Identification and Quantification of Estrogen ... - ACS Publications

Total concentrations of several known xenobiotic estrogen receptor (ER) ... in this investigation because they are known to be present in wastewater e...
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Environ. Sci. Technol. 2001, 35, 3620-3625

Identification and Quantification of Estrogen Receptor Agonists in Wastewater Effluents SHANE A. SNYDER,* DANIEL L. VILLENEUVE, ERIN M. SNYDER, AND JOHN P. GIESY Department of Zoology, National Food Safety and Toxicology Center, and Institute for Environmental Toxicology, Michigan State University, East Lansing, Michigan 48824-1311

Total concentrations of several known xenobiotic estrogen receptor (ER) agonists and natural and synthetic estrogen were measured in water by use of a combination of instrumental and bioanalytical approaches. Samples from 3 municipal wastewater treatment plants (WWTPs) in south central Michigan (upstream and effluent); 4 point source locations on the Trenton Channel of the Detroit River, MI; and 5 locations in Lake Mead, NV were analyzed. Organic compounds were extracted from 5 L water samples using solid-phase extraction disks and separated into three fractions based on polarity. Whole extracts and fractions were tested for ER agonist potency using the MVLN in vitro bioassay. ER agonist potency was characterized by comparing the magnitude of induction elicited by the extract or fraction to the maximum induction caused by 17βestradiol (E2). The greatest concentrations of ER agonists were associated with the most polar fraction (F3). Instrumental analyses and further fractionation were used to identify specific ER agonists associated with bioassay responses. Bioassay data were compared to extract concentrations in order minimize variability associated with the extraction procedure. Concentrations of endogenous estrogen, E2, and the synthetic estrogen ethynylestradiol (EE2) ranged from nondetectable to 14.6 ng/mL extract (nondetectable to 3.66 ng/L water) and represented from 88 to 99.5% of the total estrogen equivalents in the water samples analyzed. Concentrations of alkylphenols (APs) ranged from nondetectable to 148 µg/mL extract (nondetectable to 37 000 ng/L water). In general, alkylphenols contributed less than 0.5% of the total estrogen equivalents in the water samples. Both bioassay-directed fractionation results and comparison of ER agonist concentrations, adjusted for their known relative potencies, support the conclusion that E2 and EE2 were the dominant environmental estrogens in water samples from mid-Michigan and Lake Mead, NV.

Introduction Some compounds released into the environment by human activities can mimic or modulate endogenous hormones and have been termed “endocrine-disrupting” compounds (1, * Corresponding author phone: (702)567-2317; fax: (702)564-7222; e-mail: [email protected]. Current address: Southern Nevada Water Authority, 243 Lakeshore Road, Boulder City, NV 89005. 3620

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2). “Endocrine-disrupting” compounds have been defined as exogenous agents that interfere with the “synthesis, secretion, transport, binding, action, or elimination of natural hormones in the body that are responsible for the maintenance of homeostasis, reproduction, development, and/or behavior” (3). It has been hypothesized that such compounds may elicit a variety of adverse effects in both humans and wildlife, including promotion of hormone-dependent cancers, reproductive tract disorders, and reduction in reproductive fitness (1, 4-10). Much of the concern has focused on compounds that are estrogen receptor (ER) agonists. These compounds have been variously referred to as “estrogenic”, “estrogen-like”, “environmental estrogens”, or “xenoestrogens”. ER agonists and antagonists have the ability to mimic or block the functions of endogenous estrogen. Effects consistent with exposure to ER agonists have been observed in fish exposed to municipal wastewater treatment plant effluents (11, 12). Nonylphenol (NP), nonylphenol polyethoxylates (NPEs), octylphenol (OP), and synthetic and natural steroids were targeted in this investigation because they are known to be present in wastewater effluents and have been implicated as ER agonists that can cause adverse, population-level effects in aquatic organisms (11, 13-16). Methods for identifying and quantifying ER agonists in environmental samples are needed in order to assess the potential for adverse effects through an ER-mediated mechanism of action. This need was underscored by recent legislation mandating that chemicals and formulations be screened for potential to cause estrogen-like biological responses before they are manufactured or used in certain processes (Safe Drinking Water Act Amendments of 1995 Bill Number S.1316; Food Quality Protection Act of 1996 Bill Number P.L. 104-170). Halogenated aromatic hydrocarbons (HAHs) and polycyclic aromatic hydrocarbons (PAHs) are known to cause a wide range of adverse effects, including mortality, wasting syndrome, hepatotoxicity, immunotoxicity, reproductive impairment, and carcinogenicity (16-19). Some of these effects are mediated through the aryl hydrocarbon receptor (AhR) (17); however, some of these compounds can modulate the ER as well. HAHs, such as polychlorinated dibenzo-pdioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and some polychlorinated biphenyls (PCBs), have been reported to act as ER agonists in vitro (20, 21). PAHs have been reported to be both ER agonists and antagonists in vitro (22-23). Although instrumental analyses can be used to identify and quantify known ER agonists and antagonists in wastewater treatment plant (WWTP) effluents, in vitro bioassays provide useful information that can complement instrumental analyses to provide a more comprehensive characterization of a sample’s potential to modulate the ER and result in estrogenic responses. In vitro bioassays provide an integrated measure of the total potency of complex mixtures to induce particular biological responses. Thus, in vitro bioassays can account for both unknown compounds and potential nonadditive interactions among compounds. This study is based on a bioassay-directed fractionation approach to identify compounds able to modulate ER-mediated gene expression. Furthermore, bioassay-based estimates of total ER agonist potency were compared to estimates based on analytical concentrations of known ER agonists and their relative potencies (REPs) in a potency balance analysis (24, 25) to determine whether the compounds quantified could account for the magnitude of ER-mediated bioassay response observed. 10.1021/es001254n CCC: $20.00

 2001 American Chemical Society Published on Web 08/14/2001

FIGURE 1. Luciferase induction in the MVLN cell bioassay (estrogen responsive) elicited by water extracts. Response magnitude presented as percentage of the average maximum response observed for a 1000 pM 17β-estradiol standard (%-E2-max). Horizontal lines represent ( 3 SD from the mean solvent control response (set to 0%-E2-max): a. Michigan WWTPs; b. Trenton Channel; c. Lake Mead (April); and d. Lake Mead (September).

Materials and Methods Sample Collection and Fractionation. A detailed description of the analytical methodology used in this study was published previously (26). Briefly, 5 L water samples were extracted at each field site using solid-phase extraction (SPE) Empore disks. Organic extracts from these SPE disks were separated into three fractions based on polarity using normalphase high-pressure liquid chromatography (NP-HPLC) (26). In some cases, ER agonists of interest were isolated from F2 and F3 by fractionating these again with reverse-phase HPLC (RP-HPLC), with fractions collected approximately every 3 min (Supporting Information, Figure 1). NP-HPLC and RPHPLC separations were accomplished using silica and C18 analytical columns, respectively. Quality assurance and quality control measures included replicate samples, field and laboratory blanks, and spike-recovery experiments, which were described in detail previously (26). Cell Culture and Bioassay. An MCF-7 human breast carcinoma cell line, stably transfected with an ER-controlled luciferase reporter gene construct (MVLN or MCF-7-luc cells), was developed and characterized by Dr. M. D. Pons, Institut National de la Sante et de la Recherche Medicale (27). MVLN cells were cultured in 75-cm2 disposable polyethylene tissue culture flasks (Corning, Corning, NY) containing 20-25 mL of Dulbecco’s Modified Eagle Medium (DMEM) with Hams F-12 nutrient mixture (Sigma D-2906; St. Louis, MO) supplemented with 10% defined fetal bovine serum (Hyclone, Logan, UT), 27.3 I.U. insulin (Sigma I-1882)/L, and 1.0 mM sodium pyruvate (Sigma). In preparation for bioassay, cells were trypsinized from flasks or plates in which cells were 80-100% confluent. The number of cells per mL was determined microscopically by

FIGURE 2. Fine fractionation of LV Wash, Lake Mead (April), F3 extract using RP-HPLC with fluorescence detection followed by luciferase induction in the MVLN cell bioassay (estrogen responsive) by the corresponding fractions. Response magnitude presented as percentage of the average maximum response observed for a 1000 pM 17β-estradiol standard (%-E2-max). Horizontal lines represent ( 3 SD from the mean solvent control response (set to 0%-E2-max). use of a hemacytometer. MVLN cells were diluted in hormone-stripped medium [DMEM with Hams F-12 nutrient mixture, supplemented with 10% dextran-coated charcoal filtered fetal bovine serum (Hyclone), 27.3 I.U. insulin (Sigma I-1882)/L, and 1.0 mM sodium pyruvate (Sigma)] to a concentration of approximately 1.5 × 105 cells/mL. Cells were seeded into the 60 interior wells of 96-well flat bottom microplates (Packard Instruments 6005181; Meriden, CT) at VOL. 35, NO. 18, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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125 µL per well (15 000-20 000 cells per well) using a repeating pipet. To ensure homogeneity, the cell solution was continuously mixed during seeding. The 36 exterior wells of each microplate were filled with 125 µL of medium. Cells were dosed after an overnight incubation to allow for cell attachment. Extracts or fractions were dissolved in stripped medium to yield a final concentration of 1.0% extract. A 3-fold dilution of each extract or fraction was also prepared, yielding a concentration of 0.33% extract. Test wells were dosed with 125 µL 1.0% or 0.33% extract in medium to yield final in-well concentrations of 0.50% and 0.165% extract. Solvent control wells were dosed with 125 µL of medium spiked with 1.0% of the appropriate solvent to yield a final in-well concentration of 0.50% solvent. Blank wells received 125 µL of the appropriate media. Each plate tested included a minimum of three solvent control wells, three blank wells, and three replicates of each fraction tested (at both 0.50% and 0.165% levels). Dosed cells were exposed for 72 h at standard incubation conditions. Each test plate was inspected visually and differences in cell numbers and condition relative to control wells and conditions normally observed during routine culturing were noted for each well. Culture medium was then removed, and each well was rinsed twice with phosphate buffered saline (PBS) supplemented with 1.0 mM Ca2+ and Mg2+ using an eight channel vacuum manifold. Plates were inspected for cell loss during washing. Following inspection, 75 µL PBS supplemented with Ca2+ and Mg2+ was added to each well, followed by 75 µL Luc-lite reagent (Packard Instruments). Each plate was incubated for 10 min at 30 °C and then scanned with an ML 3000 microplate reading luminometer (Dynatech Laboratories, Chantilly, VA). Following the luminometer scan, 125 µL of 1.08 mM fluorescamine (Sigma) in acetonitrile (ACN) was added to each well, and plates were assayed for protein after a 15 min incubation at room temperature (28). Plates were scanned using a Cytofluor 2300 (excitation 400 nm, emission 460 nm), and responses were compared to a standard curve consisting of six concentrations of bovine serum albumin (BSA) (Sigma) ranging from 1.5 to 50 µg per well. All data were collected electronically and imported into a spreadsheet (Excel 7.0, Microsoft Inc., Seattle, WA) for data analysis. Protein content per well was calculated by regression against the BSA standard curve. Protein data were used as an index of cell number to detect outliers that were not apparent by visual inspection. Relative luminescence units (RLU) were not adjusted for protein. Sample responses in RLU were expressed as a percentage of the mean maximum response observed for standard curves developed on the same day (% E2-max) (29). The greatest response of the two extract dilutions was reported. However, for each significant response, the greatest response came from the greater extract concentration (0.5% in the well). Potency balance analyses were conducted by comparing observed bioassay response magnitudes to those predicted based on the concentrations of known ER agonists present in an extract (30). Instrumentally determined concentrations of individual compounds were multiplied by their assayspecific relative potencies. The sum of the products for all target compounds present in an extract provided an estimate of the 17β-estradiol equivalents (EEQ) in the extracts. Linear regression against a 17β-estradiol (E2) standard curve was used to predict the bioassay response magnitude for the sample. Variability in the predicted bioassay response magnitude was estimated based on the 95% confidence band for a first-order polynomial fit to the E2 standard curve (PlotIT, Scientific Programming Enterprises, Haslett, MI). Comparisons were predicated on the assumption that EEQs would behave as if they were 17β-estradiol in the bioassay. Violation of this assumption may have resulted in some error in the 3622

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predictions (29). To make an accurate comparison, it was necessary to address the potential for antagonistic and synergistic interactions. This was done by the fine fractionation of samples, by which compounds known to be antagonistic to the measurement of EEQ were separated from the active compounds.

Results and Discussion ER Agonist Activity. None of the F1 fractions induced a significant response in the MVLN assay (Figure 1). Nonpolar compounds such as PAHs, PCBs, and most organochlorine (OC) pesticides, if present, would have been contained in F1 (Figure 1) (26). Certain OC pesticides and PAHs such as chrysene, benz[a]anthracene, and benzo[a]pyrene have been reported to cause weak ER-mediated responses in vitro (23, 31). Based on the method detection limit (MDL) for E2 in the MVLN assay, concentrations of ER agonists were present in F1 at a concentration less than 0.55 ng EEQ/mL. These results support the conclusion that concentrations of nonpolar ER agonists in the surface waters and effluents examined were small. Weak ER agonists such as NP and OP were present in F2 (26). No F2 extracts elicited a significant response in the MVLN assay (Figure 1), despite the confirmed presence of NP and OP (Table 1). These results suggest that the compounds present in F2 contributed less than 0.55 ng EEQ/ mL. ER agonist potencies of NP and OP, relative to E2, for luciferase induction in MVLN cells have been reported to be 1.25 × 10-5 and 1.9 × 10-5 for NP and OP, respectively (25). When concentrations of NP and OP present in the samples (Table 1) were multiplied by their corresponding relative potencies and summed, it was concluded that these two compounds contributed less than 0.075 ng EEQ/mL for 15 of the 16 samples tested. Thus, the general lack of significant induction of the MVLN cells was consistent with the known concentrations of ER agonists (alkylphenols) present in F2. The BV-effluent sample extract contained approximately 1.90 ng of EEQ/mL, which would correspond to approximately 8.5 fmol EEQ/well in the MVLN bioassay. Based on regression against an E2 standard curve, this dose could have elicited a response as great as 53% E2-max. However, F2 extract of BV-effluent failed to induce a significant response in the MVLN bioassay (Figure 1a). This suggests that F2 of the BVeffluent sample might have contained interfering compounds that suppressed the ER agonist potency of NP and OP. The F2 extract was fractionated further, and fine fractions were analyzed in the MVLN bioassay (Supporting Information, Figure 2). No significant ER-mediated responses were observed in the fine fractions. Thus, the hypothesized interfering compounds, if present, must have properties similar to NP. The same fine fractionation was applied to the F2 extract from Black Lagoon (Supporting Information, Figure 3). Once again, no significant estrogen-like activity was observed. In general, responses of MVLN cells to F2 samples were in agreement with the potency expected based on the known concentrations and relative potencies of these compounds. F3 samples caused the greatest magnitude of ER agonist response in the MVLN bioassay. Six of the 16 F3 samples elicited significant ER-mediated responses in the MVLN bioassay (Figure 1). The greatest magnitudes of response for F3 extracts (≈80% E2-max) were observed for samples collected from the LV Wash and LV Bay in April 1997 (Figure 1c). However, samples from LV Wash and LV Bay collected in September of 1997 did not elicit significant responses in the MVLN assay (Figure 1d). The samples collected in September 1997 were obtained after a large storm event, which diluted the wastewater entering LV Wash and LV Bay (26). The difference in bioassay responses for April and September samples was paralleled by decreases in EEQs in

TABLE 1. Extract Concentrations and 17β-Estradiol Equivalents (EEQs) (ng/mL)c NP

OP

NPE

E2

EE2

NP/OP-EEQa

E2/EE2-EEQb

LV Marina Saddle Island Callville Bay

4/30/97 4/30/97 9/5/97 9/5/97 4/30/97 9/5/97

4560 3000 640 ND ND ND

172 108 ND ND ND ND

Lake Mead 36000 19400 12710 ND ND ND

10.70 8.84 0.752 1.08 ND ND

1.92 2.08 1.01 ND ND ND

0.061 0.040 0.008 NA NA NA

10.9 9.05 0.86 1.08 NA NA

WWTP Chem. B. Lagoon M. Creek

8/30/97 8/30/97 8/30/97 8/30/97

1916 3450 3740 4740

20 60 264 324

Trenton Channel 21600 29200 34700 71260

4.26 3.64 5.18 4.24

ND ND 1.44 ND

0.024 0.044 0.052 0.066

4.26 3.64 5.30 4.25

BV-upstream BV-effluent MA-upstream MA-effluent ER-upstream ER-effluent

10/8/97 10/8/97 10/8/97 10/8/97 10/8/97 10/8/97

ND 148000 ND 2065 ND 680

ND 2350 ND 64 ND ND

WWTPs ND 1160000 ND 19400 ND ND

2.50 14.6 ND 3.62 ND 1.90

ND 3.04 ND 1.43 ND ND

NA 1.90 NA 0.027 NA 0.009

2.50 14.9 NA 3.77 NA 1.90

location LV Wash LV Bay

date

a Nonylphenol and octylphenol-derived 17β-estradiol equivalents. NP/OP-EEQ ) (NP relative potency(REP) × NPconcentration) + (OPREP × OPconcentration). NPREP ) 1.25 × 10-5. OPREP ) 1.9 × 10-5. A REP estimate was not available for NPE; therefore, it was not considered when deriving EEQ estimates. b Estradiol and ethynylestradiol-derived 17β-estradiol equivalents. E2/EE2-EEQ ) (E2 REP × E2concentratration) + (EE2REP × EE2concentration). E2REP ) 1.0. EE2REP ) 0.10. c ND ) not detectable; NA ) not applicable

TABLE 2. Extract 17β-Estradiol Equivalents (EEQs) (ng/mL) and Predicted MVLN Responsese location

date

LV Wash LV Bay

4/30/97 4/30/97 9/5/97 LV Marina 9/5/97 Saddle Island 4/30/97 Callville Bay 9/5/97

FIGURE 3. Fine fractionation of LV Bay, Lake Mead (April), F3 extract using RP-HPLC with fluorescence detection followed by luciferase induction in the MVLN cell bioassay (estrogen responsive) by the corresponding fractions. Response magnitude presented as percentage of the average maximum response observed for a 1000 pM 17β-estradiol standard (%-E2-max). Horizontal lines represent ( 3 SD from the mean solvent control response (set to 0%-E2-max). the samples (Tables 1 and 2). The ER agonist potency of EE2 relative to E2 for luciferase induction in the MVLN assay previously has been reported to be approximately 0.1 (25). Based on the concentrations of E2 and EE2, and their corresponding relative potencies, samples collected from LV Wash and LV Bay in April 1997 were estimated to contain 10.9 and 9.05 ng E2/EE2-derived EEQ/mL, respectively. These concentrations should have yielded doses of approximately 50 and 41 fmol EEQ/well in the MVLN bioassay. Based on regression against an E2 standard curve, such doses would be expected to yield responses of approximately 92% and 88% E2-max, respectively. Based on the range of uncertainty in the predicted responses (Table 2) and the variability of the observed bioassay responses (Figure 1c), the responses observed for the samples collected from LV Wash and LV Bay in April 1997 were not markedly different from predicted responses. Thus, the known E2 and EE2 composition of F3 of the samples collected from LV Wash and LV Bay appeared to account for all the ER agonist potency observed. Additional

WWTP Chem. B. Lagoon M. Creek

8/30/97 8/30/97 8/30/97 8/30/97

BV-upstream BV-effluent MA-upstream MA-effluent ER-upstream ER-effluent

10/8/97 10/8/97 10/8/97 10/8/97 10/8/97 10/8/97

NP/OP- predicted E2/EE2- predicted EEQa responseb EEQc responsed Lake Mead 0.061 0 0.040 0 0.008 0 NA NA NA NA NA NA Trenton Channel 0.024 0 0.044 0 0.052 0 0.066 0 WWTPs NA NA 1.90 53 NA NA 0.027 0 NA NA 0.009 0

10.9 9.05 0.86 1.08 NA NA

92 88 35 41 NA NA

4.26 3.64 5.30 4.25

71 68 76 71

2.50 14.9 NA 3.77 NA 1.90

59 99 NA 68 NA 53

a Nonylphenol and octylphenol-derived 17β-estradiol equivalents. NP/OP-EEQ ) (NPrelative potency (REP) × NPconcentration) + (OPREP × OPconcentration). NPREP ) 1.25 × 10-5. OPREP ) 1.9 × 10-5. A REP estimate was not available for NPE; therefore, it was not considered when deriving EEQ estimates. b MVLN bioassay response magnitudes predicted based on regression of NP/OP-derived EEQ against a 17β-estradiol standard curve. Units are %E2-max. c Estradiol and ethynylestradiol-derived 17β-estradiol equivalents. E2/EE2-EEQ ) (E2REP × E2concentration) + (EE2REP × EE2concentration). E2REP ) 1.0. EE2REP ) 0.10. d MVLN bioassay response magnitudes predicted based on regression of E2/EE2-derived EEQ against a 17βestradiol standard curve. Units are %E2-max. e NA ) not applicable. Note: total EEQ ) NP/OP-EEQ + E2/EE2-EEQ. Predicted bioassay response magnitudes are not additive.

fractionation of the F3 extracts from LV Wash and LV Bay revealed that all of the ER agonist potency was associated with the fine fractions (FFs) 3 and 4, which equate roughly to the retention times of E2 and EE2 (Figures 2 and 3). FFs 3 and 4 from the F3 extract of the LV Wash sample were collected, combined, and fractionated again by RP-HPLC using a slower flow rate and solvent gradient to separate E2 and EE2 (Figure 4). ER agonist potency was observed in fine fractions where E2 and EE2 elute, and the magnitude of VOL. 35, NO. 18, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 4. Further fractionation of LV Wash, Lake Mead, F3 extract using RP-HPLC with fluorescence detection followed by luciferase induction in the MVLN cell bioassay (estrogen responsive) by the corresponding fractions. Response magnitude presented as percentage of the average maximum response observed for a 1000 pM 17β-estradiol standard (%-E2-max). Horizontal lines represent ( 3 SD from the mean solvent control response (set to 0%-E2-max). Dashed line shows chromatography of E2 and EE2 standards with no corresponding bioassay results. induction was consistent with that predicted from EEQs calculated from the measured concentrations of E2 and EE2 and their relative ER-agonist potencies (Table 2). The greater ER agonist potency of the water extracts from LV Wash and LV Bay was most likely due to increased concentrations of E2 and EE2 as a result of WWTPs discharging into the Las Vegas Wash serving a larger population of humans. Significant ER agonist potency was also associated with F3 extracts of water from three locations on the Trenton Channel of the Detroit River (B. Lagoon, Chem., and WWTP) and BV-effluent (Figure 1). From E2 and EE2 concentrations, B. Lagoon, Chem., WWTP, and BV-effluent samples were estimated to contain 5.30, 3.65, 4.25, and 14.9 ng EEQ/mL, respectively (Table 2). Based on regression against an E2 standard curve, these concentrations of EEQ were predicted to yield responses of 76%, 67%, 71%, and 99% E2-max, respectively (Table 2). Observed MVLN cell responses for these F3 samples were, however, less than predicted (Figure 1). Further fractionation and bioanalysis of F3 extracts from BV-effluent and B. Lagoon indicated that all of the observed ER agonist potency was contained in FFs 3 and 4 (Supporting Information, Figures 4 and 5). However, the magnitude of induction of the FFs was markedly different from that of the corresponding total F3 extract. This suggests that interfering compounds and/or unidentified ER (ant)agonists present in F3 might have modulated the potency of the known ER agonists. The potential presence of interfering compounds and/or unknown ER (ant)agonists was also suggested by the lack of significant response for several samples. Based on concentrations of E2 and EE2, six additional F3 samples should have elicited significant responses in the MVLN bioassay. Concentrations of EEQs calculated from concentrations of EE2 and E2 in samples collected from LV Bay (Sept. 1997), LV Marina, M. Creek, BV-upstream, MA-effluent, and EReffluent were estimated to range from 0.85 to 4.25 ng EEQ/ mL. Regression against an E2 standard curve would result in predicted responses of 35-71% E2-max in the MVLN assay for these samples. Thus, the responses were less than predicted for these samples. The reason for this observation is unknown at this time. MVLN responses for whole extracts were similar to those for F3. In those cases where F3 elicited a significant response, the whole extract also elicited a significant response (Figure 3624

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1). In five of the six cases, the whole extract response was slightly less than the response elicited by F3. This suggests that F1 and F2 may have contained some interfering compound(s) or ER antagonists that modulated the potency of the known ER agonists in the samples. However, no significant ER antagonist responses were observed (Figure 1). Because the decreases were slight, however, the results suggest that the bulk of potential interfering (antagonistic) compounds were present in F3. The extract of BV-effluent was the only sample for which the whole-extract response was greater than the corresponding F3 response. It was also the only sample for which the concentrations of NP and OP in F2 were predicted to yield significant ER agonist activity. NP and OP accounted for 11% of the total EEQ calculated for the BV-effluent extract. Thus, although F2 of the BV-effluent sample failed to elicit a significant response, NP and OP may have contributed to the response of the whole extract, such that the total extract response was greater than the F3 response. In general, however, NP and OP accounted for less than 1% of the total concentrations of sample EEQs present in samples. No significant ER activity was observed for blank samples, including field blanks, laboratory blanks, and solvent blanks. Water concentrations of these compounds have been described previously (26).

Summary The potency balance calculations based on instrumental analyses and bioassay-directed fractionation support the conclusion that E2 and EE2 were the dominant environmental estrogens in the samples. Interfering compounds or ER antagonists present in samples (predominantly in F3) may have acted to mask or dampen the potency of the known ER agonists in the MVLN bioassay. All observed responses in the MVLN bioassay were either less than, or approximately equal to, responses predicted based on the measured concentrations and relative potencies of known ER agonists. NP and OP generally contributed less than 1% of the total EEQs. Furthermore, sample fractions containing NP and OP did not elicit significant activity. For most samples, fractions containing E2 and EE2 elicited responses slightly greater than the responses of the corresponding whole extracts. Thus, among the ER agonists detected in the samples E2 and EE2 appear to be responsible for the bulk of the activity. The fact that observed responses were generally lower than predicted suggests the presence of interfering compounds. MVLN responses for whole extracts were only slightly less than those for F3 samples. This suggests that the interfering compounds may have been present in F3. Because there were few instances where MVLN responses were greater than those predicted based on concentrations of EEQs present in the extracts, the known composition can account for the magnitude of response observed. It is unlikely that there were additional ER agonists of significant concentration that were not identified. There are insufficient data to explain differences in bioactivity among the locations investigated. It should be noted that samples from the Trenton Channel of the Detroit River in Michigan received less effluent from municipal WWTPs relative to the volume of the receiving water compared to the other sites. Also, the population served by the WWTPs varied greatly among sites. Further investigations would be necessary to determine the actual loading of bioactive compounds as a function of population density. Without knowing the available fractions and bioaccumulation potential of the various compounds and dose-response relationships for target species, it is not possible to predict the potential effects of the observed concentrations of ER agonists on biota.

Supporting Information Available Figures of RP-HPLC fine fractionation and fine fractionations of BV WWTP and Black Lagoon, Trenton Channel F2 extract and BV WWTP F3 and Black Lagoon, Trenton Channel F3 extract. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Kendall, R. J.; Dickerson, R. L.; Giesy, J. P.; Suk, W. A. Principles and Processes for Evaluating Endocrine Disruptors in Wildlife; SETAC Press: Pensacola, FL, 1998. (2) Ankley, G.; Mihaich, E.; Stahl, R.; Tillitt, D.; Colborn, T.; McMaster, S.; Miller, R.; Bantle, J.; Campbell, P.; Denslow, N.; Dickerson, R.; Folmar, L.; Fry, M.; Giesy, J.; Gray, E.; Guiney, P.; Hutchinson, T.; Kennedy, S.; Kramer, V.; LeBlanc, G.; Mayes, M.; Nimrod, A.; Patino, R.; Peterson, R.; Purdy, R.; Ringer, R.; Thomas, P.; Touart, L.; Van Der Kraak, G.; Zacharewski, T. Environ. Toxicol. Chem. 1998, 17, 68-87. (3) U.S. EPA. Special report on environmental endocrine disruption: an effects assessment and analysis; EPA/630/R-96/012; Office of Research and Development: Washington, DC, February 1997. (4) Safe, S. H. Pharmacol. Ther. 1995, 67, 247-281. (5) Ahlborg, U. G.; Lipworth, L.; Titus-Ernustoff, L.; Hsieh, C.-C.; Hanberg, A.; Baron, J.; Trichopoulos, D.; Adami, H. O. Crit. Rev. Toxicol. 1995, 25, 463-531. (6) Sharpe, R. M.; Skakkebaek, N. E. Lancet 1993, 341, 1392-1395. (7) Carlsen, E.; Giwercman, A.; Keiding, N.; Skakkebaek, N. E. Environ. Health Perspect. 1995, 130, 137-139. (8) Horwitz, K. B.; Jackson, T. A.; Rain, D. L.; Richer, J. K.; Takimoto, G. S.; Tung, L. Mol. Endocrinol. 1996, 10, 1167-1177. (9) Katzenellenbogen, J. A. Environ. Health Perspect. 1995, 103, 99101. (10) Gillesby, B. E.; Zacharewski, T. R. Environ. Toxicol. Chem. 1998, 17, 3-14. (11) White, R.; Jobling, S.; Hoare, S. A.; Sumpter, J. P.; Parker, M. G. Endo. 1994, 135, 175-182. (12) Bevans, H. E.; Goodbred, S. L.; Miesner, J. F.; Watkins, S. A.; Gross, T. S.; Denslow, N. D.; Schoeb, T. Water-Resources Investigations Report 96-4266; 1996. (13) Snyder, S. A.; Snyder, E.; Villeneuve, D.; Kurunthachalam, K.; Villalobos, A.; Blankenship, A.; Giesy, J. In Analysis of Environmental Endocrine Disruptors Keith, L. H., Jones-Lepp, T. L., Needham, L. L., Eds.; American Chemical Society: Washington, DC, 2000; pp 73-95.

(14) Belfroid, A. C.; Van der Horst, A.; Vethaak, A. D.; Schafer, A. J.; Rijs, G. B. J.; Wegener, J.; Cofino, W. P. Sci. Total Environ. 1999, 225, 101-108. (15) Bennie, D. T. Water Qual. Res. J. Canada 1999, 34, 79-122. (16) Naylor, C. G.; Mieure, J. P.; Adams, W. J.; Weeks, J. A.; Castaldi, F. J.; Ogle, L. D.; Romano, R. R. JAOCS 1992, 69, 695-703. (17) Poland, A.; Knutson, J. C. Annu. Rev. Phamacol. Toxicol. 1982, 22, 517-554. (18) Neff, J. M. Polycyclic Aromatic Hydrocarbons in the Aquatic Environment: Sources, Fates, and Biological Effects; Applied Science: London, 1979. (19) Schecter, A. Dioxins and Health; Plenum Press: New York, 1994; p 710. (20) Krishnan, V.; Safe, S. Toxicol. Appl. Pharmacol. 1993, 120, 5561. (21) Anderson, M. J.; Miller, M. R.; Hinton, D. E. Aquat. Toxicol. 1996, 34, 327-350. (22) Tran, D. Q.; Ide, C. F.; McLachlan, J. A.; Arnold, S. F. Biophys. Res. Commun. 1996, 229, 102-108. (23) Clemons, J. H.; Allan, L. M.; Marvin, C. H.; Wu, Z.; McCarry, B. E.; Zacharewski, T. R. Environ. Sci. Technol. 1998, 32, 18531860. (24) Sanderson, J. T.; Giesy, J. P. In Encyclopedia of Environmental Analysis and Remediation; Meyers, R. A., Ed.; John Wiley and Sons: New York, 1998; pp 5272-5297. (25) Villeneuve, D. L.; Blankenship, A. L.; Giesy, J. P. In ToxicantReceptor Interactions; Denison, M. S., Helferich, W. G., Eds.; Taylor & Francis: 1998; pp 69-97. (26) Snyder, S. A.; Keith, T. L.; Verbrugge, D. A.; Snyder, E. M.; Gross, T. S.; Kannan, K.; Giesy, J. P. Environ. Sci. Technol. 1999, 33, 2814-2820. (27) Pons, M.; Gagne, D.; Nicolas, J. C.; Mehtali, M. Biotechniques 1990, 9, 450-459. (28) Kennedy, S. W.; Jones, S. P. Anal. Biochem. 1994, 222, 217-223. (29) Villeneuve, D. L.; Blankenship, A. L.; Giesy, J. P. Environ. Toxicol. Chem. 2000, 19, 2835-2843. (30) Hilscherova, K. M.; Machala, M.; Kannan, K.; Blankenship, A. L.; Giesy, J. P. Environ. Sci. Pollut. Res. 2001, 7, 159-171. (31) Soto, A. M.; Chung, K. L.; Sonnenschein, C. Environ. Health Perspect. 1994, 102, 380-383.

Received for review May 10, 2000. Revised manuscript received May 23, 2001. Accepted July 2, 2001. ES001254N

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