Potency of Polycyclic Aromatic Hydrocarbons (PAHs) for Induction of

Article pubs.acs.org/est

Potency of Polycyclic Aromatic Hydrocarbons (PAHs) for Induction of Ethoxyresorufin‑O‑deethylase (EROD) Activity in Hepatocyte Cultures from Chicken, Pekin Duck, And Greater Scaup Jessica A. Head,*,§ Richard W. Jeffery,‡ Reza Farmahin,‡ and Sean W. Kennedy†,‡ §

Department of Natural Resource Sciences, McGill University, Montreal, Quebec H9X 3 V9, Canada Centre for Advanced Research in Environmental Genomics, Department of Biology, University of Ottawa, Ottawa, Ontario K1N 6N5, Canada ‡ Environment Canada, National Wildlife Research Centre, Ottawa, Ontario K1A 0H3, Canada †

S Supporting Information *

ABSTRACT: The potency of tetrachlorodibenzo-p-dioxin (TCDD) and 18 polycyclic aromatic hydrocarbons (PAHs) for induction of ethoxyresorufin-Odeethylase (EROD) activity was assessed in primary hepatocyte cultures prepared from chicken (Gallus domesticus), Pekin duck (Anas platyrhynchos domesticus), and greater scaup (Aythya marila). TCDD and 8 of the PAHs induced EROD activity in a concentration-dependent manner. Seven of these were previously shown to be acutely toxic to avian embryos, while the 10 congeners that did not produce an EROD response caused limited mortality. The rank order potency of the EROD-active congeners in all three species was as follows: TCDD > dibenz[ah]anthracene > benzo[k]fluoranthene > indeno[1,2,3-cd]pyrene > benzo[a]pyrene > chrysene ≈ benz[a]anthracene ≈ benz[ghi]perylene > benzo[b]naphtho[2,3-d]thiophene. Chicken hepatoctyes were more sensitive than duck hepatocytes to EROD induction by all test compounds, but the gap in species sensitivity was 100-fold for TCDD, and generally ≤10-fold for PAHs. This study is the first to use in vitro methods to rank the AHR-mediated potency of PAHs in birds. These data may be useful for assessing risks associated with exposure to PAHs in the environment.

■

INTRODUCTION Polycyclic aromatic hydrocarbons (PAHs) are a class of structurally related contaminants that are often present as complex mixtures in the environment. PAHs are broadly distributed, and can originate from both natural sources, such as forest fires, and anthropogenic sources, such as combustion of fossil fuels, spillage of petroleum, and industrial processes.1 Birds are exposed to PAHs through ingestion of water and food, preening of feathers covered with oil, dermal absorption, and inhalation.2 Exposure of avian species to PAHs is often studied in environments polluted by accidental releases of oil3 or oil extraction.4 PAHs induce a wide range of toxic effects in vertebrates, including developmental abnormalities, immune suppression, reproductive effects, carcinogenicity, and liver toxicity.5 Effects of PAHs on birds have been studied under both laboratorybased, and environmental exposure scenarios. Laboratory studies have shown that PAHs can cause developmental abnormalities, reduced weight gain, and embryo lethality when injected into fertile eggs, or when oil is painted onto eggshells. Juvenile and adult birds exposed to PAHs exhibit reduced weight gain, endocrine disruption, immune system effects, and biochemical effects.2 Several studies report on effects in wild birds living in environments contaminated with © 2015 American Chemical Society

PAHs due to the presence of oil or oil-related products. PAHs were associated with elevated monooxygenase activity and chromosomal damage in lesser scaup (Aythya af fınis) environmentally exposed to petroleum.6 Tree swallows (Tachycineta bicolor) nesting in wetlands contaminated with oil sands processing materials had elevated thyroid hormone levels, although it is not clear if these effects are related to PAHs or other chemicals.4 Biomarkers of exposure to PAHs were elevated in harlequin ducks (Histrionicus histrionicus) living in areas affected by the Exxon Valdez oil spill up to 20 years after the accident.3 While the labile nature of PAHs often makes it difficult to establish causal links between chemical exposure and specific health outcomes in environmental samples, these studies suggest that PAHs can contribute to diverse and chronic effects in wild birds. Risk assessment for PAHs is complicated by the fact that this class of contaminants is composed of hundreds of different congeners with wide-ranging potencies. These are present in environmental samples as complex mixtures, and combinations Received: Revised: Accepted: Published: 3787

June 23, 2014 February 18, 2015 February 23, 2015 February 23, 2015 DOI: 10.1021/acs.est.5b00125 Environ. Sci. Technol. 2015, 49, 3787−3794

Article

Environmental Science & Technology and concentrations of congeners vary among samples. There are relatively few data describing the relative potencies of PAH congeners in birds. Brunstrom and colleagues describe embryo mortality following injection of PAHs into fertilized eggs of chicken, turkey, and duck,7−9 but comprehensive, environmentally relevant, data sets are generally lacking. This is presumably due to the difficulty of performing egg injection experiments for multiple congeners and with wild species. Because these difficulties are inherent to in vivo toxicity testing, additional in vitro methods for assessing the potency of PAHs in multiple species of birds are required. With this study, we take a biochemical approach to evaluating effects of PAHs in birds. The relative potency of PAHs for induction of ethoxyresorufin-O-deethylase (EROD) activity was assessed in hepatocytes cultured from three avian species. Induction of EROD activity occurs through the aryl hydrocarbon receptor (AHR), a nuclear transcription factor that is activated upon binding to PAHs and related halogenated aromatic hydrocarbons (HAHs). The activated AHR promotes transcription of responsive genes, including cytochrome P4501A (CYP1A), a xenobiotic-metabolizing monooxygenase with broad substrate specificity.10 Activity of the CYP1A enzyme is commonly measured as EROD activity. Both PAHs and HAHs are capable of inducing high levels of CYP1A expression and EROD activity, but PAHs are readily metabolized by CYP1A enzymes, whereas HAHs are resistant to metabolism. The objectives of this study were to (1) describe the AHRmediated potency of a selection of PAH congeners in multiple avian species in vitro, and (2) relate these data to current knowledge regarding the in vivo toxicity of PAHs in birds. A correlation between in vivo toxicity of PAHs and AHRmediated responses has been observed in fish cell lines, but has not been explored in birds.11 We established avian-specific relative potency values for 18 PAHs based on induction of EROD activity in primary embryo hepatocyte cultures from chicken (Gallus domesticus), Pekin duck (Anas platyrhynchos domesticus), and greater scaup (Aythya marila). Ducks were chosen as a test species because they have a high risk of exposure to PAHs through contact with oil in the aquatic environment. Additionally, in the Great Lakes region, diving ducks such as greater scaup subsist mainly on zebra mussels; a filter-feeding invasive species that can accumulate high levels of contaminants, including PAHs.12 The two domestic species, Pekin duck and chicken, were selected to provide a point of comparison for the wild scaup, and to explore differences in sensitivity among species.

Table 1. Rank Order Potency of Polycyclic Aromatic Hydrocarbons (PAHs) Tested in This Studya

a

The congeners are listed in order of their potency for induction of EROD activity in cultured chicken embryo hepatocytes. Potency is expressed as an average value relative to the potency of TCDD (RePav, see text for details). Congeners shown in grey did not induce EROD activity. In ovo toxicity data for selected congeners, as reported by Brunstrom et al. (1991), are also shown. *Data from Brunstrom et al. 1991. PAHs were injected into chicken eggs on day 7 of incubation and mortality was assessed 72 h later. In an initial study, eggs were injected with a single concentration of 300 ug/kg egg. The four most potent were then tested at multiple concentrations and LD50 were determined.

incubated in a Petersime Model XI incubator (Gettysburg, OH) set to 37 °C and 60% relative humidity. Embryos were euthanised on day 19 of incubation (2 days prehatch). Fertile unincubated Pekin duck eggs were obtained from a local commercial supplier (Les Entreprises Simetin Inc., Mirabel, PQ), and were incubated in a Curfew Model RX250 incubator at 37 °C and 80% relative humidity. Each afternoon, the incubator was left open for 10 to 15 min, and the eggs were sprayed with a mist of double distilled water. Pekin duck embryos were euthanised on day 26 of incubation (2 days prehatch). Greater scaup eggs were collected from nests in the Western Mirage Islands of Yellowknife Bay, Great Slave Lake, NWT. Eggs that had been naturally preincubated for 10 to 18 days were selected for collection. Given that these eggs were collected from the wild, they may have contained background levels of EROD-inducing chemicals. However, a previous laboratory-based study suggests that environmental levels of AHR ligands in eggs would generally not be expected to affect the potency of test compounds in cell culture.13 The greater scaup eggs were artificially incubated in a Curfew Model RX250 incubator under the same conditions as described for Pekin duck. Scaup embryos were euthanised at pipping. All procedures were conducted according to protocols approved

■

MATERIALS AND METHODS Test Chemicals. A list of the test chemicals used in this study is presented in Table 1. TCDD was provided by Dr. J. Ryan (Health Canada, Ottawa, Ontario). The 18 PAHs tested were purchased from commercial suppliers as follows (chemical abbreviations outlined in Table 1): Ant, BaA, BaP, BeP, BghiP, Chr, DahA, Fln, Flu, Phn, Per, and Pyr (Sigma, St. Louis, MO), MP, and BkF (Fluka, Milwaukee WI), Cor, and DBT (Aldrich, Milwaukee WI), IdP (Supelco, Mississauga, Ontario), and BNT (EQ Laboratories, Atlanta GA). All compounds were stated to be least 99% pure. Test chemicals were received in powdered form and dissolved in dimethyl sulfoxide (DMSO). Sources of Eggs and Incubation Conditions. Fertile unincubated White Leghorn chicken eggs obtained from the Canadian Food Inspection Agency (Ottawa, Ontario) were 3788

DOI: 10.1021/acs.est.5b00125 Environ. Sci. Technol. 2015, 49, 3787−3794

Article

Environmental Science & Technology

Figure 1. Concentration-dependent effects of TCDD and PAHs on EROD activity in cultured embryo hepatocytes from chicken (circles), Pekin duck (squares), and greater scaup (triangles). A representative curve is shown for each combination of chemical and species tested. Data points represent the average of three wells on a single plate. Error bars represent standard error. Chemical abbreviations are specified in text and in Table 1

Relative Potency (ReP). The relative potency of two different compounds is often determined by calculating the ratio of the EC50 values for each curve. Two basic assumptions are made with this method; that the slopes of the concentration−response curves are parallel, and the maximal responses (i.e., efficacies) are equivalent. Neither assumption was met in the current study (Figure 1), as is commonly the case for compounds that bind to the AHR. We therefore used a multiple point estimate (MPE) method outlined by Villeneuve and colleagues16 with some modification as previously described.17 This method makes allowances for unequal slopes and efficacies. The potency of each PAH was assessed relative to the potency of TCDD in the same species. In accordance with the MPE method, potency values were calculated at multiple points along each concentration−response curve. Because of the variability in maximal responses to different test compounds, all data were expressed as percentages of the maximal EROD response to TCDD. For example, we calculated the concentration of PAH that elicited an EROD response equal to 10% of the maximal response to TCDD (TCDD10). The same calculation was performed for potency estimates at 20%, 50%, and 80% of the maximal response to TCDD (TCDD20, TCDD50, and TCDD80). When the maximal response produced by a PAH congener was less than 50% or 80% of the maximal response to TCDD, the TCDD50 and/or TCDD80, parameters could not be determined for that congener. In these cases, an additional parameter, TCDDmax, was defined. The subscript “max” refers to the highest observed EROD response for the PAH congener, expressed as a percentage of the maximal EROD response for TCDD. ReP values for each of these parameters (ReP TCDD10 , ReP TCDD20 , ReP TCDD50 , RePTCDD80, and RePTCDDmax) were calculated by dividing the

by the Animal Care Committee at the National Wildlife Research Centre (Environment Canada, Ottawa, Canada). Preparation and Dosing of Hepatocyte Cultures. Primary cultures of chicken, Pekin duck, and greater scaup embryo hepatocytes were prepared using an adaptation of a method previously described.14 Briefly, livers from 25−30 embryos were pooled and digested for 45 min in 75 mL of a 0.5 mg/mL collagenase (Sigma) solution (chicken) or for 3 cycles of 10 min in 50 mL of 1.0 mg/mL collagenase solution (Pekin duck, greater scaup). Isolated hepatocytes were weighed and added to 48-well plates such that each well contained 60 μg protein in 0.5 mL of Waymouth’s medium (Gibco). After incubation for 24 h at 37 °C with 5% CO2, cells were exposed to 2.5 μL/well of DMSO, or of DMSO solutions of PAH or TCDD. The concentrations used varied according to the expected potency of each chemical, and ranged from nominal values of 0.0001 to 1000 nM. Three replicate wells of hepatocytes were treated with each concentration of test compound. Each 48-well plate contained a full concentration− response curve, and a minimum of 2 replicate plates were analyzed for each chemical/species pair. After the addition of test chemicals, cells were incubated for a further 24 h, after which time the medium was discarded, the cells were washed with phosphate buffered saline, and the plates were frozen on dry ice and stored at −80 °C until analysis. EROD Assays. EROD assays were carried out directly in the wells of 48-well plates as previously described.15 Data Analysis. Concentration−Response Curves. EROD activity was plotted against the logarithm of TCDD concentration or PAH concentration, and fit to a modified Gaussian curve as described elsewhere.14 Data from a minimum of two 48-well plates were used to create each curve. 3789


Article


curves. The rank order of RePavg values was nearly identical for the three species tested (Figure 2). From the most to the least

parameter value for TCDD, by the parameter value for the PAH congener (i.e., REPTCDD20 = TCDD20 for TCDD/ TCDD20 for PAH). We also calculated RePavg (the average of all ReP values) and RePrange (the range encompassing the lowest ReP and the highest ReP) (Supporting Information, SI, Table S1). EC10, EC20, EC50, and EC80 values were calculated from each concentration−response curve. These parameters were not used in the ReP estimates, but they are presented in the SI (Table S2). Relative Sensitivity (ReS). We assessed the sensitivity of hepatocytes from each species to EROD induction by the test chemicals using the lowest observed effect concentration (LOEC). This was defined as the lowest administered concentration of PAH or TCDD that caused a significant increase in EROD activity over baseline values. LOEC was used for interspecific comparisons of chemical sensitivity rather than EC50 because there were large differences in the maximal EROD response among species (Figure 1). EC50 can greatly overestimate relative sensitivity when there are large differences in maximal response.18,19 The ReS of each species was calculated relative to the response in chicken. For each congener, ReS = LOEC chicken/LOEC duck (SI Table S3). Statistics. The values from replicate 48-well plates were averaged to create a composite concentration−response curve for each chemical and species pair. TCDD10, TCDD20, TCDD50, TCDD80, TCDDmax, EC10, EC20, EC50, EC80, and maximal response values were calculated from these curves. LOEC values were determined using a one-way ANOVA (p < 0.05) followed by Dunnett’s test (p < 0.05), as previously described.18 A minimum sample size of 6 wells of hepatocytes was used to calculate the LOEC.

■

RESULTS EROD Concentration−Response Curves. Of the 18 PAHs investigated, 8 induced EROD activity in cultured hepatocytes. Representative concentration−response curves for TCDD, and these 8 PAHs are presented in Figure 1. EROD concentration−response curves were variable in terms of the height of maximal response (efficacy) and slope. In chicken hepatocytes, all curves were biphasic with a sigmoidal increase up to the point of maximal EROD activity, followed by a decrease at higher concentrations of test compound. This decrease has previously been observed,20 and is thought to be caused by competitive inhibition of the CYP1A enzyme by the inducer.21 In hepatoyctes cultured from Pekin duck and greater scaup, the 8 active congeners caused concentration-dependent increases in EROD activity, but curves were shifted to the right. In addition, the magnitude of the induction was smaller, with the largest efficacies observed in the chicken embryo hepatocytes, followed by Pekin duck, and then greater scaup (Figure 1). To allow for better visualization of the data for the two duck species, concentration−response curves with modified y-axis scaling are presented in the SI (Figures S1 and S2). In both duck species, several congeners did not produce a maximal response at the concentrations tested (SI Table S3 and Figures S1 and S2). Relative Potency. The potency of the 8 active PAHs was assessed relative to TCDD at multiple points along each curve, as described in the Methods section. The average of these ReP values (RePavg) was used to establish the rank order potency, while the range (RePrange) gave an indication of the uncertainty in ReP values caused by nonparallel concentration−response

Figure 2. Relative potency (ReP) for EROD induction by TCDD and 8 PAHs in avian embryo hepatocytes. ReP values were calculated from the EROD concentration−response curves illustrated in Figure 1. For each species, the potency of the test compound was assessed in relation to TCDD at various points on the curve (TCDD10, TCDD20, TCDD50, TCDD80, TCDDmax; see Methods for full description). These values were then converted to ReP values by dividing the value for TCDD by the value for the PAH (e.g., TCDD10/PAH10). Gray bars represent the range of ReP values calculated for each compound (RePrange), and black squares represent the average ReP value (RePavg).

potent, the order was as follows: TCDD > DahA > BkF > IdP > BaP > Chr ≈ BaA ≈ BghiP > BNT. Data are presented in this order in all figures and tables. Relative Sensitivity. The relative sensitivity (ReS) of hepatocytes cultured from chicken, Pekin duck and greater scaup to induction of EROD activity by TCDD and PAHs was assessed using LOEC values (SI Table S3). Hepatocytes from chicken were most sensitive to induction of EROD by all compounds, except for BaP, which was equipotent in the three 3790


Article


gives an indication of the uncertainty of the ReP estimates; if concentration−response curves are parallel, RePs from different points along the curve are equivalent. In the current study, the spread between the top and bottom of each RePrange value was less than 1 order of magnitude (SI Table S1). With only three exceptions (BkF, BaP, and DahA in greater scaup), the top of the range was less than 3-fold higher than the bottom of the range. This degree of variability in RePrange is similar to what was previously reported for PAHs in the H4IIE-luc bioassay,30 and lower than values reported for EROD induction in PLHC1 fish hepatoma cells.22 The rank order potencies of TCDD and the 8 active PAHs were evaluated using RePavg values, and were found to be nearly identical in the three species (Figure 2). From high to low potency, the order was: TCDD > DahA > BkF > IdP > BaP > Chr ≈ BaA ≈ BghiP > BNT. This agrees well with values that were previously reported for induction of EROD activity in chicken embryos. Machala et al. (1996) injected PAHs into chicken eggs on day 14 of incubation and assessed EROD activity 24 h later.24 They found that DahA and BkF were strong EROD-inducers, while BaA, BaP, and Chr were relatively weak inducers. The rank order potency determined from RePavg values in our study also corresponds well with RePs for EROD and other AHR-mediated responses in mammalian and fish cell lines.22,31−37 In agreement with our findings, all of these studies found that BkF and/or DahA were the most potent congeners, while Chr, BaA, and BghiP were among the least potent. IdP and BaP were generally ranked as having moderate potency. Although there are exceptions to this pattern,38 it appears that the rank order potency of PAHs for induction of AHR-mediated responses is very consistent across species, cell types, and biochemical assays, including avian cells. Relative Sensitivity (ReS). A second goal of this study was to explore differences in sensitivity to PAHs among avian species. Previous research has shown that both in vitro and in vivo responses to HAHs can vary by several orders of magnitude in birds, and that chickens are particularly sensitive.20,39,40 For example, chickens are approximately 45 times more sensitive than Japanese quail (Coturnix japonica) to the embryolethal effects of TCDD,41 and 25 times more sensitive to EROD induction by TCDD in cultured cells.42 There are relatively few data reporting on the sensitivity of duck species to HAHs, but the literature suggests that duck embryos can tolerate much higher concentrations of PCBs than chickens. Injection of 1000 or 5000 μg/kg of PCB 77 into eggs of common eider (Somateria mollissima) and mallard duck (Anas platyrhynchos domesticus), respectively, caused no observable effects, while a 200- or 1000-fold lower concentration (5 μg/kg PCB 77) caused adverse effects in chicken embryos.43 Injection of 0.4 μg/kg PCB 126 into the air cell of fertilized chicken eggs prior to incubation was associated with increased embryo mortality, deformities, and EROD activity, while a 4-fold higher concentration caused no effects in Pekin duck embryos.44 Differences in species sensitivity appear to be more moderate for PAHs. A study that injected BkF, or a mixture of 18 PAHs, into the yolk of chicken, turkey, domestic duck, and eider duck embryos early in incubation found that these four species were nearly equally sensitive to the test compounds.9 Collectively, these in vivo studies suggest that differences in sensitivity between chicken and duck are more moderate for PAHs than HAHs. This finding is reflected in the in vitro data presented here. We used the LOEC to evaluate relative

test species (Figure 3). The gap in sensitivity among species was larger for TCDD than for any of the PAHs. TCDD was

Figure 3. Lowest observed effect concentration (LOEC) of TCDD and 8 PAHs for induction of EROD activity in chicken (circles), Pekin duck (squares), and greater scaup (triangles) embryo hepatocytes. LOEC was calculated as the first concentration of test chemical that elicited an EROD response that was significantly elevated above background levels. Statistical significance was determined using a oneway ANOVA (p < 0.05) followed by Dunnett’s test (p < 0.05). LOEC is a discrete parameter, determined by the concentrations of test compound that were administered, and not calculated from the curve.

100 times more potent as an EROD-inducer in chicken hepatocytes than in hepatocytes from the two species of ducks. Almost all of the PAHs were in the range of 3−10 times more potent in chicken hepatoyctes. The two exceptions were BaP (equipotent in all three species) and DahA (approximately 30 times more potent in chicken than in greater scaup). It should be noted that the LOEC is a discrete parameter, determined by the concentrations of test compound that were administered, and not calculated from the curve. The ReS values presented here are therefore order of magnitude estimates.

■

DISCUSSION We describe concentration-dependent induction of EROD activity by TCDD and 18 PAHs in chicken and duck embryo hepatocytes. To our knowledge, this study is the first to establish in vitro potency values for PAH congeners in avian species. Of the 18 PAHs tested, 8 (BkF, DahA, IdP, BaA, BghiP, Chr, BaP, and BNT) induced concentration-dependent increases in EROD activity. These same PAH congeners have been described as producing AHR-mediated biochemical effects in other cell culture systems.22 PAHs are known to bind to the AHR23 and are associated with significant induction of EROD activity in embryonic chicken liver.7,24 Associations between EROD activity and PAHs or oil have also been reported in wild birds exposed experimentally,25 and in the environment.3,26−29 Ours is the first study to describe a concentration-dependent relationship between PAHs and EROD activity in avian cell cultures, but other studies have demonstrated a similar relationship in fish and mammalian cell lines.22 Relative Potency (ReP). The AHR-mediated potencies of the test chemicals were assessed at multiple points along each concentration−response curve according to the MPE method (as described in the Methods section). TCDD was included as a test compound in each experiment as a reference point for calculating RePs. Under the MPE method, the RePrange value 3791


Article


developmental timing and the length of exposure to PAHs may therefore be critical factors affecting the relationship between biochemical and toxicological responses. Effects of length of exposure time on EROD EC50 values have previously been observed for certain PCB congeners in avian embryo hepatocytes,50 and for biochemical responses to PAHs in various cell types.22,30 Moreover, different congeners are metabolized at different rates, potentially altering the rank order of ReP values over time.35 In general, EC50 values increase with increasing exposure time when the inducer is also metabolized by the enzyme. For example, we found exposure of chicken embryo hepatocytes to DahA for 12, 24, 48, or 72 h shifted EROD concentration−response curves to the right (unpublished data). These data imply that PAH metabolism may be important for comparisons of in vitro and in vivo data. More detailed in vivo data will be needed in order to further qualify the relationship between EROD-induction potency and in vivo responses to PAHs in birds. In broad terms, our data show that PAH congeners that are toxic to embryos in vivo, are also active as EROD-inducers in cultured cells (with the exception of BghiP). More specific relationships have been defined for PAHs in fish cell lines. EROD EC50 values for 5 PAHs in a rainbow trout liver cell line were predictive of EC50s in vivo.11 In chicken embryo hepatocytes, EROD EC50 values were significantly correlated with LD50 values for TCDD and PCB congeners,51 and similar relationships have been observed in mammalian cell lines.52 Relevance to Risk Assessment. Our data establish the rank order potency of PAH congeners for an AHR-mediated response in cultured cells. The direct applicability of these data to risk assessment may be minimal; congeners that were EROD inducers also tended to be acutely lethal to chicken embryos, but there was no clear correspondence between EROD-based ReP values and avian toxicity data. Additional LD50 values and a better understanding of how metabolism affects the relationship between in vitro and in vivo potency may help to clarify this relationship. Knowledge about species differences in susceptibilities to environmental contaminants is an important component of risk assessment for avian wildlife. Our data describing species differences in biochemical responses to TCDD and PAHs in birds add to a growing understanding that the relative potency of AHR ligands is not consistent across different species of birds,19 and correspond well with in vivo data from the literature (as discussed above). We found that the gap in sensitivity between chicken and two species of duck was greater for TCDD than it was for any of the PAHs (Figure 3). Because ReP values were defined in relation to TCDD, this resulted in RePs that were consistently orders of magnitude lower in chicken than in ducks (Figure 2). Our work cautions against the use of data from different species or taxa to develop ReP values, as has been done previously for PAHs.53 It suggests that two species exposed to a similar mixture of PAHs and HAHs in the environment may be responsive to different components of the mixture, depending on their individual sensitivities to each class of chemicals.

sensitivity to EROD-induction among the three species tested, and found that the gap in sensitivity between chicken and the two species of duck was larger for TCDD than it was for any of the PAHs (Figure 3). TCDD was 100 times more potent as an EROD inducer in hepatocytes from chicken than in hepatocytes from Pekin duck or greater scaup. This is in agreement with previously reported values for EROD induction by TCDD in hepatocytes cultured from chicken and domestic duck.20 In contrast, the difference in LOEC values between chicken and the 2 duck species was ≤10-fold for the PAH congeners tested (with the exception of DahA, which was approximately 30 times more potent in chicken than in greater scaup). One congener, BaP, was equipotent in all three species, but the magnitude of the EROD response to each test compound was lower in the duck species than in chicken. Additionally, while several PAH congers were equipotent in the two species of duck, maximal EROD activity was consistently lower in greater scaup than in Pekin duck (SI Table S3). The molecular basis for species differences in sensitivity to AHR ligands is well-described for dioxin-like HAHs, but not for PAHs. Two key amino acids in the ligand binding domain of the avian AHR are critical for determining species sensitivity to the biochemical and toxicological effects of dioxin-like HAHs.39,45,46 Future work describing interactions between PAHs and the avian AHR may explain why the degree of interspecies variability appears to be less for this class of contaminants. Association with in Vivo End Points. A broader goal of this study was to determine whether the EROD hepatocyte culture bioassay produces data that can support risk assessment for wild birds exposed to PAHs in the environment. It was therefore important to compare our ReP values against an ecologically relevant end point. Brunstrom and colleagues conducted a series of egg injection experiments with PAHs that provide some context for the ReP data presented here.7−9 In an initial experiment, a single concentration of 0.3 mg PAH/kg was injected into chicken eggs on day 7 of incubation, and mortality was assessed 72 h later.7 At this concentration, there was no mortality in 17 of 24 PAHs tested, including the 10 that did not produce an EROD response in our study. With one exception (BghiP), the congeners that were active EROD inducers in our study were also toxic to the chicken embryos at 0.3 mg PAH/kg egg. This suggests that PAH congeners that induce EROD activity in cultured cells tend to also be more acutely toxic to avian embryos in vivo. We evaluated the correspondence between in vitro and in vivo end points for each of the 8 EROD-active congeners. The two most potent EROD inducers in cultured heypatocytes (DahA and BkF) were also the two most potent congeners in avian embryos.7 For the remaining six EROD-inducing congeners (BaA, BNT, IdP, BaP, Chr, and BghiP) there was no correspondence between our results and the in vivo data. IdP, BaP, and Chr were moderate EROD-inducers in cultured cells, but at 0.3 mg/kg egg these compounds caused 55−75% mortality in chicken embryos, while both weaker ERODinducers (BaA and BNT) and more potent EROD-inducers (DahA and BkF) caused 100% mortality (Table 1). One explanation for this lack of correspondence is that while many PAHs bind to the AHR, they can also exert toxicity through mechanisms other than AHR activation.47 Some of the inconsistences between EROD EC50 data and LD50 values may also be related to the timing of exposure. PAHs can be readily metabolized, both in vivo and in cultured cells.48,49 The

■

ASSOCIATED CONTENT

* Supporting Information S

EROD concentration−response curves for Pekin duck and greater scaup are presented with modified y-axis scaling in Figures S1 and S2. Concentration−response curve parameters used in the figures and referred to in the text: Table S1, ReP 3792


Article

Environmental Science & Technology values; Table S2, EC10−EC80; Table S3, LOEC, ReS and Max EROD; and Table S4, TCDD10−TCDDmax. This material is available free of charge via the Internet at http://pubs.acs.org/.

■

(13) Head, J. A.; O’Brien, J.; Kennedy, S. W. Exposure to 3,3′,4,4′,5pentachlorobiphenyl during embryonic development has a minimal effect on the cytochrome P4501A response to 2,3,7,8-tetrachlorodibenzo-p-dioxin in cultured chicken embryo hepatocytes. Environ. Toxicol. Chem. 2006, 25, 2981−2989. (14) Kennedy, S.; Lorenzen, A.; James, C.; Collins, B. Ethoxyresorufin-O-deethylase and porphyrin analysis in chicken embryo hepatocyte cultures with a fluorescence multiwell plate reader. Anal. Biochem. 1993, 211, 102−112. (15) Kennedy, S.; Jones, S.; Bastien, L. Efficient analysis of cytochrome P4501A catalytic activity, porphyrins, and total proteins in chicken embryo hepatocyte cultures with a fluorescence plate reader. Anal. Biochem. 1995, 226, 362−370. (16) Villeneuve, D.; Blakenship, A. L.; Giesy, J. P. Derivation and application of relative potency estimates based on in vitro bioassay results. Environ. Toxicol. Chem. 2000, 19, 2835−2843. (17) Zhang, R.; Manning, G. E.; Farmahin, R.; Crump, D.; Zhang, X.; Kennedy, S. W. Relative potencies of aroclor mixtures derived from avian in vitro bioassays: comparisons with calculated toxic equivalents. Environ. Sci. Technol. 2013, 47, 8852−8861. (18) Farmahin, R.; Crump, D.; Jones, S. P.; Mundy, L. J.; Kennedy, S. W. Cytochrome P4501A induction in primary cultures of embryonic European starling hepatocytes exposed to TCDD, PeCDF and TCDF. Ecotoxicology 2013, 22, 731−739. (19) Jones, S. P.; Farmahin, R.; Kennedy, S. W. Ethoxyresorufin-Odeethylase (EROD) induction by TCDD, PeCDF and PCB 126 in bobwhite quail hepatocytes. Ecotoxicology 2014, 23 (5), 802−808. (20) Kennedy, S. W.; Lorenzen, A.; Jones, S. P.; Hahn, M. E.; Stegeman, J. J. Cytochrome P4501A induction in avian hepatocyte cultures: a promising approach for predicting the sensitivity of avian species to toxic effects of halogenated aromatic hydrocarbons. Toxicol. Appl. Pharmacol. 1996, 141, 214−230. (21) Petrulis, J. R.; Bunce, N. J. Competitive inhibition by inducer as a confounding factor in the use of the ethoxyresorufin-O-deethylase (EROD) assay to estimate exposure to dioxin-like compounds. Toxicol. Lett. 1999, 105, 251−260. (22) Villeneuve, D. L.; Khim, J. S.; Kannan, K.; Giesy, J. P. Relative potencies of individual polycyclic aromatic hydrocarbons to induce dioxinlike and estrogenic responses in three cell lines. Environ. Toxicol. 2002, 17, 128−137. (23) Piskorska-Pliszczynska, J.; Keys, B.; Safe, S.; Newman, M. The cytosolic receptor binding affinities and AHH induction potencies of 29 polynuclear aromatic hydrocarbons. Toxicol. Lett. 1986, 34, 67−74. (24) Machala, M.; Mátlová, L.; Svoboda, I.; Nezveda, K. Induction effects of polychlorinated biphenyls, polycyclic aromatic hydrocarbons and other widespread aromatic environmental pollutants on microsomal monooxygenase activities in chick embryo liver. Arch. Toxicol. 1996, 70, 362−367. (25) Peakall, D.; Norstrom, R.; Jeffrey, D. A.; Leighton, F. Induction of hepatic mixed function oxidases in the herring gull (Larus argentatus) by prudhoe bay crude oil and its fractions. Comp. Biochem. Physiol., C 1989, 94C, 461−463. (26) Trust, K.; Esler, D.; Woodin, B.; Stegeman, J. Cytochrome P450 1A induction in sea ducks inhabiting nearshore areas of Prince William Sound, Alaska. Mar. Pollut. Bull. 2000, 40, 397−403. (27) Golet, G.; Seiser, P.; McGuire, A.; Roby, D.; Fischer, J.; Kuletz, K.; Irons, D.; Dean, T.; Jewett, S.; Newman, S. Long-term direct and indirect effects of the “Exxon Valdez” oil spill on pigeon guillemots in Prince William Sound, Alaska. Mar. Ecol.: Prog. Ser. 2002, 241, 287− 304. (28) Gentes, M.-L.; Waldner, C.; Papp, Z.; Smits, J. E. G. Effects of oil sands tailings compounds and harsh weather on mortality rates, growth and detoxification efforts in nestling tree swallows (Tachycineta bicolor). Environ. Pollut. 2006, 142, 24−33. (29) Custer, T. W.; Custer, C. M.; Dickerson, K.; Allen, K.; Melancon, M. J.; Schmidt, L. J. Polycyclic aromatic hydrocarbons, aliphatic hydrocarbons, trace elements, and monooxygenase activity in birds nesting on the North Platte River, Casper, Wyoming, USA. Environ. Toxicol. Chem. 2001, 20, 624−631.

AUTHOR INFORMATION

Corresponding Author

*Phone: 514-398-7841; fax: 514-398-7990; e-mail: jessica. [email protected]. Notes

The authors declare no competing financial interest.

■

ACKNOWLEDGMENTS We thank Stephanie Jones for her expert technical advice and training. We are very grateful to Glen Fox for his enthusiasm and encouragement, and for many helpful discussions regarding this project. Significant improvements were made to this manuscript based on the comments from four anonymous reviewers. We are grateful for their input. This work was funded by the Wildlife Toxicology Research and Ecotoxicology and Wildlife Health Divisions of Environment Canada.

■

REFERENCES

(1) Douben, P. PAHs: An Ecotoxicological Perspective; John Wiley and Sons: New York, 2003. (2) Albers, P. Birds and polycyclic aromatic hydrocarbons. Avian Poult. Biol. Rev. 2006, 17, 125−140. (3) Esler, D.; Trust, K. A.; Ballachey, B. E.; Iverson, S. A.; Lewis, T. L.; Rizzolo, D. J.; Mulcahy, D. M.; Miles, A. K.; Woodin, B. R.; Stegeman, J. J.; et al. Cytochrome P4501A biomarker indication of oil exposure in harlequin ducks up to 20 years after the Exxon Valdez oil spill. Environ. Toxicol. Chem. 2010, 29, 1138−1145. (4) Gentes, M.-L.; McNabb, A.; Waldner, C.; Smits, J. E. G. Increased thyroid hormone levels in tree swallows (Tachycineta bicolor) on reclaimed wetlands of the Athabasca oil sands. Arch. Environ. Contam. Toxicol. 2007, 53, 287−292. (5) Delistraty, D. Toxic equivalency factor approach for risk assessment of polycyclic aromatic hydrocarbons. Toxicol. Environ. Chem. 1997, 64, 81−108. (6) Custer, T. W.; Custer, C. M.; Hines, R. K.; Sparks, D. W.; Melancon, M. J.; Hoffman, D. J.; Bickham, J. W.; Wickliffe, J. K. Mixed-function oxygenases, oxidative stress, and chromosomal damage measured in lesser scaup wintering on the Indiana Harbor Canal. Arch. Environ. Contam. Toxicol. 2000, 38, 522−529. (7) Brunström, B.; Broman, D.; Näf, C. Toxicity and EROD-inducing potency of 24 polycyclic aromatic hydrocarbons (PAHs) in chick embryos. Arch. Toxicol. 1991, 65, 485−489. (8) Brunströ m , B. Toxicity and erod-inducing potency of polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons (PAHs) in avian embryos. Comp. Biochem. Physiol, C: Comp. Pharmacol. 1991, 100, 241−243. (9) Brunström, B.; Broman, D.; Näf, C. Embryotoxicity of polycyclic aromatic hydrocarbons (PAHs) in three domestic avian species, and of PAHs and coplanar polychlorinated biphenyls (PCBs) in the Common Eider. Environ. Pollut. 1990, 67, 133−143. (10) Denison, M. S.; Nagy, S. R. Activation of the aryl hydrocarbon receptor by structurally diverse exogenous and endogenous chemicals. Annu. Rev. Pharmacol. Toxicol. 2003, 43, 309−334. (11) Billiard, S. M.; Bols, N. C.; Hodson, P. V. In vitro and in vivo comparisons of fish-specific CYP1A induction relative potency factors for selected polycyclic aromatic hydrocarbons. Ecotoxicol. Environ. Saf. 2004, 59, 292−299. (12) Custer, C.; Custer, T. Organochlorine and trace element contamination in wintering and migrating diving ducks in the southern Great Lakes, USA, since the zebra mussel invasion. Environ. Toxicol. Chem. 2000, 19, 2821−2829. 3793


Article


aryl hydrocarbon receptor. Proc. Natl. Acad. Sci. U. S. A. 2006, 103, 6252−6257. (47) Incardona, J. P.; Carls, M. G.; Teraoka, H.; Sloan, C. A.; Collier, T. K.; Scholz, N. L. Aryl hydrocarbon receptor−independent toxicity of weathered crude oil during fish development. Environ. Health Perspect. 2005, 113, 1755−1762. (48) Nishimoto, M.; Yanagida, G.; Stein, J.; Baird, W.; Varanasi, U. The metabolism of benzo(a)pyrene by English sole (Parophrys vetulus): comparison between isolated hepatocytes in vitro and liver in vivo. Xenobiotica 1992, 22, 949−961. (49) Näf, C.; Broman, D.; Brunströ m, B. Distribution and metabolism of polycyclic aromatic hydrocarbons (PAHs) injected into eggs of chicken (Gallus domesticus) and common eider duck (Somateria mollissima). Environ. Toxicol. 1992, 1, 1653−1660. (50) Bastien, L.; Kennedy, S. W.; Lorenzen, A. Ethoxyresorufin Odeethylase (EROD) induction by halogenated aromatic hydrocarbons (HAHs) in chicken embryo hepatocyte cultures: Time-dependent effects on the dose-response curves. Organohalogen Compd. 1997, 34, 215−220. (51) Head, J. A.; Kennedy, S. W. Correlation between an in vitro and an in vivo measure of dioxin sensitivity in birds. Ecotoxicology 2010, 19, 377−382. (52) Safe, S. Determination of 2,3,7,8-TCDD toxic equivalent factors (TEFs): Support fo the use of the in vitro AHH induction assay. Chemosphere 1987, 16, 791−802. (53) Barron, M. G.; Heintz, R.; Rice, S. D. Relative potency of PAHs and heterocycles as aryl hydrocarbon receptor agonists in fish. Mar. Environ. Res. 2004, 58, 95−100.

(30) Larsson, M.; Hagberg, J.; Giesy, J. P.; Engwall, M. Timedependent relative potency factors (REPs) for polycyclic aromatic hydrocarbons and their derivatives in the H4IIE-luc bioassay. Environ. Toxicol. Chem. 2014, 33 (4), 943−953. (31) Bols, N. C.; Schirmer, K.; Joyce, E. M.; Dixon, D. G.; Greenberg, B. M.; Whyte, J. J. Ability of polycyclic aromatic hydrocarbons to induce 7-ethoxyresorufin-O-deethylase activity in a trout liver cell line. Ecotoxicol. Environ. Saf. 1999, 44, 118−128. (32) Willett, K.; Gardinali, P.; Sericano, J.; Wade, T. L.; Safe, S. H. Characterization of the H4IIE rat hepatoma cell bioassay for evaluation of environmental samples containing polynuclear aromatic hydrocarbons (PAHs). Arch. Environ. Contam. Toxicol. 1997, 32, 442− 448. (33) Till, M.; Riebniger, D.; Schmitz, H. J.; Schrenk, D. Potency of various polycyclic aromatic hydrocarbons as inducers of CYP1A1 in rat hepatocyte cultures. Chem. Biol. Interact. 1999, 117, 135−150. (34) Fent, K.; Bätscher, R. Cytochrome P4501A induction potencies of polycyclic aromatic hydrocarbons in a fish hepatoma cell line: demonstration of additive interactions. Environ. Toxicol. Chem. 2000, 19, 2047−2058. (35) Machala, M.; Vondrácek, J.; Bláha, L.; Ciganek, M.; Neca, J. V. Aryl hydrocarbon receptor-mediated activity of mutagenic polycyclic aromatic hydrocarbons determined using in vitro reporter gene assay. Mutat. Res. 2001, 497, 49−62. (36) Bosveld, A. T. C.; de Bie, P. A. F.; van den Brink, N. W.; Jongepier, H.; Klomp, A. V. In vitro EROD induction equivalency factors for the 10 PAHs generally monitored in risk assessment studies in The Netherlands. Chemosphere 2002, 49, 75−83. (37) Larsson, M.; Orbe, D.; Engwall, M. Exposure time-dependent effects on the relative potencies and additivity of PAHs in the Ah receptor-based H4IIE-luc bioassay. Environ. Toxicol. Chem. 2012, 31, 1149−1157. (38) Ziccardi, M.; Gardner, I.; Denison, M. S. Application of the luciferase recombinant cell culture bioassay system for the analysis of polycyclic aromatic hydrocarbons. Environ. Toxicol. Chem. 2002, 21, 2027−2033. (39) Head, J. A.; Hahn, M. E.; Kennedy, S. W. Key amino acids in the aryl hydrocarbon receptor predict dioxin sensitivity in avian species. Environ. Sci. Technol. 2008, 42, 7535−7541. (40) Farmahin, R.; Wu, D.; Crump, D.; Hervé, J. C.; Jones, S. P.; Hahn, M. E.; Karchner, S. I.; Giesy, J. P.; Bursian, S. J.; Zwiernik, M. J.; et al. Sequence and in vitro function of chicken, ring-necked pheasant, and Japanese quail AHR1 predict in vivo sensitivity to dioxins. Environ. Sci. Technol. 2012, 46, 2967−2975. (41) Cohen-Barnhouse, A. M.; Zwiernik, M. J.; Link, J. E.; Fitzgerald, S. D.; Kennedy, S. W.; Hervé, J. C.; Giesy, J. P.; Wiseman, S.; Yang, Y.; Jones, P. D.; et al. Sensitivity of Japanese quail (Coturnix japonica), common pheasant (Phasianus colchicus), and white leghorn chicken (Gallus gallus domesticus) embryos to in ovo exposure to TCDD, PeCDF, and TCDF. Toxicol. Sci. 2011, 119, 93−103. (42) Hervé, J. C.; Crump, D.; Jones, S. P.; Mundy, L. J.; Giesy, J. P.; Zwiernik, M. J.; Bursian, S. J.; Jones, P. D.; Wiseman, S. B.; Wan, Y.; et al. Cytochrome P4501A induction by 2,3,7,8-tetrachlorodibenzo-pdioxin and two chlorinated dibenzofurans in primary hepatocyte cultures of three avian species. Toxicol. Sci. 2010, 113, 380−391. (43) Brunström, B. Sensitivity of embryos from duck, goose, herring gull, and various chicken breeds to 3,3′,4,4′-tetrachlorobiphenyl. Poult. Sci. 1988, 67, 52−57. (44) Jin, X.; Kennedy, S. W.; Di Muccio, T.; Moon, T. W. Role of oxidative stress and antioxidant defense in 3,3′,4,4′,5-pentachlorobiphenyl-induced toxicity and species-differential sensitivity in chicken and duck embryos. Toxicol. Appl. Pharmacol. 2001, 172, 241−248. (45) Farmahin, R.; Manning, G. E.; Crump, D.; Wu, D.; Mundy, L. J.; Jones, S. P.; Hahn, M. E.; Karchner, S. I.; Giesy, J. P.; Bursian, S. J.; et al. Amino acid sequence of the ligand-binding domain of the aryl hydrocarbon receptor 1 predicts sensitivity of wild birds to effects of dioxin-like compounds. Toxicol. Sci. 2013, 131, 139−152. (46) Karchner, S. I.; Franks, D. G.; Kennedy, S. W.; Hahn, M. E. The molecular basis for differential dioxin sensitivity in birds: role of the 3794


Potency of Polycyclic Aromatic Hydrocarbons (PAHs) for Induction of

Recommend Documents