* Department of Environmental Sciences, University of California at Riverside, Riverside, California 92507; San Francisco Estuary Institute, 7770 Pardee Lane, Oakland, California 94621;
California Department of Fish and Game Fish and Wildlife Water Pollution Control Laboratory, Rancho Cordova, California 95670; and
Department of Mathematics, University of Louisiana at Lafayette, Lafayette, Louisiana 70504
1 To whom correspondence should be addressed at Department of Environmental Science, University of California at Riverside, 3401 Watkins Dr., Riverside, CA 92521. Fax: 751-827-3993. E-mail: DANIEL.SCHLENK{at}ucr.edu.
Received April 22, 2005; accepted July 16, 2005
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ABSTRACT |
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Key Words: estrogenicity; surfactants; herbicides; rainbow trout; in vivo; vitellogenin.
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INTRODUCTION |
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Aquatic herbicides such as 2,4-dichlorophenoxyacetic acid (2,4-D), diquat dibromide, glyphosate, and triclopyr are widely used to selectively control broadleaf and woody plants in various waterways. 2,4-D is one of the oldest and most commonly used herbicides. Once it is in water, it is readily degraded to 2,4-dichlorophenol (Crosby and Tutass, 1966), which is an estrogen receptor (ER) ligand (Jobling et al., 1995
). Triclopyr has a similar chemical structure to 2,4-D and photochemically decomposes to trichloropyridinol within hours once it is in water (Petty et al., 2003
). Tricholopyridinol is also the main metabolite of chlorpyrifos, which also has weak ER activation (Andersen et al., 2002
). One study has shown that diquat does not initiate responses in estrogen receptor
and ß transactivation assays (Kojima et al., 2004
).
Alkylphenolic polyethoxylates (ApnEO, n = 640), such as nonylphenol ethoxylates (NPE) and octylphenol ethoxylates (OPE), are often used in combination with aquatic pesticides as dispersing agents, detergents, emulsifiers, and solubilizers. Alkylphenol ethoxylates (APEs) are normally present in raw sewage effluent, and their degraded products (alkylphenols) bind estrogen receptor (Routledge and Sumpter, 1996) and cause estrogenic effects in fish (Jobling and Sumpter, 1993
; Routledge et al., 1998
).
Few studies have addressed the estrogenic effects of combined exposure of pesticides and surfactants. Four herbicides were selected for evaluation based upon evidence of prior usage with surfactants and, in some cases, the potential for estrogenic activity. The objective of this study was to investigate the estrogenic potencies of four herbicides, nonionic surfactants, and the mixture of herbicides with surfactants using an in vivo rainbow trout vitellogenin assay.
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MATERIALS AND METHODS |
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Measurement of chemicals.
Measurements of 2,4-D, 4-nonylphenol (NP), nonylphenolethoxylates (NPE), triclopyr, glyphosate, and diquat were determined by the Department of Fish and Game Water Pollution Control Laboratory (Rancho Cordova, CA). For 2,4-D, water samples were acidified to pH 2 with sulfuric acid:water (1:1) and extracted by solid phase extraction (C18). The extracts were analyzed by LC-MS in API-ES negative mode. For the two surfactants, R-11 and TPA, water samples were extracted by solid phase extraction (C18) for NP and NPE. The extracts were analyzed by high-performance liquid chromatography (HPLC) with fluorescence detection and confirmed by LC-MS in API-ES negative and positive mode for NP and NPE, respectively (Huggett et al., 2003
). For glyphosate, samples were filtered and injected directly in high-performance liquid chromatography (HPLC) with post-column derivatization. For diquat, The pH of water samples was brought up to 10.5 ± 0.2 with 10% w/v NaOH (aq) or 10% v/v HCl (aq) prior to extraction using solid phase extraction (C18). A Hewlett Packard 1100 HPLC equipped with a diode array detector (DAD) was used to analyze the samples (at 308 nm for diquat). For triclopyr, water samples were acidified to pH
2 with sulfuric acid:water (1:1) and extracted by solid phase extraction (C18). The method detection limit and percentage recovery for each method were provided in Table 1.
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Range finding experimental exposure.
Rainbow trout were exposed to pesticides and surfactants at concentrations based upon the recommendations provided by the pesticide manufacturers. The measured concentrations for pesticides, glyphosate, 2,4-D, diquat dibromide, and triclopyr, were 0.11, 1.64, 2.07, and 1.25 mg/l respectively. The nominal concentrations for R-11 and TPA were 1.46 and 0.8 mg/l. For exposure to mixtures, rainbow trout were exposed to the highest concentration of each pesticide combined with each surfactant at fixed ratios of 1:2 (R-11:pesticide) or 1:4.5 (TPA:pesticide). Exposures were carried out in 20-l tanks filled with aerated water in a daily static renewal system for 7 days. In addition, rainbow trout were exposed to five concentrations of 17ß estradiol in ethanol to calculate estradiol equivalent values (EEQs) as previously described (Huggett et al., 2003; Thompson et al., 2000
; Xie et al., 2005
). Each concentration of chemical (treatment or control) had three replicates with two fish in each tank. The test water in each individual tank was monitored daily for water chemistry after every water renewal. The hardness of the test water ranged from 142 to 162 mg/l (as CaCO3); the free chlorine was <0.2 mg/l. Alkalinity ranged from 148 to 180 mg/l, and ammonia (as N-NH3) was <0.02 mg/l. Dissolved oxygen averaged 94.6% of the air saturation value, and pH values ranged from 6.0 to 8.2 mg/l. Water temperature was maintained at 16 ± 1°C. Fish were fed rainbow trout chow at approximately 1% of their body weight during the exposure. The photo period was 16 h:8 h (light:dark).
Dose-response exposure.
A 7-day dose-response exposure was carried out for 2,4-D, triclopyr, R-11, and TPA. Mixture experiments were conducted evaluating 2,4-D with R-11 or TPA and triclopyr with TPA. Each concentration (control or treatment) had three replicates with two fish in each replicate. Water chemistry, water temperature, and photo-period were similar to that of the worst-case scenario exposure (above). Fish were exposed to 2,4-D at concentrations of 0 (control), 0.00164, 0.0164, 0.164, and 1.64 mg/l. Fish were exposed to R-11 at concentrations of 0 (control), 0.0146, 0.146, 0.73, and 1.46 mg/l. Fish were exposed to TPA at concentrations of 0 (control), 0.008, 0.08, 0.4, 0.8 mg/l. For binary mixtures of pesticide 2,4-D with the two surfactants, a fixed ratio was used (the ratio of the two chemicals was kept constant, while the total concentrations of the mixture was varied). For 2,4-D + R-11, the concentrations used were 0 (control), 0.00164 mg/l (2,4-D) + 0.00089 mg/l (R-11), 0.0164 mg/l (2,4-D) + 0.0089 mg/l (R-11), 0.164 mg/l (2,4-D) + 0.089 mg/l (R-11), and 1.64 mg/l (2,4-D) + 0.89 mg/l (R-11). For binary mixture of 2,4-D + TPA, the concentrations used were 0 (control), 0.00164 mg/l (2,4-D) + 0.00048 mg/l (TPA), 0.0164 mg/l (2,4-D) + 0.0048 mg/l (TPA), 0.164 mg/l (2,4-D) + 0.048 mg/l (TPA), and 1.64 mg/l (2,4-D) + 0.48 mg/l (TPA). For binary exposure of TPA + triclopyr, the concentrations used were 0, 0.013 µg/l (TPA) + 1 µg/l (triclopyr), 0.13 µg/l + 10 µg/l (triclopyr), 1.3 µg/l (TPA) + 100 µg/l (triclopyr), and 13 µg/l (TPA) + 1000 µg/l (triclopyr).
Plasma vitellogenin levels determination.
After the exposure, the fish were euthanized in MS-222 (50 mg/l). Blood samples from rainbow trout were obtained by an incision at the caudal peduncle and collection of the blood exiting the incision. Blood was centrifuged at 10,000 rpm for 3 min at room temperature. After centrifugation, PMSF (Phenylmethyl sulphonyl fluoride; stock solution 0.1 M) was added to the plasma samples at a final concentration of 1 mM. The plasma samples were stored at 80°C until analysis.
Plasma vitellogenin levels were determined by enzyme-linked immunosorbent assay (ELISA) as previously described using a commercially available rainbow trout ELISA kit (Biosense, Bergen, Norway) (Xie et al., 2005). Total protein concentrations of the plasma samples were determined according to the methods of Bradford using bovine serum albumin as standards (0.252 mg/ml). Vitellogenin levels in the plasma samples were expressed as ng vitellogenin per mg of total protein.
Field evaluation.
To evaluate the effects of combined surfactant/pesticide exposure in a field setting, water was collected from Anderson Pond (N 40°28.070, W 12°216.372), a 10-acre pond south of Redding, California, near the Sacramento River at the northern end of the Sacramento Valley. Anderson Pond is under surveillance and treated by the California Department of Food and Agriculture for Hydrilla control. No Hydrilla has been observed in the pond since 1999, and 2004 was the final year of required observation. Triclopyr mixed with TPA was applied to control emergent water primrose. The primrose was treated in order to allow sunlight to reach the pond bottom in order to provide ideal growth conditions for any Hydrilla tubers present. The pesticide mixture was applied via 3-gallon hand sprayers. Each 3-gallon pesticide mixture consisted of 0.25 oz of TPA and 19 oz of Renovate (ai. triclopyr triethylamine salt). A total of 2.5 oz of TPA and 190 oz of Renovate was applied to two 20 by 20 meter areas of Anderson Pond. Water was collected from the middle of one of these application areas after application.
Fifty-five gallons of water was collected within the treatment area 1 h after application occurred in July of 2004. Water was also collected in February of 2005 as a negative control. The water was collected from a hand-powered Zodiac boat by submerging 1-gallon cleaned stainless steel buckets into the pond just below the surface and allowing them to fill. The water was transferred via a metal funnel into cleaned, aged plastic 5-gallon bottles (water cooler bottles). Three 5-gallon bottles were filled before the boat returned to shore, and the water was transferred to a 55-gallon Nalgene drum. The Nalgene drum was transported that day to UCR to carry out exposures with trout. Water was placed in 20-l tanks and aerated, and it was also diluted 50% with dechlorinated tap water as a second exposure concentration. Dechlorinated tap water served as a negative control, and estradiol served as a positive control (see above). Ten trout were exposed, using five replicates (two fish per tank) for 7 days, and euthanized, and tissues removed following the guidelines above. Water samples were removed for chemical analyses of triclopyr, 4-NP and nonylphenol ethoxylates after exposures were concluded.
Data analysis.
All data analyses were performed using Statistical Analysis System (SAS, version 8.2, Cary, NC). Normality was evaluated using the Shapiro test and equal variance using Levene's test. Since the assumption of normality and equal variance were violated, nonparametric tests (Kruskall-Wallis) were used to test the difference in vitellogenin levels in rainbow trout among different groups.
Estrogenicity in the unit of estradiol equivalent concentrations for chemicals was estimated from the standard curve of exposure to 17ß estradiol (E2). The estrogenicity of the mixtures was calculated based on the model of concentration addition, which assumes that mixtures act via a similar mode of action in producing an effect (Altenburger et al., 2003; Loewe and Mulschnek, 1926
).
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RESULTS |
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DISCUSSION |
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Mechanism(s) explaining the in vivo estrogenicity of 2,4-D are unknown. Possibilities include the transformation, either through biotic or abiotic means, to metabolites that are more potent ER ligands or that possess more disruptive characteristics. In aqueous solutions, 2,4-D undergoes rapid photolysis and hydrolysis to 2,4, dichlorophenol (Crosby and Tutass, 1966). However, it is slowly metabolized within fish to amino acid conjugates, with up to 90% of the parent being eliminated in the urine unchanged (James et al., 1976
; Plakas et al., 1992
). Although 2,4-dichlorophenol was not active (>1 mM) in the in vitro MCF-7 cell line assay for ER activity (Korner et al., 1998
), it did slightly bind quail ER
, but not ERß (Maekawa et al., 2004
). In trout, 2,4 dichlorophenol was shown to be an ER antagonist (Jobling et al., 1995
). Given the lack of transformation to 2,4 dichlorophenol within fish, and its antagonistic activity at the ER, it would appear that 2,4-D itself, an amino acid conjugate, or an unknown metabolite may elicit estrogenicity indirectly outside of direct ER binding. These data would be consistent with the greater than additive response observed with the APE surfactants at low concentrations. Alternatively, it is well documented that 2,4-D also contains several dioxin-like compounds which may affect the estrogenic response. However, most aryl hydrocarbon receptor (AhR) agonists tend to repress rather than enhance estrogenic activity (Safe and Wormke, 2003
). However, this may be a dose-dependent phenomenon, as previous studies comparing the estrogenic response in rainbow trout hepatocytes with varied doses of the AhR agonist, ß-naphthoflavone, indicated both enhancement and antagonism of Vtg expression following E2 treatment (Anderson et al.1996
).
Whereas 2,4-D demonstrated estrogenic activity alone, triclopyr failed to induce Vtg expression alone, but in combination with APE surfactants caused a significant increase, which was higher than would be predicted from the surfactant treatment alone. Unlike 2,4-D, triclopyr is slowly degraded in the environment, but undergoes biotransformation in fish to the metabolite, trichloropyridinol (Petty et al., 2001), which is the major metabolite of the slightly estrogenic organophosphate chlorpyrifos (Andersen et al., 2002
). Unfortunately, there have not been any reported studies showing whether trichloropyridinol is estrogenic.
For most of the pesticides examined in the current study, combination with either APE-containing surfactant enhanced estrogenic activity. This was not surprising, since it has been well established that alkylphenols and even ethoxylates induce vitellogenin and activate ER, although with low potencies. The LOEC values for Vtg induction (in terms of NP) for TPA and R-11 were similar to the values (µg/l) observed from other studies (Thorpe et al., 2001
) and similar to concentrations of alkylphenol ethoxylate and their degraded products in various surface waters (Rodgers-Gray et al., 2000). Within one day of exposure to 10150 µg/l NP, vitellogenin mRNA was detected in liver of rainbow trout, with maximum production detected after 72 h of exposure (Lech et al., 1996
). There did not seem to be a consistent difference between R-11 and TPA with regard to induction, alone or in combination with the pesticides. However, the binary mixture of 2,4-D with R-11 consistently showed higher responses in rainbow trout in terms of estrogenic activities than 2,4-D and TPA. In contrast, it was the TPA and 2,4-D mixture which provided statistically significant effects in terms of greater than additive response of EEQs at the lowest tested concentrations. Combinations with TPA also caused a statistically significant less than additive response at the high concentration. It should be noted that a similar trend was observed with the R-11 mixtures with 2,4-D, but the values were not statistically significant. The difference in responses between the mixture of 2,4-D with either surfactant may be due to the other constituents within each surfactant mixture. For example, although there were no marked differences in the content of 4-NP in the two surfactants, NPEs were approximately 20% higher in the TPA (data not shown). As these values only represent two compounds out of a multitude of isomers and other "inert" ingredients within the surfactant, it is unclear whether other compounds could be influencing the responses.
A consistent inverse sigmoidal concentration-response curve was observed when 2,4-D was combined with either surfactant. Estrogenicity was higher than the predicted sum of either compound at low concentrations, with gradual movement toward additivity in mid-range concentrations, followed by less than additive (predicted) responses in the highest concentrations. Greater than additive responses were also observed for triclopyr and TPA in the laboratory and possibly the field. Reasons for this pattern are unknown, but several reports have shown that combining environmental estrogens at sub-NOEC concentrations resulted in a dramatic enhancement of the estrogenic effect (Rajapakse et al., 2001, 2002
). U-shaped dose-response curves have been documented in many biological, toxicological, and pharmacological studies (Calabrese and Baldwin, 2001
). In one study, it was shown that certain phytoestrogens were aromatase (CYP19) inhibitors at low concentrations (<1 µM) diminishing estrogen synthesis but ER ligands at high concentrations (>1 µM) (Almstrup et al., 2002
). Alternatively, signal transduction pathways may be non-ER targets, as upregulation of coregulators may enhance the transcriptional activity of steroid hormone receptors even in the absence of the ligands (Katzenellenbogen et al., 1996
). Earlier studies examining gender ratios in fish following larval treatment with 4-NP indicated masculinization at low concentrations of exposure followed by feminization at higher concentrations (Nimrod and Benson, 1998
). One simple possibility for diminishing estrogenic activity at higher mixture concentrations may be the acute toxicity of the surfactants, which may inhibit overall protein synthesis and, hence, the estrogenic response in fish. The 72-h LC50 for 50- to 200-g rainbow trout of 4-NP was 150 to 250 µg/l (Lech et al., 1996
). Although the concentrations measured in the current study were less than this, smaller fish and a longer duration (7 d) were used, which have led to enhanced toxicity. Range-finding studies showed that 5 mg/l of each surfactant resulted in 30% of the mortality in rainbow trout within 48 h. In argument against toxicity, no mortality was observed in any of the surfactant-treated fish, and the Vtg response was clearly concentration-dependent when the fish were treated with only surfactants at the same concentrations as those used for the mixtures. The 96-h LC50 for the dimethylamine salt of 2,4-D was reported to be 100 mg/l (Tomlin, 1994
). The highest concentration utilized in the current study was 1.64 mg/l, and no mortality was observed in any treatment. Thus, it would appear that the diminished estrogenic responses at the highest concentrations are not likely due to acute toxicity. Clearly, numerous targets may be involved with these responses, and more research dedicated to the mechanisms of estrogenic synergism and antagonism with these compounds should prove fruitful.
The environmental relevance of greater than additive responses was noted by the concentration-dependent induction of Vtg in fish treated with pond water that had recently undergone triclopyr treatment. Chemical evaluation of the water indicated no detectable triclopyr, and 4-NP concentrations that were at or below laboratory-derived NOEC and LOEC values. EEQ calculations indicated estrogenicity that was similar to EEQs derived from laboratory treatments of triclopyr and TPA mixtures, which were greater than the additive responses of the individual compounds. These data indicate the mixture of triclopyr and TPA may be responsible for the estrogenicity in this sample. However, caution should be used, as water was not evaluated prior to pesticide application, and other compounds, such as natural phytoestrogens, may be present.
In summary, 2,4-D and the APE-containing surfactants R-11 and TPA were estrogenic to rainbow trout at environmentally relevant concentrations. Greater than additive responses were observed in the laboratory when 2,4-D or triclopyr were combined with the surfactant TPA. Response curves differed between the pesticides, with the 2,4-D + TPA mixture displaying an inverse dose response. The estrogenic response of triclopyr and TPA was greater than additivity at mid-range concentrations and diminished at the highest concentration. Estrogenic activity was observed in pond water treated with triclopyr and TPA that was similar to laboratory values with the combined compounds. These data suggest caution should be utilized when using NOEC and LOEC values to assess estrogenic activity for individual compounds, and that utilization of additive responses are likely inappropriate for endocrine-mediated endpoints.
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ACKNOWLEDGMENTS |
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