,8a-dihydroxy-98,10B-epoxy-7,8,9,10-tetrahydrobenzo(a)-
pyrene. 7R BaPDE I, the 7R,8S,9R,10R enantiomer of BaPDE 1 [or
(+) BaPDE I, derived from (-) BaP 7,8-dihydrodiol]; 7S BaPDE I,
7S,8R,9S,10S enantiomer of BaPDE I [or (-) BaPDE I, derived from
(+) BaP 7,8-dihydrodiol]; 7S BaPDE II, 7S,8R,9R,10R enantiomer
of BaPDE II [or (+) BaPDE II derived from (+) BaP
7,8-dihydrodiol].
376
-------
•••
.;.
a
100
7S TRANS 7R TRANS
GI-3
B
20 40 60
RETENTION TIME (mm)
Figure 8. HPLC profiles of RNA adducts formed in confluent HEC cultures
incubated with ( C)BaP.
A. In vitro markers: Elution positions of guanosine adducts
formed by in vitro reactions with BaPDE I ("GI-1,2,3") with
designation of the enantiomeric form and character of 9,10-oxide
ring opening and of a guanosine adduct formed with BaPDE II
("GII-I"). For a further description of these markers, see
text.
B. RNA isolated from HEC after 18 hr exposure of the culture
to (^C)BaP.
C. RNA isolated from HEC after 42 hr exposure of the culture
to (1JC)BaP.
377
-------
profile (Figure 8B) reveals three major radioactive peaks designated 1-3.
Comparison with in vitro markers (Figure 8A) indicated that HEC RNA peak 1
co-chromatographed with the cis 7R BaPDE I adduct. HEC RNA peak 2 elutes
with the trans 7S BaPDE II guanosine marker and peak 3 with trans 7R
BaPDE I. Similar BaP-bound RNA adducts have been detected in human and
bovine bronchial explants (Jeffrey ^t a]_., 1977) and mouse skin RNA (Moore
et^ jjK, 1977). The 42 hr time point profile (Figure 8C) is qualitatively
similar to the 18 hr profile with the exception of increased prominence of
certain minor peaks. Their retention times correspond to previously
described BaPDE-cytidine and BaPDE-adenosine adducts (Jennette et al.,
1977).
Benzo(a)pyrene-DNA Adducts in Hamster Embryo Cells--
Figure 9B presents the BaP-modified deoxynucleosides obtained after
enzymatic digestion of HEC DNA obtained from cells incubated with BaP
( C) for 21 hr. The HPLC profile revealed 4 major radioactive peaks,
designated HEC DNA-1-4, and a few minor products. Comparison of these
products with BaPDE marker adducts synthesized in vitro (Figure 9A) led to
the following assignments. Three peaks (2-4, Figure 9B) are
BaPDE-deoxyguanosine adducts resulting from addition of the two ami no group
of guanine to the 10 position of BaPDE. HEC DNA peak 2 (Figure 9B)
coincides in its elution position with the deoxyguanosine-trans 7S BaPDE I
(Figure 9A) and peak 3 with its enantiomeric pair from trans 7R BaPDE I.
The latter product was detected as the single BaP product in human and
bovine bronchial explant DNA (Jeffrey et jjl_., 1977). HEC DNA peak 4
corresponds to a deoxyguanosine-BaPDE II product. HEC DNA I in Figure 9B
elutes in the region of deoxycytidine-BaPDE adducts (Jennette et jil_., 1977)
and peaks 5-7 cochromatograph with multiple deoxyadenosine-BaPDE adducts
(dAI-1-4 in Figure 9A).
Figure 9C illustrates the effect of "post treatment incubation" on
HPLC profiles. In this type of experiment cells were exposed to
(* C)BaP for 21 hr. Following this, the radioactive medium was removed,
the cell monolayer rinsed twice with warmed medium and the cells were then
incubated in BaP-free medium for an additional 24 hr. The removal of
carcinogen from the medium was associated with a 40% reduction (Ivanovic
et _al_., 1978) in the amount of BaP ( C) bound to DNA when compared to
the plateau value (see Figure 4) seen when carcinogen was not removed.
This indicates that a DNA repair system that excises the BaP residues from
the DNA is active in hamster embryo cells. The HPLC profile (Figure 9C) of
a DNA digest of the latter sample shows that, in addition to reduction in
the total amount of BaP adducts, there is also a change in the relative
contributions of individual peaks to the total profile. Although the
changes are complex, perhaps the most striking one is a decrease in the
relative abundance of peak 3 and an increase of peak 2. This suggests that
the peak 3 adduct is excised at a more rapid rate than peak 2. Other peaks
may also undergo differential rates of excision, but this requires further
study.
378
-------
RETENTION TIME (mm 1
Figure 9.
HPLC profiles of DNA adducts formed in confluent HEC cultures
incubated with (14C)BaP.
A. In vitro markers: Elution position of deoxyguanosine and
deoxyadenosine adducts formed by in vitro reactions with BaPDE I
("dGI-1,2" with designation of the enantiomeric form and
character of 9,10-oxide ring opening; and "dA-1,2,3,4") and of
a deoxyguanosine adduct formed with BaPDE II ("dGII-I"). For a
further description of these adducts, see text.
B. DNA isolated from HEC after 21 hr exposure of the culture
to (14C)BaP.
C. DNA isolated from a culture of HEC exposed to (14C)BaP
for 21 hr and then incubated for additional 24 hr in the absence
of BaP ("Post-treatment Incubation").
379
-------
It has been reported that hamster embryo cultures are readily
transformed by BaP (Berwald and Sachs, 1965; DePaolo and Donovon, 1967) and
even better by BaPDE (Mager et !]_., 1977). On the other hand, at the
present time it has not been established that the exposure of normal human
cell cultures to BaP results in reproducible transformation. A comparison
of BaP adducts in rodent and human cell cultures is presented in Table 2.
TABLE 2. COMPARISON OF BENZO(A)PYRENE-RNA AND -DMA ADDUCTS FORMED IN
CULTURE
Hamster Embryo Cells
(Confluent, Primary Cultures)
(Ivanovic et al., 1978)
Human and Bovine
Bronchial Segments
(Jeffrey et al_., 1977)
RNA
7R BaPDE I - (Cis) - G
7S BaPDE II - (Trans)-G
7R BaPDE I - (Trans)-G
7R BaPDE I - (Cis) - G
BaPDe I - C
7S BaPDE II-(Trans)-G
7R BaPDE I -(Trans)-G
DNA
BaPDE - dC?
7S BaPDE I-(Trans)-dG
7R BaPDE I-(Trans)-dG
BaPDE II-dG
7R BaPDE I -(Trans)-dG
BaPDE - dA
RNA adducts in hamster embryo cells have a HPLC profile which is very
similar to that of human and bovine (Jeffrey et. ^1_., 1977), as well as
mouse skin RNA (Moore et al_., 1977). On the other hand, the multiple BaP
products in DNA of hamster embryo culture are in contrast to the much
simpler profile in DNA of human and bovine bronchial explants in which the
predominant product is the 7R BaPDE I-deoxyguanosine adduct
(Jeffrey et al_., 1977).
380
-------
Additional variations in profiles of BaPDE-DNA adducts have also been
seen in the mouse fibroblast 10 T ^cell line (Brown ejt a\_., 1979).
DISCUSSION
It appears, therefore, that although in all cases studied so far BaPDE
is the major BaP metabolite responsible for covalent binding to RNA and DNA
in mammalian cells, there are major differences between species and cell
types in terms of the relative abundance of the different types of
deoxynucleoside adducts that are formed. The relative importance of the
individual adducts with respect to the carcinogenic process is not known at
the present time.
To our knowledge, detailed studies similar to those described above
have not been performed with marine organisms. There is some evidence
that DNA-BaP adducts produced by fish liver microsomes, are primarily
derived from the BaPDE metabolite of BaP (Ahokas.J.T., unpublished
studies). Further studies on BaP metabolism in diverse marine organisms is
of considerable interest, particularly in relation to cancer induction in
these organisms. Such studies may provide information about the role of
various BaP metabolites in the induction of neoplasia in fish and other
marine organisms. Since, however, the pathways involved in BaP metabolism
are used for detoxificatio purposes, the fact that marine organisms
metabolize BaP and related compounds to proximate and ultimate carcinogens
does not mean that these reactions occurring in marine organisms represent
an immediate danger to humans. The diol epoxides have very short
half-lives in aqueous solution. For example, Yagi and co-workers (1977)
have reported that in tissue culture medium BaPDE I has a half-life of 8
min, and BaPDE II has a half-life of only 0.5 min. It is unlikely,
therefore, that humans would suffer hazardous exposure to ultimate
carcinogens formed by marine organisms. Once such ultimate metabolites are
bound to the DNA and other cellular macromolecules of marine organisms, it
is also highly unlikely that subsequent ingestion of these BaP-modified
macromolecules by humans would be hazardous. In contrast, bioaccumulation
of the parent compound in tissues of marine organisms could conceivably
represent a hazard to humans when such material is ingested.
The ultimate carcinogenic metabolites produced by marine organisms,
however, could lead to mutations, developmental defects, neoplasia, or
lethal effects within these species. The possible extinction of certain
marine organisms that suffer from these toxic effects could disturb marine
ecosystems and thus indirectly be harmful to the human environment. At the
same time, it may be useful to monitor the occurrence of tumors in marine
organisms to provide an index of the presence of potential human
carcinogens in the marine environment. The validity of this surveillance
system will depend, in part, on knowing the comparative similarities and
differences in metabolism of BaP and other potential carcinogens between
humans and the marine organisms being studied. This provides an additional
reason for encouraging more detailed studies on BaP metabolism and binding
to cellular macromolecules in marine organisms.
381
-------
ACKNOWLEDGMENTS
These studies were supported by Grant CA-21111-01 and contract
N01-CP-2-3234 awarded by the National Cancer Institute, Department of
Health, Education and Welfare, and Grant EPA-R-805482010 awarded by the
U.S. Environmental Protection Agency.
REFERENCES
Baird, W.M., R.G. Harvey, and P. Brookes. 1975. Comparison of the
cellular DNA-bound products of benzo(a)pyrene with the products formed
by the reaction of benzo(a)pyrene -4,5-oxide with DNA.
Cancer Res. 35:54.
Berwald, Y., and L. Sachs. 1965. In vitro transformation of normal cells
to tumor cells by carcinogenic hydrocarbons. J. Natl. Cancer Inst.
35:641.
Borgen, A., H. Darvey, N. Castagnoli, T.T. Crocker, R.E. Rasmussen, and
I.Y. Wang. 1973. Metabolic conversion of benzo(a)pyrene by Syrian
hamster liver microsomes and binding of metabolites to
deoxyribonucleic acid. J. Med Chem. 16:502.
Boyland, E. 1950. Biological significance of metabolism of polycyclic
compounds. Biochem. Soc. Symp. 5:40.
Brookes, P., and P.O. Lowley. 1964. Evidence for the binding of
polynuclear aromatic hydrocarbons to the nucleic acid of mouse skin:
relation between carcinogenic power of hydrocarbons and their binding
to deoxyribonucleic acid. Nature 202:781.
Brown, H.S., A.M. Jeffrey, and I.B. Weinstein. 1979. Formation of DNA
adducts in 10T mouse embryo fibroblasts incubated with benzo(a)pyrene
or dihydrodiol oxide derivatives. Cancer Res. 39:1673.
Brunette, D.M., and M. Katz. 1975. Interactions of benzo(a)pyrene with
cell membranes. Uptake into Chinese hamster ovary (CHO) cells and
fluorescence studies with isolated membranes. Chem. Biol. Interact.
11:1-14.
DiPaolo, J.A., P.J. Donovan. 1967. Properties of Syrian hamster cells
transformed in the presence of carcinogenic hydrocarbons. Exp. Cell
Res. 48:361.
Friedberg, E.G., K.H. Cook, T. Duncan, and K. Mortelmans. 1977. DNA
repair of enzymes in mammalian cells. In: Photochemical and
photobiological reviews, Vol. II. K.C. Smith, Ed., Plenum Press, New
York. pp. 263-322.
382
-------
Grover, P.L., A. Hewer, and P. Sims. 1972. Formation of K-region epoxides
as microsomal metabolites of pyrene and benzo(a)pyrene. Biochem.
Pharmacol. 21:2713.
Heidelberger, C. 1974. Cell culture studies on the mechanisms of
hydrocarbons on carcinogensis. In: Chemical carcinogenesis:
selected papers, Part B. P.O.P. Ts'o and J.A. DiPaolo, Eds., Marcel
Dekker Inc., New York, pp. 457-462.
Ivanovic, V., N.E. Geacintov, and I.B. Weinstein. 1976. Cellular binding
of benzo(a)pyrene to DNA characterized by low temperature
fluorescence. Biochem. Biophys. Res. Commun. 70:1172.
Ivanovic, V., N.E. Geacintov, H. Yamasaki, and I.B. Weinstein. 1978. DNA
and RNA adducts formed in hamster embryo cell cultures exposed to
benzo(a)pyrene. Biochemistry 17:1597.
Jeffrey, A.M., I.B. Weinstein, K.W. Jennette, 'K. Grzeskowiak, K. Nakanishi,
R.G. Harvey, H. Autrup, and C. Harris. 1977. Structures of
benzo(a)pyrene nucleic acid adducts formed in human and bovine
bronchial explants. Nature 269:348.
Jeffrey, A.M., K.W. Jennette, S.H. Blobstein, I.B. Weinstein, F.A. Beland,
R.G. Harvey, H. Kasai, I. Miura, and K. Nakanishi. 1976.
Benzo(a)pyrene-nucleic acid derivative found in vivo: structure of a
benzo(a)pyrenetetrahydrodiol epoxide-guanosine adduct. J. Am. Chem.
Soc. 98:5714.
Jennette, K.W., A.M. Jeffrey, S.H. Blobstein, F. Beland, R.G. Harvey, and
I.B. Weinstein. 1977. Nucleoside adducts from the in vitro reaction
of benzo(a)pyrene-7,8-dihydrodiol 9,10-oxide or benzoTa)pyrene
4,5-oxide with nucleic acid. Biochemistry 16:932.
Mager, R., E. Huberman, S.K. Yang, H.V. Gelboin, and L. Sachs. 1977.
Transformation of normal hamster cells by benzo(a)pyrene diol-epoxide.
Int. J. Cancer 19:814.
Marquardt, H., I.E. Sorgergren, P. Sims, and P.L. Grover. 1974. Malignant
transformation in vitro of mouse fibroblasts by 7,12-dimethylbenz(a)-
anthracene and 7-hydroxy-methylbenz(a)anthracene and by their K-region
derivatives. Int. J. Cancer. 13:304.
Miller, J.A. 1970. Carcinogenesis by chemicals: an overview—G.H.A.
Clowes memorial lecture. Cancer Res. 30:559.
Moore, P.O., M. Koreeda, P.G. Wislocki, W. Levin, A.H. Conney, H. Yagi, and
D.M. Jernia. 1977. In vitro reactions of the diastereomeric
9,10-epoxides of (+} and (-)-trans-7,8-dihydroxy-7,8-dihyrobenzo(a)-
pyrene with polyguanylic acid and evidence for formation of an
enantiomer of each diastereomeric 9,10-epoxide from benzo(a)pyrene in
mouse skin. ACS. Symp. Ser. 44:127.
383
-------
Selkirk, J.K., R.G. Croy, and H.V. Gelboin. 1974. Benzo(a)pyrene
metabolites: efficient and rapid separation by high pressure liquid
chromatography. Science 184:169.
Selkirk, J.K., S.K. Yang, and H.V. Gelboin. 1976. Analysis of
benzo(a)pyrene metabolism in human liver and lymphocytes and kinetic
analysis of benzo(a)ppyrene in rat liver microsomes. In: Polynuclear
aromatic hydrocarbons chemistry, metabolism, and carcinogenesis. R.
Freudenthal and P.W. Jones Eds., Raven Press, New York. pp. 153-169.
Sims, P., and P.L. Grover. 1974a. Epoxides in polycyclic aromatic
hydrocarbon metabolism and carcinogenesis. Adv. Cancer Res.
20:165-274.
Sims, P., P.L. Grover, A. Swaisland, K. Pal, and A. Hewer. 1974b.
Metabolic activation of benzo(a)pyrene proceeds by a diol-epoxide.
Nature 252:326.
Weinstein, I.B. 1976a. Molecular events in chemical carcinogenesis. In:
Advances in pathobiology, 4^ Cancer Biology II. C.M. Fenuglio and
D.W. King, Eds., Stratton Intercontinental Medical Book Corp., New
York. pp. 106-107.
Weinstein, I.B., A.M. Jeffrey, K.W. Jennette, S.H. Blobstein, R.G. Harvey,
C. Harris, H. Autrup, H. Kasai, and K. Nakanishi. 1976b.
Benzo(a)pyrene diol epoxides as intermediates in nucleic acid binding
in vitro and in vivo. Science 193:592.
Yagi, H., D.R. Thakker, D. Hernandez, M. Koreeda, and D.M. Jerina. 1977.
Absolute stereochemistry of the highly mutagenic 7,8-diol 9,10-epoxide
derived from the potent carcinogen trans-7,8-dihydroxy-7,8-dihydro-
benz(a)pyrene. J. Am. Chem. Soc. 99:2358-2359.
Yang, S.K., and H.V. Gelboin. 1977a. Benzo(a)pyrene activation and
detoxification in animal and human cells. Abstract, workshop:
Carcinogenesis studies in human cells and tissues. Aspen, Colorado,
August 14-19, 1977.
Yang, S.K., D.W. McCourt, J.C. Leutz, and H.V. Gelboin. 1977b.
Benzo(a)pyrene diol epoxides: mechanism of enzymatic formation and
optically active intermediates. Science 196:1199.
384
-------
POLYCYCLIC AROMATIC HYDROCARBONS IN THE AQUATIC
ENVIRONMENT AND CANCER RISK TO AQUATIC
ORGANISMS AND MAN
by
Jerry M. Neff,
Battelle New England Marine Research Laboratory
397 Washington Street, Duxbury, MA 02332
ABSTRACT
Published values for the concentration of polycyclic
aromatic hydrocarbons (PAHs), and in particular benzo(a)pyrene
(BaP), in fresh and marine waters and in the tissues of aquatic
animals are reviewed. All but the most heavily contaminated
fresh waters contain total PAH concentrations in the parts per
trillion or low parts per billion range. The limited data for
marine waters indicate similar concentrations. Aquatic animals
generally contain 0 to 50 yg BaP/kg dry weight. Heavily
contaminated animals may contain up to 5000 \ig BaP/kg.
Teratogenic and/or carcinogenic responses have been induced in
sponges, planaria, echinoderm larvae, teleost fish, and
amphibians by exposure to carcinogenic PAHs. There are many
reports of high incidences of cancer-like growths in natural
populations of aquatic animals and plants. In most cases the
causative agent or agents are uknown. Circumstantial evidence
implicating PAHs in these natural outbreaks of cancers has been
provided in only a few cases. It is estimated that less than
0.1% of the PAH ingested by man comes from drinking water.
Fishery products consumed by man contain PAH concentrations
similar to those in smoked and charcoal-broiled meats and green
vegetables. Aquatic PAHs represent a minor source of PAH in the
human environment.
INTRODUCTION
It has been recognized for many years that some polycyclic aromatic
hydrocarbons (PAHs) can cause cancer in laboratory mammals and possibly
man. Correlations have been made between occupational or other exposure to
PAH and the incidence of human cancer (IARC, 1973; NAS, 1972; Bridboard
^t aj_., 1976). It is now generally agreed that metabolic activation by the
mixed function oxygenase (MFO) system and sometimes by epoxide hydrase is a
necessary prerequisite for PAH-induced carcinogenesis and mutagenesis
(Jerina and Daly, 1974; Huberman jrt al_., 1976; DePierre and Ernster, 1978).
Many PAH do not yield carcinogenic metabolites, and only certain metabolites
385
-------
of carcinogenic PAH are carcinogenic. For instance, more than two dozen
metabolites of benzo(a)pyrene (BaP) have been identified in mammalian
systems (DePierre and Ernster, 1978). Yet most of the cardnogenicity of
BaP is thought to be due to the isomeric 9,10-epoxy-7,8-dihydro-7,8-dihy-
droxybenzo(a)pyrenes (Sims j2t jil_., 1974; Lehr and Jerina, 1977; Yang
et aK, 1977, 1978).
The relative carcinogenicity of several PAHs are listed in Table 1.
4-, 5-, and 6-Ring PAH are more carcinogenic than either smaller or larger
ring systems; highly angular configurations are more carcinogenic than
either linear or highly condensed ring systems. Alkylation may
substantially modify carcinogenicity of a PAH. Position of alkyl
substituents is extremely important. The degree of carcinogenicity of a
PAH is related to structure and reactivity of its major metabolites.
Newman (1976) compared the carcinogenic activity of 12 monomethylbenz(a)-
anthracenes. 7-methylbenz(a)anthracene was most active, 6-,8-, and
12-methyl isomers were slightly active, and the remaining monomethylbenz(a)-
anthracenes were inactive. The author concluded that the 7-position in
benz(a)anthracene is the position at which detoxification occurs.
TABLE 1. RELATIVE CARCINOGENICITY OF PAH TO LABORATORY MAMMALS (FROM NAS, 1972)
Compound
Carginogenicity
Compound
Carginogenicity
Anthracene
Phenanthrene
Benz (a)anthracene
7,12-Oi me-thy 1 benz(a)anthracene
Di benz(aj)an thracene
Di benz(ah)anthracene
Di benz(ac)anthracene
3enzo( a) phenanthrene
Fl uorene
3enzo(a)f1uorene
3enzo(b)f"l uorene
8enzo(c)fl uorene
Dibenzo(ag)fl uorene
Dibenzo (ah )fl uorene
01 benzo(ac)f1uorene
Fl uoranthene
Benzo(b)fluoranthene
8enzo(j)f1uoranthene
Benzo(k)f1uoranthene
8enzo(mno) fl uoranthene
Aceanthrylene
Benz(j)aceanthry1ene
= cholanthrene
3-Methylcholanthrene
Naphthacene
Pyrene
Benzo(a)pyrene
Benzo(e)pyrene
Oibenzo(al)pyrene
Oibenzo(ah)pyrene
Oibenzo(ai) pyrene
Oi benzo( cd, jk) pyrene
Indeno( 1,2,3-cd) pyrene
Chrysene
Di benzo(b,def)chrysene
Di benzo( def ,p) chrysene
Dibenzo(def ,iraio) chrysene
= anthanthrene
Perylene
Benzo( ghi) perylene
Coronene
-, not carcinogenic; _+ uncertain or weakly carcinogenic; +,
carcinogenic ++, +++, ++++, strongly carcinogenic
386
-------
The 7-position must be blocked to obtain significant carcinogenic activity.
The 5-position is that at which metabolism occurs to produce cancer.
Blockage of this position destroys carcinogenic activity. Similar results
were obtained with methylchrysenes (Hecht et _§_[., 1976). 5-methylchrysene
demonstrated very high carcinogenic activity, equal to or greater than that
of BaP. 2-methylchrysene had moderate activity; other methylchrysenes were
inactive. Thus, alkylation affects carcinogenicity of PAH by altering the
position of initial enzymatic attack on the PAH molecule by the
MFO-cytochrome P-450 system.
Different species of organisms vary substantially in sensitivity to
PAH-induced carcinogenesis. This may result from interspecific differences
in the levels of MFO-cytochrome P-450 and epoxide hydrase activity,
sterochemistry of the reactions catalyzed by enzymes from different
species, and rate at which the active metabolites are converted to less
active products. Several species of aquatic annelids, arthropods, fish,
and amphibians possess the requisite enzyme systems for metabolic activat-
ion of PAH (Neff, 1978). However, it is uncertain in most cases whether
these enzymes produce the same metabolites as those produced by mammalian
enzymes. The limited data available indicate that some aquatic animals do
produce the active metabolites necessary for carcinogenesis (e.g., the
5,6-dihydrodiol of benz(a)anthracene and the 7,8-dihydrodiol of benzo(a)-
pyrene).
Stegeman (1977) provided evidence that the MFO-cytochrome P-450 system
of fish is able to produce carcinogenic or mutagenic metabolites, at least
in vitro. He added bacteria Salmonella typhimurium strain TA-98 to a
reaction mixture containing (BaP) and MFO enzymes from scup, Stenotomus
versicolor. liver. After a suitable incubation period, the number of
bacterial survivors was reduced and the number of bacterial mutants
increased (Table 2), indicating that toxic and mutagenic metabolites had
been produced by the fish liver enzymes. Similar results were obtained by
Payne et al_. (1978) with hepatic microsomes of rainbow trout, Salmo
qairdneri, and a PAH-enriched fraction of used crankcase oil. Thus, both
marine and freshwater fish are able to produce mutagenic and, by inference,
carcinogenic metabolites from PAHs.
Laboratory Studies
Many, but not all, carcinogenic chemicals are also mutagenic or
teratogem'c. That is, they can cause abnormal growth in organisms exposed
to the chemical or birth defects in the offspring of exposed parents. Thus,
the carcinogenicity of a chemical is often inferred from its mutagenicity
or teratogenicity. Some sublethal effects of PAHs, suggestive of
mutagenesis or teratogenesis, have been described in marine organisms;
e.g., growth stimulation in algae (Boney and Corner, 1962; Boney, 1974) and
abnormal development of sea urchin embryos (Ceas, 1974; de Angel is and
Giordano, 1974).
387
-------
TABLE 2. BIOACTIVATION OF BENZO(A)PYRENE TO TOXIC AND MUTAGENIC
DERIVATIONS BY SCUP, STENOTOMUS VERSICOLQR. LIVER MIXED FUNCTION
OXYGENASES. SCUP LIVER MICROSOMES WERE INCUBATED WITH
SALMONELLA TYPHEMURIUM STRAIN TA-98 AND APPROPRIATE COFACTORS
(FROM STEGEMAN, 1977)
Incubation Bacterial Mutant
Conditions Survivors Fraction*
(xlO7) (xlO-8)
Complete 53 47.2
(12.5 ug BaP/m?,)
Minus BaP 1088 0.55
Minus NADPH 945 0.21
Mutant fraction refers to the number of His* revertants per number of
survivors.
Unfortunately, with the exception of work with amphibians, there has
been relatively little laboratory research on the carcinogenicity,
mutagenicity, and teratogenicity of PAH in aquatic organisms. Exposure of
colonial calcareous sponges, Leucosolem'a complicata and U variabilis, to
5 g/A BaP resulted in choanocyte damage and abnormal growth of the oscular
tube (Korotkova and Tokin, 1968). The solitary sponge, Sycon raphanus, was
unaffected.
Foster (1969) exposed adult planarians, Dugesia dorotocephala, to
either 3-methylcholanthrene or BaP during regeneration of excised body
parts; 9% of the planaria exposed to 3-methylcholanthrene and 7% of those
exposed to BaP developed lethal papilliform tumors and other malformations.
He suggested that the malformations and possibly also the tumors were
derived from stimulated regenerative cells—totipotent neoblasts. Off-
spring of surviving PAH-exposed adults developed abnormal growths similar
to those of the exposed adults. 3-methylcholanthrene was more carcinogen-
ic and teratogenic than BaP to the offspring (Table 3). It is tempting to
speculate, based on these results, either that planaria possess the
MFO-cytochrome P-450 system or that PAH, which accumulated in their
transparent tissues, were photooxidized to mutagenic products.
388
-------
TABLE 3. LETHAL TUMORS AND DEVELOPMENTAL ANOMALIES IN OFFSPRING OF
PLANARIA, DUGESIA DOROTOCEPHALA, TREATED WITH
3-METHYLCHOLANTHRENE OR BENZO(A)PYRENE (FROM FOSTER, 1969)
Treatment of parents
Number of
offspring
Tumors in
offspring
Malformations in
offspring
3-Methyl chol anthrene
(regenerated)
Benzo(a)pyrene
(regenerated)
Acetone control
(regenerated)
Untreated controls
(not regenerated)
40
75
65
77
12
(papiI'M form
tumors)
(nodular
tumors)
1
(nodular
tumor)
none
6 eyespots poorly de-
veloped; 4 enlarged
heads and eyes; 3 fused
eyespots
3 enlarged heads and eyes
2 small heads and eyes
2 fused eyespots
variation in pigmentation
2 enlarged heads and eyes
1 small head and eyes
1 fused eyespot
Khudolei and Sirenko (1977) induced the formation of neoplasms in the
digestive gland and hematopoietic system of fresh water mussels, Unio
pictorum, by subjecting the bivalves to 200 to 400 parts per million (ppm)
diethyl- and dimethylnitrosamines in water. However, there are no
published reports of cancer induction in bivalve molluscs by exposure to or
injection of PAHs. Such an investigation would be very informative for
several reasons. The majority of bivalve molluscs studied to date lack or
have very low MFO activity and therefore should not be able to activate PAH
to carcinogenic metabolites. Yet there are many reports of natural
populations of bivalves with high incidences of neoplastic disease. Some
of these populations are from oil-polluted habitats. If bivalves are not
sensitive to PAH-induced carcinogenesis, another causative agent for these
outbreaks of neoplasia must be sought. If PAH can cause cancer in
bivalves, a closer look at how this is accomplished may lead to new
insights into the mechanisms of PAH-induced carcinogenesis.
Bourne and Jones (1973) exposed cell cultures derived from
fibroblastic gonad cells of rainbow trout, Sal mo gairdneri (RTG-2), and
epithelioid cells of fathead minnows, Pimephales promelas (FHM), to 7, 12-
dimethylbenz(a)anthracene. The PAH contaminant inhibited mitosis and
induced an increase in the number of multinucleate cells. RTG-2 cells were
more sensitive than FHM cells to the inhibition of mitosis, although FHM
389
-------
cells showed a much greater incidence of multinucleate cells than did RTG-2
cells. In FHM cultures many cells were extremely abnormal with a
reduction in cytoplasm, clumped chromation, and basophilic cytoplasm.
These results suggest that the fish cells were converting 7,12-dimethylbenz-
(a)anthracene to reactive metabolites which bonded chemically to nuclear
proteins and DMA.
Zebra fish embryos, Brachydanio rerio, exposed to 0.56 ppm 7,12-
dimethylbenz(a)anthracene developed tail necrosis, tumor-like growths, an
enlarged pericardium, and shortened body (Jones and Huffman, 1957). Epith-
elial cells and nuclei of PAH-exposed embryos were enlarged and granular.
The growth rate of some parts of the embryos was accelerated. Ermer (1970)
painted 0.5 mg of 3-methylcholanthrene twice a week for 3 to 6 months on
the skin of three freshwater fish: three-spined stickleback, Gasterosteus
aculetus; bitterling, Rhodeus amarus; and carp, Cyprinus carpio. This
treatment produced epitheliomas (skin cancer) in G.. aculetus and F^. amarus
but not in £. carpip. Similar results were produced when BaP was used.
Epitheliomas were induced in the first two but not the third fish. He also
injected G. aculetus with 0.1 ml of 1% BaP in glycerol ten times (10 mg
BaP) and maintained the fish for four months. BaP produced injection-site
necrosis but no neoplasms.
Neukomn (1974) induced cancerous lesions in newts, Triturus cristatus,
by subcutaneous injection of several PAH into regenerative areas of the
epidermis. Epithelial hyperplasia was followed by infiltration of
underlying tissues. Neoplastic infiltration occurred within four to five
days after injection; activity continued for 12 to 20 days and culminated
in a diffuse tumor. The PAH tested varied in their carcinogem'city to the
newt in the following order of decreasing potency: dibenz(ah)anthracene,
3-methylcholanthrene, BaP, and 9,10-dimethylbenz(a)thracene. Chrysene and
benz(a)anthracene were weakly active; pyrene was inactive.
Seilern-Aspang and Kratochwil (1962, 1963) induced epitheliomas in
newts, Triturus cristatus, by subcutaneous injection of BaP. The cancer
often showed regression and differentiation into non-malignant tissue when
a regeneration process was initiated in the animals by amputating a limb or
the tail. Pizzarello and Wolsky (1966) showed that if regeneration was
initiated before injection of 3-methylcholanthrene or dibenz(ah)anthracene
into J. viridescans, the cancer did not form. However, the presence of
carcinogenic PAH in the animals retarded the process of regeneration.
These results suggest that regeneration and carcinogenesis are antagonistic
processes in the newts. Regeneration inhibits malignant growth, but the
presence of carcinogenic tendencies in the animal during regeneration
retards reconstruction of tissues lost by amputation.
Crystals of 3-methylcholanthrene, 7,12-dimethylbenz(a)anthracene, or
BaP were implanted subcutaneously in the tail of toad tadpoles, Bufo
arenarum, (de Lustig and Matos, 1971). After a few days papillomas and
lymphomas were detected in 90% of the treated animals. Tumor-like cells
invaded normal structures in the tail, displacing the newly formed tissues.
Several of the PAH-treated tadpoles developed an accessory tail or
notochord. Fluorene and paraffin were completely inactive in inducing
390
-------
cancer or accessory body structures. Subsequently, Matos and de Lustig
(1973) showed that if, the tail of the tadpole was amputated at the center
of the treated area nine days after implantation of PAH crystals to promote
the formation of a regenerative blastema, the PAH-induced cancerous nodules
were encapsulated and eventually destroyed. Incidence of PAH induced
teratogenic effects, including supernumerary fins and accessory notochords,
was increased, by a regenerative field produced by tail amputation.
7,12-dimethylbenz(a)anthracene was substantially more teratogenic than
either 3-methylcholanthrene or BaP. In a similar investigation, Ruben and
Balls (1964) induced lymphosarcomas in the African clawed toad, Xenopus
laevis, by implantation of 3-methlcholanthrene into the forelimbs or
abdomen. The presence of a regenerating limb system in the carcinogenic
environment did not diminish carcinogenic activity. Accessory limb
structures were obtained near the sites of crystal implantation in some
cases.
These studies show that carcinogenic PAHs can produce cancer-like
growths and cause teratogenesis and mutagenesis in some aquatic inverte-
brates and vertebrates. The number of species examined to date is very
low, however, and there are no reports of induction of cancer by exposure
of aquatic animals to environmentally realistic levels of carcinogenic PAHs
in the water, food, or sediments.
Field Studies
Published literature on the incidence of cancer or cancer-like growths
in tissues of natural populations of marine and freshwater invertebratres
and fish is growing rapidly. Much of the recent literature is reviewed in
several papers from two recent symposium volumes (Dawe jrt a\_., 1976;
Kraybill et _§]_., 1977). In many cases organisms from polluted environments
have a higher incidence of tumors and hyperplastic diseases than those from
uncontaminated environments. In the vast majority of cases the causative
agent or agents are unknown. A great many different inorganic and organic
chemicals can induce cancer in sensitive species and polluted aquatic
habitats nearly always contain a wide variety of potential carcinogens
(Bergel, 1974). Fish are known to be susceptible to tumor induction by
chemical gents, ionizing radiation, physical factors, and viruses (Stich
and Acton, 1976). In addition, fish neoplasms can be genetically induced.
To date, carcinogenic PAHs have not been unequivocally identified as the
causative agent for an increased incidence of cancer in any natural popul-
ation of aquatic organisms.
There are several reports of increased incidence of cancerous growths
in aquatic animals in the vicinity of an oil spill (Hodgins et^ jiU, 1977).
Yevich and Barszcz (1977) reported a high incidence of gonadal neoplasms in
soft-shell clams, Mya arenaria, from Long Cove, Searsport, ME, and of hema-
topoietic neoplasms in the same species from Harpswell, Neck, ME. Although
both sites had been contaminated with refined oil, the authors concluded
that they could not say that oil was in any way a causative factor in
inducing the neoplasms.
391
-------
Powell et _§!- (1970) induced hyperplasia of ovicells in the estuarine
bryozoan, Schizoporella unicorns, by placing normal colonies in close
proximity to coal-tar derivatives in an estuary. The authors attributed
uncontrolled growth of the reproductive structures to stimulation by
several PAH known to be present in coal tar. However, Straughan and
Lawrence (1975) found no ovicell hyperplasia in bryozoans from surface, sub-
surface, and benthic kelp fronds in the vicinity of Coal Oil Point, CA, an
area of chronic submarine oil seepage.
Brown et _al_. (1973, 1977) reported a significantly higher frequency of
tumors in 2121 freshwater fish from the polluted Fox River, IL, than in
4539 fish from relatively unpolluted Lake of the Woods, Ontario, Canada.
In fish from the Fox River, incidence of neoplasia ranged from 1.17 to
12.2% in different species with a mean of 4.38%. Fish from the unpolluted
habitat had a mean incidence of neoplasia of 1.03% with a range of 0 to
2.56% in different species. The brown bullhead, Ictalurus nebulosus, was
the most seriously infected fish in the Fox River sample. Several organic
and inorganic pollutants were detected at higher concentrations in the Fox
River than in the Lake of the Woods. Concentrations of benzene, toluene,
naphthalene, and benzanthracene in the Fox River in 1976 were 0.2, 0.1,
0.2, and 0.05 ppm, respectively. These chemicals were not detected in
Canadian waters. Several chlorinated hydrocarbons and heavy metals were
also present at higher concentrations in the Fox River than in the Lake of
the Woods. Some of the tumors were apparently caused by viruses, but the
etiology of other tumors is still unknown.
A stronger case can be made for implicating PAH as the causative agent
of a high incidence of neoplasia in tiger salamanders, Ambystoma tigrinum,
fron a 13 hectacre sewage effluent lagoon at Reese Air Force Base, TX
(Rose, 1976; 1977). Here, the incidence of neoplastic and non-neoplastic
lesions in neotem'c salamanders reached a maximum of 53% in 1975. Water
and sediment of the lagoon were analyzed for organochlorine and
organophosphate pesticides, nitrosamines, several heavy metals, and PAH.
All but the PAH were below the limits of detection or were present at
normal expected concentrations. Lagoon water contained 0.085 pg/£ BaP, and
sludge on the floor of the lagoon contained high concentrations of several
PAH, especially perylene (Table 4). The sewage effluent lagoon is quite
eutrophic and contains frequent large blooms of phytoplankton, daphnia, and
copepods. Dead planktonic organisms accumulate in large mats on the anoxic
bottom sludge. These are ideal conditions for indirect biosynthesis of PAH
by reduction of extended quinones of biological origin or of direct bio-
synthesis by anaerobic bacteria in the sludge. The high concentration of
perylene supports the former hypothesis. Other sources of this unusual PAH
assemblage are also possible.
Salamanders from the lagoon had elevated hepatic MFO activity. MFO
activity in laboratory-reared salamanders could be induced to a level
similar to that in lagoon animals by exposing them to 3-methylcholanthrene
in water (Busbee et al_., 1975, 1978). Other inducers of MFO activity—
organochlorine and organophosphate insecticides—were absent from lagoon
water. All these observations provide strong circumstantial evidence that
392
-------
PAHs are critical to the high incidence of neoplasia in salamanders from
the lagoon.
TABLE 4. PAH ISOLATED FROM SLUDGE IN A SEWAGE EFFLUENT LAGOON CONTAINING
A POPULATION OF TIGER SALAMANDER NEOTENES, AMBYSTOMA TIGRINUM,
WITH A HIGH INCIDENCE OF NEOPLASIA (FROM ROSE, 1977)
., . Concentration
Compound (yg/kg, ppb)
Perylene 300.0
Pyrene 5.8
Fluoranthene 5.7
Alkyl pyrenes ' 4.9
Benz(a)anthracene 1.4
Chrysene 1.3
Triphenylene 0.5
Benzo(a)pyrene 0.5
Benzo(e)pyrene 0.2
Anthanthrene 0.2
A cancer-like hyperplasia was reported in the marine alga, Porphyra
tenera, cultivated in coastal waters near industrial wastewater outfalls
from the city of Ohmuta, Japan (Ishio et^ !]_•, 1971; 1972a,b). The cancer
could not be induced by exposing algae to diluted waste water. However,
exposure of the leaves of the alga to bottom sediments from the outfall
region for 80 to 320 min resulted in the production of cancerous growths
within 36 days. The sediment was separated into several fractions and the
greatest carcinogenic potency was found in the neutral fraction. Two
carcinogenic compounds were isolated from this fraction. One was confirmed
to be benzanthrone and the other which had the empirical formula
C25H14 was tentatively identified as 12-hydrodibenz(cd, ghi)perylene.
Obviously, a great deal more research is required on identification of
the causative agents of neoplasia outbreaks in aquatic organisms. This
type of research requires close collaboration of histopathologists, cancer
epidemiologists, and the environmental chemists.
393
-------
Aquatic PAH and Human Cancer Risk
It has been estimated that from 50 to 90% of all human cancers are
causatively related to environmental factors, mainly chemical carcinogens
(Maugh, 1974; Wynder, 1976). The contribution of carcinogenic PAH in air,
water, and food to human cancer is completely unknown. Near ubiquity of
these compounds in the human environment indicates that PAHs could be
important causative agents of human cancer.
There are many potential sources of PAH for humans, for example, drink-
ing water, smoked, roasted or charcoal broiled foods, vegetables, vegetable
fats and oils, and air pollutants (including cigarette smoke). It has been
estimated that the amount of carcinogenic PAH consumed by man in drinking
water is typically only about 0.1% of the amount accumulated from food
(Andelman and Snodgrass, 1972). Drinking water from various sources
typically contains 0.2 to 80 ng/£ (parts per trillion) BaP and 4 to 4,000
ng/ji total PAH (Table 5). Andelman and Suess (1970) made calculations
based on four samples of drinking water and showed annual human consumption
of PAH to be about 6, 9, 22, and 70 yg for the populations served by these
water supplies. The 1970 World Health Organization Standards for Drinking
Water (WHO, 1970) recommended that the concentration of total PAH in
drinking water not exceed 0.2 yg/fc. Based on a daily human consumption of
2.5 4, human consumption of drinking water with this concentration of PAH
would result in the ingestion of 182.5 yg PAH per year. By comparison,
smoke from 100 cigarettes (five packs) contains about 264 yg total PAH and
2.4 yg BaP (Severson et ^1_., 1976).
TABLE 5. TYPICAL CONCENTRATION RANGES OF BENZO(a)PYRENE AND TOTAL PAH
IN VARIOUS FRESH WATERS (CONCENTRATIONS ARE IN ng/£)
Source
BaP
Total PAH
Reference
Groundwater (Germany)
Groundwater (Germany)
Well water (Germany)
Well water (England)
Tap water (Germany)
Reservoirs (Moscow)
Reservoirs (England)
Rainwater (Koblenz)
Lake Constance (Germany)
0.4-7.0
2-4
2-15
0.2-0.6
0.5-9.0
4-13
0.7-3.8
4-80
0.2-11.5
10.9-123.5
100-200
100-750
3.6-5.8
29.2-125.5
Bomeff & Kunte,
Hellmann, 1974
Hellmann, 1974
Lewis, 1975
Borneff & Kunte,
1969
1969
Il'nitskii & Rozhnova, 1970
9.1-43.2
200-4,000
25-234
Lewis, 1975
Hellmann, 1974
Borneff & Kunte,
1964
394
-------
Blumer (1972) suggested that petroleum pollution of the seas could
pose a cancer risk to nian through contamination of fisheries resources with
carcinogenic PAHs and of recreational beaches with tar. Sullivan (1974)
reviewed reports on BaP content of petroleum, rate of petroleum spillage in
the sea, and contamination of marine organisms with PAH. He concluded that
the amount of BaP in marine foods is higher (presumably due to oil
pollution) than in non-marine foods and may pose a cancer hazard to human
consumers of fishery products.
Available data indicate that river water may contain 0.0006 to 3.5
(parts per billion) BaP and 0.02 to 3.8 yg/z total PAH (Table 6). In most
cases there is a direct relationship between PAH concentrations in river
water and the degree of industrialization and other human activity along
the banks and adjacent flood plain. Rivers remote from human activity are
relatively uncontaminated.
TABLE 6.
TYPICAL CONCENTRATION RANGES OF BENZO(a)PYRENE AND TOTAL PAH
IN VARIOUS RIVER WATERS (CONCENTRATIONS ARE IN
Total PAH
r\ererence
Rivet-
River
River
Other
Oyster
Raskov
frcin
Sunzha
Rh i ne
Rhine
Aach a
German
River
reg i c
human
River
charge of
Thames
Trent
Severn
River
Rive',
River
a i "ijinz
at Kobler-
t Stockach
rivers
, CO USA
n, USSR remote
acti vity
below dis-
ci 1 refinery
, England
Engl and
, England
0.05-0.
0.01-0.
0.034-0.
0.0005-0.
0.073-0.
10"5-10
0.05-3.
0.17-0.
0.0053-0.
0.0015-0.
n
06
043
31
150
-4
5
28
504
043
0.73-1.50 Borneff i vjnte, 1964
C. 5-3. 00 Hellr-sr.n, -97:
1.44-3.10 Borneff i Ki.rte, 1955
0.20-1.00 Scrneff & Kunte, 1954,
Keegan, 1971
11 'nitskii e_t a]_. , 1971
Samoilovich & Red'kin,
0.8-2.35 Acheson et al- , 1976
0.025-3.79 Lewis, 1975
0.020-0.256 Lewis, 1975
1955
•963
Relatively little information is available concerning the concentrat-
ions of PAH in estuarine and oceanic waters. Apparently very few attempts
have been made to precisely identify and quantify PAH concentrations in the
oceans. Most of the data available are based on estimates of total
aromatics by UV, IR, or fluorescence techniques—methods which may be
subject to considerable interference from non-PAH material. Typical of
this type of approach is the work of Zsolnay (1977) (Table 7). UV-absorb-
ing "aromatics" represented 1 to 10% of the total hydrocarbons in the water
samples. Aromatic hydrocarbon concentrations decreased with depth in most
cases. Marty £t_a]_. (1978) measured the concentrations of aliphatic,
395
-------
alicyclic, and aromatic hydrocarbons In surface water, sea surface
microlayer, and airborne participates from the tropical Atlantic Ocean west
of the Canary and Cape Verde Islands. Aromatic hydrocarbons represented 13
to 55% of the total hydrocarbons in the samples (Table 8). PAH in the
water samples included phenanthrene, alkylphenanthrene, perylene, fluoran-
thene, and pyrene. Traces of benzofluoranthenes and benzopyrenes were also
detected. The dominant PAH in the surface microlayer and airborne aerosols
was phenanthrene (40% of the aromatic fraction). Mono-, di-, and
tri-methyl phenanthrenes were also abundant. No perylene, benzofluoranthenes,
or benzopyrenes were detected. The authors hypothesized that the phenan-
threnes were of biological origin and that the other PAHs were from
industrial smoke and petroleum spillage. It would appear, therefore, that
the concentrations of aromatics in marine surface waters are similar to
those in ground water and well water.
TABLE 7. DISTRIBUTION OF HYDROCARBONS IN MARINE AND ESTUARINE WATERS;
AROMATIC HYDROCARBON VALUES ARE RELATIVE AND ARE BASED UPON THE
USE OF PHENANTHRENE STANDARD AT 254 nm (FROM ZSOLNAY, 1977)
Date
Area
1973
Baltic
Depth
(m)
1
Total hydrocarbons
Mean & S.D.
(pg/0
13.9±3.8
Aromatic hydrocarbons
(based on UV abs.)
(ng/U
277±121
1973
Nova Scotia
to Gulf Stream
1974-1976
Sargasso Sea
off Bermuda
10-50
1 m above
sediment
1
10
25
1
30
300
1200
2000
1975
Mediterranean
5.4±1.8
3.8±1.4
4.7±1.3
3.5±1.8
6.1±1.8
0.32±0.12
0.47±0.25
0.26±0.16
0.00
0.00
16.0
52±9
47±13
30.8±10
16.8±4
8.6±9
31±10
1±0
1±1
4
0
148±36
396
-------
TABLE 8. MEAN CONCENTRATIONS OF HYDROCARBONS INSURFACE SAMPLES OF
SEAWATER, THE SEA SURFACE MICROLAYER, AND AEROSOLS COLLECTED
12 m ABOVE THE SEA SURFACE IN THE TROPICAL EASTERN ATLANTIC
OCEAN (FROM MARTY et al_., 1978)
Composition of the hydrocarbon fraction
.. Concentration of
bampie Total hydrocarbons aliphatic & aromatic
alicyclic hydrocarbons
hydrocarbons
Seawater 10 yg/* 67%, 9% n-alkanes 33% (3.3 yg/s.)
Surface microlayer 39 yg/s. 45%, 6% n-alkanes 55% (21.4 yg/£)
., 2
Aerosols 9 yg/m 87%, 2% n-alkanes 13% (1.2 yg/m )
Barbier £t _al_. (1973) analyzed the hydrocarbons in surface water off
the French coast at Brest. The concentration of total hydrocarbons was 137
vg/l. Analysis of the hydrocarbon type distribution by UV spectrophoto-
metry and mass spectrometry revealed that bicyclic aromatics represented
3.5% and PAH represented. 2.5% of the total hydrocarbons present. Thus, the
concentration of PAH in Brest seawater was approximately 3.4 yg/£. Brown
and Huffman (1976) analyzed a large number of water samples from the
Atlantic, Mediterranean, and the Indian Ocean along the Persian Gulf oil
tanker route. The mean concentration of nonvolatile hyhrocarbons in
surface water from the Atlantic Ocean was 4 yg/£ with values from 1.3 to 13
iig/A falling within one standard deviation of the mean. Mean relative
aromatic concentration was 24% of the total nonvolatile hydrocarbons.
Included in the total aromatic fraction were 1 to 4 ring aromatics plus
sulfur-containing aromatics (benzo- and dibenzothiophenes). PAH repre-
sented less than half of the total aromatic fraction, so PAH concentrations
in Atlantic Ocean surface water were about 0.4 yg/Ji with a range of 0.13 to
1.3
There are only a few reports dealing with concentration of specific
PAHs in seawater. Seawater at Brest, France, was reported to contain
traces of BaP (Saliot, 1969, as cited by Barbier et al_., 1973). Armstrong
et a!., (1977) reported that a water sample taken approximately 15 m from
In oil separator platform brine outfall in Trinity Bay, TX, contained 0.4, 0.2,
397
-------
and 0.6 ug/i naphthalene, 1-methylnaphthalene, and 2-methylnaphthalene,
respectively (Table 9). No higher molecular weight PAHs were detected in
the water despite the fact that the undiluted brine effluent contained
significant concentrations of PAH in the dimethyl naphtha!ene-dimethylphen-
anthrene range.
TABLE 9. CONCENTRATIONS OF AROMATIC HYDROCARBONS IN THE BRINE EFFLUENT
FROM AN OIL SEPARATOR PLATFORM IN TRINITY BAY, TX, AND IN THE
WATER AND SEDIMENTS 15 m FROM THE BRINE OUTFALL (FROM ARMSTRONG
et al., 1977)
Compound
Benzene
Toluene
Xylene
C2-Benzene
C^-Benzene
C,- Benzene
Naphthalene
1 -Methyl naphtha! ene
2-Methyl naphthalene
Dimethyl naphthalenes
Tri methyl naphthalenes
Ca-Naphthalenes
Biphenyl
Methyl biphenyls
Dimethyl bi phenyl s
Fluorene
Methyl f 1 uorene
Dimethyl fluorenes
Tri methyl fluorenes
Phenanthrene
Methyl phenanthrenes
Dimethyl phenanthrenes
Tri methyl phenanthrenes
Total aromatics
Brine effluent Water
(yg/i) (ug/O
3,300 1.50
3,500 3.20
2,400
3.10
650 0.80
-
300 0.40
370 0.20
500 0.60
30
260
-
15
11
30
14
63
42
-
36
63
20
-
11,760 10.50
Sediment
(ug/kg wet wt)
-
-
-
-
600
400
200
500
500
800
10,000
400
-
200
800
100
700
900
600
100
600
1,000
500
34,200
398
-------
It can be concluded that fresh and marine waters, even from relatively
polluted areas, contain enough low concentrations of PAH that they would
not pose a carcinogenic hazard to recreational users of these water bodies.
More recent studies on BaP and total PAH contamination of marine
organisms indicate that in the vast majority of cases these organisms
contain low or undetectable levels of BaP and other PAH (Table 10). BaP
concentrations in marine animals were usually in the 0 to 30 yg/kg range
except in animals collected from severely polluted locations or from the
immediate vicinity of creosoted pilings.
Cahnmann and Kuratsune (1957) measured the concentrations of eight PAH
in the tissues of oysters, Crassostrea virginica. from the harbor of Norfolk,
VA, an area fairly heavily polluted by domestic, industrial, and shipping
wastes (Table 11). Total approximate PAH concentrations ranged from 700 to
1200 yg/kg wet weight. BaP represented only 0.2 to 0.3% of the total.
Fazio (1971) measured the concentrations of nine PAH in oysters, £.
virginica, from Galveston and Aransas Bays, TX. These oysters were much
less contaminated than those from Norfolk, VA, despite the extensive oil
refinery and port activity around Galveston Bay (Table 12). Interestingly,
no BaP was detected in any of the oyster samples by a method demonstrated
to be sensitive to 2 yg/kg wet weight or less. Samples from the most
heavily contaminated station, an area in Galveston Bay closed to commercial
oyster harvesting because of elevated sewage contamination, contained a
mean total of 21.9 yg/kg wet weight of other PAHs. Oysters from
uncontaminated regions of Galveston Bay and the relatively unpolluted
Aransas Bay contained mean total PAH concentrations of 8.2 and 3.6 yg/kg,
respectively. Lower molecular weight PAHs were more abundant than higher
molecular weight PAH in both Galveston Bay and Norfolk, VA, oysters. In
both cases fluoranthene and pyrene were the most abundant PAH in oyster
tissues.
BaP concentrations as high as 1 to 5 mg/kg have been reported in some
marine animals, and repeated consumption of such heavily contaminated
animals would pose a cancer risk to humans. Heavily contaminated animals
are usually unpalatable because of their oily smell and taste. The only
report of BaP accumulation by a marine animal following an oil spill is
that of Bories et al_., (1976). Eleven days after a spill, Mytilus edulis.
contained approximately 55 yg BaP/kg. In comparison, a charcoal broiled
steak may contain up to 50 yg BaP/kg (Lijinsky, 1967), vegetable oils may
contain up to 36 yg BaP/kg (Swallow, 1976), and smoked meats may contain up
to 15 yg Bap/kg and 141 yg total BaP/kg (Panalaks, 1976). Typically food
plants contain 0.08 to 30 yg BaP/kg (Carrier, 1977). Leafy vegetables and
grains generally contain higher BaP concentrations than root vegetables.
Thus, aquatic organisms do not ordinarily contain significantly higher
concentrations of BaP and other PAHs than other human foods. However,
consumption of aquatic organisms from regions severely contaminated with
petroleum or industrial pollution should certainly be avoided.
In summary, the available evidence indicates that PAH-contaminated
water and fishery products represent minor sources of PAH toxicity to man.
399
-------
TABLE 10. SUMMARY OF ANALYSES OF THE CONCENTRATION OF BENZO(A)PYRENE IN
THE TISSUES OF MARINE ANIMALS
Organisms Location
BaP
(ug/kg
wet wt.)
Reference
Molluscs
Mussels
(HytHus edulls)
M. edulis 4
M. caTiforneanus
M. callforneanus
Oyster (Crassostrea
virginica)
C. gigas
Gaper clam
(Tresus capax)
Softshell clam
(Mya arenaria)
Butter clam
(Saxidomas qiqanteus)
Clam (unidentified)
Crustaceans
Blue crab (Calli-
nectes sapidus) Chesapeake Bay
Unidentified crab Ran"tan Bay, N.J.
Shrimp (Penaeus sp.) Palacios, Tx.
Fa1mouth, Mass.
Little Slppewissett
Wild Harbor
Oregon Bays
Vancouver Harbor, B.C.
25 stations between Bodega
Head - San Diego, Calif.
West coast Vancouver Island, B.C.
Long Island Sound
Chincoteague, Va.
Tillmook Bay, Or.
Oregon Bays
Oregon Bays
Coos Bay, Or.
Chincoteague, Va.
<0.5
0.5
<0.1-30.2
<0.1-63
<0. 1-8.2
<0.1-0.2
2
0.2
<0. 1-3.21
<0. 1-6.66
0.29-1.05
0.3
<5
3
Fish
Cod (Gadus sp.) Atlantic, 40 km off Toms River, N.J. <1
Menhaden (Brevoortia
tyrannusl Raritan Bay, N.J. 1.5
Flounder
(unidentified) Long Branch, N.J. <2
Pancirov 4 Brown^ 1977
Pancirov & Brown. 1977
Mix et al., 1977
DunnT STich, 1975
Dunn & Young, 1976
Dunn & Stich, 1975
Pancirov 4 Brown, 1977
Pancirov 4 Brown, 1977
Mix et al-, 1977
Mix et al.., 1977
Wx Si Si-, '977
Mix et aj[, 1977
Pancirov & Browrx 1977
Pancirov & Brown , 1977
Pancirov & Brown, 1977
Pancirov & Brown, 1977
Pancirov 4 Brown, 1977
Pancirov & Brown
Pancirov 4 Brown, 1977
400
-------
TABLE 11. CONCENTRATION OF SELECTED PAH IN THE TISSUES OF OYSTERS,
CRASSOSTREA VIRGINICA. FROM THE HARBOR OF NORFOLK, VA
(FROM CAHNMANN AND KURATSUNE, 1957)
Compound
Fluoranthene
Pyrene
Chrysene
Benz(a)anthracene
8enzo(k)f1uoranthene
Benzo(a)pyrene
3enzo(e)pyrene
Benzo(ghi)perylene
Approximate concentration
wet weight
600 - 1000
100 - 160
20 - 40
<10
8 - 12
2 - 6
<20
1 - 5
TABLE 12. CONCENTRATIONS OF SELECTED PAH IN THE TISSUES OF OYSTERS,
CRASSOSTREA VIRGINICA FROM THE TEXAS, GULF COAST (FROM FAZIO,
1981)
Compound
Phenanthrene
Fluoranthene
Pyrene
Chrysene
Benzc (b )fluoranthene
Benzc (k Jfluoranthene
Benzo(a)pyrene
Benzo (e )pyrene
Perylene
Concentration (yg/kg wet vrt)
Station A Station B Station C
mean range mean
-
1.7 1.7 3.0
0.9 0.8-1.0 1.8
0.5 0.3-0.6 0.2
0.3 0.0-0.5 1.2
0.1
-
0.2 0.0-0.4 1.2
0.7
range
1.6-5.6
1.3-5.9
0.0-1.7
0.5-1.8
0.0-0.6
0.0-4.0
0.0-3.5
mean
2.2
7.8
6.5
0.6
2.2
0.4
-
2.1
0.1
range
0.0-6.9
2.8-14
3.8-13
0.0-2.0
0.0-4.1
0.0-1.1
0.8-4.2
0.0-1.5
Station A, Aransas Bay oyster reefs relatively unccntaminated from domestic and petroleum
operations; Station 8, Galveston Bay oyster reefs approved for commercial harvesting;
Station C, Galveston Bay oyster reefs closed to commercial harvesting due to domestic
sev/age contamination.
401
-------
REFERENCES
Acheson, M.A., R.M. Harrison, R. Perry, and R.A. Well ings. 1976. Factors
affecting the extraction and analysis of polynuclear aromatic
hydrocarbons in water. Water Res. 10:207-212.
Andelman, J.B., and J.E. Snodgrass. 1972. Incidence and significance of
polynuclear aromatic hydrocarbons in the water environment. CRC Crit.
Rev. Environ. Contr. 4:69-83.
Andelman, J.B., and M.J. Suess. 1970. Polynuclear aromatic hydrocarbons
in the water environment. Bull. WHO 43:479-508.
Armstrong, H.W., K. Fucik, J.W. Anderson, and J.M. Neff. 1977. Effects of
oilfield brine effluent on benthic organisms in Trinity Bay, TX.
American Petroleum Institute, Washington, DC. API Publication No.
4291. 82 pp.
Barbier, M., D.Joly, A. Saliot, and D. Tourres. 1973. Hydrocarbons from
seawater. Deep-Sea Res. 20:305-314.
Bergel, F. 1974. Carcinogenic hazards in natural and manmade
environments. Proc. Roy. Soc. London. Ser. B. 185:165-181.
Blumer, M. 1972. Oil contamination and the living resources of the sea.
In: Marine pollution and sea life. M. Ruivo, Ed., Fishing News
(Books), London, pp. 476-481.
Boney, A.D. 1974. Aromatic hydrocarbons and the growth of marine algae.
Mar. Pollut. Bull. 5:185-186.
Boney, A.D., and E.D.S. Corner. 1962. On the effects of some carcinogenic
hydrocarbons on the growth of spore!ings of marine red algae. J. Mar.
Biol. Assn. U.K. 42:579-585.
Bories, G., J. Tulliez, J.C. Peltier, and R. Fleckinger. 1976. Change in
the concentration of aliphatic and naphthenic hydrocarbons, as well as
3,4-benzopyrene, in mussels from a coastal zone polluted by a spill of
fuel oil. C.R. Acad. Sci. Paris. Ser. D 282:1641-1644 (French).
Borneff, J., and H. Kunte. 1964. Carcinogenic substances in water and
soil. Part XVI: Evidence of polycyclic aromatic hydrocarbons in
water samples through direct extraction. Arch. Hyg. (Berlin)
148:585-597 (German).
Borneff, J., and H. Kunte. 1965. Carcinogenic substances in water and
soil. XVII. Concerning the origin and estimation of the polycyclic
aromatic hydrocarbons in water. Arch. Hyg. (Berlin) 149:226-243
(German).
402
-------
Borneff, J., and H. Kunte. 1969. Carcinogenic substances in water and
soil. XXVI. Routine method for analysis of polycyclic aromatic
hydrocarbons in water. Arch. Hyg. (Berlin) 153:220-229 (German).
Bourne, E.W., and R.W. Jones. 1973. Effects of 7,12-dimethylbenz(a)anth-
racene (DMBA) in fish cells in vitro. Trans. Am. Microsc. Soc.
92:140-142.
Bridboard, K., J.F. Finklea, J.K. Wagoner, J.B. Moran, and P.Caplan. 1976.
Human exposure to polynuclear aromatic hydrocarbons. In:
Carcinogenesis~a comprenhensive survey. Vol. 1. Polynuclear
aromatic hydrocarbons. Chemistry, metabolism, and carcinogenesis. R.
Freudenthal and P.W. Jones, Eds. Raven Press, New York. pp.319-324.
Brown, E.R., J.J. Haxdra, L. Keith, I. Greenspan, J.B.G. Kwapinski, and P.
Beamer. 1973. Frequency of fish tumors found in a polluted water
shed as compared to nonpolluted Canadian water. Cancer Res.
33:189-198.
Brown, E.R., T. Sinclair, L. Keith, P.Beamer, J.J. Hazdra. V.Nair, and 0.
Callaghan. 1977. Chemical pollutants in relation to diseases in
fish. In: Aquatic pollutants and biologic effects with emphasis on
neoplasia. H.F. Kraybill, C.J. Dawe, J.C. Harshbarger, and R.G.
Tardiff, Eds., Ann. N.Y. Acad. Sci. 298:535-546.
Brown, R.A., and H.L. Huffman, Jr. 1976. Hydrocarbons in open ocean
waters. Science 191:847-849.
Busbee, D., D. Colvin, I. Muijsson, F. Rose, and E. Cantrell. 1975.
Induction of aryl hydrocarbon hydroxylase in Ambystoma tigrinum.
Comp. Biochem. Physiol. 50C:33-36.
Busbee, D.L., J. Guyden, T. Kingston, F.L. Rose, and E.T. Cantrell. 1978.
Metabolism of benzo(a)pyrene in animals with high aryl hydrocarbon
hydroxylase levels and high rates of spontaneous cancer. Cancer Let.
4:61-67.
Cahnmann, H.J., and M. Kuratsune. 1957. Determination of polycyclic
aromatic hydrocarbons in oysters collected in polluted water. Anal.
Chem. 29:1312-1317.
Carrier, R.F. 1977. Plant interactions. In: Environmental, health, and
control aspects of coal conversion: an information overview. Vol. II
H.M. Braunstein, E.D. Copenhaver, and H.A. Pfuderer, Eds., Oak Ridge
National Laboratory, Oak Ridge, TN. ORNL/EIS-95. 118 pp.
Ceas, M.P. 1974. Effects of 3-4 benzopyrene on sea urchin egg
development. Acta Embryol. Exper. 3:267-272.
403
-------
Dawe, C.J., D.G. Scarpelli, and S.R. Wellings, Eds. 1976. Tumors in
aquatic animals. In: Progress in experimental tumor research. Vol.
20. S. Karger, Basel, Switzerland. 438 pp.
de Angelis, E., and G.G. Giordano. 1974. Sea urchin egg development
under the action of benzo(a)pyrene and 7,12-dimethylbenz(a)anthracene,
Cancer Res. 34:1275-1280.
de Lustig, E.S., and E.L. Matos. 1971. Teratogenic effects included in
tail of Bufo arenarium tadpoles following treatment with carcinogens.
Experientia 27:555-556.
De Pierre, J.W., and L. Ernster. 1978. The metabolism of polycyclic
hydrocarbons and its relationship to cancer. Biochem. Biophys. Acta
473:149-186.
Dunn, B.P., and H.F. Stich. 1975. The use of mussels in estimating
benzo(a)pyrene contamination of the marine environment. Proc. Soc.
Exper. Biol. Med. 150:49-51.
Dunn, B.P., and D.R. Young. 1976. Baseline levels of benzo(a)pyrene in
southern California mussels. Mar. Pollut. Bull. 7:231-234.
Ermer, M. 1970. Studies with carcinogens in short-lived fish species.
Zool. Anz. 184:175-193 (German).
Fazio, T. 1971. Analysis of oyster samples for polycyclic hydrocarbons.
Proc. 7th national shellfish sanitation workshop. Food and Drug
Administration, Division of Shellfish Sanitation. Washington, DC.
pp. 283-243.
Foster, J.A. 1969. Malformations and lethal growth in planaria treated
with carcinogens. Natl. Cancer. Inst. Monogr. 31:683-691.
Hecht, S.S., M. Loy, and D. Hoffman. 1976. On the structure and
carcinogenicity of the methylchrysenes. In: Carcinogenesis—a
comprehensive survey. Vol. I. Polynuclear, aromatic hydrocarbons.
Chemistry, metabolism, and carcinogenicity. R. Freucenthal and P.W.
Jones, Eds. Raven Press, New York. pp. 325-340.
Hellman, von H. 1974. Occurrence and origin of so-called carcinogens and
other polycyclic hydrocarbons in water. Deut. Gewasserkund. Mitteil.
18:155-157 (German).
Hodgins, H.O., B.B. McCain, and J.W. Hawkes. 1977. Marine fish and
invertebrate diseases, host disease resistance, and pathological
effects of petroleum. In: Effects of petroleum on arctic and
subarctic marine environments and organisms. D.C. Malins, Ed.,
Academic Press, New York. pp. 95-174.
404
-------
Huberman, E., L. Sachs, S.K. Yang, and H.V. Gelboin. 1976. Identification
of mutagenic metabolites of benzo(a)pyrene in mammalian cells. Proc.
Natl. Acad. Sci. 73:607-612.
Il'nitskii, A.P., K.P. Ershova, A.Y. Khesina, L.G. Rozhkova, V.G. Klubkov,
and A.A. Korolev. 1971. Stability of carcinogens in water and
efficacy of methods of decontamination. Hyg. Sanit. 36:9-13
(Russian).
Il'nitskii, A.P., and L.G. Rozhova. 1970. Pollution of reservoirs by
carcinogenic hydrocarbons. Vop. Onhol. 16:78 (Russian) (Cited by
Andelman and Suess, 1972; op.cit.).
International Agency for Research on Cancer. 1973. Monograph on the
evaluation of carcinogenic risk of the chemical to man: Certain
polycyclic aromatic hydrocarbons and heterocyclic compounds. Vol. III.
WHO, Geneva, Switzerland.
Ishio, S., K. Kawabe, and T. Tomiyama. 1972a. Algal cancer and its causes.
I. Carcinogenic potencies of water and suspended solids discharged to
the River Ohmuta. Bull. Jap. Soc. Sci. Fish. 38:17-24 (Japanese).
Ishio, S., H. Nagagawa, and T. Tomiyama. 1972b. Algal cancer and its
causes. II. Separation of carcinogenic compounds from sea bottom mud
polluted by wastes of the coal chemical industry. Bull. Jap. Soc.
Sci. Fish. 38:571-576 (Japanese).
Ishio, S., T. Yano, and R. Nakagana. 1971. Algal cancer and causal
substances in wastes from the coal chemical industry. In: Proc. 5th
international water pollution research conference. Pergamon Press,
London pp.1-8.
Jernia, D.M., and J.W. Daly. 1974. Arene oxides: a new aspect of drug
metabolism. Science 185:573-578.
Jones, R.W., and M.N. Huffman. 1957. Fish embryos as bioassay material in
testing chemicals for effects on cell division and differentiation.
Trans. Am. Microsc. Soc. 76:177-183.
Keegan, R.E. 1971. The trace fluorometric determination of polynuclear
aromatic hydrocarbons in natural water. Doctoral Dissertation,
Graduate School of Chemistry, University of New Hampshire, Durham, NH.
(Cited by Andleman and Snodgrass, 1972; op. cit.).
Khudolei, V.V., and O.A. Sirenko. 1977. Tumor development in the bivalve
mollusk Um'o pictorum induced by N-nitroso compounds. Bull. Exper.
Biol. Med. 83:684-686.
405
-------
Korotkova, G.P., and B.P. Tokin. 1968. Stimulation of the process of
somatic embryogenesis in some porifera and coelenterata. I. Effect
of carcinogenic agents on some Porifera. Acta Biol. Hungary
19:465-474.
Kraybill, H.F., C.J. Dawe, J.C. Harshbarger, R.G. Tardiff, Eds. 1977.
Aquatic pollutants and biologic effects with emphasis on neoplasia.
Ann. N.Y. Acad. Sci. 298:604.
Lehr, R.E., and D.M. Jernia. 1977. Metabolic activations of polycyclic
hydrocarbons. Structure-activity relationships. Arch. Toxicol.
39:1-6.
Lewis, W.M. 1975. Polynuclear aromatic hydrocarbons in water. Water
Treat. Exam. 24:243-277.
Lijinsky, W. 1967. Detection of carcinogenic chemicals in the
environment. Cancer Bull. 19:63-64.
Marty, J.C., A. Saliot, and M.J. Tissier. 1978. Inventory, distribution,
and orgin of aliphatic and polycyclic aromatic hydrocarbons in
seawater, the surface microlayer, and marine aerosols in the eastern
tropical Atlantic Ocean. C.R. Acad. Sci. Paris, Ser. D. 286:833-836
(French).
Matos, E.L., and E.S. de Lustig. 1973. Teratogenic effects of carcinogen
implantation in a regenerative field in Bufo arenarium tadpoles.
Teratology 8:167-174.
Maugh, T.H. 1974. Chemical carcinogens: a long-neglected field blossoms,
Science 183:940-944.
Mix, M.C., R.T. Riley, K.I. King, S.R. Trenholm, and R.L. Schaffer. 1977.
Chemical carcinogens in the marine environment. Benzo(a)pyrene in
economically-important bivalve mollusks from Oregon estuaries. In:
Fate and effects of petroleum hydrocarbons in marine ecosystems and
organisms. D.A. Wolfe, Ed., Pergamon Press, New York. pp. 421-431.
National Academy of Sciences, NAS. 1972. Particulate polycyclic organic
matter. Washington, DC. 361 pp.
Neff, J.M. 1978. Polycyclic aromatic hydrocarbons in the aquatic
environment: sources, fates, and biological effects. American
Petroleum Institute, Washington, DC. 350 pp.
Neukomn, S. 1974. The newt test for studying certain categories of
carcinogenic substances. In: Experimental model systems in
toxicology and their significance to man. Proc. Europ. Soc. Study
Drug. Tox., Int. Conf. 15:228-235.
406
-------
Newman, M.S. 1976. Carcinogenic activity of benz(a)anthracenes. In:
Carcinogenesis--a comprehensive survey. Vol. I. Polynuclear aromatic
hydrocarbons. Chemistry, metabolism, and carcinogenesis.
R. Freudenthal and P.W. Jones, Eds. Raven Press, New York. pp.
203-208.
Panalaks, T. 1976. Determination and identification of polycyclic
aromatic hydrocarbons in smoked and charcoal-broiled food products by
high pressure liquid chromatography and gas chromatography. J.
Environ. Sci. Health 811:299-315.
Pancirov, R.J., and R.A. Brown. 1977. Polynuclear aromatic hydrocarbons
in marine tissues. Environ. Sci. Technol. 11:989-992.
Payne, J.F., I. Martins, and A. Rahimtula. 1978. Crankcase oils: are
they a major mutagenic burden in the aquatic environment. Science
200:329-330.
Pizzarello, D.J., and A. Wolsky. 1966. Carcinogenesis and regeneration in
newts. Experientia 22:387-388.
Powell, N.A., C.S. Sayce, and D.F. Tufts. 1970. Hyperplasia in an
estuarine bryozoan attributable to coal tar derivatives. J. Fish.
Res. Bd. Canada 27:2095-2096.
Rose, F.L. 1976. Tumorous growths of the tiger salamander, Ambystoma
tigrinum, associated with treated sewage effluent. In: Tumors in
aquatic animals. C.J. Dawe, D.G. Scarpelli, S.R. Well ings,
Eds. Prog. Exper. Tumor Res. 20:251-262.
Rose, F.L. 1977. Tissue lesions of tiger salamander (Ambystoma tigrinum):
relationship to sewage effluents. In: Aquatic pollutants and
biologic effects with emphasis on neoplasia. H.F. Kraybill, C.J.
Dawe, J.C. Hershbarger, and R.G. Tardiff., Eds. Ann. N.Y. Acad. Sci.
298:270-279.
Ruben, L.N., and M. Balls. 1964. The implantation of methylcholanthrene
crystal into regenerating and non-regenerating forelimbs of Xenopus
laevis. J. Morphol. 115:239-254.
Saliot, A. 1969. Contribution to the study of organic compounds dissolved
in seawater. These 3 erne cycle. Fac. Sci. Paris 106 pp. (French)
(Cited by Barbier et al_., 1973; op. cit.).
Samoilovich, L.N., and Y.R. Red'kin. 1968. 3,4-benzopyrene pollution of
the River Sunzha caused by the petrochemical industry in Grozny. Gig.
Sanit. 33:6 (Russian)(Cited by Andelman and Suess, 1972; op. cit.)-
407
-------
Seilern-Aspang, F., and K. Kratochwil. 1962. Induction and differentia-
tion of an epithelial tumor in the newt (Triturus cristatus). J.
Embryo!. Exper. Morphol. 10:337-356.
Seilern-Aspang, F., and K, Kratochwil. 1963. Spontaneous healing of an
infiltrating and metastasizing epithelial tumor of Triturus cristatus.
Arch. Geschwulstoforsch. 21:292-300.
Severson, R.F., M.E. Snook, H.C. Higman, O.T. Chortyk, and F.J. Akin.
1976. Isolation, identification, and quantisation of polynuclear
aromatic hydrocarbons in tobacco smoke. In: Carcinogenesis—a
comprehensive survey. Vol. I. Polynuclear aromatic hydrocarbons--
chemistry, metabolism, and carcinogenesis. R. Freudenthal and P.M.
Jones, Eds. Raven Press, New York. pp. 253-270.
Sims, P., P.L. Grover, A. Swaisland, K. Pal, and A. Hewer. 1974.
Metabolic activation of benzo(a)pyrene proceeds by a diol-epoxide.
Nature (London). 252:326-328.
Stegeman, J.J. 1977. Fate and effects of oil in marine animals. Oceanus
20:59-66.
Stich, H.F., and A.B. Acton. 1976. The possible use of fish tumors in
monitoring for carcinogens in the marine environment. In: Tumors in
aquatic animals. C.J. Dawe, D.G. Scarpelli, and S.R. Wei lings, Eds.
Prog. Exper. Tumor. Res. 20:44-54.
Straughan, D., and D.M. Lawrence. 1975. Investigation of ovicell
hyperplasia in bryozoans chronically exposed to natural oil seepage.
Water Air Soil Pollut. 5:39-45.
Sullivan, J.B. (1974). Marine pollution by carcinogenic hydrocarbons.
In: Marine pollution monitoring (petroleum). U.S. Natl. Bur.
Standards Special Publ. 409:261-263.
Swallow, W.H. 1976. Survey of polycyclic aromatic hydrocarbons in
selected foods and food additives in New Zealand. N.Z. J. Sci.
19:407-412.
World Health Organization. 1970 European Standards for drinking water,
2nd ed., revised. WHO, Geneva, Switzerland. 58 pp.
Wynder, E.L. 1976. Nutrition and Cancer. Fed. Proc. 35:1309-1315.
Yang, S.K., D.W. McCourt, J.C. Leutz, and H.V. Gelboin. 1977.
Benzo(a)pyrene diol epoxides: mechanism of enzymatic formation and
optically active intermediates. Science 196:1199-1201.
408
-------
Yang, S.K., P.P. Roller, and H.V. Gelboin. 1978. Benzo(a)pyrene
metabolism: Mechanism in the formation of epoxides, phenols,
dihydrodiols, and the 7,8-diol-9,10-epoxides. In: Carcinogenesis--a
comprehensive survey. Vol. III. Polynuclear aromatic hydrocarbons:
Second International Symposium on Analysis, Chemistry, and Biology.
P.W. Jones and R.I. Freudenthal, Eds. Raven Press, New York. pp.
285-302.
Yevich, P.P., and C.A. Barszcz. 1977. Neoplasia in soft-shell clams (Mya
arenaria) collected from oil-impacted sites. In: Aquatic pollutants
and biologic effects with emphasis on neoplasia. H.F. Kraybill, C.J.
Dawe, J.C. Harshbarger, and R.G. Tardiff, Eds. Ann. N.Y. Acad. Sci.
298:409-426.
Zsolnay, A. 1977. Inventory of nonvolatile fatty acids and hydrocarbons
in the oceans. Mar. Chem. 5:465-475.
409
------- |