ToxSci Advance Access originally published online on January 10, 2006
Toxicological Sciences 2006 90(2):349-361; doi:10.1093/toxsci/kfj082
| ||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||
Comparison of In Vitro and In Vivo Estrogenic Activity of UV Filters in Fish
,1
* University of Applied Sciences Basel, Institute of Environmental Technology, St. Jakobs-Strasse 84, CH-4132 Muttenz, Switzerland; University of Zürich, Institute of Plant Biology, Limnology, Seestrasse 987, CH-8802 Kilchberg, Switzerland; Springborn Smithers Laboratories (Europe) AG, Seestrasse 21, CH-9326 Horn, Switzerland; and Swiss Federal Institute of Technology (ETH), Department of Environmental Sciences, CH-8092 Zürich, Switzerland
1 To whom correspondence should be addressed at University of Applied Sciences Basel, Institute of Environmental Technology, St. Jakobs-Strasse 84, CH-4132 Muttenz, Switzerland. Fax: +41-61–467–42–90. E-mail: karl.fent{at}bluewin.ch.
Received July 28, 2005; accepted December 8, 2005
 |
ABSTRACT |
|---|
 TOP ABSTRACT INTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
In this work, we evaluate whether in vitro systems are good predictors for in vivo estrogenic activity in fish. We focus on UV filters being used in sunscreens and in UV stabilization of materials. First, we determined the estrogenic activity of 23 UV filters and one UV filter metabolite employing a recombinant yeast carrying the estrogen receptor of rainbow trout (rtER
) and made comparisons with yeast carrying the human hER for receptor specificity. Benzophenone-1 (BP1), benzophenone-2 (BP2), 4,4-dihydroxybenzophenone, 4-hydroxybenzophenone, 2,4,4-trihydroxy-benzophenone, and phenylsalicylate showed full dose–response curves with maximal responses of 81–115%, whereas 3-benzylidene camphor (3BC), octylsalicylate, benzylsalicylate, benzophenone-3, and benzophenone-4 displayed lower maximal responses of 15–74%. Whereas the activity of 17ß-estradiol was lower in the rtER than the hER assay, the activities of UV filters were similar or relatively higher in rtER, indicating different relative binding activities of both ER. Subsequently, we analyzed whether the in vitro estrogenicity of eight UV filters is also displayed in vivo in fathead minnows by the induction potential of vitellogenin after 14 days of aqueous exposure. Of the three active compounds in vivo, 3BC induced vitellogenin at lower concentrations (435 µg/l) than BP1 (4919 µg/l) and BP2 (8783 µg/l). The study shows, for the first time, estrogenic activities of UV filters in fish both in vitro and in vivo. Thus we propose that receptor-based assays should be used for in vitro screening prior to in vivo testing, leading to environmental risk assessments based on combined, complementary, and appropriate species-related assays for hormonal activity.
Key Words: UV filters; Pimephales promelas; vitellogenin; in vitro-in vivo comparison; fish estrogen receptor ; human estrogen receptor .
|
INTRODUCTION |
|---|
|
TOPABSTRACTINTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
Numerous studies have focused on compounds that are agonists for estrogen receptors
and ß (ER, ERß) (Routledge and Sumpter, 1997
; Schultz et al., 2000; Sohoni and Sumpter, 1998; Soto et al., 1991). These include a wide range of different compounds having different molecular structures and binding affinities that may induce effects in wildlife (Jobling et al., 1998; Vos et al., 2000). With regard to the considerable number of chemicals that are and will be assessed for possible estrogenic activity, in vitro systems play an important role for identification and first screening of estrogenic compounds that interact with the ER. Within the framework on the Endocrine Disrupters Testing and Assessment (EDTA) of the OECD, animal and non-animal tests are proposed (OECD, 2002). For assessing possible ecotoxicological effects, a fish in vivo screening assay is planned, but no in vitro assays using fish-based systems have been proposed (OECD, 2004). Considering the vast number of compounds to be tested for aquatic systems, it is important to employ appropriate in vitro systems for fish (Ackermann et al., 2002; Le Guével and Pakdel, 2001; Pakdel et al., 2000). In this regard, the question arises as to what extent in vitro systems can mimic in vivo activity of estrogenic compounds. In vitro systems are cost-effective tools and allow for rapid screening of a large number of compounds, but they have limitations, which may result in unreliable predictions. Therefore, the most appropriate way to determine the endocrine-disrupting activity of chemicals seems to include both in vitro and in vivo assays, as no single assay may be best suited to determine the hormonal activity of a compound. In our present work, we address this question, focusing on important chemicals in personal-care products found in the aquatic environment. Sunscreens and cosmetics including lipsticks, skin lotions, hair sprays, hair dyes, shampoos, and numerous other products contain increasing amounts of compounds protecting from ultraviolet (UV) radiation. Either organic UV filters, or inorganic micropigments (ZnO, TiO2) scattering and reflecting UV light, or combinations of both, are applied. Increased sunlight-protection factors are being used for preventing negative effects on the human skin, which generally requires higher percentages of UV filters in the products. Combinations of different UV filters are increasingly employed for absorbing UVA, UVB, and UVC light.
Inputs of UV filters into the aquatic system occur directly via recreational activities (bathing) into surface water, and indirectly via wastewater. Ultraviolet filters are photostable, many of them highly lipophilic (log Kow 3–7) and relatively stable in the aquatic environment (Balmer et al., 2005; Poiger et al., 2004), which makes these compounds critical for bioaccumulation. Residues of several UV filters have been detected in human milk (Hany and Nagel 1995) and in fish (Balmer et al., 2005; Nagtegaal et al., 1997), in the latter between 21–3100 ng/g lipid, and also in lakes and wastewater, with maximum concentrations up to 125 ng/l (Poiger et al., 2004) and 19 µg/l (Balmer et al., 2005), respectively.
At present, the estrogenicity of UV filters in fish remains elusive, and the ecotoxicological risk for aquatic life is not known. Estrogenic activity in vitro has been shown for some UV filters in MCF-7 cells (Schlumpf et al., 2001), recombinant cell lines (Mueller et al., 2003; Schreurs et al., 2002), and recombinant yeast systems carrying the human ER (Kunz and Fent, unpublished; Routledge and Sumpter, 1997; Schultz et al., 2000). Estrogenic activity has also been observed experimentally in vivo in rats (Durrer et al., 2005; Schlumpf et al., 2001; Seidlová-Wuttke et al., 2004). In fish, high concentrations of 3-benzylidene camphor, 4-methyl-benzylidene camphor, and octyl-methoxycinnamate (Holbech et al., 2002; Inui et al., 2003) were found to be estrogenic after short-term exposure. Contrary to these studies, no estrogenicity was observed at 10 µM octyl-methoxycinnamate, benzophenone-3, homosalate, octyl dimethyl-p-aminobenzoic acid, butyl methoxydibenzoylmethane, and 1 µM 4-methyl-benzylidene camphor after short-term exposure of transgenic zebrafish (Schreurs et al., 2002). Therefore the estrogenic activity of UV filters at low aqueous concentrations remains unclear.
The objectives of this study were to elucidate whether commonly used UV filters are estrogenic in vitro and in vivo in fish, to compare the in vitro activities in two in vitro systems carrying either a fish or the human ER, and to compare the in vitro activity with the in vivo activity. We test the hypothesis that the estrogenic activity of chemicals is best assessed by the use of a tiered approach using a combination of in vitro and in vivo assays of the same phyla. As the rtER has a different activity toward known estrogenic compounds than the hER (Le Guével and Pakdel, 2001; Pakdel et al., 2000; Petit et al., 1995), the question arises whether fish-based in vitro systems should be used for assessing estrogenicity in fish. Direct comparison of fish in vitro and in vivo activity demonstrates that the estrogenic activity in vivo may be partially predictable from the in vitro activity, although in vitro screening tends to overestimate the number of estrogenic compounds due to lack of or low metabolism. This indicates the need for a tiered approach, combining in vitro and in vivo assessments of hormonal activity of UV filters for ecological risk assessment.
|
EXPERIMENTAL SECTION |
|---|
|
TOPABSTRACTINTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
Chemicals
17ß-Estradiol (E2) was purchased from Fluka AG (Buchs, Switzerland). Ultraviolet filters (Table 1) were obtained as follows. Benzophenone-1 (BP1), benzophenone-2 (BP2), benzophenone-3 (BP3), benzophenone-4 (BP4), 4'-hydroxybenzophenone (4HB), 4,4'-dihydroxybenzophenone (4DHB), 2,4,4'-trihydroxybenzophenone (THB), 4-aminobenzoic acid (PABA), benzylsalicylate (BS), phenlysalicylate (PS), octyl salicylate (OS), octocrylene (OC), and octyl dimethyl PABA (OD-PABA) were from Fluka AG; octyl-methoxycinnamate (OMC), 3-(4'-methylbenzylidene-camphor) (4MBC), 3-benzylidene-camphor (3BC), and homosalate (HMS) were from Merck (Glattbrugg, Switzerland). Ethoxylated ethyl-4-aminobenzoate (PEG-25 PABA), a polymer consisting of ethyl 4-aminobenzoate and oxirane, was purchased from Induchem (Volketswil, Switzerland), and isopentyl-4-methoxycinnamate (IMC) was from Haarmann & Reimer (Holzminden, Germany). Bisimidazylate (BIM) was purchased form T. H. Geyer (Friedrichsthal, Germany). 4-tert-Butyl-4'-methoxydibenzoylmethane (BM-DBM) and 2-phenyl-5-benzimidazole-sulfonic-acid (PBS) were purchased from Aldrich (Fluka AG, Buchs Switzerland). 2,2-Methylenbis-phenol (ECL) was purchased from Ciba Speciality Chemicals (Basel, Switzerland), and Uvinul A plus B (UAB), a mixture of 35% 2-(4-diethylamino-2-hydroxybenzoyl)-benzoic acid hexylester and 65% OMC, was a gift from BASF AG (Wädenswil, Switzerland). All compounds were >99% pure. Stock solutions were made in ethanol and stored in the dark at 4°C. Analytical grade ethanol (EtOH, free of UV filters) was purchased from T. J. Baker (Stehelin AG, Basel, Switzerland). Bidestilled water was produced using a Jencons Autostill double D-ionstill destillator (Renggli AG, Rotkreuz, Switzerland).
|
Experiments In Vitro in Yeast
Recombinant yeast assay expressing the rainbow trout estrogen receptor alpha (rtERa assay).
We investigated estrogenic activity of UV filters in vitro by applying a quantitiative ß-galactosidase assay in liquid culture of recombinant yeast expressing the estrogen receptor of rainbow trout (rtER
) that was kindly provided by F. Pakdel, University of Rennes. We slightly modified the previously described assay procedure (Le Guével and Pakdel, 2001; Petit et al., 1995). In general, the assay is based on transactivation of rtER and induction of ß-galactosidase leading to a color change. The estrogen-inducible expression system used is described in detail in (Le Guével and Pakdel, 2001). In brief, the yeast (Saccharomyces cerevisiae) genome carries a stably integrated DNA sequence of the rainbow trout estrogen receptor (rtER). Yeast cells also contain expression plasmids carrying two estrogen-responsive elements (ERE) upstream of the yeast proximal iso-1-cytochrome c gene promoter fused to the lacZ gene (encoding the enzyme ß-galactosidase). Thus, the induction is strictly dependent on the presence of rtER and estrogens (Petit et al., 1995). When an active ligand (i.e., 17ß-estradiol or an estrogenic UV filter) binds to the receptor, ß-galactosidase is synthesized and secreted into the medium, leading to a catalytic hydrolysis of o-nitrophenyl-ß-D-galactopyranosid (ONPG) and resulting in the development of a yellow color, which was measured as absorbance at 405 nm.
Preparation of rtER assay media and yeast growth.
The assay media were prepared as previously published (Le Guével and Pakdel, 2001; Petit et al., 1995) and described (F. Pakdel, personal communication), with the following amendments: Complete Minimal Dropout Medium (CM) was prepared by adding 2% D-glucose instead of 1%. In addition to the CM medium, we used YPD growth medium (2% peptone enzymatic digest from meat, 2% D-glucose, and 1% yeast extract). Thus prior to the assay, yeast cell growth was calculated as described (Petit et al., 1995), but with the modification that yeast colonies from CM medium were inoculated in Erlenmeyer flasks containing 15 ml of YPD growth medium instead of CM medium, which lead to increasing growth rates and better assay performance.
rtERa assay procedure.
The whole assay was performed as described in detail elsewhere (Le Guével and Pakdel, 2001; Petit et al., 1995), but instead of hemolysis tubes in clear polystyrene, 96-well microplates (Greiner Bio-One, Huber AG, Basel, Switzerland) were used, leading to small modifications of the assay procedure according to Schultis and Metzger (2004). Thereby the centrifugation step after cell lysis was excluded and the lysed suspension, instead of the supernatant alone, was transferred to the flat-bottom 96-well plate. The protein measure (Petit et al. 1995) was replaced by quantifying yeast turbidity (A620), to assess and correct for yeast growth and as a control for cytotoxicity. Cytotoxicity was manifested by significantly reduced yeast growth or even cell lysis, and it was determined by absorbance at 620 nm. High concentrations of some UV filters that lead to cytotoxicity were omitted from curve fitting and calculations.
For screening of the UV filters, a 96-well V-bottomed microtiter plate was filled with 100 µl/well S. cerevisiae cells in YPD culture. Three rows contained serially diluted positive control E2, one row the ethanol blank, and four rows the analyzed UV filter in quadruplicates with increasing concentrations, resulting in dose–response curves. After cell lyses the lysed suspension was transferred to a new flat-bottom 96-well plate (Greiner Bio-One, Huber AG, Basel, Switzerland), ONPG was added, and the estrogenic activity was measured as previously described (Le Guével and Pakdel, 2001; Petit et al., 1995; Schultis and Metzger, 2004).
The hER recombinant yeast was kindly provided by J. Sumpter, Brunel University, and the assay was performed according to Routledge and Sumpter (1996) and Kunz and Fent (unpublished). The yeast (S. cerevisiae) genome carries a stably integrated DNA sequence of the human estrogen receptor (hER), and it also contains expression plasmids carrying EREs, regulating the expression of the reporter gene lacZ (encoding the enzyme ß-galactosidase). Thus, when an active ligand (i.e., E2 or an estrogenic UV filter) binds to the receptor, ß-galactosidase is synthesized and secreted into the medium, leading to a color change of chromogenic substrate chlorophenol red ß-D-galactopyranoside (CPRG) from yellow to red.
Experiments In Vivo in Fish
Fish.
The 14-days fish experiments were conducted using juvenile, sexually undifferentiated fathead minnows (Pimephales promelas), between 2 and 3 months of age and with a total body length between 19 and 27 mm. This fish species has been chosen because of its frequent use in the field of endocrine disrupters and established techniques including vitellogenin (VTG) antibodies. The experimental procedure and duration was similar to that of Panter et al. (2002), who showed that estrogens and antiestrogens are detectable after 14 days of exposure by virtue of the VTG response.
Mixed-sex juvenile fathead minnows were received from the cultivator (Aquatic Research Organisms, Hampton NH, USA) and adapted for a minimum of 14 days in our laboratory in aquaria prior to the experiment. Fish were fed with Tetramin pellets (Tetra GmbH, Melle, Germany) twice a day with a quantity equivalent to 1% of body weight prior to the onset of experiments. During the experiments, fish were fed with brine shrimp (Artemia salina, Argent Chemical Labs, Redmond WA, USA) at a feeding rate of 1% of body weight twice a day.
Exposure.
Fish were held in well-aerated reconstituted tap water medium (total hardness 160 mg/l as CaCO3, total alkalinity 30 mg/l as CaCO3, conductivity 500 µs/cm) and a 16/8 h light/dark cycle at 25° ± 1°C. The studies were conducted using a 24-h static-renewal procedure with daily renewal of total aquaria water. For exposure, 10 randomly selected fish were each placed in stainless-steel tanks (10 liter) and exposed to individual UV filters for 14 days. Not all UV filters evaluated in vitro could be analyzed in vivo. To have a reasonable number of in vivo experiments, UV filters were selected as follows: either because they exhibited maximal estrogenic activities in our in vitro assays (BP1, BP2, 4DHB) or because they possessed submaximal (BP3, BP4, 3BC) or no (4MBC, OMC) estrogenic activity in our in vitro assays but were reported to be estrogenic by other studies, and because of their frequent use.
The first experiment was performed with 4MBC, 3BC, BP1, and BP2, and the second experiment was carried out with BP3, BP4, OMC, and 4DHB. In both experiments two controls, solvent control (SC, 1 ml ethanol in 10 liters of water) and positive control for estrogenic activity (100 ng/l E2), were included. Stock solutions of each chemical were prepared freshly in ethanol prior to the start of the experiment and added daily to the experimental water by mixing. The following nominal concentrations of UV filters were used: 10, 100, 500, 1000, and 5000 µg/l for BP1, BP3, BP4, OMC, and 4DHB, respectively; 10, 100, 500, and 1000 µg/l for 4MBC and 3BC, respectively; and 10, 100, 500, 1000, and 10,000 µg/l for BP2.
The concentrations were selected on the basis of environmental residues and included higher levels in order to span a large concentration range. Toxic side effects (i.e., lethargy, uncoordinated swimming, loss of equilibrium, hyperventilation) were observed for fish exposed to 5000 µg/l BP3 and 1000 µg/l 4MBC, and the experiments were stopped at day 8 of exposure.
Physicochemical measurements and biological observations.
Physicochemical parameters were determined daily. pH and oxygen saturation ranged between 7.2–7.9 and 6.5–8.3 mg/l, respectively, throughout the exposure period. Mortalities and abnormal behavior were recorded daily, and dead fish were removed from the tanks as soon as they were identified. On day 14 all fish were anesthesized with buffered tricaine methane sulfonate (MS-222, 100 mg/l with 200 mg NaHCO3/l). Subsequently individual fish were measured, weighted, transferred into labeled Eppendorf tubes, frozen, and stored at –20°C for homogenization and VTG analysis.
Vitellogenin analysis.
Fish were defrosted at 4°C and individually homogenized in ice-cold assay buffer (Biosense, Bergen, Norway) in a 1:2 ratio wet weight:buffer volume, using a Ultra Turax homogenizer (IKA, Huber + Co. AG, Reinach, Switzerland). The homogenates were centrifuged at 10,000 x g for 3 min at room temperature using a microcentrifuge (Eppendorf centrifuge 5415 D, Vaudaux-Eppendorf AG, Schönenbuch, Switzerland). The supernatant was withdrawn and immediately used for vitellogenin (VTG) analysis or frozen at –80°C until required for VTG analysis. Whole-body homogenates were assayed for VTG using a quantitative heterologous carp enzyme-linked immunosorbent assay, which has been shown to be highly reliable for VTG determination in the fathead minnow (Panter et al., 2002; Tyler et al., 1999). The commercially available quantitative carp vitellogenin ELISA kit, which is based on a sandwich ELISA format (Biosense), was used for determination of VTG in whole-body homogenates of individual fish and was conducted as described by Biosense. Purified carp VTG from blood plasma (Biosense) was used as a standard for quantitation according to the provider's description.
Analytical chemistry.
For the duration of the experiment, four aliquots of 250 ml exposure waters were taken from the two highest and the two lowest concentrations of each UV filter and controls at the beginning (0 h) and prior to water renewal (24 h). The aliquots of the same concentration of UV filter were pooled for each UV filter at each concentration and time point in brown glass flasks, preserved by acidification using HCl to pH 2–3, and stored at 4°C until analysis. Chemical analyses of UV filter concentrations were carried out by high performance liquid chromatography (HPLC) and UV detection (Kunz et al. unpublished). Briefly, 25 or 250 ml of water samples, depending on sample concentration, were extracted and concentrated by solid phase extraction (SPE). The 2500x concentrated eluent was then analyzed by HPLC-DAD.
Data Processing and Statistical Analysis
Recombinant yeast assay.
The absorbance measurement at 405 nm (ONPG) and 620 nm (turbidity) for the rtER assay allowed for subsequent correction for turbidity (yeast growth) as follows:
 |
For all UV filters, the maximal response relative to the standard (=100%) were calculated. Thereby the height of the UV filter dose–response curve was expressed as a percentage of the maximal effect produced by the dose–response curve of E2.
High concentrations of some UV filters that inhibited growth of the yeast, or even lysed cells were omitted from curve fitting and calculations. For curve-fitting and EC50 calculations (GraphPad Software Inc., San Diego, CA, USA), the corrected absorbance values versus the logarithm of concentration were plotted, whereby the best fit from a number of nonlinear regression models was selected for final data analysis. In this study, we used the Hill equation (or sigmoidal dose–response with variable slope) to fit full dose–response curves, which reached the same height (
80% maximal response) as the corresponding standard E2. Moderate (30–80% maximal response) and submaximal (<30% maximal response) dose–response curves were fitted using the best fit from a number of non-linear regression models. Coefficient of determination (R2), residuals and 95% confidence intervals were calculated, and the runs test was carried out to verify that the fitted curve represents data correctly. Estrogenic potencies were calculated for all active UV filters. Thereby the EC50 of UV filters with full dose–response curves was divided by the EC50 of the E2 standard. For UV filters with submaximal dose–response curves, estrogenic potencies were estimated based on their EC50 values, despite differences in curve steepness and height when compared to the standards. In this way, a good approximation of the estrogenic potencies of submaximal UV filters was achieved.
Fish experiment.
After testing the data distribution for normality by using the Kolmogorov-Smirnov test, means of wet weight and total length of individual fish were calculated, and data were analyzed by analysis of variance (ANOVA) followed by a Dunnett's Multiple Comparison test to compare the treatment means with respective controls. Means of VTG concentrations of individual fish were calculated, and data were analyzed with the non-parametric Kruskal-Wallis test, followed by a Dunn's Multiple Comparison test to compare the treatment means with respective controls. Statistical comparisons with the control were made using the SC as the overall control. The results are given as mean ± standard error of mean (SEM). Differences were considered significant at p 
0.05. All computations were performed with PRISM 4.0 (GraphPad Software Inc.).
|
RESULTS |
|---|
|
TOPABSTRACTINTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
Estrogenic Activity of UV Filters In Vitro
Ten of 23 analyzed UV filters and the UV filter metabolite 4HB were found to possess estrogenic activity in the recombinant yeast assay expressing the rainbow trout ER
. BP1, BP2, 4DHB, THB, 4HB, and PS were full rtER agonists exhibiting full dose–response curves. They had maximal responses of 81–123% as compared to E2 (Fig. 1). Moderate, but clear dose–response curves were found for BS, BP3, and BP4, characterized by lower maximal responses of 43–74% (Fig. 1). Submaximal dose–response curves were observed with 3BC and OS with maximal responses of 27% and 15%, respectively (Fig. 1). Table 2 shows the relative potencies of the UV filters compared to E2, as determined by their half-maximal induction activities (EC50). The most potent UV filter was BP1, which was only 87 times less potent then E2. Estrogenicity decreased in the following order 4HB > 3BC > BS > BP2 > BP3 > THB > PS > 4DHB > BP4 > OS. The estrogenic potencies were in the range of 390 to 24,750 times lower than E2. The remaining 12 UV filters, namely 4MBC, BIM, BM-DBM, ECL, HMS, IMC, OC, PABA, OD-PABA, PEG25-PABA, OMC, PBS, and UAB were inactive up to 2.5 x 10–2 M. In every assay we checked for potential cytotoxicity caused by the UV filter by routinely measuring yeast growth (620 nm) besides ß-galactosidase activity (405 nm). Hence UV filter concentrations, which lead to reduced yeast cell growth or complete growth inhibition, were omitted from data analysis for hormonal activities. At high concentrations slight cytotoxicity occurred for BS (5.04 x 10–5 M), BP4 (10.0 x 10–4 M), 4DHB (1.00 x 10–3 M) and PS (1.00 x 10–4 M). Hormonal activity was therefore assessed at non-cytotoxic concentrations only.
|
|
Comparison with hER
.The same UV filters found active in our present study with recombinant yeast expressing the rainbow trout ER
were previously found active in a recombinant yeast system expressing the human estrogen receptor alpha (hER). With this system, we investigated 17 UV filters and one metabolite for their multiple hormonal activities such as estrogenicity, antiestrogenicity, androgenicity, and antiandrogenicity in vitro (Kunz and Fent, unpublished). The only exception was OS that exhibited minimal estrogenic activity in the rtER only. The activities of BP1, 3BC, and the salicylates were relatively higher with rtER than with hER, but they were lower for the remaining benzophenones, displaying lower EC50 values with hER. BP1 and 4HB showed strongest activities in both receptor systems. The rankings of the other UV filters differed between the receptor systems, however. In the hER assay benzophenone derivatives were the most potent compounds. Benzophenone-, camphor-, and salicylate derivatives were most potent in the rtER assay. The maximal responses of estrogenic compounds in the rtER assay were in most cases higher than in the hER, with only BP2 and 4DHB as exceptions (Table 2). A direct comparison of the two assays is shown in Figure 2. BP1 as the most potent UV filter in both assays displayed an estrogenic activity only 87 times less than that of E2 with rtER and 5000 times less with hER. In particular, the relative activity of 3BC was higher in the rtER assay. In contrast to the relatively higher activity of UV filters in the rtER assay, the hER assay was 62 times more sensitive toward E2.
|
Estrogenic Activity of UV Filters in Fish In Vivo
Measured exposure concentrations.
To determine actual effect concentrations and to get an estimate of concentration decrease, concentrations of UV filters in aquaria waters were measured at the beginning of exposures (0 h) and 24 h later, prior to water renewal at the lowest and the two highest exposure concentrations. Concentrations decreased during exposure, but to a variable extent for different compounds. Table 3 shows that actual concentrations determined by HPLC analysis were close to nominal. After 24 h before water renewal, concentrations decreased to various degrees (0–32%) depending on compound and concentration. The different concentration decreases are a result of different physicochemical properties of UV filters (lipophilicity) and uptake by fish.
|
Effects of UV filters on fish survival, weight, and length.
No mortality was observed in control, solvent control (SC), and positive control (E2) exposed fish in either experiment. The UV filters did not affect survival during exposure, except at the highest concentrations of 4MBC, 3BC, and BP1. After 8 days of exposure, two fish died at 753 µg/l 4MBC (survival 80%), and the experiment was stopped. At 953 µg/l 3BC and 4919 µg/l BP1, one fish each was found dead at day 12 and day 11, respectively. Fish at 8783 µg/l BP2 showed some signs of edema at the end of exposure.
During the 14-day exposures, all control, SC, and E2 fish grew as determined by increase in wet weight and total body length (Table 4). At low concentrations of UV filters no significant differences from controls were observed. Length gain was significantly decreased for 435 and 953 µg/l 3BC and 4MBC (Table 4), however. No difference occurred in wet weight and mean length in the SC and E2. 3BC and 4MBC led to a dose-related decrease in the weight gain and body length at 435 and 953 µg/l, and 415 and 753 µg/l, respectively. No decreases in body weight gain and body length were observed with all other UV filters at all exposure concentrations, except for 4919 µg/l BP1 after 14 days of exposure.
|
Estrogenicity of UV filters.
Significant VTG induction occurred in fish exposed to ng/l E2. Mean whole-body VTG content was 2600 µg/ml and was highly induced compared to the water and solvent control having a residual level of 0.3 µg/ml. Dose-dependent increases in VTG were observed in fish exposed to 3BC, BP1, and BP2 (Fig. 3). 3BC showed higher VTG induction and at lower concentration compared to BP1 and BP2. Dose-related significant VTG induction occurred at 3BC concentrations of 435 µg/l (407 µg VTG/ml) and 953 µg/l (1753 µg VTG/ml). Concentration-related VTG induction was also found after exposures to higher concentrations of BP1 and BP2. Although increased at medium concentrations, VTG induction was significant only at the highest concentrations of BP1 and BP2, namely at 4919 µg/l BP1 (907 µg VTG/ml) and 8783 µg/l BP2 (1504 µg VTG/ml). The UV filters BP3, BP4, and 4DHB did not result in a significant VTG induction at all exposure concentrations in fish, although they showed submaximal estrogenic activity in vitro. 4MBC and OMC, which showed no estrogenicity in vitro, were not estrogenic in vivo. Therefore, three of five UV filters that exhibited estrogenic activity in vitro were also estrogenic in vivo.
|
|
DISCUSSION |
|---|
|
TOPABSTRACTINTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
In the present study we show that it is most appropriate to determine the endocrine-disrupting activity of chemicals both in vitro and in vivo, preferably in a tiered approach, as no single assay may be best suited to determine the hormonal activity of a compound and because of species differences. In this way, we demonstrate for the first time that as many as 10 of 23 commonly used UV filters are estrogenic in an in vitro yeast assay carrying a fish ER (rtER). Compared to our results in the recombinant yeast carrying the hER
, where we investigated 17 UV filters and one metabolite for estrogenicity, antiestrogenicity, androgenicity, and antiandrogenicity in vitro (Kunz and Fent, unpublished), we found that all compounds, except OS, were equally estrogenic in both assays, despite lower activity of E2 in the rtER assay. In fish we demonstrated that three of eight UV filters were estrogenic in vivo. Comparing in vitro activities in two systems with fish in vivo activity, we found the rtER in vitro data more accurate than the hER data for prediction of the in vivo activity. Hence estrogenic activity of chemicals is best assessed by the use of a tiered approach with a combination of in vitro and in vivo assays of the same species.
In Vitro Activity in the rtER Assay
We found 10 of 23 compounds exhibiting estrogenic activities in the rtER assay with different maximal responses and dose–response curves. The range of moderate to full dose–response curves can be explained by the molecular structures. Ultraviolet filters displaying full dose–response curves are characterized by at least one ring-substituted hydroxyl group. They display lower maximal responses with increasing molecular symmetry. Additional substituents on the phenolic ring have a diminishing effect on the maximal estrogenic responses in both ER systems (Kunz and Fent, unpublished; Routledge and Sumpter, 1997). This is the case for BP3, BP4, BS, OS, and 3BC, which are substituted with methoxy-, sulfonic acid-, benzyl, octyl, or camphor groups, possibly indicating only partial agonism. Like those compounds with very large molecular structures that prevent uptake into cells, inactive UV filters had, with very few exceptions, only one non-hydroxylated ring that was connected and/or attached to other substituents such as ethoxy, alkyl-, amino-, cyano-, or methoxy-groups, which were shown to significantly decrease the chemical's affinity for the rtER, as previously shown for hER (Blair et al., 2000). The structural basis of the estrogenic UV filters found in our study is in line with recent findings on structure–activity relationships of structurally similar chemicals in hER-systems (Miller et al., 2001; Routledge and Sumpter, 1997; Schultz et al., 2000).
Comparison Between Rainbow Trout and Human ER
We were interested in elucidating whether the structural differences between human and rainbow trout ER (Petit et al., 1995, 2000) were responsible for functional differences. Homologies in amino acid sequences between hER and rtER are variable, depending on domain (Pakdel et al., 1990). The most highly conserved region is the C domain (92% homology), which is responsible for DNA binding and dimerization (Petit et al., 2000). Whereas rtER and hER have similar binding affinities to an estrogen response element, the rtER C domain is responsible for a weaker DNA binding stability. The E domain shares 60% similarity with rtER and hER and contains the hormone-binding domain. Petit et al. (1995) found that the rtER has a lower affinity for E2 than the hER. This was further demonstrated for 17ß-estradiol, estrone, and 17ß-ethinylestradiol (Le Guével and Pakdel, 2001) and is confirmed by the present study; the E2 concentration necessary to induce 50% activity was 62 times higher in the rtER assay. The weaker magnitude of E2 stimulation mediated by rtER is attributed to the lower DNA-binding stability and not to structural differences between the two ER (Petit et al., 2000).
As for most of the UV filters, relative sensitivities of rtER and hER systems varied very little (one order of magnitude), indicating that the main difference between the two receptors is their sensitivity for E2 (Table 2). Differences between the two ER occurred for the salicylate derivatives (PS, BS, OS), which showed several times higher activity in rtER, and for 3BC, which showed as much as 1300 times higher activity than in the hER. Maximal responses were generally higher in the rtER assay, except for BP2 and 4DHB. This indicates a higher relative sensitivity and weaker partial agonism in the rtER of compounds showing submaximal activity.
Thus, in contrast to the lower activity of E2, the activity of some UV filters is relatively higher in the rtER assay. This cannot be fully explained by a lower DNA-binding stability found by Petit et al. (2000) for estradiol, but is rather attributed to structural differences of the two ER and the molecular structure of the UV filters interacting with the ER. This points toward a slightly different substrate binding specificity of the fish and human ER based on differences in the binding domain of the two receptors. In addition, differences in the transactivation process such as dimerization and DNA-binding capacity may also account in part for the different relative sensitivities. Forthcoming studies focusing on UV filter receptor binding and influence on the transactivation process will elucidate the reasons for the differing sensitivities of the fish and hER.
Comparison with Other In Vitro Studies
The estrogenicity of UV filters found in our study with rtER is consistent with results obtained in vitro in human ER systems, although relative sensitivities may differ. Estrogenicity of some salicylate and camphor derivatives have been reported in mammalian systems such as recombinant yeast (Kunz and Fent, unpublished; Miller et al., 2001; Mueller et al., 2003), receptor binding assays (Blair et al. 2000; Mueller et al., 2003; Schlumpf et al., 2004), proliferation of MCF-7 cells (Schlumpf et al., 2001, 2004), and reporter gene induction in transfected cell lines (Schreurs et al., 2002; Suzuki et al., 2005; Yamasaki et al., 2003). The estrogenicity of BP1, BP2, BP3, and 3BC found in our rtER assay is consistent with findings in MCF-7 cells (Schlumpf et al., 2001, 2004), reporter hER/HeLa cells (Yamasaki et al., 2003), MCF7 reporter cells (Suzuki et al., 2005), and HEK293 cells (Schreurs et al., 2005). Our results are also consistent with the hER cell assay for BP3, but not for 4MBC in the HEK 293 reporter gene assay. In addition, 3BC, HMS, and 4MBC showed activity in the hER assay (Schreurs et al., 2002).
In fish, at least two ER subtypes, ER and ERß, occur, and in zebrafish, a third form has been reported (Menuet et al., 2002). At present it is not known to what extent UV filters interact with these receptors. Reasons for differences between results obtained in our study with the rtER assay and other in vitro assays are, first, that UV filters may be active toward the ERß, but not the ER. In the MCF-7 and other human cells, active UV filters may interfere with both hERs. 4MBC was estrogenic in the MCF-7 cells (Schlumpf et al., 2001), but it did not exhibit estrogenic activity toward the rtER in the present study, a finding similar to that previously reported from our experiments with the hER (Kunz and Fent, unpublished), based on the fact that 4MBC binds preferably to the ERß (Schlumpf et al., 2004). This is also the case for HMS (Schreurs et al., 2002). In addition, yeast has only a low capability for metabolism; therefore metabolites of UV filters binding to the ER are not identified by the rtER assay. The differences may also depend in part on different binding activities of the ERs of different species and on differences between in vitro assays and their varying capabilities to activate a chemical metabolically. This leads to the conclusion that species differences in the estrogenic activity occur and that one in vitro assay alone is not sufficient for assessing the estrogenicity of chemicals and fully characterizing their estrogenic potential. Moreover homologous in vitro systems are more reliable for predicting in vivo activity.
Comparison of In Vivo Activities
Our in vivo experiments demonstrate that of eight analyzed UV filters, three—3BC, BP1, and BP2—showed estrogenic activity in fathead minnows. 3BC led to dose-dependent induction of VTG at lower concentrations (435, 953 µg/l) than BP1 (4919 µg/l) and BP2 (8783 µg/l). 4MBC, OMC, BP3, BP4, and 4DHB did not induce VTG up to the highest concentrations in the range between 753 µg/l (4MBC) and 5010 µg/l (4DHB). Schreurs et al. (2002) observed no estrogenicity in transgenic zebrafish exposed for 96 h at 10 µM of OMC (2.90 mg/l), OD-PABA (2.77 mg/l), HMS (2.62 mg/l), BP3 (2.28 mg/l), and 1 µM of 4MBC (0.25 mg/l), which is consistent with our data with these UV filters analyzed at similar concentrations but for a longer period of time. Injection of high concentrations of 3BC (27, 68, 137 mg/kg and higher) induced VTG in rainbow trout (Holbech et al., 2002). In medaka, estrogenic activity of 4MBC and OMC was observed only at about 200 times higher concentrations (Inui et al., 2003), but it was not found at 20 times higher concentrations than in our study. This may be related to species-specific differences in VTG induction. The relative degree of VTG induction is species-specific, as shown for rainbow trout that reacted with higher VTG induction to endocrine disrupters than roach (Routledge et al., 1998).
The UV filters 4MBC, BP3, 4DHB, and OMC exhibited estrogenicity in vivo in rats (Mueller et al., 2003; Schlumpf et al., 2001, 2004; Yamasaki et al., 2003), but this was not observed in our fish study. The differences can be explained by species differences in metabolism and different affinities to the ERs, as 4MBC and HMS preferably bind to the ERß (Schlumpf et al., 2004; Schreurs et al., 2002). Whether this is the case in fish is not known. Most likely, the differences are based on the different metabolic capabilities of fish compared to rats, but also on lower exposure concentrations. In our fish experiments, UV filter levels in water were lower than in rats exposed to UV filters via feed.
Comparison of In Vitro and In Vivo Activity
The estrogenic activity in vitro was matched in vivo for most UV filters. 4MBC and OMC exhibited neither estrogenic activity in in vitro transactivation assays carrying either the hER or rtER nor in fish in vivo. 3BC, BP1, and BP2 showing activity in vitro were also active in vivo. Both in vitro and in vivo, they possessed the highest potencies of the tested UV filters. The in vitro activity of BP1 (EC50 7.9 x 10–7 M) was higher than that of 3BC (1.2 x 10–5 M), whereas BP2 was the least active of these three compounds, both in vitro and in vivo (Figs. 1–3
). The in vivo activity of 3BC was higher than expected from its in vitro potency in the rtER assay, where it was the second most potent UV filter after BP1. Being only 87 times less active than E2 in the rtER assay, BP1 showed only weak in vivo activity in fathead minnows. This might be explained by its higher metabolism and lower lipophilicity (and lower bioaccumulation potential) compared to the more lipophilic 3BC. Furthermore, the relatively higher estrogenicity of 3BC in vivo might be based on the higher binding activity of 3BC to the ERß than to ER of fathead minnows, as 3BC binds preferentially to human recombinant ERß, and only slightly to ER (Schlumpf et al., 2004). There are no indications that metabolites of 3BC are more active than the parent compound. On the basis of the rtER assay, the relatively low in vivo activity of BP2 is consistent with our in vitro data. This might also be the reason why 4DHB, BP3, and BP4 possess lower rtER potencies. The estrogenic activity of most benzophenones and salicylates seems to be abolished in vivo because of metabolism.
In comparisons of the potency rankings of UV filters for the rtER and the hER assay, the data clearly demonstrate that the rtER in vitro data are more accurate than the hER data in predicting the in vivo activity. This indicates that hormonal activity of UV filters should be assessed by a suite of species-related in vitro and in vivo assays in which the in vitro assay should be able to predict to most potent compounds for further in vivo testing. Differences in in vitro and in vivo activities, which we nevertheless observed in our fish-based assays, are attributable to metabolism, and also to different activities to different ERs in fish. Our approach using rtER in vitro and fathead minnow in vivo may cover species differences in fish. Perhaps using the same fish species (rainbow trout) in the in vivo assay as in the in vitro assay would have resulted in more comparable results between the in vitro and in vivo assays.
Environmental Consequences
In the environment only a few UV filters such as OC, 4MBC, BP3, and BM-DBM have been analyzed to date. In lake water, BP3, 4MBC, and OC occurred at concentrations of 80–125, 60–80, and 22–27 ng/l, respectively, in the upper layer of a bathing lake (Poiger et al. 2004), but they were lower in other lakes (Balmer et al., 2005). Concentrations in treated wastewater were 0.06–2.7 (4MBC), 0.01–0.7 (BP3), 0.01–0.1 (OMC), and 0.01–0.27 µg/l (OC) (Balmer et al., 2005). Residues of 4MBC, OMC, BP3, and HMS were also found in muscle tissue of fish from a German lake at levels between 21 and 3100 ng/g lipid (sum of all UV filters 2 µg/g in perch and 0.5 µg/g in roach), and between 25 and 166 ng/g lipid in 10 whitefish from Swiss lakes (Balmer et al., 2005). However, 3BC, BP1, and BP2, which were found in our study to be estrogenic, have not yet been analyzed in aquatic systems. If they were in the same range, VTG induction after short-term exposure to a single UV filter would probably not pose a hazard to fish. However, different UV filters may act additively (Heneweer et al., 2005), as indicated for other endocrine disrupters (Routledge et al., 1998). Moreover, long-term exposure to UV filters may affect fish reproduction at much lower concentrations.
As it is not known to what extend these UV filters occur in the environment and in fish, comprehensive hazard and risk assessment is premature. Forthcoming studies should determine environmental concentrations of estrogenic UV filters and to relate them to effect concentrations. For hazard and risk assessment, potential effects on reproduction, fecundity, and fertility in fish are necessary, as are bioaccumulation studies. Moreover, UV filters may have multiple hormonal activities such as antiestrogenicity, androgenicity, and antiandrogenicity, in addition to estrogenicity (Kunz and Fent, unpublished). Whether these multiple hormonal activities are reflected in vivo in fish, and whether reproduction effects occur, is now under investigation in our laboratory.
|
CONCLUSIONS |
|---|
|
TOPABSTRACTINTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
Considering the vast number of compounds to be tested for possible endocrine activity, it is important to employ appropriate in vitro systems. They are cost effective and allow for rapid screening of a large number of compounds, but they have limitations that may result in unreliable predictions. In the present study we show that it is most appropriate to determine the endocrine-disrupting activity of chemicals both in vitro and in vivo, as no single assay appears to be best suited to determine the hormonal activity of a compound, and because there are species differences. We propose that receptor-based assays with related or even the same species should be used for in vitro screening prior to in vivo testing. In this tiered approach, the predictive power of in vitro systems is enhanced, and cost intensive in vivo studies can be reduced by employing species-specific in vitro assays. This leads to the conclusion that an environmental risk assessment should be based on combined, complementary, and appropriate species-related in vitro and in vivo assays for hormonal activity.
|
ACKNOWLEDGMENTS |
|---|
We thank Farzad Pakdel, Université de Rennes I, Rennes, France, for providing the recombinant rtER
yeast cells, John Sumpter, Brunel University, Uxbridge, UK, for providing the recombinant hER yeast cells, Friedrich Jüttner, University of Zürich, and Hans-Rudolf Schmutz, Chemistry Department (FHBB), for support. This work was supported by the Swiss National Science Foundation (NRP50, contract 4050–066554 to K. Fent). |
REFERENCES |
|---|
|
TOPABSTRACTINTRODUCTIONEXPERIMENTAL SECTIONRESULTSDISCUSSIONCONCLUSIONSREFERENCES |
|---|
Ackermann, G. E., Schwaiger, J., Negele, R. D., and Fent, K. (2002). Effects of long-term nonylphenol exposure on gonadal development and biomarkers of estrogenicity in juvenile rainbow trout (Oncorhynchus mykiss). Aquatic Toxicol. 60, 203–221.[Medline]
Balmer, M., Buser, H. R., Müller, M. D., and Poiger, T. (2005). Occurrence of some organic UV filters in wastewater, in surface waters, and in fish from Swiss lakes. Environ. Sci. Technol. 39, 953–962.[Medline]
Blair, R. M., Fang, H., Branham, W. S., Hass, B. S., Dial, S. L., Moland, C. L., Tong, W., Shi, L., Perkins, R., and Sheehan, D. M. (2000). The estrogen receptor relative binding affinities of 188 natural and xenochemicals: Structural diversity of ligands. Toxicol. Sci. 54, 138–153.
Durrer, S., Maerkel, K., Schlumpf, M., and Lichtensteiger, W. (2005). Estrogen target gene regulation and coactivator expression in the rat uterus after developmental exposure to the ultraviolet filter 4-methylbenzylidene camphor. Endocrinology 146, 2130–2139.
Hany, J., and Nagel, R. (1995). Nachweis von UV-Filtersubstanzen in Muttermilch. aus Rheinland-Pfalz Deut. Lebensm.-Rundsch. 91, 341–345.
Heneweer, M., Musse, M., Van den Berg, J., and Sanderson, T. (2005). Additive estrogenic effects of mixtures of frequently used UV filters on pS2-gene transcription in MCF-7 cells. Toxicol. Appl. Pharmacol. 208, 170–177.[CrossRef][Web of Science][Medline]
Holbech, H., Norum, U., Korsgaard, B., and Bjerregaard, P. (2002). The chemical UV-filter 3-benzylidene camphor causes an oestrogenic effect in an in vivo fish assay. Pharmacol. Toxicol. 91, 204–208.[CrossRef][Web of Science][Medline]
Inui, M., Adachi, T., Takenaka, S., Inui, H., Nakazawa, M., Ueda, M., Watanabe, H., Mori, C., Iguchi, T., and Miyatake, K. (2003). Effect of UV-screens and preservatives on vitellogenin and choriogenin production in male medaka (Oryzias latipes). Toxicology 194, 43–50.[CrossRef][Web of Science][Medline]
Jobling, S., Nolan, M., Tyler, C. R., Brighty, G., and Sumpter, J. P. (1998). Widespread sexual disruption in wild fish. Environ. Sci. Technol. 32, 2498–2506.[CrossRef]
Le Guével, R., and Pakdel, F. (2001). Streamlined beta-galactosidase assay for analysis of recombinant yeast response to estrogens. BioTechniques 30, 1000–1004.[Medline]
Menuet, A., Pellegrini, E., Anglade, I., Blaise, O., Laudet, V., Kah, O., and Pakdel, F. (2002). Molecular characterisation of three estrogen receptor forms in zebrafish: Binding characteristics, transactivation properties, and tissue distributions. Biol. Reprod. 66, 1881–1892.
Miller, D., Wheals, B. B., Beresford, N., and Sumpter, J. P. (2001). Estrogenic activity of phenolic additives determined by an in vitro yeast bioassay. Environ. Health Persp. 109, 133–138.[Web of Science][Medline]
Mueller, S. O., Kling, M., Firzani, P. A., Mecky, A., Duranti, E., Shields-Botella, J., Delansorne, R., Borschard, T., and Kramer, P. J. (2003). Activation of estrogen receptor a and ERb by 4-methylbenzylidene-camphor in human and rat cells: Comparison with phyto- and xenoestrogens. Toxicol. Lett. 142, 89–101.[CrossRef][Web of Science][Medline]
Nagtegaal, M., Ternes, T. A., Baumann, W., and Nagel, R. (1997). UV-Filtersubstanzen in Wasser und Fischen. UWSF-Z. Umweltchem. Ökotoxikol. 9, 79–86.
OECD (2002). Detailed review paper on appraisal of test methods for sex hormone disrupting chemicals. OECD Environment Directorate, Paris.
OECD (2004). Detailed review paper on fish screening assays for the detection of endocrine active substances. OECD Environment Directorate, Paris.
Pakdel, F., Le Gac, F., Le Goff, P., and Valotaire, Y. (1990). Full-length sequence and in vitro expression of rainbow trout estrogen receptor cDNA. Mol. Cell. Endocrinol. 71, 195–204.[CrossRef][Web of Science][Medline]
Pakdel, F., Métivier, R., Flouriot, G., and Valotaire, Y. (2000). Two estrogen receptor (ER) isoforms with different estrogen dependencies are generated from the trout ER gene. Endocrinology 141, 571–580.
Panter, G. H., Hutchinson, T. H., Länge, R., Lye, C. M., Sumpter, J. P., Zerulla, M., and Tyler, C. R. (2002). Utility of a juvenile fathead minnow screening assay for detecting (anti-)estrogenic substances. Environ. Toxicol. Chem. 21, 319–326.[CrossRef][Web of Science][Medline]
Petit, F., Valotaire, Y., and Pakdel, F. (1995). Differential functional activities of rainbow trout and human estrogen receptor expressed in the yeast Saccharomyces cerevisiae. Eur. J. Biochem. 233, 584–592.[Web of Science][Medline]
Petit, F. G., Valotaire, Y., and Pakdel, F. (2000). The analysis of chimeric human/rainbow trout estrogen receptors reveals amino acid residues outside of P- and D-boxes important for the transactivation function. Nucleic Acids Res. 28, 2634–2642.
Poiger, T., Buser, H. R., Balmer, M., Bergqvist, P. A., and Müller, M. D. (2004). Occurrence of UV filter compounds from sunscreens in surface waters: Regional mass balance in two Swiss lakes. Chemosphere 55, 951–963.[Medline]
Routledge, E. J., Sheahan, D., Desbrow, C., Brighty, G. C., Waldock, M., and Sumpter, J. P. (1998). Identification of estrogenic chemicals in STW effluent. 2. In vivo responses in trout and roach. Environ. Sci. Technol. 32, 1559–1565.[CrossRef]
Routledge, E. J., and Sumpter, J. P. (1996). Estrogenic activity of surfactants and some of their degradation products assessed using a recombinant yeast screen. Environ. Toxicol. Chem. 15, 241–248.[CrossRef]
Routledge, E. J., and Sumpter, J. P. (1997). Structural features of alkylphenolic chemicals associated with estrogenic activity. J. Biol. Chem. 272, 3280–3288.
Schlumpf, M., Cotton, B., Conscience, M., Haller, V., Steinmann, B., and Lichtensteiger, W. (2001). In vitro and in vivo estrogenicity of UV screens. Environ. Health Perspect. 109, 239–244.[Web of Science][Medline]
Schlumpf, M., Schmid, P., Durrer, S., Conscience, M., Maerkel, K., Henseler, M., Gruetter, M., Herzog, I., Reolon, S., Ceccatelli, R. (2004). Endocrine acitivity and developmental toxicity of cosmetic UV filters—an update. Toxicology 205, 113–122.[CrossRef][Web of Science][Medline]
Schreurs, R. H., Lanser, P., Seinen, W., and Van der Burg, B. (2002). Estrogenic activity of UV filters determined by an in vitro reporter gene assay and in vivo transgenic zebrafish assay. Arch. Toxicol. 76, 257–261.[CrossRef][Web of Science][Medline]
Schreurs, R. H. M. M., Sonneveld, W., Jansen, J. H. J., Seinen, W., and Van der Burg, B. (2005). Interaction of polycyclic musks and UV filters with the estrogen receptor (ER), androgen receptor (AR), and progesterone receptor (PR) in reporter gene bioassays. Toxicol. Sci. 83, 264–272.
Schultis, T., and Metzger, J. W. (2004). Determination of estrogenic activity by LYES-assay (yeast estrogen screen-assay assisted by enzymatic digestion with lyticase). Chemosphere 57, 1739–1745.[Medline]
Schultz, T. W., Seward, J. R., and Sinks, G. D. (2000). Estrogenicity of benzophenones evaluated with a recombinant yeast assay: Comparison of experimental and rules-based predicted activity. Environ. Toxicol. Chem. 19, 301–304.[CrossRef]
Seidlová-Wuttke, D., Jarry, H., and Wuttke, W. (2004). Pure estrogenic effect of benzophenone-2 (BP2) but not of bisphenol A (BPA) and dibutylphtalate (DBP) in uterus, vagina, and bone. Toxicology 205, 103–112.[Medline]
Sohoni, P., and Sumpter, J. P. (1998). Several environmental oestrogens are also anti-androgens. J. Endocrinol. 158, 327–339.[Abstract]
Soto, A. M., Justicia, H., Wray, J. W., and Sonnenschein, C. (1991). p-Nonyl-phenol: An estrogenic xenobiotic released from "modified" polystyrene. Environ. Health Persp. 92, 167–173.[Web of Science][Medline]
Suzuki, T., Kitamura, S., Khota, R., Sugihara, K., Fujimoto, N., and Ohta, S. (2005). Estrogenic and antiandrogenic activities of 17 benzophenone derivatives used as UV stabilizers and sunscreens. Toxicol. Appl. Pharmacol. 203, 9–17.[CrossRef][Medline]
Tyler, C. R., Van Aerle, R., Hutchinson, T. H., Maddix, S., and Trip, H. (1999). An in vivo testing system for endocrine disrupting in fish early life stages using induction of vitellogenin. Environ. Toxicol. Chem. 18, 337–347.[CrossRef]
Vos, J. G., Dybing, E., Greim, H. A., Ladefoged, O., Lambré, C., Tarazona, J. V., Brandt, I., and Vethaak, A. D. (2000). Health effects on endocrine-disrupting chemicals on wildlife, with special reference to the European situation. Crit. Rev. Toxicol. 30, 71–133.[CrossRef][Web of Science][Medline]
Yamasaki, K., Takeyoshi, M., Yakabe, Y., Sawaki, M., and Takatsuki, M. (2003). Camparison of the reporter gene assay for ER-alpha antagonists with the immature rat uterotrophic assay of 10 chemicals. Toxicol. Lett. 142, 119–131.[Medline]
CiteULike 
Connotea 
Del.icio.us What's this?
This article has been cited by other articles:
|
|
 |
|
  P. Y. Kunz, T. Gries, and K. Fent The Ultraviolet Filter 3-Benzylidene Camphor Adversely Affects Reproduction in Fathead Minnow (Pimephales promelas) Toxicol. Sci., October 1, 2006; 93(2): 311 - 321. [Abstract] [Full Text] [PDF]
|
|
| ||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||