Page 228 - The Toxicology of Fishes

208 The Toxicology of Fishes

CYP1A is clearly the predominant enzyme responsible for each oxygenation of BaP, but other CYP
isoforms are likely contributors. Studies with recombinant mammalian CYP isoforms indicate metabolic
activation of several PAHs by CYP1B1 as well as CYP2C9, CYP2C19, and CYP3A4, albeit at slower
rates than CYP1A and CYP1B enzymes (Shimada et al., 1999). CYP1A forms are readily inducible in
fish by PAH and related planar polycyclic molecules, and this likely is a major factor in the sensitivity
of fish to PAH-type carcinogens. In addition, other phase I enzymes may catalyze oxidation reactions
(see above). Inhibition of cytotoxicity with indomethacin within trout cell lines that lack EROD activity
suggests that prostaglandin-H synthase may contribute to the activation of BaP (Schirmer et al., 2000).
Clearly cooxidative pathways should be explored in more detail, particularly in fish that possess signif-
icant quantities of unsaturated fatty acids which are prone to oxidation.
Other metabolites with potential biological activity include quinones, which have been suggested to be
derived from initial phenolic metabolites. BaP has been shown to be converted to several quinone metab-
olites, including the 1,6- and 3,6-quinones derived from the 1-hydroxy and 3-hydroxybenzo(a)pyrene
precursors (Stegeman, 1981). Although numerous biotransformation studies have identified quinones as
BaP metabolites (Stegeman and Hahn, 1994; Willett et al., 2000; Yuan et al., 1997), few have explored
the potential effects of these metabolites on cell function. Lemaire et al. (1994) reported that BaP quinones
were capable of creating hydroxyl radical in hepatic microsomes of flounder and perch. Oxidative damage
in larval turbot was observed following BaP treatment with parallel formation of BaP quinone metabolites.
Although quinones and the (–)-7,8-dihydrodiol can lead to greater toxicity of BaP, it should be noted
that the majority of the metabolites of BaP are nontoxic. Examples include phenols (at the 1 and 3
positions) which are primarily conjugated as glucuronides and potentially sulfate derivatives (Buhler
and Williams, 1989; Stegeman and Hahn, 1994). Glutathione conjugates presumably of the various arene
oxides (4,5, 7,8, and 9,10) have also been reported in fish (Gallagher et al., 1996). The 4,5 and 9,10
dihydrodiols of BaP also do not appear to undergo bioactivation.
Tremendous species differences exist in the bioactivation of BaP in fish. Liver microsomes of channel
catfish, which appear less susceptible to PAH-induced liver cancer, produce lower levels of (+)-7,8-expoxide
(based on lower concentrations of (–)-7,8-dihydrodiol) than cancer-prone brown bullhead catfish (Pangrekar
and Sikka, 1992; Yuan et al., 1997). In vivo studies by Willett et al. (2000) confirmed the species differences
in (–)-7,8-dihydrodiol formation. However, hepatic EROD activities did not correlate with 7,8-BaP dihy-
drodiol formation, indicating that another enzyme system may be involved in BaP oxygenation (Willett et
al., 2000). Ploch et al. (1998) also reported that, although in vitro DNA binding by BaP was directly related
to hepatic EROD activity in the two species, in vivo DNA adducts were not related to hepatic EROD. These
results are consistent with studies in carp, which have significantly higher rates of phase I metabolite
formation compared to brown bullhead but are more resistant to PAH-induced cancers (Sikka et al., 1990).
The effects of other pollutants and environmental conditions on other metabolic pathways of PAHs,
or all xenobiotics, are largely unknown. Dieldrin significantly enhances the biliary elimination of BaP
but does not alter the metabolite profile (Barnhill et al., 2003). Tributyltin significantly inhibits BaP
biotransformation in Arctic char (Padros et al., 2003). Rainbow trout acclimated to hypersaline conditions
had a higher level of phase I metabolites and a shift from dihydrodiol to phenolic metabolites with
3-hydroxybenzo(a)pyrene predominant (Seubert and Kennedy, 2000).
In addition to phase I enzymes, species-dependent differences in phase II pathways may also contribute
to differences in BaP-induced toxicities between species. Although significant epidemiological and
biochemical evidence exists in humans and rodents (GSTM1), few studies have demonstrated a corre-
lation between adverse effects and species-specific phase II metabolism in fish; for example, hepatic EH
and GST activities in channel catfish were higher than in brown bullhead (Gallagher et al., 1996; Willett
et al., 2000). When comparing overall biliary metabolite profiles in each species, significant differences
were not observed in quantity or profile of conjugated BaP metabolites (Willett et al., 2000). Likewise,
English sole from contaminated sites in Puget Sound appear to be more sensitive to liver lesions resulting
from PAH exposure than starry flounder, but overall GST-catalyzed conjugation of BPDE is threefold
higher in English sole (Gallagher et al., 1998). Although glucuronides represent the highest percentage
of conjugated BaP metabolites in fish, few studies have identified specific metabolites or evaluated
species differences in hepatic UGTs and whether this pathway contributes to species differences in BaP
toxicity.