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692 The Toxicology of Fishes

may be used as a biomarker of effect (i.e., oncogenic activation) and potentially susceptibility (Kurelec,
1997). The latter was indicated in studies with carp (Cyprinus carpio) pretreated with an MDR inhibitor
(verapamil) and then treated with various PAHs, resulting in significantly enhanced accumulation of
PAHs and CYP1A activity compared to fish that did not receive verapamil (Kurelec, 1997). Similar
studies were carried out with invertebrates showing enhanced susceptibility to chemical toxicity when
MDR was inactivated or in animals that expressed lower endogenous concentrations (Kurelec, 1992).
Characterizing MDR expression in fish species has been the subject of several recent studies (for a
review, see Strum and Segner, 2005). MDR appears to be endogenously expressed in numerous tissues,
including renal proximal tubules of Fundulus heteroclitus (Miller, 1995), bile canaliculi, exocrine pan-
creas, lumenal surface of the intestinal epithelium, interrenal tissue, branchial blood vessels, gas gland,
pseudobranch, and the gill transverse septa in Poecilia reticulata (Hemmer et al., 1995). In cholangio-
cellular carcinomas of PAH-contaminated winter flounder, expression of hepatic MDR was not observed
using immunohistochemistry (Bard et al., 2002); however, significant MDR expression was observed in
the intestine of fish from the contaminated site. Overexpression of MDR was observed in homogenates
of livers bearing tumors from a population of Fundulus heteroclitus from a creosote-contaminated site
on the Elizabeth River in Virginia (Cooper et al., 1999). MDR was low in healthy, injured, and extrafocal
tissue of cancerous livers from the Atlantic flounder (Platichthys flesus) but increased in transitional
stages of foci toward the cell types persisting during progression toward carcinomas (Kohler et al., 1998).
In other controlled laboratory studies, mutagenic benzo(a)pyrene metabolites as well as other xenobiotics
were shown to induce expression of MDR in the gut of channel catfish following treatment (Doi et al.,
2001; Kleinow et al., 2000). Basal levels of expression were highest in brain, kidney, gills, heart, and
intestine of turbot (Scophthalmus maximus) (Tutundjian et al., 2002). Results from these studies and
ones in invertebrates (Kurelec, 1992, 1997) provide strong evidence that overexpression of MDR,
particularly in the intestines of fishes, may be a component of a genetically based mechanism providing
a selective advantage in pollution-resistant populations of predatory aquatic organisms. Clearly, because
of the discrepancies and variability of expression in animals within the same species, more characteriza-
tion studies are necessary to further develop this nonspecific indicator of effect and susceptibility in fish.

Phase II Biotransformation Enzymes
Following initial phase I transformation, which either exposes or adds single polar atoms (i.e., OH) to
enhance water solubility and excretion, phase II processes tend to augment this process by adding large
endogenous polar molecules (Figure 16.3). Expression of several phase II enzymes in mammals is
regulated via the xenobiotic response element (XRE) and thus would be expected to be induced by
planar aromatic hydrocarbons and serve as potential verification for exposure to Ah receptor agonists.
Examples in fish include uridine diphosphate (UDP)–glucuronosyltransferase (UDPGT), at least one
isoform of which is induced by treatment of several Ah receptor agonists such as PAHs and PCBs (Forlin
et al., 1996; George, 1994). Studies in the field appear to verify laboratory studies with a direct
relationship between UDPGT activity in fish that accumulate PCBs or organochlorines (Van der Oost
et al., 2003). In eelpout (Zoarces viviparous), seasonal variation in responses were much larger in phase
I enzymes than in UGT (Ronisz et al., 1999). For a more in-depth discussion regarding purification and
regulation, see George (1994) and Chapter 4 of this text.
Another phase II family of enzymes that can be upregulated following exposure to planar aromatic
hydrocarbons in mammals is the glutathione S-transferase (GST) family. In fish, the expression of GST
activity (chlorodinitrobenzene [CDNB] dehalogenation) in response to acute doses of Ah agonists does
not appear to result in significant elevation (George, 1994); however, if fish are chronically exposed to
PCBs, GST activity has been shown to be significantly elevated (Forlin et al., 1996). Indeed, numerous
contradictory field studies have shown induction, no change, and in some instances repression of activity
at various polluted sites (for a review, see George, 1994). This inconsistency in response may be related
to the relatively simple assay for GST that has been used in a majority of these studies and which does
not differentiate between major GST isoforms (with the exception of the class of GSTs that lack CDNB
activity). In fact, in studies characterizing the effects of various inducers on GSTs of plaice (Pleuronectes
platessa), expression of a GST gene structurally homologous to a class isoform was repressed by planar