TOXICOLOGICAL SCIENCES 105(1), 5–23 (2008)
Advance Access publication December 20, 2007
Nonadditive effects of PAHs on Early Vertebrate Development:
Mechanisms and Implications for Risk Assessment
Sonya M. Billiard†,§,1Joel N. Meyer,† Deena M. Wassenberg,‡ Peter V. Hodson,§ and Richard T. Di Giulio*,†
*Health Canada, Health Products and Food Branch, Bureau of Chemical Safety, Ottawa, Ontario K1A 0L2, Canada; †Integrated Toxicology Program, Nicholas
School of the Environment, Duke University, Durham, North Carolina 27708-0328; ‡College of Biological Sciences, University of Minnesota, St Paul, Minnesota
55108; and §School of Environmental Studies, Queen’s University, Kingston, Ontario K7L 3N6, Canada
Received September 4, 2007; accepted December 11, 2007
Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous
environmental contaminants. Traditionally, much of the research
has focused on the carcinogenic potential of specific PAHs, such as
benzo(a)pyrene, but recent studies using sensitive fish models have
shown that exposure to PAHs alters normal fish development. Some
PAHs can induce a teratogenic phenotype similar to that caused by
planar halogenated aromatic hydrocarbons, such as dioxin. Conse-
of compounds. Unlike dioxins, however, the developmental toxicity
of PAH mixtures is not necessarily additive. This is likely related to
their multiple mechanisms of toxicity and their rapid biotransfor-
mation by CYP1 enzymes to metabolites with a wide array of
structures and potential toxicities. This has important implications
for risk assessment and management as the current approach for
complex mixtures of PAHs usually assumes concentration addition.
In this review we discuss our current knowledge of teratogenicity
caused by single PAH compounds and by mixtures and the
importance of these latest findings for adequately assessing risk of
PAHs to humans and wildlife. Throughout, we place particular
to be a sensitive and rapid developmental model to elucidate effects
of hydrocarbon mixtures.
Key Words: PAHs; DLCs; developmental toxicity; synergism;
AHR; CYP1A; risk assessment.
Polycyclic aromatic hydrocarbons (PAHs) are environmental
contaminants derived from petroleum and generated by the
incomplete combustion of organic compounds. As a class,
PAHs enter the environment through natural sources such as oil
seeps and forest fires and through a variety of anthropogenic
activities. These include the burning of fossil fuels and wood,
production of coke and charcoal, metal smelting, petroleum
refining, and petroleum spills (Douben, 2003; Latimer and
Zheng, 2003). The United Nations Environmental Program has
identified PAHs as a class of pollutants of global concern as
they pose a real and potential risk to human health and the
environment. In the United States, temporal trends suggest that
environmental contamination by PAHs is steadily climbing,
tracking patterns in urbanization, and combustion of fossil fuels,
particularly by automobiles (Van Metre and Mahler, 2005; Van
Metre et al., 2000). This is in contrast to reports of downward
trends of environmental contamination by the planar halogenated
aromatic hydrocarbons, i.e., the dioxin-like compounds (DLCs).
PAHscompriseanimportantclass ofcontaminants that invariably
occur as mixtures at United States Environmental Protection
Agency Superfund hazardous waste sites and other polluted
environments. Widespread PAH contamination at these sites,
coupled with their known (i.e., carcinogenic) and suspected
Disease Registry’s priority list (ATSDR, 1997, 2005). These data
suggest that PAHs pose a significant threat to human health and
exposure to the less prevalent DLCs (Van Metre and Mahler,
2005). Nonetheless, with the exception of their well-established
role in carcinogenesis, the adverse developmental effects of PAH
exposure in sensitive fish species have received attention only
recently (Barron et al., 2004; Billiard et al., 2006; Hawkins et al.,
Risk Assessment Models
While it is clear that some PAHs are toxic and that emissions
are increasing, risk assessment of this class of environmental
contaminants can be complicated. Because most environmental
exposures are from complex mixtures of PAHs, we focus more
on the issues and impacts associated with combinations of
PAHs and less so on individual compounds. Current models
used to estimate risks of mixtures assume dose or concentration
1To whom correspondence should be addressed at Health Canada, Health
Products and Food Branch, Bureau of Chemical Safety, 251 Sir Frederick
Banting Driveway, Postal Locator 2204C, Ottawa, Ontario K1A OL2, Canada.
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additivity (Barron et al., 2004; Basu et al., 2001). Historically,
additivity borrows from the concept of toxic equivalency factors
(TEFs) which were developed to assess the risk of DLC
mixtures to humans and wildlife with similar modes of action
(Ontario Ministry of the Environment, 1985; Safe, 1993; Van
den Berg et al., 1998). For additivity, exposure-response curves
for chemically related compounds should be parallel and exhibit
a similar range of response and efficacy (i.e., similar maxima).
Nevertheless, these prerequisites are not always met, even for
DLC congeners (Parrott et al., 1995; Van den Berg et al., 2006).
One criteria for applying the TEF approach to DLCs is the
assumption that effects are mediated by binding to the aryl
hydrocarbon receptor (AHR) (Van den Berg et al., 1998, 2006),
nuclear translocation and heterodimerization with the aryl
to dioxin-responsive elements (DREs) to regulate expression of
cytochrome P450 (CYP) enzymes, such as CYP1A1/2 and
CYP1B1, the aryl hydrocarbon receptor repressor (AHRR), and
phase II conjugation enzymes (Nebert and Dalton, 2006).
whether toxic equivalencies (TEQ ¼
concentrations)) can or should be used to predict the toxicity of
complex mixtures. PAHs do not meet the criteria for TEFs
developed for DLCs (i.e., structural similarity to polychlorinated
dibenzo-p-dioxins/dibenzo furans (PCDD/Fs), persistent, and
bioaccumulative) (Van den Berg et al., 1998, 2006). Further,
studies reviewed herein will demonstrate that the role of the
AHR in PAH toxicity is very complex. An exception where
a TEQ approach might be justified would be for carcinogenic
PAHs which require CYP-mediated bioactivation or toxification
(Incardona et al., 2006). However, PAH-type AHR agonists are
rapidly metabolized by the enzymes that they induce, including
CYPs such as CYP1A. Certainly, using the induction of CYP1A
catalytic activity as a surrogate toxic end point to derive TEFs for
PAH mixtures is problematic if not altogether flawed (Whyte
etal.,2000).While CYP1Ainductionisnota toxic endpoint,the
their toxicity to developing organisms (Guiney et al., 1997; Heid
et al., 2001; Safe, 1990, 1993). Thus, measurement of CYP1A
activity as a biomarker of exposure is oftentimes used as a
surrogate biomarker of effect or as an ‘‘early warning system’’
(Payne et al., 1987). In the strictest sense and as with any
biomarker of exposure, CYP1A induction is not necessarily
indicative of PAH toxic action. As we caution in this review,
PAHs that cause developmental (i.e., noncarcinogenic) toxicity.
while others may exert their toxicity independent of this activity
and may actually have their toxicity reduced by this activity.
The idea for this review began with our observations that
some PAHs mimicked dioxin, causing developmental and
P(TEF 3 measured
cardiovascular toxicity in the exquisitely sensitive fish early life
stages (ELS) model. However, these PAH effects could not be
predicted from AHR binding or CYP1A-induction potencies.
Furthermore, studies indicated that unlike DLCs, the potencies
of PAHs in mixtures were not always additive and that the
TEQ approach fell short for these combinations. In this review,
we present evidence to suggest that multiple mechanisms
underscore PAH developmental effects. It follows then that the
toxic potency of a given PAH mixture will be driven by its
composition (including both AHR- and non-AHR–binding
PAHs). Thus, current models for assessing the risk of PAHs
during fish and vertebrate development are inadequate, and
a new paradigm for characterizing hazard is required.
In making this case, we explore alternate mechanisms of
PAH toxicity as a critical step toward reducing the uncertainty
Institute of Environmental Health Sciences workshop (2002) on
the role of environmental agents as important risk factors for
cardiovascular disease emphasized the advantages of using
alternative, nonmammalian models for the study of chemically
induced cardiovascular malformations. While sensitivity varies
of DLCs and some PAHs are remarkably similar, likely because
the functional properties of the AHR and developmental sig-
a transparent chorion which allows for direct, noninvasive
observation during embryogenesis. These characteristics make
mechanisms of chemically induced vascular defects in verte-
brates (Antkiewicz et al., 2006; Carney et al., 2006b; Goldstone
and Stegeman, 2006). Topics discussed herein include: (1) car-
diovascular toxicity of AHR agonists during development, (2)
chemically resistant populations, (3) altered CYP1A activity
studies, (4) possible mechanisms of PAH developmental tox-
icity, and (5) implications for risk assessment.
DLCs AND PAHs ARE CARDIOTERATOGENS DURING
EARLY VERTEBRATE DEVELOPMENT
In vertebrates, the cardiovascular system is one of the most
sensitive targets of dioxin action. Cardiac dysfunction in fish
ELSs is also a recently discovered subset of effects following
chronic PAH exposure (Billiard et al., 2006; Incardona et al.,
2004, 2006; Wassenberg and Di Giulio, 2004b; Wassenberg
et al., 2005). For that reason, mechanisms underlying PAH
teratogenicity are often compared to those of DLCs (Billiard
et al., 1999, 2006; Incardona et al., 2005). The role of the AHR
pathway in mediating dioxin toxicity has been the subject of
extensive study, and a complete review is beyond the scope of
this manuscript. However, it is instructive to briefly discuss the
role of the AHR pathway in mediating teratogenesis by DLCs
BILLIARD ET AL.
in vertebrates because of the apparent similarities in toxicity of
The Cardiovasculature is a Sensitive Target of Dioxin
Although the majority of studies have examined cardiac
effects of dioxin exposure in birds and fish, a recent study
showed that the fetal murine heart is especially vulnerable to
dioxin toxicity when challenged in utero (Thackaberry et al.,
2005b). Specifically, reduced heart somatic index and in-
hibition of cardiomyocyte proliferation were observed. In the
same study, postnatal or lactational exposure of pups exposed
in utero showed increased heart-to-body weight and decreased
heart rate (i.e., bradycardia; Thackaberry et al., 2005b). This
proposed compensatory cardiac hypertrophy was also observed
in an earlier study in dioxin-treated mice (Lin et al., 2001).
Similarly, altered heart size and decreased myocyte number or
growth have been demonstrated in fish and birds—in these
species dioxin exposure also increases vascular apoptosis
and permeability (Antkiewicz et al., 2005; Guiney et al., 2000;
Hill et al., 2004; Hornung et al., 1999; Invitski et al., 2001;
Rifkind, 2006; Toomey et al., 2001; Walker and Catron,
2000). Thus, the suite of dioxin-induced cardiovascular defects
are fairly similar across vertebrate taxa (Thackaberry et al.,
2005b). These effects are supported mechanistically by micro-
array studies in hearts of dioxin-exposed zebra fish embryos
(Danio rerio) and fetal mice where expression of cardiac-
specific genes, such as those encoding for cell cycle regula-
tion and proliferation, are downregulated following exposure
(Carney et al., 2006a; Thackaberry et al., 2005a). Cardiovascular
gene expression profiles of dioxin-exposed zebra fish embryos
were also consistent with cardiomyopathy observed in these
fish (Goldstone and Stegeman, 2006; Handley-Goldstone et al.,
Fish appear to be the most sensitive of all vertebrate classes
to dioxin-induced teratogenicity (Peterson et al., 1993). The
relative sensitivity of fish to dioxin-induced ELS mortality
(Elonen et al., 1998) is in the range of the most sensitive avian
and mammalian species. Accordingly, our understanding of
mechanisms underlying dioxin toxicity has been advanced
using fish ELS as model developmental systems (Goldstone
and Stegeman, 2006). Exposure of salmonid ELS to dioxin
causes an overt toxicity syndrome called blue sac disease
(BSD), characterized by increased rates of mortality, pericar-
dial and yolk sac edemas, regional ischemia, subcutaneous
hemorrhages, craniofacial deformities, and arrested growth
(Spitsbergen et al., 1991). Signs of toxicity closely resembling
BSD have been described in several fish species including
rainbow trout (Oncorhynchus mykiss), zebra fish, Japanese
medaka (Oryzias latipes), and killifish (Fundulus heteroclitus)
exposed to DLC-type AHR agonists (Chen and Cooper, 1999;
Elonen et al., 1998; Helder, 1981; Toomey et al., 2001; Walker
and Peterson, 1991; Wannemacher et al., 1992). Several
studies have shown that fish embryonic vasculature is the
primary physiologic target for dioxin-induced embryotoxicity
(Belair et al., 2001; Dong et al., 2002; Guiney et al., 1997;
Henry et al., 1997; Hornung et al., 1999). In embryonic fish,
pericardial edema is secondary to inhibition of angiogenesis,
reduced blood flow, and circulatory failure (Bello et al., 2004;
Hornung et al., 1999). In birds, inhibition of blood vessel
formation (Ivnitski et al., 2001) is associated with both reduced
expression and sensitivity to vascular endothelial growth factor
(VEGF) (Ivnitski-Steele et al., 2004, 2005). Similar toxicolog-
ical profiles across taxa are a strong argument that embryonic
stages of zebra fish are an appropriate model of effects in
For a more comprehensive review of dioxin-induced
cardiovascular toxicity, the reader is referred to Carney et al.
(2006b) and Goldstone and Stegeman (2006).
Role of the AHR Pathway in Developmental Toxicity of DLCs
signaling pathway can disrupt normal cardiovascular develop-
ment during the ELS of vertebrates (Carney et al., 2006b;
Goldstone and Stegeman, 2006). There are well-established
positive relationships among the AHR-binding affinity of
DLCs, their potency for CYP1A induction, and their toxicity to
vertebrate models, including mammals, birds, and fishes
(Guiney et al., 1997; Heid et al., 2001; Safe, 1990, 1993). In
mammals, activated AHR is mandatory for acute and
teratogenic response to dioxin as evidenced by the resistance
of receptor null mice to toxicity (Fernandez-Salguero et al.,
1996; Gonzalez and Fernandez-Salguero, 1998; Mimura et al.,
1997; Peters et al., 1999; Rifkind 2006; Walisser et al., 2004).
Genetic ablation of the Ahr is also associated with aberrant
vascular patterning or phenotypes in nontreated mice, under-
scoring the importance of this pathway in normal vascular
development (Lahvis et al., 2000; Lund et al., 2003; McMillan
and Bradfield, 2007; Rifkind, 2006; Walisser et al., 2004).
Recent advances using morpholino antisense technology
have confirmed the obligatory role of the AHR pathway in
dioxin-induced ELS toxicity by embryonic knockdown of
target gene expression in zebra fish (see below). However, there
are some distinct differences compared to mammalian AHR and
unique challenges, particularly when it comes to elucidating
a physiological or developmental role for this pathway in fish.
Unlike mammals which have only one AHR (AHR1), fish
have several AHR isoforms belonging to two clades (AHR1
and AHR2) which originated from an independent duplication
event prior to divergence of teleosts and separation of piscine
and mammalian lines (Hahn et al., 1997; Karchner et al.,
2005). Thus, it has been suggested that AHR function in fish is
shared or divided among several AHRs (Hahn, 2001; Karchner
et al., 2005). Although zebra fish AHR1(A) more closely
resembles the single mammalian AHR protein (Karchner et al.,
1999), this receptor does not bind DLC (dioxin) or PAH
Perturbation or atypical modulation of the AHR
PAH DEVELOPMENTAL TOXICITY
benzo-(a)pyrene (BaP) ligands and, furthermore, exhibits
negligible transactivation activity in vitro (Andreasen et al.,
2002b; Karchner et al., 2005). However, recent data suggest that
PAHs can act via AHR1A to induce CYP1A in vivo; i.e., knock
down of zebra fish embryo ahr1a partially protected against
pyrene-induced toxicity and reduced induction of CYP1A in
hepatic tissue (Incardona et al., 2006). Additionally, a functional
AHR protein, AHR1B, has been characterized and is expressed
early in zebra fish development, prompting speculation that it
plays more of an endogenous or physiological role in ELS of
fish (Karchner et al., 2005). On the other hand, AHR2 protein
categorically mediates toxic outcome following dioxin insult in
fish (Andreasen et al., 2002a; Antkiewicz et al., 2006; Karchner
et al., 2005; Prasch et al., 2003).
Using morpholinos to knock down translation of AHR2
protein in developing zebra fish prevents CYP1A expression
and toxicity typical of dioxin exposure (Carney et al., 2004; as
reviewed in Carney et al., 2006b). We have also shown that
ahr2 knockdown protects zebra fish from PCB126, a potent
dioxin-like AHR agonist (Billiard et al., 2006). Along with
AHR2, ARNT1 protein is obligatory for cardiac toxicity in
dioxin-exposed zebra fish embryos (Antkiewicz et al., 2006;
Prasch et al., 2006). AHR?ARNT cooperativity in develop-
mental signaling and dioxin toxicity has also been shown in
mutant mice homozygous for a lower expressing, or hypo-
morphic, Arnt allele (Walisser et al., 2004). Because AHR2
morphants are normal compared to noninjected embryos, one
might conclude that there is no functional role of this receptor
during fish development (Billiard et al., 2006; Dong et al.,
2004; Incardona et al., 2006; Prasch et al., 2003; Teraoka et al.,
2003). However, the fact that an AHR2 morphant phenotype
has not been observed in zebra fish studies could be more
a consequence of rapid degradation of the antisense morpho-
lino in vivo or that residual AHR expression in knockdown
embryos is capable of supporting normal development (Carney
et al., 2006b; Karchner et al., 2005).
showing that dioxin toxicity to mammals is AHR dependent,
the role of CYP1A induction and activity in toxicity is
ambiguous for vertebrates in general. Roles for CYP-mediated
toxicity, including generation of reactive oxygen species (ROS)
and/or reactive metabolites, are discussed in detail below (see
‘‘Possible mechanisms of developmental PAH toxicity’’
section). Of the CYP enzymes, CYP1A is maximally induced
following AHR activation by environmental contaminants in
fish as well as in mammals (Goldstone and Stegeman, 2006;
Jo ¨nsson et al., 2007a,b; Timme-Laragy et al., 2007; Wang
et al., 2006). For these reasons, we gave most weight to the
CYP1A subfamily in this review. However, by definition,
AHR-mediated toxicity depends on one or more of the AHR-
regulated gene targets. Thus, the toxicological implication of
other CYP1s, particularly for rapidly metabolized PAHs, is
also discussed where appropriate.
While fish studies are consistent with the literature
What is the toxicological significance of CYP1A activity?
Recent evidence calls for a review of the classically held
paradigm that CYP1A induction is simply a biomarker of
exposure or adverse effect (Rifkind, 2006). While induction of
CYP1A in mammals is tightly linked to AHR activation, it is
not necessarily predictive of the expression of other, and
currently unknown, AHR-regulated genes that mediate toxic
response to DLCs (Brunnberg et al., 2006). In mammalian
studies, there is no evidence that cyp1a1 or 1b1 null mice are
protected against dioxin-induced developmental toxicity. In
contrast, CYP1A2 sequestered dioxin in hepatic tissue of
pregnant mice or dams, thus protecting against embryotoxicity
and teratogenesis (Dragin et al., 2006; Ma and Lu, 2007).
Guiney et al. (1997) provided the first mechanistic evidence
in fish that CYP1A could be the downstream target of dioxin-
activated AHR and that this induction might mediate de-
velopmental toxicity. Dose-response curves for CYP1A
induction in the vascular endothelium of lake trout sac fry
were correlated with dose-response curves for mortality, and
overt signs of toxicity were preceded by CYP1A induction
(Guiney et al., 2000). Dioxin-induced apoptosis was also
correlated with embryotoxicity and colocalized with CYP1A
expression in medaka and killifish embryos (Cantrell et al.,
1998; Toomey et al., 2001). Recently, Jo ¨nsson et al. (2007b)
demonstrated similar dose-response curves for CYP1A, 1B1,
1C1, and CYP1C2 expression and abnormalities (e.g.,
pericardial edema) in zebra fish embryos exposed to PCB126.
Most studies report that chemical inhibition of CYP1A
activity, or morpholino knock down of CYP1A protein, is
protective in the case of DLCs (Billiard et al., 2006; Cantrell
et al., 1996; Dong et al., 2004; Teraoka et al., 2003;
Wassenberg and Di Giulio, 2004b). As a result, it has been
hypothesized that CYP1A enzyme activity causes develop-
mental toxicity of DLCs (Billiard et al., 2006; Cantrell et al.,
1996; Dong et al., 2004; Teraoka et al., 2003; Wassenberg and
Di Giulio, 2004b). However, several recent studies suggest the
opposite, that dioxin activation of the AHR pathway causes
teratogenicity by a CYP1A-independent mechanism in zebra
fish (Antkiewicz et al., 2006; Carney et al., 2004). Regarding
the two morpholino studies that were at odds with respect to
CYP1A-mediated teratogenesis, Carney et al. (2004) suggested
that signs of TCDD toxicity are not manifested until 96 hours
post-fertilization (hpf) in zebra fish and that this explained the
‘‘lack of response’’ at 72 hpf observed by Teraoka et al. (2003).
While AHR activation seems to be a prerequisite for dioxin
toxicity, the identity of the trigger gene or genes regulating
teratogenicity remains unknown (Carney et al., 2004). This
brings us back to the original question first proposed almost
two decades ago: What gene targets downstream of the AHR/
ARNT pathway elicit dioxin toxicity? This question, still
unanswered, is the target of a concerted effort currently
underway in the field of toxicogenomics (e.g., Antkiewicz
et al., 2005; Carney et al., 2006a; Goldstone and Stegeman,
2006; Handley-Goldstone et al., 2005).
BILLIARD ET AL.
PAHs Mimic Developmental and Cardiotoxic Effects of
Dioxin but Mixture Toxicity is Nonadditive
As with dioxin, chronic exposure of ELS of fish species to
some PAHs leads to CYP1A induction, edemas, hemorrhaging,
cardiac dysfunction, mutations, heritable changes in progeny,
morphological deformities, neuronal cell death, anemia, re-
duced growth, and increased mortality rates (Barron et al.,
2003, 2004; Billiard et al., 2006; Brinkworth et al., 2003;
Colavecchia et al., 2004, 2006; Incardona et al., 2004,
2005; Marty et al., 1997; Wassenberg and Di Giulio, 2004a;
Wassenberg et al., 2005). However, evidence for PAH-induced
developmental and cardiovascular toxicity is not restricted to
fish. The Nebert laboratory first demonstrated that BaP was
teratogenic to mice (as reviewed in Nebert, 1989). Since then,
several studies have demonstrated that this class of compounds
are developmental toxicants in both humans and laboratory
species (as reviewed by Ramesh et al., 2004). Crude oil has
also been shown to induce developmental effects in rats and
birds, effects that were attributed to the PAH fraction of
petroleum hydrocarbonmixtures (Feuston etal.,1997; Hoffman
and Gay, 1981). In their studies with ducks, Hoffman and Gay
(1981) observed that addition of individual PAHs to crude oil
greatly enhanced its embryotoxic effects. Thus, additive or
synergistic effects could account for the developmental toxicity
of crude oil even when individual PAHs are present at low
concentrations (Hoffman and Gay, 1981). More recently,
prenatal exposure to of rats and human, PAHs has been
associated with adverse effects on fetal growth and/or the
developing vasculature (Choi et al., 2006; Sanyal and Li, 2007).
Further, studies have indicated that PAHs are a significant risk
factor for pathogenesis of cardiovascular disease in humans (as
reviewed in Korashy and El-Kadi, 2006).
Because some PAHs are agonists for the AHR and
moderately induce CYP1A (Billiard et al., 2002; Table 1),
a commonly held assumption is that PAH teratogenicity is also
AHR mediated. If this hypothesis is correct, we would expect
that those compounds with low or negligible binding affinity to
the AHR would be less toxic than those PAHs with moderate to
high affinity for the AHR protein. In addition, toxic interactions
among PAH-type AHR agonists should be strictly additive, and
the potency of complex PAH mixtures could be estimated
using the TEQ approach for dioxin-like AHR agonists (Walker
et al., 1996; Zabel et al., 1995).
However, data regarding the role of the AHR pathway in
causing PAH cardiotoxicity during development do not support
this hypothesis. A recent study suggests that cardiovascular
toxicity caused by exposure of fish ELS to weathered crude oil
is independent of both AHR1 (AHR1A) and AHR2 (Incardona
et al., 2005). Further, CYP1A activity is protective against
toxicity caused by exposure to some PAHs. Another confound-
ing factor is that PAH-type AHR agonists and CYP1A
inhibitors typically co-occur in environmental mixtures. We
have observed nonadditive, synergistic cardiovascular interac-
tions between PAH-type AHR agonists and CYP1A inhibitors.
Our recent work with a sediment extract from a PAH-polluted
Superfund site also suggests that the interactive effects of
PAH mixtures may drive the developmental toxicity of these
compounds (Wassenberg and Di Giulio, 2004b; Wassenberg
et al., 2005). This is a departure from responses observed with
dioxin-like AHR agonists where CYP1A inhibition generally
reduced toxicity or had no effect (Cantrell et al., 1996; Dong
et al., 2002).
Although there is a strong, positive relationship between the
ability of PAHs to bind to the AHR and to induce CYP1A
(Billiard et al., 2002; Bols et al., 1999; Piskorska-Pliszczynska
et al., 1986), the relationships between AHR binding and
CYP1A activity and their roles in PAH toxicity are far less
clear. The most potent AHR agonists are not necessarily toxic
(Billiard, 2002; Billiard et al., 2002), and PAHs do not always
behave in an additive fashion (Basu et al., 2001; Hodson et al.,
2007; Wassenberg and Di Giulio, 2004b; Wassenberg et al.,
2005). In studies with rainbow trout embryos, the rank order of
potency of PAHs for CYP1A induction (Basu et al., 2001) did
not predict the potency of these same compounds for causing
BSD (Billiard, 2002). Surprisingly, the most potent CYP1A in-
ducer, benzo[k]fluoranthene (BkF), was nontoxic to trout ELS.
Similarly, Wassenberg and Di Giulio (2004b) demonstrated
that BaP by itself was virtually nontoxic to larval killifish,
despite its high potency for inducing CYP1A. In contrast,
alkyl-PAHs such as retene (7-isopropyl-1-methylphenanthrene)
are weaker AHR agonists than B[k]F but far more toxic to fish
ELS (Billiard et al., 1999). Overall, AHR-binding affinity
underestimates PAH potency for causing dioxin-like terato-
genesis to fish ELS (Barron et al., 2004; Billiard et al., 2002).
Thus, it appears that potency of PAHs for binding to the AHR
and activating the cyp1a gene alone is not enough to predict
toxicity and that CYP1A induction equivalency factors for
PAHs are neither additive nor appropriate for estimating
toxicities of mixtures of PAHs (Basu et al., 2001).
A key difference between the embryotoxicities of PAHs and
DLCs is the rapid metabolism and excretion of PAHs (Table 1),
a process based largely on CYP1 enzyme activity (Brown
et al., 2002; Niimi and Palazzo, 1986). An altered capacity to
metabolize and excrete these compounds could be a factor in
Key Pharmacokinetic and Toxicological Differences between
Planar Halogenated DLCs and PAHs
Potent AHR agonists
Potent CYP1A inducers
Poor substrates for CYP1A
Generate ROS by uncoupling CYP1A
Weak to moderate AHR agonists
Weak to moderate CYP1A inducers
Excellent substrates for CYP1A
Metabolized to reactive intermediates
Some are cardiotoxic
PAH DEVELOPMENTAL TOXICITY
resistance to PAH toxicity that has been observed in some
populations of fish chronically exposed to contaminated
environments. In the following section, we summarize our
studies of a Fundulus population inhabiting an estuary polluted
by PAHs from an adjacent U.S. Superfund site. This population
has adapted to pollutants present in its environment over many
generations of exposure, providing a case study in evolutionary
ecotoxicology and a powerful model for understanding de-
velopmental toxicity through examination of mechanisms
INSIGHTS GAINED FROM CHEMICALLY RESISTANT
Multiple populations of fish have been discovered inhabiting
sites highly contaminated with DLCs, PAHs, and mixtures
thereof (reviewed by Wirgin and Waldman, 2004). These
populations are defined ‘‘resistant’’ by the criteria that:
? They survive chronic exposure to levels of contaminants
in their habitats that would prove fatal to most naı ¨ve fish and
? They are resistant to the toxic effects of laboratory
exposure to DLCs and PAHs (e.g., Meyer et al., 2002, 2003a;
Meyer and Di Giulio, 2003; Nacci et al., 1999, 2002; Ownby
et al., 2002; Prince and Cooper, 1995; Roark et al., 2005;
Wirgin and Waldman, 2004).
Several groups of investigators have asked what biochemical
or molecular changes are responsible for the resistant
phenotype of these fish (reviewed by Hahn, 1998, 2001;
Wirgin and Waldman, 1998, 2004). These studies are relevant
to this review in that a better understanding of mechanisms of
adaptation to AHR agonists should yield insight into the basic
mechanisms of toxicity of AHR agonists. A common
characteristic of the populations studied has been altered
CYP1A mRNA and protein expression and activity (Arzuaga
and Elskus, 2002; Bello et al., 2001; Courtenay et al., 1999;
Elskus et al., 1999; Meyer et al., 2002, 2003b). While ‘‘basal’’
or ‘‘constitutive’’ expression is perhaps a misleading term to
describe activity in chronically exposed populations, CYP1A
expression in wild fish from contaminated sites is only two- to
threefold higher than in reference site fish, rather than markedly
higher as would be expected based on the level of contaminant
exposure. More strikingly, sensitivity to CYP1A induction is
markedly reduced or undetectable in fish from these resistant
populations after laboratory exposure to prototypical inducers.
Furthermore, lack of CYP1A induction has been observed in
several other populations that have not been directly tested for
toxicity resistance (Brammell et al., 2004; Fo ¨rlin and Celander,
1995; Monosson and Stegeman, 1991), suggesting that
recalcitrance to CYP1A induction is common to fish
populations chronically exposed to AHR agonists.
These results, consistent across several species of fish at
multiple sites, suggest that low CYP1A inducibility might
be an adaptive response, a hypothesis advanced by several
researchers (Elskus, 2001; Elskus et al., 1999; Nacci et al.,
2002). For example, extracts of sediments from the contam-
inated Elizabeth River are highly teratogenic to killifish
embryos reared from parents from relatively uncontaminated
reference sites. The predominant effects of exposures to these
extracts are cardiovascular defects. In contrast, embryos from
parents of killifish indigenous to the Elizabeth River site are
remarkably resistantto these
suggested that downregulation of AHR-mediated gene tran-
scription and upregulation of antioxidant defenses were
associated with the resistance exhibited by the Elizabeth River
population (Meyer et al., 2002, 2003a). In a related study,
Nacci et al. (2002) showed that the Elizabeth River killifish
exhibited low CYP1A inducibility and produced few DNA
adducts after laboratory exposure to BaP.
These studies provide correlative evidence suggesting that
low CYP1A inducibility is protective in the context of ELS
toxicity. The hypothesis that high CYP1A activity mediates
toxicity of high concentrations of DLCs and PAHs also
appeared mechanistically attractive based on the capacity of
CYP1A to activate PAHs to reactive metabolites and generate
ROS in the presence of DLCs, as discussed elsewhere in this
article. Nonetheless, while the correlation between resistance
and altered CYP1A expression is widespread, making a causal
link theoretically plausible, no cause-and-effect relationship
has been demonstrated. Furthermore, one of the resistant
populations exhibited refractory CYP1A induction in response
to exposure to DLCs but not PAHs (Courtenay et al., 1999), and
in hepatocytes from another fish population, CYP1A was highly
refractory to induction by DLCs, but only less refractory to
induction by PAHs (Bello et al., 2001). These results suggest
that the modes of toxicity of DLCs and PAHs are not identical.
There is currently no direct evidence to support the
hypothesis that a lack of CYP1A inducibility in resistant fish
populations is protective in the context of exposure to high
levels of PAHs. For killifish embryos from a reference site
(Kings Creek), we observed no correlation between CYP1A
inducibility and their subsequent survival when exposed to
Elizabeth River environmental samples contaminated with
lethal concentrations of PAHs (Fig. 1, adapted from Meyer
et al., 2002). Additionally, killifish from this site on the
Elizabeth River were highly resistant to the toxicity of those
sediments, and resistance was more heritable than the lack of
CYP1A inducibility (Meyer et al., 2002, 2003b). These data
suggest that co-occurrence of low inducibility and chemical
resistance in Elizabeth River fish does not reflect a causal
relationship, i.e., resistance was likely conferred by some other
mechanism or combination of mechanisms.
In summary, there is a common co-occurrence of poor
CYP1A inducibility and resistance to embryotoxicity in wild
fish populations inhabiting sites highly contaminated with
AHR agonists. However, current evidence from experiments
with indigenous fish populations does not support the
BILLIARD ET AL.
hypothesis that lack of CYP1A induction confers protection in
PAH-contaminated environments. It may be that altered
expression of other AHR-regulated genes provides protection
to these resistant fish populations. In this case, lack of CYP1A
inducibility may not be a toxicologically protective adaptation
but rather a marker of AHR pathway inhibition. While
regulation of other AHR pathway genes has thus far received
less attention in resistant fish populations, mRNA inducibility
of at least one additional AHR-regulated gene, AHRR, is
diminished in at least two resistant populations (Meyer et al.,
2003b; Roy et al., 2006). mRNA expression levels in the
absence of a laboratory exposure were not significantly
different from those observed in a reference population in
any of three tested resistant populations (Karchner et al., 2002;
Meyer et al., 2003b; Roy et al., 2006).
Wild populations of fish resistant to PAH-induced embry-
otoxicity are a complicated model system. The resistant
phenotype could reflect a number of changes across more than
one genetic or biochemical parameter. However, the value of
studying such populations is that they demonstrate environ-
mentally relevant responses to DLC and PAH contamination.
Whatever the mechanisms of adaptation that operate in these
resistant populations, they reflect processes that unequivocally
protect against realistic, multigenerational environmental
exposures to PAHs. Currently, the evidence obtained from
these populations suggests that lack of CYP1A inducibility,
while correlated with PAH resistance, is not responsible for
PAH resistance. A protective role for altered an AHR pathway,
however, appears likely. One possibility is that epigenetic
alterations in AHR pathway gene expression are involved. This
phenomenon has been observed in various mammalian systems
(Jin et al., 2004; Mulero-Navarro et al., 2006; Schnekenburger
et al., 2007; Takahashi et al., 1998; Tokizane et al., 2005),
although evidence so far does not support a role for promoter
region methylation in preventing CYP1A inducibility in
resistant fish populations (Arzuaga et al., 2004; Timme-Laragy
et al., 2005).
Finally, contaminated sites also contain numerous other
known and unknown chemical contaminants which may further
complicate observed synergistic cardiovascular interactions
between PAH-type AHR agonists and CYP1A inhibitors
INSIGHTS GAINED FROM ALTERED CYP1A ACTIVITY
To directly test the hypothesis that lowered CYP1A activity
is protective of PAH-derived toxicity, pharmacological agents,
mRNA antisense morpholinos, and CYP1A1 knockout mice
have been used to modulate or prevent CYP1A activity in
animals exposed to PAHs. Experiments with killifish embryos
coexposed to extracts of PAH-contaminated sediments or pure
PAHs with a suite of CYP1A inhibitors demonstrated that
CYP1A inhibition increased the incidence of heart deformities
rather than decreasing embryotoxicity. This suggested that
many AHR agonists and CYP1A inhibitors together were
synergistic, independent of the mechanism of action or
chemical structure of the inhibitor (Wassenberg and Di Giulio,
2004a,b; Wassenberg et al., 2005).
An inherent limitation of studies with CYP1A inhibitors is
that they may have other activities in addition to the inhibition
of CYP1A. For example, a-naphthoflavone (ANF) has been
characterized as a CYP1A inhibitor, an AHR antagonist and an
AHR agonist (Merchant and Safe, 1995; Merchant et al., 1992;
Testa and Jenner, 1981). Piperonyl butoxide is a nonspecific
P450 inhibitor, so in vivo it is likely to inhibit other enzymes in
addition to CYP1A (Franklin, 1977; Murray and Reidy, 1990;
Testa and Jenner, 1981). A further complication is that the
nature of these interactions may change with exposure or dose.
Inhibitors will have their own unique subset of sublethal and
lethal modes of action which can interact with acute and
chronic mechanisms of PAH toxicity. For example, rainbow
trout embryos exposed to a range of retene concentrations
showed an exposure-dependent increase in the prevalence and
severity of signs of BSD (Billiard et al., 1999; Brinkworth
et al., 2003). As with killifish, a coexposure to low
concentrations of ANF greatly enhanced the toxicity of retene.
In contrast, at higher concentrations of ANF, the toxicity of all
concentrations of retene decreased so that rates of mortality and
BSD were not different from control (Hodson et al., 2007).
This exposure-dependent synergism and antagonism of toxicity
by one inhibitor of CYP1A activity may be related to exposure-
dependent changes in the rate of metabolism and tissue
We have also compared toxicity of PAHs with and without
CYP1A inhibitors to wild type versus ahr2a or cyp1a zebra
fish morphants. Similar to the results with chemical inhibitors
induction in embryos and posthatch survival in Elizabeth River sediment pore
water for King’s Creek larvae (adapted from Meyer et al., 2002). Larvae
previously characterized for EROD induction after exposure to the inducer
3-methylcholanthrene as embryos were exposed to a 1:10 dilution of Elizabeth
River sediment pore water in individual glass scintillation vials (i.e., one part
sediment pore water, nine parts clean artificial sea water) 2 days posthatch.
Spearman’s rank order correlation coefficient: r ¼ ? 0.0008, p ¼ 0.9228.
Lack of correlation between Ethoxyresorufin-O-deethylase (EROD)
PAH DEVELOPMENTAL TOXICITY
in killifish embryos, knock down of CYP1A protein expression
by morpholino injection increased the toxicity of a PAH-type
AHR agonist (BNF) and the toxicity of a mixture of BNF with
the CYP1A inhibitor, ANF (Billiard et al., 2006). In contrast,
cyp1b1 knockdown did not alter synergistic developmental
toxicity of PAHs in zebra fish (Timme-Laragy et al., 2007).
Homozygous CYP1A1 knockout mice showed less liver
damage and survived the acute effects of injection of the PAH
BaP for 3 days longer than did those that were heterozygous for
CYP1A1 (Uno et al., 2001). However, these CYP1A1
knockout mice also showed fourfold higher levels of BaP-
DNA adducts than did those heterozygous for CYP1A1. That
is, the acute lethality of BaP was reduced by a lack of CYP1A1
whereas genotoxicity was actually increased (Uno et al., 2001).
In contrast, but more in parallel with what we have observed in
fish, in a more recent study this group found that BaP
administered in the diet caused lethality in CYP1A1 knockout
mice at a dose that was not lethal to CYP1A1-expressing mice
(Uno et al., 2004). They proposed that CYP1A1 activity is
critical for the detoxification of orally administered BaP in
mice, rather than enhancing the toxicity of BaP, as they
While they did not look at effects of PAH toxicity, Dubey
et al. (2003, 2005) found that CYP1A1 activity was
cardioprotective in human cardiac fibroblast and smooth
muscle cells because it catalyzes the conversion of estradiol
to anti-mitogenic metabolites that act independently of the
estrogen receptor. Treatment of these cells with CYP1A1
inhibitors removed the cardioprotective effects of estradiol.
Further exploration of cardioprotective effects of CYP1A1
activity on estrogen during development may be fruitful.
The cooccurrence of AHR agonists and CYP1A inhibitors is
typical of PAH-contaminated environmental mixtures. Fluo-
ranthene (FL) and the heterocyclic PAHs, carbazole (CB) and
dibenzothiophene (DBT), are common components of complex
mixtures of PAHs (e.g., coal tar, crude oil). Each has been
characterized in vivo and in vitro as a CYP1A inhibitor, and
each increases embryotoxicity when combined with PAH-type
AHR agonists(Wassenbergand DiGiulio2004a,b;Wassenberg
et al., 2005; Willett et al., 1998, 2001). Hence, synergistic
interactions of these compounds with other PAHs may be
responsible for the observed embryotoxicity of complex
mixtures. FL, CB, and DBT have only recently been char-
acterized as CYP1A inhibitors, and their potential interactions
with other oxidative enzymes or their ability to interfere with
AHR signaling is not yet known (Wassenberg et al., 2005;
Willett et al., 1998, 2001). The synergy that we have observed
when assessing risks posed by exposure to PAH contamination
and also provides additional insights concerning mechanisms
underlying greater-than-additive toxicities caused by exposure
to PAH mixtures.
While the role of CYP1A activity has not been clearly
elucidated in the toxicity of either DLCs or PAHs, there is
a clear difference between these two classes of compounds
when CYP1A is inhibited during development. Lowered
CYP1A activity or knock down of CYP1A protein in embryos
dosed with DLCs resulted in either reduced toxicity or had no
change in toxicity (Billiard et al., 2006; Cantrell et al., 1996;
Carney et al., 2004; Dong et al., 2004; Teraoka et al., 2003;
Wassenberg and Di Giulio, 2004b). In contrast for PAHs,
inhibition of CYP1A was manifested as either an increase or
decrease in fish embryotoxicity, depending on ANF exposure
concentration (Hodson et al., 2007; Wassenberg and Di Giulio,
2004b). However, cyp1a knockdown in zebra fish embryos
markedly enhanced toxicity of BNF alone and in combination
with ANF (Billiard et al., 2006). This disparity suggests that
the mechanism of toxicity of the PAH-type AHR agonists
differs from that of DLCs and that the relationship among AHR
agonism, CYP1A activity, and PAH-mediated embryotoxicity
is quite complex.
POSSIBLE MECHANISMS OF DEVELOPMENTAL
Although developmental toxicity has been associated in fish
models with exposure to crude oil, retene (C-4 alkyl-
phenanthrene), and PAH-contaminated sediments, not all
PAHs tested cause this syndrome. For example, in contrast to
injected BaP and BkF, waterborne BaP and BkF were potent
CYP1A inducers, but nontoxic to fish embryos within the
limits of their solubilities (Billiard, 2002; Wassenberg and Di
Giulio, 2004b), perhaps because toxicity was limited kineti-
cally by low rates of uptake (see ‘‘AhR-mediated toxicity’’
section below). Conversely, phenanthrene did not induce
CYP1A but was embryotoxic without signs of BSD (Hawkins
et al., 2002). As discussed above, effects can be potentiated or
antagonized by specific mixtures, suggesting nonlinear toxic
interactions. In this section, we summarize recent research that
points to four hypotheses (Fig. 2) to explain BSD: narcosis,
cardiac-mediated toxicity (direct interference with cardiac
development), AHR-mediated toxicity (altered transcription
of AHR pathway genes during ontogeny), and CYP1A-
mediated toxicity (toxicity due to oxygenation of PAHs by
While a narcotic mechanism for PAH toxicity has been
proposed for the development of water and sediment quality
guidelines (Di Toro et al., 2000), this model has not been
thoroughly evaluated for the chronic toxicity of PAHs. Narcosis
usually describes chemicals with baseline or minimum toxicity
for which the exact biochemical mechanism is unknown
(Meador, 2006). The proposed target site of action of narcotics
is the lipid membrane bilayer, and thus the potency of narcotic-
acting chemicals is directly related to their octanol-water
BILLIARD ET AL.
partition coefficients or Kow(Incardona et al., 2004; McCarty
and Mackay, 1993). Historically, the strong correlations
between water-lipid partitioning and toxicity were interpreted
to mean that narcotics dissolve in lipid membranes and
interfere with membrane function to cause an anesthetic or
narcotic-like effect. However, a review by Campagna et al.
(2003) demonstrated that many anesthetics with similar
structures and similar affinities for lipid actually caused
markedly different biochemical, neurological, and anesthetic
effects. Each anesthetic had unique molecular targets and
interfered with neuronal functioning in a different way. Hence,
in aquatic toxicology, narcosis as a general mechanism most
properly describes the relationships among water-lipid parti-
tioning, the kinetics of bioaccumulation, and the acute lethality
of organic compounds. As shown by Campagna et al. (2003),
narcosis encompasses a wide array of mechanisms, but these
are difficult to discern in aquatic toxicology because of the
dominating effect of water-lipid partitioning on exposure and
measured toxicity. Therefore, it is not surprising that the
sublethal toxicity of PAHs which interact with specific
receptors can only be approximated from Kow.
One test of the narcosis model is the concentration of
chemical present in fish that are intoxicated; for lethality,
concentrations are typically in the range of 3–7 mmol/kg (wt/
wt), the acute critical body residue (McCarty and Mackay,
1993). For example, waterborne phenanthrene was embryo-
toxic to trout when tissue concentrations exceeded 1 mmol/kg
(Billiard, 2002; Hawkins et al., 2002). While phenanthrene was
lethal to embryos, it was not a CYP1A inducer and did not
cause signs of BSD. Taken together, these data suggested
Flow chart of potential mechanisms of developmental toxicity or BSD of PAHs to ELS of fish.
PAH DEVELOPMENTAL TOXICITY
a nonspecific, narcotic mode of action, in contrast to retene,
which caused mortality and severe BSD in trout embryos at
tissue concentrations that were almost undetectable (<0.01
mmol/kg; Hawkins et al., 2002). However, as discussed in the
next section, a study by Incardona et al. (2004) concluded that
three-ringed PAHs such as phenanthrene were not narcotic to
zebra fish embryos. Rather, embryotoxic effects were second-
ary to disruption of cardiac conduction. The narcosis model
also cannot account for the nonadditive, synergistic cardiovas-
cular interactions between PAH-type AHR agonists and
CYP1A inhibitors observed with killifish (Wassenberg and
Di Giulio, 2004b). Similarly, when larval trout were coexposed
to waterborne phenanthrene and the CYP1A inducer BNF, the
toxicity of phenanthrene increased dramatically, while tissue
concentrations of phenanthrene decreased to <0.01 mmol//kg.
This suggested a switch from a general (narcotic) to a specific
mode of action and a potential role in toxicity for metabolites
of phenanthrene created by CYP1A oxygenation (Hawkins
et al., 2002). Finally, owing to their rapid metabolism, typical
bioconcentration factors for most PAHs are <1000. Hence, at
waterborne concentrations that are chronically toxic (for e.g.,
retene: 0.04–0.1lM; Brinkworth et al., 2003), they would not
bioaccumulate to concentrations consistent with the critical
body residues required for narcosis (Billiard, 2002; Wassenberg
and Di Giulio, 2004b).
While the toxicity of phenanthrene appears to fit the narcosis
model (Hawkins et al., 2002), there is also recent evidence that
it may act by directly targeting cardiac function. Research by
Incardona et al. (2004, 2005) demonstrated that the tricyclic
PAHs fluorene, phenanthrene, and DBT were directly cardio-
toxic to zebra fish embryos and that this dysfunction was
independent of narcosis. The effects mimicked the atrio-
ventricular conduction block associated with the silent heart
(sih) mutation in zebra fish embryos, and the proposed cause
was blockage of cardiac ion channels. This highly specific
mode of action seems to be independent of CYP1A
oxygenation or induction because knock down of ahr2
expression by morpholinos did not reduce toxicity. Larger
PAHs appeared to act by different mechanisms, with chrysene
causing strong CYP1A induction but not cardiotoxicity and
pyrene causing CYP1A induction and pathology in peripheral
blood vessels similar to dioxin (Incardona et al., 2004). Tests
with four to five component mixtures containing the same total
PAH concentrations, but each with a different mix of PAHs,
showed that only the mixture with high proportions of
phenanthrene and DBT was cardiotoxic. This was consistent
with a specific mode of action of individual compounds rather
than a general mode of action associated with the sum of all
PAHs. Incardona et al. (2005) also demonstrated that the
concentrations of individual tricyclic PAHs found in weathered
Alaska North Slope crude oil were sufficient to explain cardiac
dysfunction in developing zebra fish exposed to the water-
accommodated fraction. However, the overall response of the
fish reflected the presence of other compounds having
Morpholino knock down of AHR1 (AHR1A; Andreasen
et al., 2002a; Karchner et al., 2005) and AHR2 blocked
CYP1A induction in zebra fish but did not protect against
toxicity of phenanthrene and DBT, suggesting that cardiac
dysfunction was independent of the AHR receptor (Incardona
et al., 2005). In contrast, pyrene toxicity was localized in the
peripheral vasculature and was clearly AHR dependent because
toxic effects were decreased or delayed in AHR2 morphants.
DBT and phenanthrene are usually classified as weak AHR
agonists or noninducers (e.g., Basu et al., 2001; Billiard, 2002;
Hawkins et al., 2002; Wassenberg et al., 2005). However, for
mixtures of PAHs, there may yet be a role of CYP1A induction
in their toxicity. As indicated above, coexposure of larval trout
to phenanthrene and to BNF, a strong CYP1A inducer,
significantly increased the toxicity of phenanthrene, although
the specific effects on the heart were not examined (Hawkins
et al., 2002). Taken together, these data suggest multiple
mechanisms of toxicity that vary with the structure of PAH and
with complex interactions in mixtures of AHR agonists and
Another possible mechanism that has been proposed for
dioxin-induced cardiovascular toxicity (reviewed in Goldstone
and Stegeman, 2006) that could also apply to PAH is cross-talk
of activated AHR with homologous signaling pathways (i.e.,
hypoxia) through negative regulation or competition for
reciprocal transcription/cellular cofactors. A correctly patterned
vasculature for oxygen and nutrient transport to developing
tissues is essential for normal development and survival of the
factor-1 (HIF-1), a heterodimeric basic-helix-loop-helix tran-
scription factor composed of two subunits, HIF-1a and HIF-1b
(otherwise known as ARNT), mediates changes in gene
expression of proangiogenic growth factors. These include
VEGF, critical for normal vascular development. Only a few
studies have looked at this in fish embryos and while they do
not necessarily support a role for this particular interaction in
dioxin-exposed zebra fish embryos (Handley, 2003; Handley-
Goldstone et al., 2005; Prasch et al., 2004), other nuclear cross-
talk scenarios have not been tested, and signaling antagonism
during development has not been explored specifically for
PAHs. While a possible interaction between developmental
signaling pathways has been mentioned here because it
underscores cardiotoxic effects similar to dioxin and PAHs, it
also highlights a potential role of the AHR pathway discussed
There is evidence both for and against a role of the AHR in
PAH-induced BSD, likely because the mechanisms of toxicity
BILLIARD ET AL.
vary considerably among PAHs. For AHR-mediated toxicity,
PAHs may act through the AHR without direct involvement of
CYP1s. Some four- to six-ringed unsubstituted PAHs that bind
strongly to the AHR receptor and are potent CYP1A inducers
(e.g., BaP, BkF, chrysenes; Basu et al., 2001; Billiard, 2002;
Incardona et al., 2005) were nontoxic to ELS within the limits
of their solubility (Billiard, 2002; Incardona et al., 2005;
Wassenberg et al., 2005). For these PAHs, AHR binding and
gene activation do not seem sufficient by themselves to cause
BSD, but they may still be part of the mechanism of toxicity if
toxicity involves induction of biotransformation of the PAH to
toxic metabolites. In fact, the toxicity of these large PAHs such
as BaP may simply be kinetically limited. Sundberg et al.
(2006) demonstrated embryotoxicity of BaP in trout larvae
exposed at fertilization to injections of single doses of BaP
(0.040–8.0 lg/g) in triolein. CYP1A induction was found at
concentrations of 3 lg/g or greater and an increased frequency
of deformities at 5 lg/g or greater. The toxicity of injected BaP
suggests that the amount of BaP that crosses the chorion during
waterborne exposures may be insufficient to cause toxicity.
Our recent studies demonstrate that AHR2 zebra fish
morphants are protected from synergistic embryotoxicity
caused by coexposure to the PAH-type AHR agonist, i.e.,
BNF and to the CYP1A inhibitor ANF (Billiard et al., 2006).
These data provide the first evidence that synergistic,
embryotoxic effects of PAHs might be mediated in part by
the AHR signaling pathway (Billiard et al., 2006). Recent
studies by Incardona et al. (2005, 2006) used morpholinos to
block the synthesis of the AHR protein in zebra fish exposed to
an array of different PAHs. Developmental defects in zebra fish
exposed as embryos to petrogenic PAH mixtures were
independent of AHR1 (AHR1A) and AHR2, likely because
the tricyclic PAHs (e.g., DBT and phenanthrene) used are weak
AHR agonists or non-CYP1A inducers, as indicated above.
The authors concluded that cardiotoxicity was an indirect effect
stemming from direct impacts on cardiac conduction. In
contrast, benz[a]anthracene caused CYP1A induction and
embryotoxicity to zebra fish that was first evident as a failure
of the developing heart to complete looping, with signs of BSD
following cardiotoxicity. Toxicity was antagonized by ahr2 but
not cyp1a knockdown, indicating an AHR2-dependent, but
CYP1A-independent mechanism, as observed with dioxin
(Incardona et al., 2006). The molecular basis for this AHR
dependence has not been described.
For some individual PAHs and PAH mixtures, CYP1A
enzyme activity plays several roles in embryotoxicity. In zebra
fish embryos, CYP1A, 1B1, and 1C1 mRNA are inducible by
AHR ligands as early as 24–48 hpf prior to the onset of toxicity
(Andreasen et al., 2002b; Timme-Laragy et al., 2007).
Immunolocalized CYP1A protein is first detected in zebra
fish vasculature and later in cardiac, hepatic, and kidney tissues
(Andreasen et al., 2002b). Here, the key process may be CYP1-
mediated metabolism of PAHs to toxic intermediates. For
example, pyrene caused CYP1A induction, pericardial and
yolk sac edema, spinal curvature, and reduced peripheral
circulation in zebra fish embryos. Toxicity was antagonized by
ahr2 and cyp1a knockdown, and partially antagonized by
ahr1a knockdown, indicating a role of CYP1A enzyme activity
in the mechanism of toxicity (Incardona et al., 2005, 2006). In
contrast, there was a protective role for CYP1A activity in PAH
versus DLC teratogenesis and synergism in zebra fish embryos
coexposed to PAH-type AHR agonists and CYP1A inhibitors,
as shown by increased toxicity after cyp1a knockdown or
inhibition (Billiard et al., 2006; Wassenberg and Di Giulio,
2004b; Wassenberg et al., 2005); ahr2 morphants were
protected from cardiovascular defects under the same con-
ditions (Billiard et al., 2006).
How do we reconcile these divergent observations? Under
normal circumstances, functional CYP1A would confer pro-
tection to embryos exposed to PAH-type AHR agonists by
metabolizing parent PAHs to reactive intermediates with
relatively low toxicity when conjugated via phaseII, metabo-
lism. This model predicts that when CYP1A activity or
synthesis of CYP1A protein is stimulated or inhibited, there
would be a change in the relative proportions and amounts of
toxic metabolites and observed toxicity. In other words,
toxicity is a function of the rate at which specific toxic
metabolites are produced and excreted. Protection would only
be observed if there was an alternate AHR-dependent
metabolic pathway for PAHs, assuming that an AHR agonist
alters expression of other DRE-driven genes downstream of the
AHR/ARNT transcription complex. Hence, AHR knockdown
could block alternate pathways of PAH metabolism and protect
against synergistic toxicity between BNF (AHR agonist,
CYP1A inducer) or ANF (CYP1A inhibitor). Alternatively,
PAHs could be metabolized and excreted by an AHR-
independent pathway in AHR2 morphants. Otherwise, in the
absence of metabolism, PAHs would accumulate to toxic
concentrations, as with the narcosis model. Another hypothesis
to explain the increased toxicity of PAHs when CYP1A is
inhibited is that lowering the rate of metabolism extends the
half-life of the PAHs, allowing parent PAH-type AHR agonists
to persist longer and behave more like DLCs. While the
specific mechanism as to how prolonged agonism confers
toxicity has not elucidated, it deserves further investigation
(Timme-Laragy et al., 2007).
Some of these hypotheses have been tested by measuring the
products of metabolism of retene by trout in the presence and
absence of CYP1A inhibitors. In juvenile trout coexposed to
retene and ANF, the excretion of retene metabolites in bile was
progressively reduced by higher concentrations of ANF
(Hodson et al., 2007). In the tissues of larval trout exposed
only to retene, there was a broad spectrum of metabolites that
included di-hydroxy- and mono-hydroxy derivatives, with very
low concentrations of parent retene. With increasing exposure
PAH DEVELOPMENTAL TOXICITY
to ANF, there was a progressive reduction in the number and
size of metabolite peaks and a corresponding increase in parent
retene concentrations. At intermediate exposures to ANF, di-
hydroxy metabolites were virtually absent, but there was
a higher concentration of mono-hydroxy metabolites, and
toxicity increased dramatically. At high exposures of ANF,
tissue metabolites of retene were virtually absent (parent retene
was the predominant form), and retene toxicity was almost
completely antagonized (Hodson et al., 2007). Partial in-
hibition of CYP1A by ANF seemed to favor the production of
the more toxic metabolites of retene while complete inhibition
prevented metabolism altogether and eliminated toxicity. These
results suggest that CYP1A enzyme activity prevents the
accumulation of both parent PAH and the most toxic of the
metabolites. Thus, the role of CYP1A enzyme activity in BSD
caused by some PAHs may be to prevent the generation of
While this discussion has focused on CYP1A, the recent
discovery of PAH-inducible CYP1B and CYP1C proteins in
fish (e.g., Wang et al., 2006) suggests that PAH metabolism
and toxicity following AHR binding by PAHs involves
multiple reactions. However, the potential for ANF to modulate
PAH metabolism and toxicity by CYP1B and CYP1C are
unknown because the tissue distribution, PAH specificity, and
metabolites generated by each CYP protein are not yet worked
out. These are clearly important research needs if the
mechanisms of different PAHs are to be understood.
Similarly, we recently showed that ANF alone acts as a weak
AHR2 agonist (Timme-Laragy et al., 2007). Furthermore, when
this dose of ANF was combined with BNF—a mixture that we
have previously shown to inhibit BNF-induced CYP1A enzyme
activity and synergize toxicity (Billiard et al., 2006)—mRNA
induction of AHR-mediated cyp1 genes (including cyp1a,
cyp1b1, and 1c1) was greatly enhanced and preceded the onset
of deformities in zebra fish embryos in contrast to inhibition of
CYP1A protein as measured by in vivo activity of ethoxyresorufin-
o-deethylase (EROD) (Timme-Laragy et al., 2007).
Based on these data, we propose that the key difference in
the role of CYP1A in mediating toxicity between PAH and
DLC AHR agonists is that PAHs are much more metabolically
labile than DLCs. Thus, CYP1A and other enzymes metabolize
the wide variety of PAHs in different ways, resulting both in
differential production of toxic metabolites and differential
toxicokinetics. For example, while chrysene, pyrene, and
benz(a)anthracene caused CYP1A induction, the tissue distri-
bution and intensity of induction varied considerably among
the three compounds, implying different rates of uptake and
partitioning among tissues (Incardona et al., 2006). Hence,
there is a need to understand the toxicokinetics of PAHs and
the relative role of metabolites versus parent compounds.
Hornung et al. (2007) mimicked maternal transfer of more
persistent chemicals, such as DLCs, by acute waterborne
exposure of postfertilized medaka eggs to B[a]P. This study
demonstrated that multiple embryonic tissues of fish embryos
have the ability to metabolize PAHs even prior to cardiac
or hepatic development. Redistribution of parent PAH and
metabolic intermediates occurs throughout embryonic devel-
opment. This is coincident with a switch from endogenous
feeding as the yolk sac is resorbed and development of the
cardiovascular system (Hornung et al., 2007). Mobilization and
transfer of sequestered chemical from the yolk could partly
explain why the embryonic vasculature is a sensitive, first
target. This also confirms other studies showing that the
AHR pathway is operational early in fish development
(Andreasen et al., 2002b; Jo ¨nsson et al., 2007a,b; Timme-
Laragy et al., 2007).
One would expect that if metabolism drives PAH toxicity,
there should be a clear time dependency of effects. In studies
with rainbow trout continuously exposed from eggs to retene,
CYP1A activity of embryonic fish increased continually during
development as tissue PAH concentrations decreased. These
experiments demonstrated that retene was metabolized by
CYP1A enzymes and that induction preceded the onset of
toxicity (Brinkworth et al., 2003).
Alkyl-substituted PAHs comprise up to 95% of total PAHs
in environmentally relevant mixtures such as crude oil (Fig. 3).
In a review and evaluation of several developmental PAH
toxicity models, Barron et al. (2004) proposed that alkyl-
phenanthrenes were the primary PAHs driving the toxicity of
Alaska North Slope Crude to herring and salmon embryos after
the Exxon Valdez oil spill (Carls et al., 1999, Heintz et al.,
1999, 2002). The toxicity of sediment extracts derived from oil
sands and enriched with alkylated PAHs indicates that alkyl
substitution is a significant contributor to embryotoxicity
(Colavecchia et al., 2004; Rhodes et al., 2005).
Alkyl-substituted phenanthrenes such as retene induce
CYP1A activity and cause signs of BSD in ELS of rainbow
trout, zebrafish, and medaka (Billiard et al., 1999, Brinkworth
et al., 2003; Hawkins et al., 2002 Kiparissis et al., 2002). As
discussed above, these data indicated that the toxicity of retene
is tied to its metabolism. But why would retene metabolites be
more toxic than parent? The process of metabolism likely
generates ROS, but this would not explain why toxicity is
enhanced when CYP1A metabolism is partially inhibited.
Another explanation, already introduced above, is the gener-
ation of specific metabolites with a uniquely high toxicity.
Tabash (2003) demonstrated that retene metabolites generated
in vivo by retene-exposed juvenile trout, or in vitro by liver S-9
fractions from juvenile trout exposed to BNF, were primarily
benzylic alcohols, i.e., metabolites with hydroxylations on the
alkyl side chains. These alcohols are unique to alkyl-PAHs
since hydroxylation of unsubstituted PAHs can only form
phenols. It is also possible that that the presence of alkyl side
chains causes the formation of specific and highly toxic
phenols because hydroxylations are restricted to specific
BILLIARD ET AL.
For other alkyl-substituted homologs of phenanthrene, both
the size and substitution pattern of alkyl groups determines
potency for BSD in medaka; compounds substituted in the 1,7
positions are the most toxic to embryos (Kiparissis et al.,
2002). Some alkyl-PAHs, including alkylated napthalenes and
anthracenes, do not cause BSD in medaka (Kiparissis et al.,
2002; Turcotte et al., 2005).
There is strong evidence of a role for oxidative stress in PAH
synergistic toxicity. Upregulated antioxidant defenses appear to
be acting in both the short-term and heritable adaptation to
PAH toxicity in the Elizabeth River. Killifish indigenous to this
site display elevated hepatic total glutathione levels (Bacanskas
et al., 2004). Laboratory-raised offspring from this population
appear better equipped to withstand oxidative stress, as
evidenced by increased resistance to a model pro-oxidant,
tert-butyl hydroperoxide, higher basal total oxy-radical scav-
enging capacity, glutathione concentrations, and MnSOD
protein levels (Meyer et al., 2003a). Furthermore, treatment
of reference site killifish embryos to the model pro-oxidant
cumene hydroperoxide resulted in a similar phenotype as seen
in PAH and DLC toxicity, whereas cotreatment of embryos
with vitamin-E-acetate showed potential for reducing PAH
synergistic toxicity (Wassenberg, 2004).
Vitamin E is thought to be maternally loaded into the yolk
sac of salmonids. Exposure of larval rainbow trout to the alkyl-
PAH retene caused a 20% decrease in whole body vitamin E
concentrations (Bauder et al., 2005). Coexposure to retene and
to 10 lg/l vitamin E restored antioxidant levels to control levels
and significantly reduced the severity of BSD (Bauder et al.,
2005). However, in the same study, known pro-oxidants such
as paraquat, tert-butyl hydroperoxide, and carbon tertrachloride
did not increase tissue concentrations of hydroperoxides or cause
signs of BSD (Bauder et al., 2005). These paradoxical results
could be a function of exposure conditions, chemical uptake,
different modes of action between retene and the pro-oxidants,
or the limits of detection of the hydroperoxide method.
In summary, it is clear that multiple mechanisms govern the
adverse developmental effects of PAHs and that the toxic
potency of complex mixtures is unquestionably a function of
their components and how they interact. The accumulation by
fish of potent CYP1A inducers (e.g., BaP) means that all other
aromatic constituents accumulated from a mixture may be
subject to oxygenation to metabolites that are more or less toxic
than the parent compounds. Similarly, the presence of low
molecular weight PAHs that inhibit CYP1A activity, or that
competitively bind to the AHR (e.g., 2-aminoanthracene,
Wassenberg and Di Giulio, 2004b), will modulate CYP1A
activity and the metabolism and toxicity of other PAHs.
Moreover, the loss of low molecular weight PAHs with
weathering of crude oil may cause a progressive shift in the
extent of compound interactions. Overall, due to their complex
nature and multiple cellular targets, the developmental toxicity
of PAHs is not easily predicted from hydrophobicity (i.e.,
narcotic mode of action) or the AHR activity of individual
PAHs (Incardona et al., 2006). Understanding the mechanism
underlying PAH effects is paramount for understanding and
characterizing their risk. This includes characterizing the
regulation and contribution of other AHR-mediated genes in
the ELS fish model to PAH toxicity.
Hydrocarbon content of Alaska North Crude Oil (adapted from Wang et al., 2002, 2003).
PAH DEVELOPMENTAL TOXICITY
IMPLICATIONS FOR RISK ASSESSMENT
PAHs invariably occur in the environment as mixtures, and
assessing the risks of complex mixtures such as PAHs is
challenging. For example, in sediment samples from the PAH-
contaminated Elizabeth River, VA site discussed earlier, BaP
(AHR agonist) and FL (CYP1A inhibitor) comprised 11 and
26%, respectively, of total PAHs measured (18 individual
compounds) (Vogelbein and Unger, 2003). The marked
synergy between PAHs acting as AHR agonists and as CYP1A
inhibitors described herein suggests that for some PAH
mixtures, the common assumption of additivity may greatly
underestimate these risks. This is consistent with recent
arguments that widely accepted models of PAH toxicity are
oversimplified and that TEQ models do not predict their
interactive, developmental effects (Incardona et al., 2006).
Elucidating the mechanistic basis for PAH interactions will act
to reduce the degree of uncertainty in such risk assessments.
Unquestionably, a detailed chemical characterization of PAH
mixture constituents, beyond the standard typical 16 non-
substituted PAHs, at each particular site would be required to
predict nonadditive effects. Biological testing of selected
complex mixtures would improve the confidence in risk
assessments that are based primarily on analytical data.
Superfund Basic Research Grant (P42 ES10356); Environ-
mental Health Sciences Training Grant (T32 ES07031) to
R.T.D.; Discovery Grant from the Natural Sciences and
Engineering Research Council of Canada to P.V.H.
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