Reconstitution Studies of Pesticides and Surfactants Exploring the

increase the in vivo estrogenic activity of two herbicides in rainbow trout.(5) With .... Codes for Reconstituted Assays Using AP/APEOS Mixtures a...
0 downloads 0 Views 1MB Size
Download PDF
Article pubs.acs.org/est

Reconstitution Studies of Pesticides and Surfactants Exploring the Cause of Estrogenic Activity Observed in Surface Waters of the San Francisco Bay Delta Daniel Schlenk,*,† Ramon Lavado,† Jorge Eduardo Loyo-Rosales,‡ Wesley Jones,† Lindley Maryoung,† Navneet Riar,† Inge Werner,§ and David Sedlak‡ †

Department of Environmental Sciences, University of California, Riverside, California 92521, United States Department of Civil and Environmental Engineering, University of California, Berkeley, California 94720, United States § Swiss Centre for Applied Ecotoxicology, Eawag/EPFL, 8600 Dubendorf, Switzerland ‡

S Supporting Information *

ABSTRACT: To evaluate the potential role of endocrine disruption in the decline of pelagic fishes in the San Francisco Bay Delta of California, various surface water samples were collected, extracted, and found to elicit estrogenic activity in laboratory fish. Chemical analysis of the estrogenic samples indicated 2 pesticides (bifenthrin, diuron), 2 alkyphenols (AP), and mixtures of 2 types of alkyphenol polyethoxylates (APEOs). Evaluation of estrogenic activity was further characterized by in vitro bioassays using rainbow trout hepatocytes (Oncorhynchus mykiss) and in vivo studies with Japanese medaka (Oryzias latipes). In the in vitro bioassays, hepatocytes exposed to the pesticides alone or in combination with the AP/APEO mixtures at concentrations observed in surface waters failed to show estrogenic activity (induction of vitelloginin mRNA). In the in vivo bioassays, medaka exposed to individual pesticides or to AP/APEO alone did not have elevated VTG at ambient concentrations. However, when the pesticides were combined with AP/APEOs in the 7-day exposure a significant increase in VTG was observed. Exposure to a 5-fold higher concentration of the AP/APEO mixture alone also significantly induced VTG. In contrast to earlier studies with permethrin, biotransformation of bifenthrin to estrogenic metabolites was not observed in medaka liver microsomes and cytochrome P450 was not induced with AP/APEO treatment. These results showed that mixtures of pesticides with significantly different modes of action and AP/APEOs at environmentally relevant concentrations may be associated with estrogenic activity measured in water extracts and feral fish that have been shown to be in population decline in the San Francisco Bay Delta.

1. INTRODUCTION Several species of pelagic fish have been experiencing a precipitous decline in populations since 2001 in the eastern Delta of San Francisco (SF) Bay.1 Recent studies in one of the species (striped bass, Morone saxatilis) indicated the potential for endocrine disrupting compounds contributing to observed reproductive dysfunction.2 To determine the cause, estrogenic activity-guided chemical fractionation was conducted on water extracts collected from a number of waterways in the Central Valley and SF Delta where consistent activity was noted in water extracts from several locations.3 Chemical evaluations of the extracts for 95 analytes failed to indicate individual pharmaceuticals, personal care products, or steroid hormones as causative agents.3 Several alkylphenols, alkylphenol ethoxylates, triazine herbicide degradates, and diuron were detected in estrogenic samples but at concentrations that failed to elicit estrogenic activity in reconstitution studies when they were administered as individual compounds3 or were below published thresholds for estrogenic activity in fish.4 © 2012 American Chemical Society

Alkylphenol ethoxylate containing surfactants are often used with pesticides to enhance the bioavailability of pesticides and have been shown to significantly increase the in vivo estrogenic activity of two herbicides in rainbow trout.5 With multiple mechanisms of action, in vivo estrogenic activity of akylphenols has been shown to be greater than what would be predicted from in vitro bioassays, which only use estrogen receptor activation.10 Whereas previous studies have evaluated the impacts of individual estrogenic compounds in aquatic organisms, relatively few have examined the effects of mixtures in whole animals. Consequently, the purpose of this study was to further investigate the sources of estrogenic activity in water at the SF Bay Delta and test the hypothesis that mixtures of pesticides, alkylphenols, and alkylphenol ethoxylates may be Received: Revised: Accepted: Published: 9106

April 27, 2012 July 16, 2012 July 20, 2012 August 10, 2012 dx.doi.org/10.1021/es3016759 | Environ. Sci. Technol. 2012, 46, 9106−9111

Environmental Science & Technology

Article

responsible in part for estrogenic activity observed in fish bioassays used in earlier investigations. Additional studies were also included that examined the potential bioactivation of pesticides to estrogenic metabolites by the mixture.

observed in any of the animals indicating juvenile animals without endogenous expression of vitellogenin.6 The hepatocytes were isolated by enzymatic digestion with trypsin followed by mechanical disaggregation and gradient centrifugation with Percoll as described previously.3 Following hepatocyte isolation, the cells were seeded in 48-well plates with a density of 1 × 106 cells/well and allowed to settle for 2 h prior to treatment with reconstituted ethanol extracts. Monolayers of cells were confirmed using light microscopy and the cytotoxicity assay described below. Each treatment had three replicates as well as negative and positive (17β-estradiol) controls, and solvent controls (0.01% ethanol). Cells were incubated for 24 h at 18 °C and then resuspended in PBS buffer, centrifuged at 5200g for 5 min and the pellets washed twice with PBS. Cells used for PCR were immediately processed for total mRNA extraction. To ascertain that exposure concentrations used in this study were not cytotoxic, the MTT [(3-(4,5-dimethylthiazol-2-yl)2,5-diphenyltetrazolium] test was conducted for ethanol (carrier solvent), reconstituted extracts and 17β-estradiol (positive control), prior to the vitellogenin-mRNA (VTGmRNA) expression assay. Cell viability was determined by the MTT reduction assay adapted from Mosmann.22 2.4. In Vitro Measurement of Vitellogenin (VTG) mRNA Expression in Rainbow Trout Hepatocytes. Vitellogenin (VTG) mRNA was measured by quantitative PCR (qPCR) used previously in our laboratory.3 Total mRNA was extracted from cells using SV Total RNA Isolation Kit from Promega (Madison, WI) following the manufacturer’s instructions. VTG mRNA was quantified by qPCR by using iScript One-step RT-PCR kit with SYBR Green from Bio-Rad (Hercules, CA, USA) and as a sense primer tVit-364 5′CCCACTGCTGTCTCTGAAACAG-3′ and as antisense primer tVit-565 5′-GACAGTTATTGAGATCCTTCTCTTG3′ from rainbow trout. β-actin was used as housekeeping gene and as sense primer 5′-GTCCTTCATGATTCTCTGCTGA-3′ and antisense primer 5′-ACTCGGGTTCATTTGCATAAACA-3′. Each primer (250 nM, VTG or β-actin) was added to 25 μL PCR reactions containing SYBR Green RT-PCR Reaction Mix (Bio-Rad, Hercules, CA, USA), 100 ng mRNA sample and iScript Reverse Transcriptase for One-Step RT-PCR from BioRad. Thermocycling parameters were as follows: 10 min at 50 °C (cDNA synthesis); 5 min at 95 °C (iScript Reverse transcriptase inactivation); 40 cycles of 10 s at 95 °C and 30 s at 56 °C. Fluorescence data were collected at the end of each cycle. Following the amplification reaction, a melting curve analysis was carried out between 60 and 95 °C, fluorescence data were collected at 0.1 °C intervals. A reaction excluding reverse-transcriptase was included as a negative control. The C(t) was selected to be in the linear phase of amplification. All real-time reactions were done in an iCycler-MyIQ Single Color Real-Time PCR Detection System (Bio-Rad) and data analysis was done using IQ5 (Bio-Rad). The efficiency of the reaction was determined using a dilution series and was determined to be 95 ± 6%. Data was presented as fold-induction relative to solvent-control cells. The limits of detection were 0.15 ng/L. 2.5. In Vivo Measurement of VTG Protein Expression in Medaka. To evaluate the estrogenic activity of extracts from sites as well as compounds identified in active fractions or sites, in vivo bioassays were carried out using adult male Japanese medaka (O. latipes). The fish were >120 d old posthatch (18.8 ± 1.3 mm, 64.8 ± 14.7 mg) and obtained from an ongoing culture at the University of California, Riverside, maintained at

2. METHODS 2.1. Sampling Methods. Water samples (10 L) were obtained from 7 locations within the SF Bay Delta in April of 2008 (Table 1). Samples were extracted by solid phase extraction (Oasis HLB, Waters, Milford, MA) and eluted with methanol for chemical and bioassay evaluation as previously described.3 Table 1. Sampling Sites in the Sacramento/San Joaquin Delta River site ID 340 405 508 602 711 815 902 915

location

latitude

longitude

Napa River at Delta Carquinez Strait, just west of Benicia army dock Suisun Bay, off Chipps Island Grizzly Bay, northeast of Suisun Slough Sacramento River at the tip of Grand Island San Joaquin River Old River at mouth of Holland Cut Old River−Western arm at railroad bridge

38-05′51′′ N 38-02′22.9′′ N

122-15′43.9′′ W 122-09′01.8′′ W

38-02′43.8′′ N 38-06′50.4′′ N

121-55′07.7′′ W 122-02′46.3′′ W

38-10′43.7′′ N

121-39′55.1′′ W

38-01′09.1′′ N

121-34′55.9′′ W

37-56′33′′ N

121-33′48.6′′ W

2.2. Chemical Analyses. Analytes included 31 pesticides, 17α-estradiol, 17β-estradiol, estrone, estriol, progesterone, medroxyprogesterone, testosterone, androstenedione, nonylphenol, nonylphenol monoethoxylate, nonylphenol diethoxylate, octylphenol, octylphenol monoethoxylate, and octylphenol diethoxylate (Table S1 of the Supporting Information). Grab water samples were collected in previously baked 4-L amber glass bottles. Samples were immediately packed in containers with ice and transported to the laboratory and processed for water quality measurements (chemical and estrogenicity analysis). Chemical analysis (experimental section of the Supporting Information) involved filtration, solid phase extraction (SPE), and GC-MS/MS analysis for steroid hormones and nonionic detergents and their degradation products (i.e., nonylphenol, octylphenol, octylphenol mono-, diethoxylates, and nonylphenol monoand diethoxylates) using modifications to previously published methods.3 Positive controls for bioassays consisted of 100 ng/L E2-amended river water and dechlorinated tap water. 2.3. In Vitro Bioassay: Primary Hepatocyte Culture Preparation and Exposure. To evaluate the in vitro estrogenic effects of the detected compounds individually and in mixtures, exposures to rainbow trout primary hepatocytes were carried out. Following exposures, vitellogenin mRNA was measured using quantitative PCR. Primary hepatocytes were isolated from juvenile rainbow trout (Oncorhynchus mykiss) of approximately 5 months (15 ± 3 cm) that were purchased from Jess Ranch Fishing Lakes (Apple Valley, CA) and held in a flow-through living-stream system with carbon-filtered dechlorinated municipal water at 13−15 °C. The fish were fed daily with commercial fish food (Nelson and Sons Silvercup, Murray, UT). Gonadal morphology indicative of gender was not 9107

dx.doi.org/10.1021/es3016759 | Environ. Sci. Technol. 2012, 46, 9106−9111

Environmental Science & Technology

Article

Table 2. Codes for Reconstituted Assays Using AP/APEOS Mixtures and Pesticidesa Exposure ID C AP/APEOS NP NPEOS OP OPEOS pesticides bifenthrin diuron

SC

A1

A5

91 ng/L 606 ng/L 13 ng/L 84 ng/L

455 ng/L 3030 ng/L 65 ng/L 420 ng/L

P1

P5

B1

B5

1 ng/L 41 ng/L

5 ng/L 205 ng/L

1 ng/L

5 ng/L

D1

D5

41 ng/L

205 ng/L

a

C, Control (water); SC, solvent control (0.01% ethanol); NP, nonylphenol; NPEOS, nonylphenol ethoxylates; OP, octylphenol; OPEOS, octylphenol ethoxylates; B, Bifenthrin, D: Diuron.

25 ± 2 °C and a light cycle of 16:8 h light/dark photoperiod. Dechlorinated tap water was used for all stock cultures and experiments. Water quality parameters were constantly monitored, and fish were fed live Artemia sp. nauplii ad libitum twice daily. Fish were not fed during exposure. Adult male Japanese medaka were held in 1000 mL glass jars (3 fish per jar, three replicates per treatment) under static conditions for 7 days. For the site comparisons, animals were exposed to reconstituted ethanol extracts, the solvent carrier ethanol (0.01% v/v), and to the positive control (100 ng/L E2) and negative control (dechlorinated tap water). Microbial buildup was not observed on containers from any treatment. For the chemical and mixture evaluations, exposures also were carried out for 7 days, but chemical concentrations were refreshed daily from measured stock solutions. The measured values were 98 ± 4% of nominal concentrations. On the last day of exposure, the fish were anesthetized with 1 g/L tricaine methane sulfonate (MS-222) and the livers were excised and pooled from each individual replicate and frozen at −80 °C until VTG analysis. The VTG protein level in the liver was quantified by enzymelinked immunosorbent assay (ELISA) using commercially available kits for Japanese medaka (Biosense, Bergen, Norway). The VTG concentrations were normalized to total protein in liver homogenate determined with Coomassie blue protein assay (Pierce Biotechnology, Rockford, IL). The limit of detection for the VTG assay was 0.1 ng/μg protein. For site− site comparisons, estrogenic activity was expressed as 17βestradiol equivalents (EEQs), which was calculated using an E2 dose−response curve.3,5 2.6. Reconstitution Studies. Using the highest ambient combined concentrations of pesticides and AP/APEOs determined in water samples from the SF Bay Delta (1×), Japanese medaka were exposed to the AP/APEO mixture with and without pesticides (Table 2). Because the estrogenic activity of individual APs and APEOs in Japanese medaka was previously characterized,4 the APs and APEOs were combined as a surfactant mixture for this study. An arbitrary 5-fold higher concentration of each compound was also used for concentration−response comparisons. Although some of the compounds have been detected at 5-fold higher concentrations in other waterbodies,20,21 only the 1× values were considered environmentally relevant for this study. Exposure duration was 7 days with daily renewal. Similarly, in vitro studies were also carried out using the same treatment combinations for 24 h. Each treatment was replicated 4 times with negative (solvent) and positive controls (100 ng/L E2). 2.7. Biotransformation Studies. To evaluate the potential mechanism of enhanced biotransformation of bifenthrin to

estrogenic metabolites by AP/APEOs, animals were treated as described above with the ambient AP/APEO mixture for 7 days. The animals were euthanized with MS-222, the livers removed, pooled, and microsomal preparations were carried out with subsequent freezing at −80 °C.18,22 Bifenthrin biotransformation was evaluated in liver microsomes using a modification of Nillos et al.18 14C-Testosterone hydroxylase was measured as previously described.22 2.8. Statistical Analyses. Data were presented as mean ± standard deviation (SD). Statistical differences between treatments (p < 0.05) were determined using analysis of variance (ANOVA). Prior to performing ANOVA, the experimental data were checked for homogeneity of variance using Bartlett’s test. Treatment differences were assessed with Bonferroni’s and Tukey’s Multiple Comparison Tests (GraphPad Prism v.4.03; GraphPad Software Inc., San Diego, CA).

3. RESULTS In vivo estrogenic activity was observed in most water extracts from the SF Bay Delta in April of 2008 (Table 3). Chemical Table 3. Estradiol Equivalents (EEQs) in the Sacramento/ San Joaquin Delta River Sites in April 2008 by in Vivo Bioassays (Data Are Presented As Mean ± SD, 4 Replicates of 3 Animals Per Exposure)a

a

site ID

VTG (ng/μg)

in vivo EEQs (ng/L)

340 405 508 602 711 815 902 915

20.1 ± 1.3 139 ± 6.0 bdl 25.8 ± 3.7 121 ± 5.0 15 ± 6.2 bdl 42 ± 6.2

0.90 ± 0.03 25.65 ± 4.01 bdl 1.05 ± 0.10 12.79 ± 1.65 0.80 ± 0.12 bdl 2.02 ± 0.32

LOD of VTG (0.1 ng/μg), bdl = below detection limit.

analyses of the water samples indicated the presence of diuron, bifenthrin, 4-nonylphenol, 4-octylphenol, and the corresponding mono- and diethoxylates ranging from below detection to 379 ng/L (Table S2 of the Supporting Information). Linear regression analyses failed to show relationships between individual compounds and biological activity (data not shown). When evaluating individual compounds, significant increases in vitellogenin protein were only observed in the 5-fold higher AP/APEO mixture (A5) and the positive control (E2) (Figure 1). However, when compounds were combined significant elevations were observed. When bifenthrin and diuron were 9108

dx.doi.org/10.1021/es3016759 | Environ. Sci. Technol. 2012, 46, 9106−9111

Environmental Science & Technology

Article

Central Valley and upper SF Bay Delta, bioassay-guided fractionation was performed on samples where estrogenic activity was routinely observed throughout 2006−2007.3 Sitespecific signatures of estrogenic activity were observed, but subsequent measurements of 95 compounds including steroid hormones, pharmaceutical, industrial, and agricultural chemicals failed to identify causative agents in bioactive fractions. To further explore the SF Bay Delta, evaluation of seven additional locations within the SF Bay Delta were carried out and results indicated that all but one location had measurable in vivo estrogenic activity. The site which had the highest activity (Sacramento River) was 12 mi upstream of a site had consistently higher activity in 2006 and 2007.3 Fractionation of the water extracts yielded ng/L concentrations of diuron, simazine, hydroxyatrazine, alkylphenol ethoxylates, and several other pharmaceutical compounds. Reconstitution studies with the individual compounds from that study failed to show significant in vitro or in vivo activity.3 Likewise, neither in vitro nor in vivo estrogenic activity was observed after treatment with a combined exposure of the 6 alkylphenols and alkylphenol ethoxylates at ambient concentrations. Alkylphenol ethoxylates are widely used in North America as surfactants and industrial detergents.8 In some applications, they are combined with pesticides to enhance bioavailability.5 The concentrations observed in the SF Bay Delta were comparable to concentrations observed in Back River, a subestuary of Chesapeake Bay, which showed that APEs were primarily derived from wastewater.20 However, the concentrations observed in this batch of samples from the SF Bay Delta would require a 10 to 50% contribution from typical WWTP effluent,20 and effluent discharge of this magnitude does not occur in these sites. Effluent from the largest WWTP in the area typically contributes no more than 2.5% of the Sacramento flow in April when the samples were collected (Mueller-Solger, Personal Communication; Interagency Ecological Program; Department of Water Resources). The degradation products of the ethoxylates, alkylphenols, have been shown to have estrogenic activity with multiple mechanisms of action including weak binding to the estrogen receptor, enhanced synthesis of gonadotropin, and the inhibition of steroid conjugation enzymes.9 Threshold values for in vivo estrogenic activities of 4-nonyl- and octylphenols are typically in the 1 μg/L range with the corresponding ethoxylates much less active.4 Because concentrations from the SF Bay Delta samples were in the ng/L range, it was not surprising that the AP/APEO combination at ambient concentrations did not elicit an estrogenic response, but the 5-fold higher concentration did cause estrogenic activity in vivo. The greater response in vivo to the AP/APEO mixture relative to in vitro activity was consistent with previous studies with sheepshead minnows, which had greater response in the whole animal compared to cellular responses.10 The greater in vivo activity is likely due to the multiple molecular targets of AP/ APEOs that either enhance endogenous hormone concentrations or lead to effects outside of direct interactions of the AP/APEOs with the ER. When hepatocytes or fish were exposed to the pyrethroid insecticide, bifenthrin at 1 ng/L or 5 ng/L, estrogenic activity was also not observed. Previous studies in Japanese medaka and zebrafish at 10 μg/L concentrations of bifenthrin induced production of vitellogenin.11,12 Exposure of fathead minnow to 140 ng/L also caused significant induction of Vtg mRNA.13 Although bifenthrin is estrogenic in laboratory exposures to

Figure 1. Effects of AP/APEOS mixtures (A) and pesticides both as mixtures (P) and individual compounds (B, D) on VTG production by in vivo Japanese medaka bioassays. See Table 2 for codes. Data are presented as mean ± SD. Different letter means significant differences (One-way ANOVA, Tukey’s Test, p < 0.05).

exposed as a mixture together (P1) with AP/APEs (A1), estrogenic activity was higher than the AP/APE mixture alone (A1), or when bifenthrin alone (B1) was combined with AP/ APEs (Figure 2). In contrast, when diuron alone (D1) was

Figure 2. Effects of combined AP/APEOS mixtures and pesticides on VTG production by in vivo Japanese medaka bioassays. See Table 2 for codes. Data are presented as mean ± SD. Different letter means significant differences (One-way ANOVA, Tukey’s Test, p < 0.05).

combined with AP/APEs (A1) significant estrogenic activity was observed, but the activity was not as high as the full combinations. In vitro evaluations, though positive for E2 were not affected by any of the treatments (Figure S1 of the Supporting Information). Biotransformation studies evaluating the potential mechanism of bifenthrin biotransformation failed to detect in vitro biotransformation in liver microsomes from Japanese medaka, but biotransformation of bifenthrin was observed in human (3.5 ± 1.5% Conversion) and rat liver (8 ± 4% Conversion) microsomes. Treatment of medaka with ambient concentrations of AP/APEOs for 7 days failed to induce hepatic testosterone hydroxylase (Figure S2 of the Supporting Information).

4. DISCUSSION To determine the cause of estrogenic activity previously observed in water extracts from the surface waters of the 9109

dx.doi.org/10.1021/es3016759 | Environ. Sci. Technol. 2012, 46, 9106−9111

Environmental Science & Technology

Article

model fish species, exposure to environmentally relevant concentrations which are approximately 100−1000 fold less do not appear to cause estrogenic activity in fish when exposure occurs as the single compound. Similarly, individual exposures to ambient concentrations of diuron or to 5-fold higher concentrations failed to elicit estrogenic effects in either assay. Diuron is the third most applied herbicide in the state of California and surface water concentrations peak during rain storm events.21 Diuron has been shown to be antiestrogenic, antiandrogenic, and disrupt amphibian steroidogenesis at relatively high concentrations (0.3−7.3, 15.6−7.3, and 14.5 mg/L, respectively).14 However, diuron is degraded in the environment to 3,4-dichloroanilide and undergoes biotransformation to 3,4-dichloroacetanilide via CYP1A within vertebrates.15 Each metabolite/degradate weakly binds to the bovine androgen receptor with relative binding affinities between 0.002 and 0.01% of the affinity of dihydrotestosterone.16 Again, given the nanogram per liter concentrations observed in these specific SF Bay Delta samples, estrogenic activity was not expected. Although estrogenic activity was not observed with the AP/ APEO treatment, the combination of the surfactants with the two pesticides at ambient concentrations caused in vivo estrogenic activity. Mechanisms for the enhancement of in vivo activity but not in vitro activity are unclear but, like the AP/APEO response, may point to molecular targets other than the direct interaction of the parent compounds with the ER. Potential mechanisms include (1) enhanced absorption of the pesticides, (2) the induction of biotransformation enzymes that convert each pesticide to more potent ER ligands, (3) inhibition of estradiol clearance through inhibition of steroid catabolism, or (4) upstream enhancement of steroid biosynthesis. Because APs/APEOs enhance the bioavailability of active ingredients including pesticides,5 uptake of diuron into the animal may have been augmented. However, given the hydrophobic nature of bifenthrin, absorption of this compound would likely be rapid even without the surfactants. In addition, enhanced absorption should have led to an increased estrogenic signal within the hepatocytes, which was not observed in the mixture studies. Whether greater uptake occurs with the mixture in the whole animal relative to isolated cells is unclear. Without measurements of tissue concentrations of the pesticides with and without the surfactants, this mechanism cannot be discounted. With regard to biotransformation in fish, lower environmentally relevant doses of nonylphenol has been shown to induce several cytochrome P450 isoforms, but in higher concentrations NP inhibits expression and catalytic activities of these enzymes.17 Another pyrethroid, permethrin was stereoselectively metabolized to metabolites via cytochrome P450 that were more potent ER agonists.18 Similarly, diuron is activated to endocrine active metabolites by cytochrome P450 as well.15,16 Experiments with liver microsomes from Japanese medaka failed to show detectable biotransformation of bifenthrin, and liver microsomes from fish treated with the AP/APEO mixture failed to induce or inhibit cytochrome P450 activities. The catalytic activities of 6β and 16β testosterone hydroxylase have been previously observed in heterologously expressed CYP3As from medaka23 and the baseline activity observed in the current study was comparable to activities observed in medaka liver microsomes (∼30 pmol/min/mg).24 Heterotropic and homotropic allosteric cooperativity with nonylphenol was observed with medaka CYP3A with low

concentrations (10−100 nM) stimulating catalytic activity and exposure to higher concentrations of NP (>100 nM) inhibiting activity25. Consequently, the failure of induction and/or inhibition of CYP biotransformation by alkylphenols and their ethoxylates suggests that enhanced formation of estrogenic metabolites (from bifenthrin or diuron) may not be a plausible mechanism explaining the elevated estrogenic response in the fish. Alternatively, the AP/APEO mixture may alter conjugation reactions leading to presumably higher plasma concentrations of steroid hormones.17 Plasma levels of E2 increased 3-fold in Atlantic Salmon (Salmo salar) following treatment with 50 ug/L NP.19 Whether E2 elevation occurs in medaka following exposure to the mixtures is unknown but, if E2 is present in high concentrations, it could be responsible for the enhanced vitellogenin observed in the present study. However, the additional impact of diuron and bifenthrin clearly cause additional augmentation. The impact of these mixtures on steroid conjugation and/or biosynthesis, particularly at upstream feedback loop targets, should be further explored. In summary, mixtures of pesticides with alkylphenols and alkylphenol ethoxylates at ambient concentrations in the SF Bay Delta caused significant estrogenic activity in fish but not in isolated cells. Because many of the same compounds were observed in previous bioassay guided fractionation studies, an association between these compounds and estrogenic activity in fish is present. Additional studies characterizing the mechanisms of chemical interactions and the potential reproductive outcomes would help determine whether a causal relationship exists between these compounds and fish declines within the SF Bay Delta.

■

ASSOCIATED CONTENT

S Supporting Information *

Table of analytes analyzed in water samples from Sacramento/ San Joaquin Delta River; table of steroids, AP/APEOS, and pesticides in the Sacramento/San Joaquin Delta sites; and figures of effects of AP/APEO mixtures. This material is available free of charge via the Internet at http://pubs.acs.org.

■

AUTHOR INFORMATION

Corresponding Author

*E-mail: [email protected]. Notes

The authors declare no competing financial interest.

■

ACKNOWLEDGMENTS This work was supported by the CALFED Bay Delta Program (Grant No. U055C031) and the UCR Agricultural Experimental Station (D.S.). We thank David Crain at the California Fish and Game Laboratory at Rancho Cordova Laboratory for the analyses of pesticides, and the laboratory of Jay Gan at UC Riverside for the provision of bifenthrin and analytical verification of nominal chemistry values.

■

REFERENCES

(1) Thomson, J. R.; Kimmerer, W. J.; Brown, L. R.; Newman, K. B.; Mac Nally, R.; Bennett, W. A.; Feyrer, F.; Fleishman, E. Bayesian change point analysis of abundance trends for pelagic fishes in the upper San Francisco Estuary. Ecol. Appl. 2010, 20, 1431−1448. (2) Spearow, J. L.; Kota, R. S.; Ostrach, D. J. Environmental contaminant effects on juvenile striped bass in the San Francisco Estuary, California, USA. Environ. Toxicol. Chem. 2011, 30 (2), 393− 402.

9110

dx.doi.org/10.1021/es3016759 | Environ. Sci. Technol. 2012, 46, 9106−9111

Environmental Science & Technology

Article

(3) Lavado, R.; Loyo-Rosales, J. E.; Floyd, E.; Kolodziej, E. P.; Snyder, S. A.; Sedlak, D. L.; Schlenk, D. Site-specific profiles of estrogenic activity in agricultural areas of California’s inland waters. Environ. Sci. Technol. 2009, 43 (24), 9110−9116. (4) Balch, G.; Metcalfe, C. Developmental effects in Japanese medaka (Oryzias latipes) exposed to nonylphenol ethoxylates and their degradation products. Chemosphere 2006, 62 (8), 1214−1223. (5) Xie, L. T.; Thrippleton, K.; Irwin, M. A.; Siemering, G. S.; Mekebri, A.; Crane, D.; Berry, K.; Schlenk, D. Evaluation of estrogenic activities of aquatic herbicides and surfactants using an rainbow trout vitellogenin assay. Toxicol. Sci. 2005, 87 (2), 391−398. (6) Carlson, D. B.; Williams, D. E. Sex-specific vitellogenin production in immature rainbow trout. Environ. Toxicol. Chem. 1999, 18 (10), 2361−2363. (7) Mossman, T. Rapid colorimetric assay for cellular growth and survivals: Application to proliferation and cytotoxicity assays. J. Immunol. Methods 1983, 65, 55−63. (8) Nimrod, A. C.; Benson, W. H. Estrogenic responses to xenobiotics in channel catfish (Ictalurus punctatus). Mar. Environ. Res. 1996, 42 (1−4), 155−160. (9) Rempel, M. A.; Schlenk, D. Effects of environmental estrogens and antiandrogens on endocrine function, gene regulation, and health in fish. Int. Rev. Cell Mol. Biol. 2008, 267, 207−252. (10) Folmar, L. C.; Hemmer, M. J.; Denslow, N. D.; Kroll, K.; Chen, J.; Cheek, A.; Richman, H.; Meredith, H.; Grau, E. G. A comparison of the estrogenic potencies of estradiol, ethynylestradiol, diethylstilbestrol, nonylphenol and methoxychlor in vivo and in vitro. Aquat. Toxicol. 2002, 60 (1−2), 101−110. (11) Jin, S.; Yang, F.; Liao, T.; Hui, Y.; Xu, Y. Seasonal variations of estrogenic compounds and their estrogenicities in influent and effluent from a municipal sewage treatment plant in China. Environ. Toxicol. Chem. 2008, 27, 146−153. (12) Wang, L. M.; Liu, W.; Yang, C. X.; Pan, Z. Y.; Gan, J. Y.; Xu, C.; Zhao, M. R.; Schlenk, D. Enantioselectivity in estrogenic potential and uptake of bifenthrin. Environ. Sci. Technol. 2007, 41 (17), 6124−6128. (13) Beggel, S.; Connon, R.; Werner, I.; Geist, J. Changes in gene transcription and whole organism responses in larval fathead minnow (Pimephales promelas) following short-term exposure to the synthetic pyrethroid bifenthrin. Aquat. Toxicol. 2011, 105 (1−2), 180−188. (14) Orton, F.; Lutz, I.; Kloas, W.; Routledge, E. J. Endocrine disrupting effects of herbicides and pentachlorophenol: in vitro and in vivo evidence. Environ. Sci. Technol. 2009, 43 (6), 2144−2150. (15) Abass, K.; Reponen, P.; Turpeinen, M.; Jalonen, J.; Pelkonen, O. Characterization of diuron N-demethylation by mammalian hepatic microsomes and cDNA-expressed human cytochrome p450 enzymes. Drug Metab. Dispos. 2007, 35 (9), 1634−1641. (16) Bauer, E. R. S.; Daxenberger, A.; Petri, T.; Sauerwein, H.; Meyer, H. H. D. Characterisation of the affinity of different anabolics and synthetic hormones to the human androgen receptor, human sex hormone binding globulin and to the bovine progestin receptor. APMIS 2000, 108 (12), 838−846. (17) Arukwe, A.; Forlin, L.; Goksoyr, A. Xenobiotic and steroid biotransformation enzymes in Atlantic salmon (Salmo salar) liver treated with an estrogenic compound, 4-nonylphenol. Environ. Toxicol. Chem. 1997, 16 (12), 2576−2583. (18) Nillos, M. G.; Chajkowski, S.; Rimoldi, J. M.; Gan, J.; Lavado, R.; Schlenk, D. Stereoselective biotransformation of permethrin to estrogenic metabolites in fish. Chem. Res. Toxicol. 2010, 23 (10), 1568−1575. (19) Kortner, T. M.; Mortensen, A. S.; Hansen, M. D.; Arukwe, A. Neural aromatase transcript and protein levels in Atlantic salmon (Salmo salar) are modulated by the ubiquitous water pollutant, 4nonylphenol. Gen. Comp. Endrocrinol. 2009, 164 (1), 91−99. (20) Loyo-Rosales, J. E.; Rice, C. P.; Torrents, A. Fate and istribution of the octyl-and nonlyphenol ethoxylates and some carboxylated transformation products in the Back River, Maryland. J. Environ. Mon. 2010, 12, 614−621. (21) Green, P. G.; Young, T. M. Loading of the herbicide diuron into the California Water System. Environ. Eng. Sci. 2006, 23 (1), 545−551.

(22) Lavado, R.; Rimoldi, J. M.; Schlenk, D. Mechanisms of fenthion activation in rainbow trout (Oncorhynchus mykiss) acclimated to hypersaline environments. Toxicol. Appl. Pharmacol. 2009, 235, 143− 152. (23) Kullman, S. W.; Hamm, J. T.; Hinton, D. E. Identification and characterization of a cDNA encoding cytochrome P450 EA from the fresh water teleost Japanese medaka (Oryzias latipes). Arch. Biochem. Biophys. 2000, 380, 29−38. (24) Kullman, S. W.; Kashiwada, S.; Hinton, D. E. Analysis of medka cytochrome P450 #A homotropic and heterotropic cooperativity. Mar. Env. Res. 2004, 58, 469−473.

9111

dx.doi.org/10.1021/es3016759 | Environ. Sci. Technol. 2012, 46, 9106−9111