Vertical Profiles of Dioxin-like and Estrogenic Activities Associated with a Sediment Core from Tokyo Bay, Japan

Transcription

1 Environ. Sci. Technol. 2000, 34, Vertical Profiles of Dioxin-like and Estrogenic Activities Associated with a Sediment Core from Tokyo Bay, Japan KURUNTHACHALAM KANNAN,*, DANIEL L. VILLENEUVE, NOBUYOSHI YAMASHITA, TAKASHI IMAGAWA, SHINYA HASHIMOTO, AKIRA MIYAZAKI, AND JOHN P. GIESY National Food Safety and Toxicology Center, Department of Zoology, Institute for Environmental Toxicology, Michigan State University, East Lansing, Michigan 48824, National Institute for Resources and Environment, 16-3 Onogawa, Tsukuba , Japan, and Tokyo University of Fisheries, Konan, Minato-ku, Tokyo 108, Japan In vitro bioassays were used to measure dioxin-like and estrogenic activities associated with florisil fractions of extracts from a sediment core collected from Tokyo Bay, Japan. Florisil fractions 2 (F2) and 3 (F3) elicited significant dioxin-like responses in vitro. Dioxin-like activities of F2 samples were correlated with the vertical profile of PAH concentrations (R 2 ) 0.85). Contribution of PAHs to Ah receptor-mediated activities in sediments was greater than those by PCDDs/DFs, PCBs, and PCNs. The dioxin-like activity of F3 samples suggests the presence of relatively polar, Ah receptor-active compounds in the Tokyo Bay sediment core. Significant estrogenic activities, which may be related to the presence of certain estrogenic PAHs, were observed for F2 samples. Estrogen equivalents (E2- EQs) calculated from the concentrations and relative potencies of known estrogenic compounds in F2 were greater than bioassay-derived E2-EQs. This suggests that complex interactions between estrogenic and antiestrogenic compounds (PAHs, PCDD/DFs, and PCNs) may have modulated the activity. F3 samples were toxic to MVLN cells; therefore, their estrogenic activities could not be estimated. Introduction Environmental matrixes such as sediments contain complex mixtures of residues of organic compounds of both natural and anthropogenic origin. The concentrations and toxic potencies of compounds present in such complex mixtures can range over several orders of magnitude and can be modulated by interactions (synergism, antagonism, etc.) among chemicals. This complicates hazard evaluation for complex mixtures of xenobiotics present in environmental matrixes. Instrumental analytical techniques are available to identify and quantify some compounds in complex * Corresponding author phone: (517) ; fax: (517) ; kuruntha@pilot.msu.edu. Michigan State University. National Institute for Resources and Environment. Tokyo University of Fisheries. mixtures, but there are many compounds for which neither methods nor standards are available. Furthermore, instrumental analyses provide little information on the biological effects of complex mixtures and do not account for possible interactions among individual chemicals. In vitro bioassays are useful tools for characterizing complex mixtures of contaminants, which act through a known mechanism of action. In vitro bioassays provide a biologically relevant, integrated measure of the combined potency of all compounds in a sample (1, 2). When combined with instrumental analysis, sample fractionation techniques, and mass-balance analysis, in vitro bioassays can be used to identify specific compounds or classes of compounds associated with observed biological activity (3-6). In this study, florisil fractions of extracts from a sediment core collected from Tokyo Bay, Japan, were analyzed using in vitro bioassays in order to evaluate vertical profiles of dioxin-like and estrogenic activities. Two in vitro bioassays were used. In vitro luciferase assay with recombinant rat hepatoma cells (H4IIE-luc; 7) was used to screen for compounds capable of modulating aryl hydrocarbon receptor (AhR)-mediated gene expression. In vitro luciferase assay with recombinant MCF-7 human breast carcinoma cells (MVLN; 8) was used to screen extracts for compounds that can modulate gene expression through an estrogen receptor (ER)-mediated mechanism. Where possible, bioassay-derived potencies, relative to a 2,3,7,8-tetrachlorodibenzo-p-dioxin or 17-β-estradiol standard (TCDD-EQ or E2-EQ), were compared to relative potency estimates calculated by multiplying the concentrations of known dioxin-like or estrogenic compounds (reported elsewhere: 9) by assay-specific relative potencies and summing the total (TEQ or EEQ). This type of mass-balance analysis (3) was used to evaluate whether the known composition of the extracts could account for the magnitude of dioxin-like or estrogenic activity observed. TCDD-EQ or E2-EQ estimates significantly greater than TEQ or EEQ estimates would suggest either the presence of unidentified dioxin-like or estrogenic compounds or the synergistic interactions between components of the extract. TCDD-EQ or E2-EQ estimates significantly less than TEQ or EEQ estimates would suggest the presence of antagonists or interfering compounds in the extracts. This type of massbalance analysis has been applied in the identification and characterization of dioxin-like and estrogenic compounds in both surficial sediment and surface waters (6, 10). This study applied the same mass-balance principles to help characterize the historical profile of dioxin-like and estrogenic activities associated with a dated sediment core. Materials and Methods Samples and Fractionation. A sediment core was collected in May 1995 from the northern part of Tokyo Bay (35 35 N and E) using an acrylic tube (120 cm long and 11 cm i.d.). The core was sliced at 2-cm intervals for up to 20 cm and then at 5-cm intervals for up to 93 cm using a clean stainless steel slicer. Each section was freeze-dried and stored in a refrigerator until analysis. A detailed description of the sample collection, extraction, and fractionation procedure has been provided elsewhere (9). Briefly, sediments were Soxhlet extracted using dichloromethane (DCM) and hexane (3:1, 400 ml). Extracts were treated with acid-activated copper granules to remove sulfur. Concentrated extracts were passed through 10 g of activated Florisil packed in a glass column (10 mm i.d.) for fractionation. The first fraction (F1; nonpolar), eluted with 100 ml of hexane, contained PCBs. PAHs and certain organochlorine pesticides were eluted in the second ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 17, /es001044a CCC: $ American Chemical Society Published on Web 07/27/2000

2 FIGURE 1. Luciferase induction in H4IIE-luc (dioxin-responsive) cell bioassay elicited by Tokyo Bay sediment core extract fractions 1 (F1), 2 (F2), and 3 (F3) and procedural blank. Response magnitude presented as percentage of the maximum response observed for a 3130 pm 2,3,7,8-tetrachlorodibenzo-p-dioxin (% TCDD max). Horizontal line equals 3 SD above the mean solvent control response (set to 0% TCDD max). fraction (F2; midpolar) with 100 ml of 20% DCM in hexane. The APs such as nonylphenol (NP) and octylphenol (OP) were eluted in the third fraction (F3; polar) with 100 ml of 50% DCM in methanol. PCNs were eluted in F1 and F2, while PCDDs/DFs were eluted in all the three fractions, with a predominant proportion in F2. Cell Culture and Bioassay. H4IIE-luc cells are rat hepatoma cells, which were stably transfected with a luciferase gene under control of dioxin-responsive elements (DREs) (7). MVLN cells are human breast carcinoma cells stably transfected with a luciferase reporter gene under control of estrogen-responsive elements (EREs) of the Xenopus vitellogenin A2 gene (8, 11). Culturing conditions for both cell lines have been described previously (6). MVLN and H4IIE-luc cells were cultured in 100-mm disposable Petri plates and incubated at 37 C in a humidified 95:5 air:co 2 atmosphere. Cells for bioassay were plated into the 60 interior wells of 96-well culture plates (250 µl/well) at a density of approximately cells/well. Cells were incubated overnight prior to dosing. Test wells were dosed with 2.5 µl of the appropriate florisil fraction. Samples were tested using three replicate wells. Control wells received appropriate solvents. Sample responses, expressed as mean relative luminescence units (RLU), from three replicate wells were converted to a percentage of the mean maximum response observed for standard curves generated on the same day (% E2 max and % TCDD max for 17-β-estradiol and TCDD standards, respectively). Significant responses were defined as those outside the range defined by three times the standard deviation (expressed in % standard max) of the mean solvent control response (0% standard max). Dose-response relationships were examined for sediment core extracts from selected depths to estimate potencies relative to 17-βestradiol (E2) and TCDD. Dose-responses consisted of six concentrations prepared by 3-fold dilution from the final extract. Details regarding the derivation of relative potency estimates from bioassay results have been described elsewhere (1, 5, 6, 12, 13). Results and Discussion Dioxin-like Activity. None of the F1 extracts elicited a significant increase in luciferase expression in H4IIE-luc cells (Figure 1). On the basis of the detection limit for TCDD in the H4IIE-luc bioassay, approximately 5.0 pg of TEQ/g dry wt would be needed to produce a significant response. Thus, these results suggest that the total concentration of TEQs in F1 samples was less than 5.0 pg/g dry wt. Tri- and tetrachloronaphthalenes elute in F1, but they have not been shown to elicit significant activity in H4IIE-luc cells (12). PCBs were the only AhR agonists known to be present in F1. PCB congeners that elicit dioxin-like activity, including non- and mono-ortho-pcbs, were detected in the order of, on average, CB118 ( ng/g) > CB105 ( ng/g) > CB77 ( ) > CB156 < ng/g) > CB126 ( ng/g) > CB169 (< ng/g) (Table 1). Congener-specific PCB concentrations were multiplied by their H4IIE-specific potencies relative to TCDD in order to estimate the concentration of TEQs in F1 (14). The greatest concentration of TEQs contributed by PCBs was 1.2 pg/g, dry wt, which was observed at a depth of cm and corresponded to the maximum total PCB concentration among the sediment core sections (Table 1; Figure 2). Despite being fifth in order of abundance, pentachlorobiphenyl congener 126 (3,3,4,4,5-P5CB) accounted for 80-90% of the total TEQs contributed by PCBs throughout the sediment core (Figure 3). The vertical profile of TEQs contributed by PCBs (PCB- TEQs) was similar to their concentration profile, which increased from the 1940s, peaked in the 1980s, and then gradually declined (Figure 2). PCB-TEQ concentrations in surface sediments corresponding to the 1990s were 3-fold less than the maximum value observed for sediment deposited in the 1980s. Interestingly, instrumental analysis also suggested the presence of non-ortho-coplanar congener 77 (3,3,4,4 -T4CB) at concentrations of 4 pg/g, dry wt, at cm depth. Concentrations of other non-ortho-congeners CB 126 (3,3,4,4,5-P5CB) and CB 169 (3,3,4,4,5,5 -H6CB) were 0.1 and <0.1 pg/g, respectively, at cm depth. These results suggest the occurrence of certain coplanar PCB congeners in sediment deposited in the early 1900s, prior to widespread industrial production and use. Overall, the TEQ concentration contributed by PCBs was less than that which would be expected to elicit luciferase induction in H4IIE cells. Thus, the lack of significant H4IIE-luc response to F1 supports the results of instrumental analysis (9). F2 samples for sediment from depths of 0-70 cm of the sediment core exhibited significant dioxin-like activity (Figure 1). Magnitudes of induction as great as 90% TCDD max were observed. F2 samples for sediment from depths greater than 70 cm were not active (Figure 2). Penta- through octachloronaphthalenes, PCDDs/DFs, and PAHs were the target analytes present in F2 (9). Several PCN congeners that have been shown to elicit luciferase activity in H4IIE-luc cells were found in sediment core extracts (Figure 3; 12). TEQs contributed by PCNs (PCN-TEQs) were generally less than those contributed by PCBs. However, at depths of cm, PCN- TEQs were similar to or greater than those of PCB-TEQs (Table 1). The greatest PCN-TEQ concentration, 0.44 pg/g, dry wt, was observed at a depth of cm, and the vertical profile of PCN-TEQs resembled the concentration profile of PCNs (Figure 2). TEQs contributed by PCNs in surface sediments were 4-fold less than the greatest PCN-TEQ concentration, observed at cm depth (Table 1). PCN congeners 66/ 67 (1,2,3,4,6,7-/1,2,3,5,6,7-H6CN) accounted for 51-86% of the total PCN-TEQs (Figure 3). PCN-TEQ concentrations were less than the 5.0 pg/g, dry wt, concentration required to yield a significant response in the H4IIE-luc assay. This suggests that PCNs alone were not responsible for the AhR-mediated activity associated with F2 of sediment core extracts. TEQ concentrations contributed by various 2,3,7,8- substituted PCDD/DF congeners (PCDD/DF-TEQs) ranged from 4.2 to 336 pg/g, dry wt (Table 1). These concentrations were 2-3 orders of magnitude greater than those contributed VOL. 34, NO. 17, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

3 TABLE 1. Concentration (pg/g, dry wt) Profiles of TEQs Contributed by PCDDs, PCDFs, PCNs, PCBs, and PAHs at Selected Sections of a Sediment Core from Tokyo Bay, Japan a PCDDs PCDFs PCDD/DFs PCNs PCBs 0.38 na b PAHs a H4IIE relative potencies were used for PCDDs, PCDFs, PCBs, PAHs, and PCNs (12, 14, 16). b na, not analyzed. FIGURE 2. Vertical profile of 2,3,7,8-tetrachlorodibenzo-p-dioxin equivalents (TEQs) contributed by PCDD/DFs, PCBs, PCNs, and PAHs in a sediment core collected from Tokyo Bay, Japan. All the TEFs were based on H4IIE rat hepatoma cell bioassays. PAH-TEFs were from ref 16, whereas those of PCBs and PCDD/DF TEFs were from ref 14. PCN TEFs were from ref 12. by PCBs and PCNs (Table 1) and were great enough to elicit significant responses in the H4IIE-luc bioassay. PCDD/DF- TEQ concentrations were correlated with bioassay response magnitudes (% TCDD max), but the correlation was weak (see Figure 5 in Supporting Information). Despite the fact that PCDF concentrations were less than those of PCDDs, PCDFs accounted for 70-87% of the total PCDD/DF-TEQs throughout the core. The PCDF congener 1,2,3,4,6,7,8-H7CDF accounted for 50-70% of the total PCDD/DF-TEQs at depths of 0-50 cm. In deeper sections (50-90 cm), PCDF congeners 1,2,3,6,7,8-H6CDF, 1,2,3,7,8-P5CDF, and 2,3,7,8-TCDF accounted for greater than 10% of the total TEQs (Table 1; Figure 3). Concentrations of PCDD/DF-TEQs in surface sediments were 3-fold less than those observed at a depth of cm. The great contribution of 1,2,3,4,6,7,8-H7CDF to PCDD/DF-TEQs could be due its great relative potency in H4IIE bioassays as compared to the reported values of consensus TEF (14, 15). Overall, the results suggest that PCDDs/DFs may have contributed significantly to the dioxinlike activity observed for F2 samples. Several PAHs, including benzo[k]fluoranthene (BkF), benzo[a]pyrene (BaP), benzo[b]fluoranthene (BbF), chrysene (Chr), benz[a]anthracene (BaA), indeno[1,2,3,-cd]pyrene (IP), dibenz[a,h]anthracene (DaA), and anthracene (Ant) have been shown to elicit dioxin-like responses or induce cytochrome P4501A1 activity in vitro (16-18). H4IIE-specific potencies, relative to TCDD, have been reported for a number of the PAHs quantified in this study (16). These were used to estimate TEQs contributed by PAHs (PAH-TEQs). Relative potencies of Bkf, BaP, BbF, Chr, BaA, IP, and DaA were , , , , , , and , respectively (16). H4IIE-luc responses elicited by F2 extracts were significantly correlated with sediment PAH concentrations, PAH-TEQs, and PAH profile (see Figure 6 in Supporting Information). The correlation between PAH-TEQs and H4IIE-luc response magnitudes (% TCDD max) was greater than that for PCDD/DF-TEQs. Calculated PAH-TEQs were 5-50 times greater than PCDD/DFs-TEQs (Table 1). Concentrations of PAH-TEQs ranged from 11.5 pg/g at cm depth to 1800 pg/g at cm depth (Figure 2). The concentration of PAH-TEQs in surface sediments was 333 pg/g, dry wt, which was 5.4-fold less than the highest observed at cm depth. Averaged across all depths, BbF accounted for approximately 43% of the total PAH-TEQs, followed by BkF and IP at 33% and 20%, respectively (Figure 3). These results suggest that, although PCDDs/DFs may have significantly contributed to the responses elicited by F2 samples, AhR-active PAHs in F2 samples probably account for the majority of the response observed. Dose-response curves were obtained for selected F2 samples by analyzing them at six different dilutions (3-fold serial dilutions). Bioassay-derived TCDD-EQs were then estimated from the dose-response relationships (13). TCDD- EQ concentrations, at selected depths, ranged from 27 to 992 pg/g, dry wt. TCDD-EQ estimates were 2-12-fold less than those predicted from instrumental TEQs (Table 2). Nevertheless, instrumental TEQs and bioassay-derived TCDD- EQs were correlated. These results suggest that interactions among less active or inactive compounds present in the mixture may have modulated the activity. These results are ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 17, 2000

4 FIGURE 3. Contribution (%) of dioxin-like congeners to 2,3,7,8- tetrachlorodibenzo-p-dioxin equivalents of PCDDs, PCDFs, PCBs, PCNs, and PAHs in a sediment core at cm depth from Tokyo Bay, Japan. The depth of the cm core was selected because this section had the greatest concentration of most of the compounds analyzed. TABLE 2. Predicted and Estimated Dioxin-like Activities (pg/g, dry wt) in Fraction 2 (F2) of Tokyo Bay Sediment Core Extracts at Selected Depths depth (cm) F2 bioassay TCDD-EQs a F2 (instrumental TEQs) b a Bioassay-derived TCDD-EQs were calculated as EC - X TCDD/EC - X sample where X is the maximum response magnitude observed for the sample. Uncertainty in the TCDD-EQ estimates, due to deviation from parallelism to the standard curve, was less than 3-fold (13). b Instrumental TEQ values for F2 represent those from PAHs. TEQs from PCDDs/ DFs were not included. similar to that obtained for riverine sediments from the Czech Republic (19) in which concentrations of TEQs were greater concentrations of TCDD-EQs. Certain alkylated PAHs have been shown to interact with AhR agonists and elicit antagonistic activity in H4IIE bioassays (20). The results of this study are in accordance with the earlier studies reporting great contribution of PAHs to dioxin-like activity in sediments (5, 19, 21). Several F3 extracts also elicited significant luciferase activity in H4IIE-luc cells (Figure 1), although the magnitude of induction was generally less than that elicited by F2 samples. A small portion of PCDDs/DFs (<10%) eluted in F3 (<10% of the total concentration). Assuming 10% of the total PCDD/DF-TEQs eluted in F3, PCDD/DF-TEQ concentrations in F3 samples associated with the upper 18 cm of the core would have ranged from 12 to 33 pg/g dry wt (Table 1). On the basis of regression against the TCDD standard curve, such concentrations of TEQs would be expected to yield bioassay responses of approximately 25-45% TCDD max. Observed responses were similar to or less than this predicted magnitude of response (Figure 1). This suggests that PCDD/ DFs eluted in F3 may be able to account for at least a portion FIGURE 4. Luciferase induction in MVLN (estrogen responsive) cell bioassay elicited by Tokyo Bay sediment core extract fractions 1 (F1), 2 (F2), and 3 (F3) and procedural blank. Response magnitude presented as percentage of the maximum response observed for a 1000 pm 17-β-estradiol (% E2 max). Horizontal line equals 3 SD above the mean solvent control response (set to 0% E2 max). of the activity induced by F3 samples associated with the upper 18 cm of the core. Assuming 10% of the total PCDD/ DF-TEQs eluted in F3, PCDD/DF-TEQ concentrations in F3 samples associated with core sections deeper than 18 cm would have ranged from 0.4 to 6.8 pg/g dry wt (Table 1). On the basis of regression against the TCDD standard curve, 6.8 pg/g TEQ would be expected to yield a response of approximately 17% TCDD max, while most F3 samples associated with sections deeper than 18 cm would not be expected to yield a significant response. Responses for F3 samples associated with depths greater than 18 cm were markedly higher than predicted (Figure 1). This suggests that PCDD/DFs probably do not account for all the activity associated with F3, particularly for greater depths in the sediment core. Thus, although PCDD/DFs may contribute to the activity observed, the results suggest the presence of additional, unidentified, relatively polar AhR agonists in Tokyo Bay sediments. This agrees with earlier studies, which reported the presence of acid labile, polar AhR agonists in surficial sediment from Korea (5, 6). A recent study has suggested that the AhR may be capable of binding a wider range of structures than previously suspected (22). Metabolic products of marine biota, such as brevitoxin-6 (produced by the dinoflagellate Ptychodiscus brevis), are one potential source of polar AhR-active compounds in marine environment (22). Similarly, natural products, such as harmane and topolone, derived from wood pulp, and tryptophan derivatives have also been shown to induce cytochrome P4501A1 (22). Thus, the cause of the dioxin-like activity associated with F3 samples remains a topic for further investigation. Estrogenic Activity. Florisil fractions F1, F2, and F3 from each of the 25 sediment extracts were screened for their ability to promote ERE-mediated gene expression in MVLN cells (Figure 4). Only one of the 25 F1 samples, the sample associated with the cm section, elicited a significant estrogenic response in the MVLN bioassay (Figure 4). The magnitude of response was less than 10% E2 max. On the basis of regression against the E2 standard curve, this response would be associated with approximately 2.0 pg E2- EQ/g, dry wt, while most F1 samples would have contributed less than 1.6 pg E2-EQ/g, dry wt, to the total E2-EQ concentration of the associated core section. The lack of VOL. 34, NO. 17, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

5 TABLE 3. Ranges of Concentrations (ng/g, dry wt), TEQs (pg/g, dry wt), and E2-EQs (pg/g, dry wt) of Target Analytes in a Sediment Core from Tokyo Bay a concn TEQs E2-EQs PCDDs PCDFs PCNs PCBs PAHs nonylphenol < a Relative potencies were used for PCDDs, PCDFs, PCBs, PAHs, and PCNs (12, 14, 16, 17). estrogenic response for most F1 samples was expected and consistent with the polarity of known estrogen agonists. F2 samples were significantly more estrogenic than F1 samples. Estrogenic activity was observed in sediment as deep as cm (Figure 4). Response magnitudes as great as 22% E2 max were observed. On the basis of regression against the standard E2 curve, approximately 25 pg of E2- EQ/g, dry wt, would be required to yield a response of 20% E2 max. If present, organochlorine pesticides such as toxaphene, chlordecone, endosulfan, and p,p -DDT would elute in F2. Such pesticides have been shown to elicit weak estrogenic responses in vitro (23). However, such responses have been shown to occur at concentrations greater than 1 µg/g. Concentrations of organochlorine pesticides in Tokyo Bay sediments were not expected to be great enough to contribute to the estrogenic activity observed for F2 samples. A recent study reported DDT concentrations of <2 ng/g, dry wt, in Tokyo Bay sediment (24). Similarly, concentrations of organochlorine pesticides in sediment from Masan Bay, Korea, were insufficient to account for estrogenic activity associated with F2 samples (5). Thus, organochlorine pesticides do not appear to be a likely cause of the estrogenic activity observed. MVLN responses in this study were significantly correlated with the concentration of PAHs detected in the sediment core (see Figure 7 in Supporting Information). Some PAHs have been shown to elicit estrogenic responses in vitro (17). Estrogenic potencies of chrysene, benz[a]anthracene, and benzo[a]pyrene, relative to E2, have been reported to be , , and (17). On the basis of these relative potency values, EEQs contributed by PAHs were in the range of pg/g, dry wt, at depths of cm and pg/g, dry wt, at depths of 0-25 cm. Predicted response magnitudes, based on PAH-EEQ estimates, were consistently greater than the MVLN response magnitudes observed. Furthermore, PAH-EEQs (up to 160 pg/g) were greater than bioassay-derived E2-EQ estimates for F2 samples (up to 25 pg/g). These results suggest that interactions with antiestrogens or other interfering compounds may have modulated the activity of estrogenic PAHs. Dioxin-like compounds such as PCDD/DFs, which have been shown to be antiestrogenic both in vitro and in vivo, were known to be present in F2 samples (25, 26). Thus, interactions between compounds in the F2 samples could account for the discrepancies. Although the cell bioassay used by Clemons et al. (17) was analogous to the MVLN bioassay, there may be cell-line specific differences in the relative estrogenic potencies of PAHs. Overall, the results suggest that estrogenic PAHs account for at least a portion of the estrogenic activity of F2. This agrees with earlier reports of PAHs contributing to estrogenic activity in sediments (5, 20). F3 samples were toxic to MVLN cells, therefore their estrogenic potency could not be determined. MVLN cells are relatively sensitive to the presence of toxic components in extracts (6,20). Nonlyphenol (NP) and octylphenol (OP) were the target analytes present in F3. NP and OP are weak estrogenic compounds and their MVLN-specific potencies, relative to E2, have been reported to be and , respectively (27). On the basis of the measured concentrations of NP, E2-EQs contributed by NP were estimated to be pg/g, dry wt, at the top 10 cm. Such concentrations of E2-EQs would be expected to elicit a significant response in the MVLN bioassay. However, any estrogenic activity attributable to NP and OP were obscured by the toxic effects of the F3 samples. Overall, these results suggest the utility of in vitro bioassays in assessing dioxin-like and estrogenic potential of sediments. Mass balance calculations suggest that PAHs account for a considerable portion of both the dioxin-like and estrogenic activity of the Tokyo Bay sediment core (Table 3). Bioassayderived TCDD-EQs were significantly correlated with instrumental TEQs, although instrumental values were fold greater. These results suggested the existence of antagonistic interactions among various compounds present in F2. Similarly, potential antagonistic interactions between estrogenic PAHs and antiestrogenic compounds such as TCDD, PCNs, and hydroxylated PCBs in F2 were suggested by the discrepancy between instrumental PAH-EEQs and bioassay responses. Next to PAHs, PCDDs/DFs were the greatest contributors to TEQs in the Tokyo Bay sediment core. While PAHs can be toxic to benthic organisms, they would not be expected to biomagnify like PCDD/DFs. On this basis, PCDD/DFs are compounds of concern in the marine food chain in Tokyo Bay. The mass-balance profile of dioxin-like activity in F3 suggests the presence of unidentified, relatively polar compound that can act through AhR-mediated mechanism. This adds to a growing body of evidence for the presence of unidentified dioxin-like compounds in sediments. Although the mass-balance analyses used in this study have limitation, as discussed earlier (5), risk assessment based solely on instrumental analyses may not accurately reflect actual hazards. Although in vitro bioassays cannot be directly extrapolated to determine the risk for adverse effects, they point out additional sources of uncertainty, which should be considered. Acknowledgments This work was supported by grants from the Chlorine Chemistry Council of the Chemical Manufacturers Association (United States) and National Institute for Resources and Environment (NIRE, Japan). Supporting Information Available Figures showing relationships between PCDD/DF-TEQs measured from instrumental analysis and H4IIE-luc bioassay responses (Figure 5), PAH-TEQs measured from instrumental analysis and H4IIE-luc bioassay responses (Figure 6), and PAH concentrations and MVLN bioassay responses (Figure 7) (4 pages). This material is available free of charge via the Internet at Literature Cited (1) Villeneuve, D. L.; Khim, J. S.; Kannan, K.; Giesy, J. P. Aquat. Toxicol. In press. (2) Hilcherova, K.; Machala, M.; Kannan, K.; Blankenship, A. L.; Giesy, J. P. Environ. Sci. Pollut. Res. In press. (3) Sanderson, J. T.; Giesy, J. P. Wildlife toxicology, functional response assays. In Encyclopedia of Environmental Analysis and Remediation; Meyers, R. A., Ed.; John Wiley & Sons: New York, 1998; pp (4) Khim, J. S.; Kannan, K.; Villeneuve, D. L.; Koh, C. H.; Giesy, J. P. Environ. Sci. Technol. 1999, 33, (5) Khim, J. S.; Villeneuve, D. L.; Kannan, K.; Koh, C. H.; Giesy, J. P. Environ. Sci. Technol. 1999, 33, (6) Khim, J. S.; Villeneuve, D. L.; Kannan, K.; Koh, C. H.; Giesy, J. P. Environ. Toxicol. Chem. 1999, 18, ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 17, 2000

6 (7) Sanderson, J. T.; Aarts, J. M. M. J. G.; Brouwer, A.; Froese, K. L.; Denison, M. S.; Giesy, J. P. Toxicol. Appl. Pharmacol. 1996, 137, (8) Pons, M.; Gagne, D.; Nicolas, J. C.; Mehtali, M. BioTechniques 1990, 9, (9) Yamashita, N.; Kannan, K.; Imagawa, T.; Villeneuve, D. L.; Hashimoto, S.; Miyazaki, A.; Giesy, J. P. Environ. Sci. Technol. 2000, 34, (10) Snyder, S.; Villeneuve, D. L.; Snyder, E. M.; Giesy, J. P. Environ. Sci. Technol. Submitted for publication. (11) Demirpence, E.; Duchesne, M. J.; Badia, E.; Gagne, D.; Pons, M. J. Steroid Biochem. Mol. Biol. 1993, 46, (12) Villeneuve, D. L.; Kannan, K.; Khim, J. S.; Falandysz, J.; Blankenship, A. L.; Giesy, J. P. Arch. Environ. Contam. Toxicol. In press. (13) Villeneuve, D. L.; Blankenship, A. L.; Giesy, J. P. Environ. Toxicol. Chem. In press. (14) Giesy, J. P.; Jude, D. J.; Tillitt, D. E.; Gale, R. W.; Meadows, J. C.; Zajieck, J. L.; Peterman, P. H.; Verbrugge, D. A.; Sanderson, J. T.; Schwartz, T. R.; Tuchman, M. L. Environ. Toxicol. Chem. 1997, 16, (15) Van den Berg, M.; Birnbaum, L.; Bosveld, A. T. C.; Brunström, B.; Cook, P.; Feeley, M.; Giesy, J. P.; Hanberg, A.; Hasegawa, R.; Kennedy, S. W.; Kubiak, T.; Larsen, J. C.; van Leeuwen, F. X. R.; Liem, A. K. D.; Nolt, C.; Peterson, R. E.; Poellinger, L.; Safe, S.; Schrenk, D.; Tillitt, D.; Tysklind, M.; Younes, M.; Wærn, F.; Zacharewski, T. Environ. Health Perspect. 1998, 106, (16) Willet, K. L.; Gardinali, P. R.; Sericano, J. L.; Wade, T. L.; Safe, S. H. Arch. Environ. Contam. Toxicol. 1997, 32, (17) Clemons, J. H.; Allan, L. M.; Marvin, C. H.; Wu, Z.; Mccarry, B. E.; Bryant, D. W.; Zacharewski, T. R. Environ. Sci. Technol. 1998, 32, (18) Villeneuve, D. L.; DeVita, W. M.; Crunkilton, R. L. Identification of cytochrome P4501A inducers in complex mixtures of polycyclic aromatic hydrocarbons. In Environmental Toxicology and Risk Assessment, 7th ed.; Little, E. E., DeLonay, A. J., Greenberg, B. M., Eds.; ASTM STP 1333; American Society for Testing and Materials: Philadelphia, PA, 1998; pp (19) Hilcherova, K.; Kannan, K.; Kang, Y. S.; Holoubek, I.; Machala, M.; Masunaga, S.; Nakanishi, J.; Giesy, J. P. Environ. Toxicol. Chem. Submitted for publication. (20) Hilcherova, K.; Kannan, K.; Holoubek, I.; Giesy, J. P. Environ. Toxicol. Chem. Submitted for publication. (21) Anderson, J. W.; Zeng, E. Y.; Jones, J. M. Environ. Toxicol. Chem. 1999, 18, (22) Washburn, B. S.; Rein, K. S.; Baden, D. G.; Walsh, P. J.; Hinton, D. E.; Tullis, K.; Denison, M. S. Arch. Biochem. Biophys. 1997, 343, (23) Soto, A. M.; Chung, K. L.; Sonnenschein, C. Environ. Health Perspect. 1994, 102, (24) Nakada, N.; Isobe, T.; Nishiyama, H.; Okuda, K.; Tsutsumi, S.; Yamada, J.; Kumata, H.; Takada, H. Bunseki Kagaku 1999, 48, (25) Krishnan, V.; Safe, S. Toxicol. Appl. Pharmacol. 1993, 120, (26) Anderson, M. J.; Miller, M. R.; Hinton, D. E. Aquat. Toxicol. 1996, 34, (27) Villeneuve, D. L.; Blankenship, A. L.; Giesy, J. P. Interactions between environmental xenobiotics and estrogen receptormediated responses. In Toxicant-receptor interactions; Denison, M. S., Helferich, W. G., Eds.; Taylor and Francis: Philadelphia, PA, 1998; pp Received for review February 25, Revised manuscript received June 16, Accepted June 21, ES001044A VOL. 34, NO. 17, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY