Cytochrome P4501A (CYP1A), a xenobiotic metabolizing enzyme found in allvertebrates, is highly induced following exposure to a number of organic contaminants.Several populations of teleosts residing in highly contaminated areas have been found toexhibit resistance to the toxic effects of contaminants, a condition characterized byreduced expression of CYP1A.Within this work I demonstrated that expression of CYP1A mRNA, protein, andactivity in caged rainbow trout (Oncorhynchus mykiss) was an effective biomarker ofpolychlorinated biphenyl (PCB) contamination. Furthermore, through the use of bothlaboratory and field studies, I demonstrated that several species inhabiting a PCBcontaminated site exhibited either acquired (Ameiurus natalis) or natural (Lepomiscyanellus) resistance to the CYP1A inducing effects of PCBs. Further studiescharacterized the response of several other Lepomis species to CYP1A inducingcompounds, demonstrating that the natural resistance of L. cynaellus is a characteristicshared by at least two other members of the genus. Lepomis species were relativelyinsensitive to CYP1A induction following PCB exposure yet exhibited highly inducedCYP1A levels following exposure to another CYP1A inducer, the model polyaromatichydrocarbon benzo[a]pyrene (BaP), suggesting a number of species within the genusLepomis may display natural resistance to certain classes of CYP1A inducingcompounds.Additional studies using responsive and resistant populations of killifish wereused to examine the consequences of resistance on fish physiology. Thyroid hormones,known to be altered by PCBs in mammals, were variable but did not differ significantlybetween responsive and resistant fish following PCB exposure. Treatment with PCBssuppressed production of the egg yolk precursor protein vitellogenin in primaryhepatocytes of responsive fish. Studies examining the developmental impacts of toxicantexposure demonstrated altered aspects of development in PCB responsive but notresistant Fundulus heteroclitus embryos exposed to polybrominated diphenyl ethers(PBDEs), compounds structurally related to PCBs. PBDE exposure in juvenile Ictaluruspunctatus failed to induce CYP1A or uridine diphosphate glucuronyltransferase(UDPGT) activity indicating PBDEs do not impact these commonly measuredtoxicological endpoints. The findings of this work describe novel pollutant responses in anumber of species with varying degrees of pollutant sensitivity and contribute to theunderstanding of toxicant induced alterations in teleost physiology.

Many physiological responses to dioxin-like compounds (DLCs), including polychlorinated dibenzodioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) are mediated by the aryl-hydrocarbon receptor (AhR). In birds, activation of the AhR stimulates the transcription of cytochrome P4501A (CYP1A) genes, including CYP1A4 and CYP1A5, and ultimately leads to expression of biotransformation enzymes, including ethoxyresorufin-O-deethylase (EROD). It is well established that potencies of different DLCs range over several orders of magnitude. There is also a wide variation among birds in their responsiveness to DLCs both in efficacy and threshold for effects. A molecular basis for this differential sensitivity has been suggested. Specifically, a comparison of the AhR ligand-binding domain (LBD) indicated that key amino acid residues are predictive of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) sensitivity. Based on sequencing of the AhR LBD from numerous avian species a sensitive classification scheme has been proposed (in order of decreasing sensitivity, chicken (type I; sensitive) > Common pheasant (type II; moderately sensitive) > Japanese quail (type III; insensitive)). A series of egg injection studies with White-leghorn chicken (<i>Gallus gallus domesticus</i>), Common pheasant (<i>Phasianus colchicus</i>) and Japanese quail (<i>Coturnix japonica</i>) were performed to determine whether molecular and biochemical markers of exposure to DLCs are predictive of the proposed classification scheme. In addition, I was interested in determining whether this classification scheme applies to other DLCs, specifically dibenzofurans. Determining which species are "chicken- like", "pheasant-like" and "quail-like" in their responses to DLCs should allow more refined risk assessments to be conducted as there would be less uncertainty about the potential effects of DLCs in those species for which population-level studies do not exist.<p> Several concentrations of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), 2,3,4,7,8-pentachlorodibenzofuran (PeCDF), or 2,3,7,8-tetrachlorodibenzofuran (TCDF) (triolein vehicle) were injected into the air cells of Japanese quail, Common pheasant and chicken eggs. Liver from 14 d post-hatch chicks was harvested for analysis of CYP1A4 and CYP1A5 mRNA abundance by quantitative real-time PCR (Q-PCR), and EROD activity. Lowest observed effective concentration (LOEC) and relative potency (ReP) values for CYP1A mRNA abundance and EROD activity were determined and used to make comparisons of sensitivity between each species and DLC potency within each species.<p> The TCDD is widely considered to be the most potent DLC and this is supported by the rank order of LOEC values for CYP1A5 mRNA abundance in White-leghorn chicken (TCDD > PeCDF > TCDF). CYP1A4 mRNA abundance and EROD activity in White-leghorn chicken were significantly increased in the lowest dose exposure groups of each of the three DLCs, so the potency of these compounds based on these endpoints was not established. Interestingly, TCDD was not the most potent DLC in Common pheasant and Japanese quail. In Common pheasant, PeCDF is the most potent as a CYP1A4 mRNA inducer, followed by TCDD and TCDF. However, TCDF was the most potent EROD activity inducer for Common pheasant, followed by PeCDF, and then TCDD. No significant increases were found in CYP1A5 mRNA abundance in pheasant within the tested dose ranges for all the three DLCs. No significant increases in either CYP1A5 mRNA abundance or EROD activity were found in Japanese quail. In addition, PeCDF and TCDF, but not TCDD, significantly increased CYP1A4 mRNA abundance.<p> According to the predicted relative sensitivity by comparing the AhR LBD amino acid sequences, the White-leghorn chicken is more responsive to DLCs than the Common pheasant which is more responsive than the Japanese quail. By comparing the relative sensitivity calculated based on the LOEC values from my study, the sensitivity order to TCDD and TCDF support the proposed molecular based species sensitivity classification scheme (chicken > pheasant > quail), while pheasant is almost as sensitive as chicken to PeCDF ( pheasant ¡Ý chicken > quail).<p> Taken together, the data suggest that TCDD is the most potent DLC in White-leghorn chicken, but not in Common pheasant, or or Japanese quail. The data suggest that in type II avian species PeCDF may be more potent than TCDD. In addition, I found in my study that different biomarkers have different responses, which depends on species and chemicals as well. These data provide further insight into avian sensitivities to DLCs.</p>

A rainbow trout liver cell line, RTL-W1, was used to evaluate the toxic potential of a heavy fuel oil (HFO) HFO 7102, and its fractions, which together with the HFO are referred to as the oil samples. The fractions were F2, F3, F3-1, F3-2 and F4 and had been prepared by low-temperature vacuum distillation by collaborators at Queen's University. For presentation to the cells, HFO 7102 and its fractions were made into High Energy-Chemically Enhanced Water Accommodated Fractions (HE-CEWAFs). The procedure for this involved adding Corexit 9500 to the oil samples, mixing them on a vortex, and letting the phases settle. The HE-CEWAFs were added to RTL-W1 cell cultures, and at various times afterwards cell viability and CYP1A induction were monitored. Cell viability was evaluated with two dyes, Alamar Blue, which monitors energy metabolism, and 5-carboxfluorescein diacetate acetoxymethyl ester (CFDA AM), which measures plasma membrane integrity. With both indicator dyes, Corexit 9500 was cytotoxic but the concentrations eliciting cytotoxicity varied with the cell culture media. In Leibovitz's L-15 with fetal bovine serum (FBS), which was the medium used for studying CYP1A induction, Corexit 9500 was only cytotoxic at concentrations of 0.1 % (v/v) and greater. For the oil samples, F3-2 at 1 mg/ml and F4 at 10 mg/ml, which were the highest testable concentrations for each, no loss of cell viability was observed over 24 h. The other oil samples were cytotoxic only at their highest testable concentrations, which ended being between 1 and 10 mg/ml. CYP1A induction was monitored in RTL-W1 as catalytic activity and as the level of CYP1A (P4501A) protein. The catalytic activity was assayed as 7-ethoxyresorufin o-deethylase (EROD) activity; the CYP1A protein level, by western blotting. The positive control was 2, 3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), which strongly induced both EROD activity and CYP1A protein. Corexit 9500 by itself neither induced EROD activity nor CYP1A protein. All the oil samples induced both EROD activity and CYP1A protein. With both endpoints, the most potent fraction was F3; the least potent, F3-2. As the induction of CYP1A is associated with the development of blue sac disease (BSD) and mortality in early life stages of fish, the results suggest that HFO 7102 and its fractions have the potential to reduce recruitment of young into adult fish populations. CYP1A induction by F3 was studied further, again through EROD activity and western blotting. As the F3 concentration was increased, EROD activity increased but declined at high concentrations, whereas CYP1A protein continued to increase. This suggests the presence of compounds in F3 that at high concentrations inhibit the catalytic activity of CYP1A. When F3 was presented to RTL-W1 cultures together with TCDD, CYP1A protein was induced but not EROD activity. Again this suggests that F3 contains inhibitor(s) of CYP1A as well as inducers. When cultures were exposed to either F3 or TCDD for 24 h and then followed by western blotting for up to 6 days after F3 or TCDD removal, CYP1A levels declined in F3 cultures but not in TCDD cultures. This suggests that RTL-W1 were able to inactivate CYP1A inducer(s) in F3 through metabolism. Overall the results suggest that the pattern of CYP1A induction by F3, and by extension, HFO involves complex interactions between the many chemical components in these mixtures. Likely the most important chemicals are the polycyclic aromatic hydrocarbons (PAHs).

Many physiological responses to dioxin-like compounds (DLCs), including polychlorinated dibenzodioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) are mediated by the aryl-hydrocarbon receptor (AhR). In birds, activation of the AhR stimulates the transcription of cytochrome P4501A (CYP1A) genes, including CYP1A4 and CYP1A5, and ultimately leads to expression of biotransformation enzymes, including ethoxyresorufin-O-deethylase (EROD). It is well established that potencies of different DLCs range over several orders of magnitude. There is also a wide variation among birds in their responsiveness to DLCs both in efficacy and threshold for effects. A molecular basis for this differential sensitivity has been suggested. Specifically, a comparison of the AhR ligand-binding domain (LBD) indicated that key amino acid residues are predictive of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) sensitivity. Based on sequencing of the AhR LBD from numerous avian species a sensitive classification scheme has been proposed (in order of decreasing sensitivity, chicken (type I; sensitive) > Common pheasant (type II; moderately sensitive) > Japanese quail (type III; insensitive)). A series of egg injection studies with White-leghorn chicken (<i>Gallus gallus domesticus</i>), Common pheasant (<i>Phasianus colchicus</i>) and Japanese quail (<i>Coturnix japonica</i>) were performed to determine whether molecular and biochemical markers of exposure to DLCs are predictive of the proposed classification scheme. In addition, I was interested in determining whether this classification scheme applies to other DLCs, specifically dibenzofurans. Determining which species are "chicken- like", "pheasant-like" and "quail-like" in their responses to DLCs should allow more refined risk assessments to be conducted as there would be less uncertainty about the potential effects of DLCs in those species for which population-level studies do not exist.<p> Several concentrations of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), 2,3,4,7,8-pentachlorodibenzofuran (PeCDF), or 2,3,7,8-tetrachlorodibenzofuran (TCDF) (triolein vehicle) were injected into the air cells of Japanese quail, Common pheasant and chicken eggs. Liver from 14 d post-hatch chicks was harvested for analysis of CYP1A4 and CYP1A5 mRNA abundance by quantitative real-time PCR (Q-PCR), and EROD activity. Lowest observed effective concentration (LOEC) and relative potency (ReP) values for CYP1A mRNA abundance and EROD activity were determined and used to make comparisons of sensitivity between each species and DLC potency within each species.<p> The TCDD is widely considered to be the most potent DLC and this is supported by the rank order of LOEC values for CYP1A5 mRNA abundance in White-leghorn chicken (TCDD > PeCDF > TCDF). CYP1A4 mRNA abundance and EROD activity in White-leghorn chicken were significantly increased in the lowest dose exposure groups of each of the three DLCs, so the potency of these compounds based on these endpoints was not established. Interestingly, TCDD was not the most potent DLC in Common pheasant and Japanese quail. In Common pheasant, PeCDF is the most potent as a CYP1A4 mRNA inducer, followed by TCDD and TCDF. However, TCDF was the most potent EROD activity inducer for Common pheasant, followed by PeCDF, and then TCDD. No significant increases were found in CYP1A5 mRNA abundance in pheasant within the tested dose ranges for all the three DLCs. No significant increases in either CYP1A5 mRNA abundance or EROD activity were found in Japanese quail. In addition, PeCDF and TCDF, but not TCDD, significantly increased CYP1A4 mRNA abundance.<p> According to the predicted relative sensitivity by comparing the AhR LBD amino acid sequences, the White-leghorn chicken is more responsive to DLCs than the Common pheasant which is more responsive than the Japanese quail. By comparing the relative sensitivity calculated based on the LOEC values from my study, the sensitivity order to TCDD and TCDF support the proposed molecular based species sensitivity classification scheme (chicken > pheasant > quail), while pheasant is almost as sensitive as chicken to PeCDF ( pheasant ¡Ý chicken > quail).<p> Taken together, the data suggest that TCDD is the most potent DLC in White-leghorn chicken, but not in Common pheasant, or or Japanese quail. The data suggest that in type II avian species PeCDF may be more potent than TCDD. In addition, I found in my study that different biomarkers have different responses, which depends on species and chemicals as well. These data provide further insight into avian sensitivities to DLCs.</p>

A rainbow trout liver cell line, RTL-W1, was used to evaluate the toxic potential of a heavy fuel oil (HFO) HFO 7102, and its fractions, which together with the HFO are referred to as the oil samples. The fractions were F2, F3, F3-1, F3-2 and F4 and had been prepared by low-temperature vacuum distillation by collaborators at Queen's University. For presentation to the cells, HFO 7102 and its fractions were made into High Energy-Chemically Enhanced Water Accommodated Fractions (HE-CEWAFs). The procedure for this involved adding Corexit 9500 to the oil samples, mixing them on a vortex, and letting the phases settle. The HE-CEWAFs were added to RTL-W1 cell cultures, and at various times afterwards cell viability and CYP1A induction were monitored. Cell viability was evaluated with two dyes, Alamar Blue, which monitors energy metabolism, and 5-carboxfluorescein diacetate acetoxymethyl ester (CFDA AM), which measures plasma membrane integrity. With both indicator dyes, Corexit 9500 was cytotoxic but the concentrations eliciting cytotoxicity varied with the cell culture media. In Leibovitz's L-15 with fetal bovine serum (FBS), which was the medium used for studying CYP1A induction, Corexit 9500 was only cytotoxic at concentrations of 0.1 % (v/v) and greater. For the oil samples, F3-2 at 1 mg/ml and F4 at 10 mg/ml, which were the highest testable concentrations for each, no loss of cell viability was observed over 24 h. The other oil samples were cytotoxic only at their highest testable concentrations, which ended being between 1 and 10 mg/ml. CYP1A induction was monitored in RTL-W1 as catalytic activity and as the level of CYP1A (P4501A) protein. The catalytic activity was assayed as 7-ethoxyresorufin o-deethylase (EROD) activity; the CYP1A protein level, by western blotting. The positive control was 2, 3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), which strongly induced both EROD activity and CYP1A protein. Corexit 9500 by itself neither induced EROD activity nor CYP1A protein. All the oil samples induced both EROD activity and CYP1A protein. With both endpoints, the most potent fraction was F3; the least potent, F3-2. As the induction of CYP1A is associated with the development of blue sac disease (BSD) and mortality in early life stages of fish, the results suggest that HFO 7102 and its fractions have the potential to reduce recruitment of young into adult fish populations. CYP1A induction by F3 was studied further, again through EROD activity and western blotting. As the F3 concentration was increased, EROD activity increased but declined at high concentrations, whereas CYP1A protein continued to increase. This suggests the presence of compounds in F3 that at high concentrations inhibit the catalytic activity of CYP1A. When F3 was presented to RTL-W1 cultures together with TCDD, CYP1A protein was induced but not EROD activity. Again this suggests that F3 contains inhibitor(s) of CYP1A as well as inducers. When cultures were exposed to either F3 or TCDD for 24 h and then followed by western blotting for up to 6 days after F3 or TCDD removal, CYP1A levels declined in F3 cultures but not in TCDD cultures. This suggests that RTL-W1 were able to inactivate CYP1A inducer(s) in F3 through metabolism. Overall the results suggest that the pattern of CYP1A induction by F3, and by extension, HFO involves complex interactions between the many chemical components in these mixtures. Likely the most important chemicals are the polycyclic aromatic hydrocarbons (PAHs).

Chronically contaminated killifish from Newark Bay (NB) NJ, and New Bedford Harbor (NBH) MA, have developed resistance to halogenated aromatic hydrocarbons that bind to and activate the aryl hydrocarbon receptor (AHR). To study the mechanisms of resistance, adult killifish were exposed to halogenated and non-halogenated AHR ligands and enzymatic and toxicological endpoints were measured in adult and embryonic fish. The chlorinated and non-chlorinated AHR ligands 3,34,4-tetrachlorobiphenyl (PCB77) and benzo-a-pyrene (B[a]P) induced cytochrome P450 1A (CYP1A) in reference site, but not in NB killifish. Expression of CYP3A (not part of the AHR gene battery) was inducible only in Flax Pond killifish. Basal expression of the phase II enzyme glutathione-s-transferase (GST) was higher in NB killifish. These results suggest that NB killifish are resistant to CYP1A induction by chlorinated and non-chlorinated AHR ligands. Higher basal GST activity observed in NB killifish could be protective against toxicity caused by contaminants found in this site. Activation of AHR and induction of CYP1A, by AHR ligands has been associated with the toxic effects caused by these chemicals. To determine the association between resistance to CYP1A induction and the toxicity caused by AHR ligands, CYP1A activity, developmental deformities and reactive oxygen species (ROS) production were measured in reference site and contaminated (NB and NBH) killifish embryos exposed to AHR ligands. 3,34,45-pentachlorobiphenyl (PCB126) and 3-methylcholantherene (3-MC) induced CYP1A, and ROS production in reference site embryos. NB and NBH embryos were resistant to PCB126 induction of CYP1A, but responded to 3-MC. Killifish embryos from NB and NBH were resistant to PCB126 induced deformities. PCB126 and 3-MC did not increase ROS production in NB or NBH killifish embryos. Alpha-naphthoflavone (ANF) (an AHR/CYP1A inhibitor) blocked PCB126 mediated deformities and CYP1A induction in reference site embryos, but increased ROS production. The P450 inhibitor, piperonyl butoxide (PBO) was able to block PCB126 mediated induction of CYP1A activity and ROS production. These results suggest that PCB126 induced deformities are dependent on activation of AHR and CYP1A induction. In chronically contaminated killifish populations, loss of sensitivity to coplanar PCBs and PAHs could be through reduced expression of AHR, or altered DNA sequence or methylation status of the CYP1A gene promoter. Hepatic AHR expression, measured by photoaffinity labeling, was lower in NB killifish than reference site animals, suggesting that NB killifish express less AHR protein. DNA sequence analysis did not reveal considerable differences between contaminated and reference site populations, however additional DNA fragments were observed in some promoters but not in others. The methylation of the CYP1A promoters was studied using methylation sensitive restriction enzymes and no differences were detected between reference site and NB killifish. Treatment with the DNA methyltransferase inhibitor AzaC did not restore CYP1A induction by PCB126 in NB killifish. These studies suggest that resistance to activation of AHR and induction of xenobiotic activating enzymes (CYP1A and CYP3A) in combination with increased expression of conjugating enzymes (GST) protects chronically contaminated killifish against these chemicals.

A exposição dos peixes a poluentes provoca danos nos organismos que podem ser identificados precocemente através de respostas biológicas. O presente estudo visou avaliar os efeitos do naftaleno e benzo(a)pireno em pampos da espécie Trachinotus carolinus. Foram avaliados os efeitos citogenotóxicos, histopatológicos e bioquímicos após exposições às concentrações de 0,9 M; 2,7 M e 8,1 M de NAP e BAP por períodos de 12, 24, 48 e 96 horas. O NAP causa quebra no DNA de eritrócitos de pampos em concentrações de 8,1 M e a partir de 12 horas de exposição. O BAP revelou ser genotóxico a partir da menor concentração e de 24 horas. A mutagenicidade de ambos os poluentes, avaliada através da indução de formação de micronúcleos e anormalidades nucleares eritrocitárias, também ocorre a partir de curtos períodos de exposição e freqüências de MN e ANE estão relacionadas com a duração da exposição. O período de exposição aos HPAs foi determinante na intensidade e severidade das lesões observadas nos tecidos dos peixes. A especificidade de CYP1A, observada segundo análise imunohistoquímica, ocorreu de maneira dose-dependente e evidenciada principalmente nos maiores períodos experimentais. Os poluentes orgânicos, nas condições experimentais utilizadas, não provocaram alteração significativa na atividade das enzimas catalase e GST da espécie. Os biomarcadores, citogenotóxicos e histopatológicos utilizados neste estudo, demonstraram ser ferramentas eficientes para aferir a toxicidade, genotoxicidade e mutagenicidade de NAP e BAP como também sua relação dose-resposta na espécie T. carolinus. / Effects of exposure of fish to pollutants can be identified through stress responses. The present study aims to evaluate the effects of naphthalene and benzo(a)pyrene in Florida pompanos, Trachinotus carolinus. Evidences from citogenotoxical, histopathological and biochemical studies showed that alterations caused by exposures to 0.9 M, 2.7 M and 8.1 M of NAP and BAP occurred within 12 to 96 hours. NAP at 8.1 M induced erythrocyte DNA strand breaks in pompanos since early periods of exposure. Genotoxic effects of BAP at the lowest concentration were documented soon after 24 hours of exposure. Mutagenotoxicity of both pollutants, as seen by the induction of MN and ENA, was revealed since early periods and their frequencies are related to the duration of exposure. Exposures to these PAHs, for longer periods, resulted in increased frequency and severity of lesions observed in fish tissues. Specificity of CYP1A, observed through immunohistochemical analyses, was related to the dose of the pollutants and mainly at longer periods of exposure. These organic pollutants, under the experimental conditions, did not interfere with the activity of liver catalase and GST of the species. The citogenotoxic and histopathologic biomarkers used in this study proved to be efficient tools to ascertain the toxicity, genotoxicity and mutagenesis of NAP and BAP, as well as their dose related response, in the species T. carolinus.

A exposição dos peixes a poluentes provoca danos nos organismos que podem ser identificados precocemente através de respostas biológicas. O presente estudo visou avaliar os efeitos do naftaleno e benzo(a)pireno em pampos da espécie Trachinotus carolinus. Foram avaliados os efeitos citogenotóxicos, histopatológicos e bioquímicos após exposições às concentrações de 0,9 M; 2,7 M e 8,1 M de NAP e BAP por períodos de 12, 24, 48 e 96 horas. O NAP causa quebra no DNA de eritrócitos de pampos em concentrações de 8,1 M e a partir de 12 horas de exposição. O BAP revelou ser genotóxico a partir da menor concentração e de 24 horas. A mutagenicidade de ambos os poluentes, avaliada através da indução de formação de micronúcleos e anormalidades nucleares eritrocitárias, também ocorre a partir de curtos períodos de exposição e freqüências de MN e ANE estão relacionadas com a duração da exposição. O período de exposição aos HPAs foi determinante na intensidade e severidade das lesões observadas nos tecidos dos peixes. A especificidade de CYP1A, observada segundo análise imunohistoquímica, ocorreu de maneira dose-dependente e evidenciada principalmente nos maiores períodos experimentais. Os poluentes orgânicos, nas condições experimentais utilizadas, não provocaram alteração significativa na atividade das enzimas catalase e GST da espécie. Os biomarcadores, citogenotóxicos e histopatológicos utilizados neste estudo, demonstraram ser ferramentas eficientes para aferir a toxicidade, genotoxicidade e mutagenicidade de NAP e BAP como também sua relação dose-resposta na espécie T. carolinus. / Effects of exposure of fish to pollutants can be identified through stress responses. The present study aims to evaluate the effects of naphthalene and benzo(a)pyrene in Florida pompanos, Trachinotus carolinus. Evidences from citogenotoxical, histopathological and biochemical studies showed that alterations caused by exposures to 0.9 M, 2.7 M and 8.1 M of NAP and BAP occurred within 12 to 96 hours. NAP at 8.1 M induced erythrocyte DNA strand breaks in pompanos since early periods of exposure. Genotoxic effects of BAP at the lowest concentration were documented soon after 24 hours of exposure. Mutagenotoxicity of both pollutants, as seen by the induction of MN and ENA, was revealed since early periods and their frequencies are related to the duration of exposure. Exposures to these PAHs, for longer periods, resulted in increased frequency and severity of lesions observed in fish tissues. Specificity of CYP1A, observed through immunohistochemical analyses, was related to the dose of the pollutants and mainly at longer periods of exposure. These organic pollutants, under the experimental conditions, did not interfere with the activity of liver catalase and GST of the species. The citogenotoxic and histopathologic biomarkers used in this study proved to be efficient tools to ascertain the toxicity, genotoxicity and mutagenesis of NAP and BAP, as well as their dose related response, in the species T. carolinus.

Les développements actuels liés à l'ingénierie tissulaire et aux microtechnologies permettent aujourd'hui de proposer de nouveaux outils de criblages in vitro pour les études de toxicité. Nous proposons de développer une biopuce à cellules mimant un organe in vitro. Afin de valider notre approche, nous présentons l'influence du microenvironnement sur la culture de lignées cellulaires rénales (MDCK) et hépatiques (HepG2/C3A). <br />Dans cette étude, nous avons testé trois débits (0, 10 et 25 µL/min) et trois ensemencements cellulaires. Enfin, nous avons soumis notre biopuce à trois chargements de chlorure d'ammonium (0, 5 et 10 mM) afin de démontrer le potentiel de ce modèle pour de futures applications liées à la toxicité. L'activité cellulaire en biopuce a été suivie par la prolifération des cellules, les consommations de glucose et de glutamine, les productions d'albumine et d'ammoniac et enfin, par l'activité enzymatique de détoxification des CYP 1A.<br />En condition dynamique, il a été observé une augmentation des consommations et des productions cellulaires au regard des conditions statiques. L'activité de détoxification des CYP 1A a été également accrue. En présence du chlorure d'ammonium les réponses cellulaires furent similaires en biopuce au regard des conditions de culture standard en Pétri. De plus, le chlorure d'ammonium a semblé induire l'activité des CYP 1A en biopuce. <br />Par cette étude, nous montrons la pertinence de notre biopuce pour des tests de toxicité in vitro en condition dynamique. Ce nouveau modèle de culture cellulaire in vitro pourra à terme être applicable aux études de criblages dans les industries chimiques, pharmaceutiques et cosmétiques.

<p>A gill filament-based assay for the cytochrome P450 1A (CYP1A)-catalysed activity ethoxyresorufin <i>O</i>-deethylase (EROD) was developed in rainbow trout (<i>Oncorhynchus mykiss</i>) and applied to Atlantic salmon (<i>Salmo salar</i>), Arctic charr (<i>Salvelinus alpinus</i>), Atlantic cod (<i>Gadus morhua</i>), saithe (<i>Pollachius virens</i>), and spotted wolffish (<i>Anarhichas minor</i>). Exposure to waterborne β-naphthoflavone (βNF; 10^-6 M) induced branchial EROD activity in all species but the spotted wolffish. In rainbow trout exposed to low concentrations of benzo[a]pyrene (BaP; 10^-9 M) and the textile dye indigo (10^-8 M) the gills responded more rapidly than the liver to BaP, and indigo induced branchial but not hepatic EROD activity.</p><p>A CYP1A-dependent BaP adduct formation was shown in gills of fish exposed to waterborne ³H-BaP, i.e. the adduct formation was enhanced by βNF and blocked by ellipticine (CYP1A inhibitor). The predominant location for BaP adducts was the secondary lamellae (most exposed part of the gill filament), whereas the CYP1A enzyme was also present in the primary lamellae of the gill filament. Hence, in addition to the cell-specific expression of CYP1A an important determinant for the localisation of adducts seemed to be the bioavailability of BaP. This idea is supported by the fact that the CYP1A enzyme was induced only in secondary lamellae by BaP (10^-7 M) and indigo (10^-6 M), whereas it was induced in both primary and secondary lamellae by 3,3´,4,4´,5-pentachlorobiphenyl (10^-8 M). Apparently, readily metabolised inducers (BaP and indigo) are biotransformed in the secondary lamellae.</p><p>My results show that gill filament EROD activity is a sensitive biomarker of exposure to waterborne dioxin-like pollutants, and that the assay has potential for use in monitoring. Furthermore, the results suggest that readily metabolised dioxin-like compounds absorbed via the gills may undergo first-pass metabolism in the gill cells and therefore remain undetected by monitoring of EROD activity in the liver.</p>