||<p>The research presented in this thesis focused on the mechanistic aspects of the toxic and biochemical effects of PCBs in flounder <em>(Platichthys flesus),</em> with the aim to provide a scientific basis for the suggested involvement of PCBs in the aetiology of diseases observed in flounder. Therefore, the first goal was to study the biochemical effects of PCBs in flounder by investigating the inducibility of cytochrome P4501A and associated EROD activity and glutathione-S-transferase activity upon administration to PCBs and TCDD in vivo. Secondly, the mechanism of action of PCBs in flounder was investigated by identifyingg the hepatic microsomal Ali receptor pathway. In addition, the potency of individual PCB congeners and a commercial PCB mixture to specifically inhibit the catalytic CYP1A activity was studied using in vitro techniques. Finally, the endocrine disrupting effect of PCBs was studied by analysing the retinoid and thyroid hormone levels in flounder after acute and chronic exposure to either a commercial PCB mixture or contaminated harbour sludge.<br/>summary of results<p>Short-time exposure of flounder to the technical PCB mixture Clophen A50 showed minor effects on total hepatic cytochrome P450 concentrations and only a relatively slight induction in EROD activity, indicating that flounder is not very sensitive towards i.p. and oral administration of Clophen A50 ( <strong>chapter</strong> 3 and 5). The most potent PCB congener (CB-126), in terms of mammalian derived TCDD toxic equivalents (TEQ), was found to be more potent as inducer of CYP1A protein and activity in flounder, but not to the extend as was expected based on TCDD TEQs ( <strong>chapter</strong> 5). On the other hand, high induction of the hepatic microsomal CYP1A content and associated EROD activity were observed in flounder orally exposed to TCDD ( <strong>chapter</strong> 4). Interestingly, TCDD induced EROD activity in flounder could be inhibited by co-treatment with Clophen A50. In contrast, administration of flounder with a combination of TCDD and CB-126 resulted in an additive effect on EROD activity ( <strong>chapter</strong> 5). With respect to induction of hepatic CYP1A protein content, administration of combinations of both Clophen A50 and CB-126 with TCDD resulted in an additive effect compared to exposure of flounder to the individual compounds. These results indicated a direct inhibition of CYP1A catalytic activity by residual PCBs present in the microsomal suspension, rather than inhibition of transcription of the CYP1A gene or translation of CYP1A mRNA. Evidence for direct inhibition of EROD activity by individual PCB congeners and the technical PCB mixture Clophen A50 is presented in <strong>chapter</strong> 6.<p>The relative sensitivity of flounder towards TCDD exposure was established by comparing the flounder computed no effect level (CNEL) for EROD activity to -lowest observed -adverse - effect levels (LOAELs) from other fish species and the rat (chapter 4). The rat was clearly more sensitive towards TCDD exposure, but other fish species, such as carp and rainbow trout, were only slightly more sensitive than flounder.<p>A number of biotransformation enzymes other then CYP1A are also found or suggested to be modulated by the Ali receptor signal transduction pathway. Among them are certain forms of the phase 11 enzyme system such as glutathione-S-transferase (GST), In contrast to some other fish species and mammals, hepatic cytosolic GST activity in flounder was not altered upon exposure to TCDD (chapter 4). Moreover, no effect on GST activity was observed in flounder after Clophen A50 exposure, whereas a slight inhibition was observed upon CB-126 administration ( <strong>chapter</strong> 5). In contrast, a strong inhibition of GST activity was observed when flounder were treated with combinations of TCDD and either CB-126 or Clophen A50 ( <strong>chapter</strong> 5). An explanation for these observations was not found, but it was suggetsed that GST substrate inhibition by PCB-metabolite residues could have occurred.<p>To better understand the contradiction in induction pattern of the flounder CYP1A enzyme system towards exposure to either PCBs or TCDD, the hepatic Ah receptor pathway in flounder was characterised ( <strong>chapter</strong> 6). In addition, the potency of a number of PCB congeners and Clophen A50 to inhibit the CYP1A catalytic activity in vitro was studied. In chapter 6, evidence for the presence of low levels of Ali receptor (1-7 fmol/mg protein) in flounder hepatic cytosol was presented using protamine sulphate and hydroxylapatite analysis and velocity sedimentation on sucrose gradient. The level of Ali receptor in flounder was similar to levels of receptor reported in some other fish species, but much lower than Ali receptor levels reported in mammals. Additional evidence for the presence of the cytosolic Ali receptor in flounder liver was provided by first-strand cDNA synthesis and subsequent amplification of flounder poly A+ RNA using RT-PCR. The specificity of the 690 bp RT-PCR reaction product was established by southern blotting and hybridisation. Subsequent sequencing of the RT-PCR product showed that its deduced amino acid sequence was 75% identical to the killifish species <em>Fundulus heteroclitus</em> AhR-2 sequence and 61% identical to the <em>Fundulus</em> AhR-1 sequence, which is more similar to mammalian AhRs. Binding of the liganded Ah receptor to the DRE using a complementary pair of synthetic oligonucleotides, corresponding to wild type Ali receptor site of DRE3, could be demonstrated using rat and guinea pig derived cytosol. In contrast, DRE binding of liganded Ali receptor could not be demonstrated using flounder cytosol. These data show that the hepatic AhR pathway is only marginally present in flounder But since a good induction of CYP1A was observed in flounder exposed to TCDD, the flounder AhR pathway is functional and the apparent low responsiveness of flounder CYP1A towards PCB exposure can not be attributed to a non-functional Ah receptor pathway.<p>A more plausible explanation for the low CYP1A inducibility upon PCB exposure in flounder is direct inhibition of CYP1A catalytic activity by residual PCBs present in the microsomal suspension. Evidence for substrate inhibition by PCB congeners was provided in chapter 6. All of the PCB congeners tested, as well as Clophen A50, were capable of in vitro inhibition of the flounder CYP1A catalytic activity in a competitive way. The competitive inhibition of CYP1A activity by PCBs occurred at similar PCB concentrations in flounder as in rat. CB- 126 was found to be the most potent inhibitor of EROD activity whereas CB- 153 was least potent. The inhibition constants (K <sub>i</sub> ) of the tested PCBs were close to the Michaelis constant (K <sub>m</sub> ) for ethoxyresorufin. These studies also suggested a higher catabolic efficiency of the flounder CYP1A enzyme system towards ethoxyresorufin as compared to the rat CYP1A system.<p>Short-term exposure of flounder towards TCDD did not result in either retinoid or thyroid alterations ( <strong>chapter</strong> 4). In contrast alterations in retinoid and thyroid hormone levels were observed in flounder exposed to Clophen A50. But such changes were not dependent on the dose of PCB administered ( <strong>chapter</strong> 3). These results would suggest that PCBs are less endocrine disrupting in flounder as they are in mammals. However, in a chronic exposure experiment in which flounder were exposed to contaminated harbour sludge for three years, retinol levels in both plasma and liver were reduced (chapter 7). In addition, a negative correlation between hepatic retinol concentrations and CYP1A protein levels was observed, indicating involvement of PHAH inducible enzymes.<p><strong>Conclusions</strong><p>From these studies, the following main conclusions can be drawn:<p>1 . The commercial PCB mixture Clophen A50 only cause minor effects on hepatic microsomal CYP1A protein levels and associated EROD activity in flounder, even at Clophen A50 concentrations as high as 500 mg/kg body weight. This low responsiveness towards Clophen A50 exposure is observed despite a functional hepatic Ah receptor pathway in flounder.<p>2. The commercial PCB mixture Clophen A50 inhibit the TCDD induced CYP1A activity. This finding can, at least to some extend, be attributed to competitive inhibition of the CYP1A activity by residual PCB congeners present in the reaction mixture.<p>3. In contrast to rodents, short-term exposure of flounder towards PCBs or TCDD does not induce meaningful retinoid and thyroid hormone disrupting effects.<p>4. Long-term exposure of flounder towards contaminated harbour sludge causes a marked decline of plasma and hepatic retinoid levels.<p>5. Exposure up to 3 weeks to PCBs and TCDD does not induce any gross pathological effects in flounder, even at high doses of PHAHs administered.<p>Comparing the present results with the outcome of studies on mammals and a range of other fish species, the conclusion can be drawn that flounder is relatively insensitive to PCB exposure regarding the parameters studied. Hence this study does not provide evidence for the involvement of PCBs in the development of the lesions observed in flounder from Dutch coastal and estuarine areas.<p>One may speculate about the nature of other factors involved in the diseases concerned, like pollutants other than PCBs. PAHs for example, are known to bind covalently to DNA after metabolisation, thus possibly initiating carcinogenesis. One may also assume that synergistic actions may occur between certain pollutants. However, the present studies showed that both individual PCB congeners and complex mixtures of PCBs are competitive inhibitors of EROD activity in vitro. If such effects also occur in the wild, PCBs may even antagonise the toxic potential of other pollutants. For instance, inhibition of CYP1A activity by residual PCB congeners might result in reduced formation of reactive PAH metabolites and thus reducing the risk for covalent DNA binding and tumour formation. Hence, this would also decrease the likelihood that PAH initiate carcinogenesis. Furthermore, non-chemical background stressors may also be involved, such as physical disturbance and pathogenic micro-organisms in the case of infectious diseases.<p>A remark should be made about the usefulness of EROD activity as biomarker for monitoring exposure of fish to environmental pollutants such as PHAHs and PAHs. As a spinoff of the observed competitive inhibition of EROD activity by PCBs, one should reconsider the value of EROD activity as biomarker. On the other hand, induction CYP1A protein levels is not inhibited by PCBs. Therefore, CYP1A protein levels might be a better biomarker than EROD activity for exposure assessment of fish to aromatic compounds.<p>Finally, in view of the recent developments in science, more emphasis should be put on developing appropriate tools to study the role of environmental pollutants, such as PCBs or related compounds in carcinogenesis. For example, the RAS gene sequence has been identified in flounder (see appendix). Whereas in humans expression of the oncogen is characteristic for a number of tumors, this relation has yet to be elucidated in fish. Nevertheless, the study of alterations in the expression of onco- and suppressor genes might be a useful approach in further studies.