Differing Species Responsiveness of Estrogenic Contaminants in Fish

Article pubs.acs.org/est

Differing Species Responsiveness of Estrogenic Contaminants in Fish Is Conferred by the Ligand Binding Domain of the Estrogen Receptor Shinichi Miyagawa,†,◆ Anke Lange,‡,◆ Ikumi Hirakawa,†,§,∥ Saki Tohyama,†,⊥ Yukiko Ogino,† Takeshi Mizutani,† Yoshihiro Kagami,§ Teruhiko Kusano,§ Masaru Ihara,# Hiroaki Tanaka,# Norihisa Tatarazako,∇ Yasuhiko Ohta,∥ Yoshinao Katsu,○ Charles R. Tyler,*,‡ and Taisen Iguchi*,† †

Okazaki Institute for Integrative Bioscience, National Institute for Basic Biology, National Institutes of Natural Sciences, and Department of Basic Biology, The Graduate School for Advanced Studies (SOKENDAI), Okazaki, Aichi 444-8787, Japan ‡ Biosciences, College of Life and Environmental Sciences, University of Exeter, Stocker Road, Exeter, EX4 4QD, U.K. § Ecogenomics, Inc. Kurume, Fukuoka 839-0864, Japan ∥ The United Graduate School of Veterinary Science, Yamaguchi University, and Department of Veterinary Medicine, Faculty of Agriculture, Tottori University, Koyama, Tottori 680-8553, Japan ⊥ Graduate School of Nutritional and Environmental Sciences, University of Shizuoka, Shizuoka 422-8526, Japan # Research Center for Environmental Quality Management, Kyoto University, Ohtsu, Shiga 520-0811, Japan ∇ Environmental Quality Measurement Section, Research Center for Environmental Risk, National Institute for Environmental Studies, 16-2 Onogawa, Tsukuba, Ibaraki 305-8506, Japan ○ Department of Biological Sciences, Hokkaido University, Sapporo, Hokkaido 060-0810, Japan S Supporting Information *

ABSTRACT: Exposure to estrogenic endocrine disrupting chemicals (EDCs) induces a range of adverse effects, notably on reproduction and reproductive development. These responses are mediated via estrogen receptors (ERs). Different species of fish may show differences in their responsiveness to environmental estrogens but there is very limited understanding on the underlying mechanisms accounting for these differences. We used custom developed in vitro ERα reporter gene assays for nine fish species to analyze the ligand- and species-specificity for 12 environmental estrogens. Transcriptonal activities mediated by estradiol-17β (E2) were similar to only a 3-fold difference in ERα sensitivity between species. Diethylstilbestrol was the most potent estrogen (∼10-fold that of E2) in transactivating the fish ERαs, whereas equilin was about 1 order of magnitude less potent in all species compared to E2. Responses of the different fish ERαs to weaker environmental estrogens varied, and for some considerably. Medaka, stickleback, bluegill and guppy showed higher sensitivities to nonylphenol, octylphenol, bisphenol A and the DDT-metabolites compared with cyprinid ERαs. Triclosan had little or no transactivation of the fish ERαs. By constructing ERα chimeras in which the AF-containing domains were swapped between various fish species with contrasting responsiveness and subsequent exposure to different environmental estrogens. Our in vitro data indicate that the LBD plays a significant role in accounting for ligand sensitivity of ERα in different species. The differences seen in responsiveness to different estrogenic chemicals between species indicate environmental risk assessment for estrogens cannot necessarily be predicted for all fish by simply examining receptor activation for a few model fish species.

■

INTRODUCTION Steroid hormones, including estrogen, play important roles in the reproductive biology of vertebrates and their biological effects are principally mediated through steroid receptors, ligand-regulated transcription factors. Steroid receptor proteins are divided into six functionally independent domains termed A to F extending from the amino to carboxyl terminus1 and their functions include DNA-binding, dimerization, ligand binding and transcriptional activation.2 ER transcriptional activation is known to require specific ligand-inducible activation functions © 2014 American Chemical Society

AF-1 and AF-2, located in the N-terminal (A/B) domain and ligand binding- or E-domain (LBD), respectively, and both are required for full ER activity.3,4 Two estrogen receptor (ER) subtypes (ERα and ERβ) have been cloned from a variety of vertebrate species. Teleost fish Received: Revised: Accepted: Published: 5254

January 16, 2014 March 28, 2014 April 1, 2014 April 1, 2014 dx.doi.org/10.1021/es5002659 | Environ. Sci. Technol. 2014, 48, 5254−5263

Environmental Science & Technology

Article

functional domains might account for ligand sensitivity of ERα in different species. As models for this work, we chose medaka and two cyprinid fish, carp and roach, which were found to have relatively highly estrogen responsive (sensitive, medaka) and relatively poorly responsive (insensitive, carp, and roach) ERαs.

express three distinct types of ER (ERα, ERβ1, and ERβ2) and ERβ2 appears to be closely related to ERβ1, reflecting a gene duplication event.5 Although functions of each ER subtype in fish are not well established, ERα subtype is indispensable for fertility in mammals.6 ERs can recruit coactivators in the presence of estrogens and xenoestrogens, and then activate gene transcription associated with alterations in the tertiary structure following the ligand binding. There is global concern about the presence of endocrine disrupting chemicals (EDCs) in the environment and their ability to interfere with normal hormone signaling pathways. A wide range of EDCs have been identified that include a variety of industrial and agricultural compounds, notably alkylphenols, pesticides, plasticizers, and bisphenols. They enter the aquatic environment where wildlife is especially at risk of exposure from discharges emanating from wastewater treatment works (WwTWs) and/or agricultural runoff. Exposure to EDCs is known to induce a range of effects in fish including delayed onset of sexual maturation, reduced gonadal growth, gonadal deformations, inhibition of spermatogenesis, reduced sperm counts, lowered egg production, skewed sex ratios, and increased prevalence of intersex.7,8 A considerable body of research has been focused on compounds that induce feminized responses that include elevated concentrations of blood vitellogenin in male and immature female fish, induction of a female-like ovarian cavity in the testis of males, and intersex in wild fish. These responses have been associated with the proximity of these fish to point source discharges of estrogenic WwTW effluents and plausible links drawn between the feminized responses and some of the chemicals contained in these WwTW effluents including steroid estrogens and various industrial chemicals, such as the alkylphenols, 4-nonylphenol (NP), and 4-tert-octylphenol (4tOP). Feminization in fish has furthermore been induced experimentally by exposure to some of these compounds and also to certain pesticides such as the organochlorine pesticide dichlorodiphenyltrichloroethane (DDT).9−13 An analysis of the literature would indicate that different species of fish show differences in their responsiveness and thus potentially their susceptibility to the effects of environmental estrogens.14,15 Understanding of the mechanisms of action of EDCs can likely help account for why these differences in responsiveness between species occur. Here, we used custom developed in vitro ERα reporter gene assays for nine test- and wildlife fish species including, (bluegill (Lepomis macrochirus), common carp (Cyprinus carpio), fathead minnow (Pimephales promelas), goldfish (Carassius auratus), guppy (Poecilia reticulata), medaka (Oryzias latipes), roach (Rutilus rutilus), three-spined stickleback (Gasterosteus aculeatus), and zebrafish (Danio rerio) to analyze the ligand- and species-specificity for 12 environmental estrogens. The test compounds included bisphenol A (BPA), the alkylphenols NP and 4tOP; triclosan, a polychlorinated phenoxy phenol used as a antimicrobial agent; equilin, an equine estrogen used in hormone replacement therapies in the U.S., UK, and Japan; diethylstilbestrol (DES), a synthetic estrogen; and DDT, that is still used as an insecticide in tropical regions of Africa, South Eastern Asia, and South America. In the reporter assay, ligand-binding is assumed to be the most likely step that attributes to differential sensitivities among fish species, although there is little direct evidence to support this. By constructing chimeras in which the AFcontaining domains (A/B or LBD) were swapped between species, we further investigated whether differences in these

■

MATERIALS AND METHODS Animals and Reagents. Medaka, common carp, goldfish and guppy were purchased from a local commercial supplier. Zebrafish were kindly provided from Prof. Nagahama (Ehime University, Matsuyama, Japan) and fathead minnows were raised at the University of Exeter (UK). Three-spined sticklebacks were collected in Hokkaido (Japan) and bluegills were collected at Lake Biwa (Shiga Prefecture, Japan). 17βEstradiol (E2), diethylstilbestrol (DES) and equilin were obtained from Sigma (St. Louis, MO); bisphenol A (BPA), 4-nonylphenol (NP), 4-tert-octylphenol (4tOP), triclosan, DDT and its metabolites (purities >97.0% for o,p′-DDE; >98.5% for p,p′-DDE; >99.0% for p,p′-DDD; >99.5% for o,p′DDT, p,p′-DDT and o,p′-DDD) were from Kanto-Kagaku (Tokyo, Japan). Construction and Cloning of Estrogen Receptors. The full-coding regions of bluegill ERα (EF191017), carp ERα (AB334722), fathead minnow ERα (AY775183), goldfish ERα (AY055725), guppy ERα (AB621910), medaka ERα (AB033491), roach ERα (AB190289), stickleback ERα (AB330740) and zebrafish ERα (AB037185) were amplified by PCR and subcloned into the mammalian expression vector pcDNA3.1 (Life Technologies, Carlsbad, CA). For the cloning of guppy ERα, RNA isolated from the ovary were reverse transcribed into cDNA which served as template for PCR using degenerate oligonucleotides as described previously.15 Phylogenetic Tree of ERs. The deduced amino acid sequences of DBD and LBD including hinge region (D domain) were aligned using the Clustal X program. Alignments with questionable gaps were removed. A neighbor-joining tree was constructed from this alignment using a 1000 replicate bootstrap analysis. FigTree version1.3.1 was used to draw and view the rooted neighbor-joining tree. ER sequences were obtained from the Genebank database. The accession numbers of the sequences used in the tree are provided in the Supporting Information (SI). Construction of Chimera ERs. Chimera ERs in which the LBD or A/B domain of different species were swapped were constructed using the In-Fusion cloning kit according to the manufacturer’s protocols (Takara, Ohtsu, Japan). The sequences of primers used to create the chimeras are provided in Table S1 (SI) and the following ERα chimeras were created: m(carp LBD)ERα (medaka ERα containing carp LBD), m(roach LBD)ERα (medaka ERα containing roach LBD), c(medaka LBD)ERα (carp ERα containing medaka LBD), r(medaka LBD)ERα (roach ERα containing medaka LBD), m(carp A/B)ERα (medaka ERα containing carp A/B domain), m(roach A/B)ERα (medaka ERα containing roach A/B domain), c(medaka A/B)ERα (carp ERα containing medaka A/B domain) and r(medaka A/B)ERα (roach ERα containing medaka A/B domain). The numbers of amino acid residues corresponding to each domain are numbered in Figure S1 (SI). The obtained clones were sequenced to verify that no additional nucleotides were introduced during subcloning. Transactivation Assay. HEK293 cells were used as the host cells for the transactivation assay and the assay was reported previously15 and details are provided in the SI. The 5255

dx.doi.org/10.1021/es5002659 | Environ. Sci. Technol. 2014, 48, 5254−5263

EC50 (μM) 95% CI (μM)b RP (%)c





p,p′-DDT

5256

o,p′-DDD

o,p′-DDE

o,p′-DDT

2.068 1.223−3.498 0.013

2.893 1.103−7.585 0.010

3.106 1.343−7.184 0.009

2.395 0.947−6.056 0.012

2.712 1.078−6.826 0.010

2.76 × 10−4 (1.66−4.60)x10−4 100


E2

p,p′-DDE

0.325 0.160−0.661 0.087


OP


2.65 × 10−4 (2.10−3.35)x10−4 100

0.301 0.156−0.579 0.095


NP

p,p′-DDD

0.121 0.086−0.169 0.206

1.895 1.311−2.739 0.015


BPA

0.177 0.117−0.269 0.150

0.7011 0.424−1.159 0.038

0.885 0.622−1.258 0.030

1.334 0.699−2.545 0.020

1.68 1.024−2.755 0.016

2.042 1.315−3.172 0.013

0.101 0.077−0.132 0.247

0.425 0.336−0.537 0.059

bluegill 2.49 × 10−4 (1.83−3.37)x10−4 100

goldfish

2.84 × 10−4 (1.52−5.31)x10−4 100


E2

0.0656 0.036−0.120 0.247

0.4289 0.335−0.549 0.038

0.426 0.289−0.627 0.038

0.7258 0.580−0.908 0.022

1.551 1.122−2.146 0.010

2.248 1.353−3.734 0.007

1.62 × 10−4 (0.97−2.70)x10−4 100

0.043 0.025−0.072 0.342

0.040 0.018−0.090 0.361

0.121 0.090−0.162 0.120

1.45 × 10−4 (0.71−2.96)x10−4 100

guppy

1.362 0.996−1.860 0.012

2.745 1.509−4.993 0.006

2.929 1.265−6.783 0.006

3.076 1.510−6.264 0.005

2.957 1.230−7.105 0.006

2.368 1.278−4.389 0.007

1.66 × 10−4 (0.85−3.22)x10−4 100

0.488 0.340−0.703 0.036

0.369 0.208−0.654 0.047

0.750 0.567−0.992 0.023

1.74 × 10−4 (1.12−2.70)x10−4 100

roach

2.993 1.428−6.276 0.012

3.086 1.528−6.231 0.012

3.137 1.339−7.350 0.011

2.791 1.184−6.577 0.013

3.098 0.370−25.93 0.012

2.037 0.116−35.890 0.018

3.58 × 10−4 (2.20−5.81)x10−4 100

0.991 0.706−1.391 0.036

1.211 0.818−1.794 0.029

2.641 1.540−4.528 0.013

3.55 × 10−4 (2.26−5.58)x10−4 100

carp

0.3269 0.237−0.451 0.102

0.7238 0.554−0.945 0.046

1.146 0.876−1.500 0.029

2.613 1.550−4.403 0.013

3.06 1.532−6.11 0.011

3.085 1.615−5.893 0.011

3.32 × 10−4 (2.22−4.96)x10−4 100

0.066 0.036−0.123 0.552

0.074 0.039−0.137 0.497

0.309 0.226−0.424 0.118

3.66 × 10−4 (2.40−5.57)x10−4 100

stickleback

zebrafish

2.216 1.427−3.441 0.007

2.906 1.556−5.426 0.005

2.833 1.563−5.136 0.005

2.646 1.427−4.903 0.006

2.406 1.272−4.552 0.006

2.664 1.222−5.807 0.006

1.48 × 10−4 (1.20−1.82)x10−4 100

0.293 0.185−0.464 0.041

0.533 0.364−0.782 0.022

0.664 0.485−0.910 0.018

1.20 × 10−4 (0.81−1.78)x10−4 100

Table 1. Gene Transcriptional Activities for 12 Environmental Estrogenic EDCs Mediated by ERαs from Nine Fish Speciesa medaka

0.1086 0.078−0.151 0.156

0.492 0.371−0.652 0.034

0.715 0.488−1.047 0.024

0.9749 0.774−1.228 0.017

1.752 1.231−2.494 0.010

2.359 1.457−3.820 0.007

1.70 × 10−4 (1.29−2.23)x10−4 100

0.037 0.023−0.059 0.465

0.036 0.022−0.060 0.471

0.226 0.172−0.297 0.075

1.70 × 10−4 (1.23−2.35)x10−4 100

fathead minnow

1.659 1.171−2.350 0.014

2.439 1.480−4.020 0.010

2.700 1.206−6.047 0.009

2.676 1.533−4.671 0.009

2.841 1.489−5.418 0.008

2.288 1.107−4.732 0.010

2.36 × 10−4 (1.50−3.72)x10−4 100

0.341 0.266−0.436 0.066

0.219 0.131−0.364 0.103

0.762 0.597−0.973 0.030

2.25 × 10−4 (1.50−3.39)x10−4 100

Environmental Science & Technology Article


1.54 × 10−4 (1.37−1.74)x10−4 249.125 1.98 × 10−2 (1.71−2.29)x10−2 1.945

5.83 × 10−5 (0.24−1.40)x10−4 529.180

6.07 × 10−3 (4.54−8.12)x10−3 5.077




DES

EQU

TCS

2.673 1.625−4.398 0.007

2.65 × 10−3 (1.79−3.93)x10−3 6.925

2.17 × 10−5 (1.20−3.92)x10−5 845.870

1.83 × 10−4 (1.02−3.30)x10−4 100

guppy

2.94 × 10−3 (1.90−4.55)x10−3 7.929

2.70 × 10−5 (1.50−4.87)x10−5 864.024

2.33 × 10−4 (1.27−4.29)x10−4 100

roach

1.00 × 10−2 (0.85−1.18)x10−2 3.534

4.90 × 10−5 (2.69−8.93)x10−5 723.639

3.55 × 10−4 (2.31−5.46)x10−4 100

carp

3.16 0.7844−12.73 0.011

6.75 × 10−3 (5.57−8.18)x10−3 5.247

2.35 × 10−5 (1.76−3.13)x10−5 1508.735

3.54 × 10−4 (2.27−5.53)x10−4 100

stickleback

2.612 1.302−5.241 0.007

3.29 × 10−3 (2.36−4.60)x10−3 5.634

4.85 × 10−5 (3.17−7.42)x10−5 382.759

1.86 × 10−4 (1.18−2.91)x10−4 100

zebrafish

8.02 × 10−3 (0.58−1.12)x10−2 2.350

7.70 × 10−6 (0.52−1.15)x10−5 2447.734

1.89 × 10−4 (1.45−4.51 × 10−4 100

medaka

4.04 × 10−3 (3.15−5.18 × 10−3 6.155

4.40 × 10−5 (2.63−7.34)x10−5 566.098

2.49 × 10−4 (1.44−4.30)x10−4 100

fathead minnow

a

EC50s and 95% confidence intervals for each compound were derived from dose-response curves fitted on data normalized between 0 and 100%, where zero and one hundred were defined as the smallest and largest value in each data set, respectively. For each species, potencies are presented relative to the response to E2. Data are presented as mean ± SEM from three independent assays each consisting of three technical replicates per concentration tested. b95% CI: 95% confidence interval of EC50. cRP: relative potency = (EC50 E2/EC50 chemical X) × 100.

2.984 1.862−4.783 0.013

3.84 × 10−4 (2.58−5.72)x10−4 100

3.08 × 10−4 (1.49−6.36)x10−4 100


bluegill

goldfish

E2

Table 1. continued

Environmental Science & Technology Article

5257



Article

Figure 1. EC50s of teleost ERα transactivation in relation to the evolutionary relationship between species. The x-axis represents molar EC50 concentrations.

however, all induced the transactivation of all fish ERαs examined (Table 1 and Figure S4, SI). At 10−7 M, ERαs from medaka, stickleback, bluegill and guppy showed higher sensitivities to these three chemicals compared with other fish ERαs. Carp ERα was the least sensitive to NP and 4tOP compared with other fish ERαs and the differences in EC50 between medaka and carp ERα were 30-fold for NP and 23-fold for 4tOP. BPA did not activate carp ERα transactivation at 10−6 M. Collectively, these data show cyprinids, such as carp and roach, were less sensitive to these estrogenic EDCs compared with the other fish species tested. Triclosan induced luciferase activity in bluegill, guppy, medaka, and stickleback only at μM concentrations whereas cyprinid ERαs were not activated by triclosan at any of the tested concentrations (Table 1). Transactivation of Fish ERα Exposed to DDT and Its Metabolites. With the exception of p,p′-DDD, the guppy ERα was the most sensitive to the DDT metabolites tested. Carp ERα required the highest concentrations of o,p′-DDT, the o,p′metabolites and p,p′-DDE to induce a response (Table 1 and Figure S4, SI). Of the o,p′-metabolites, o,p′-DDT was more effective in activating the ER-transcriptional activity compared with o,p′DDD and o,p′-DDE and bluegill, guppy, medaka, and stickleback ERαs were more sensitive to the o,p′-metabolites compared with the other fish ERαs examined. In these species, the differences in EC50s between these two compounds and o,p′-DDT were at least 2.2-fold (o,p′-DDE; stickleback) and 3.5-fold (o,p′-DDT; stickleback). Similarly for the o,p′-metabolites, medaka, stickleback, bluegill, and guppy ERαs were more sensitive to the p,p′forms than other fish ERαs. In these species again, p,p′-DDT was a stronger activator of fish ERα compared to p,p′-DDD, but overall, the differences in EC50s for the different p,p′metabolites was much lower compared to the o,p′-forms. The maximum fold-difference in responses between p,p′-DDD and p,p′-DDT was 3.1-fold (guppy). Species and Phylogeny Differences in ERα Response to Estrogenic Chemicals. The results for the transactivation of ERαs from the nine different teleost species indicate some interesting findings when considering the evolutionary relationships between these species (Figure 1). For all species, the EC50

rationale for use of a mammalian cell line to compare the ER transactivation between fish species was to avoid complications that might arise in using a fish cell line due to endogenous factors. ERs interact with endogenous factors (cofactors and general transcriptional factors) and this is likely to occur with different affinities depending on how related the species are for the ER studied and the fish host cell system. Use of a cell line derived from a more closely related species, therefore, might introduce a bias in transcriptional activities and make any comparison of the data more difficult. Data Analysis. Results are presented as mean ± SEM from three separate experiments each consisting of three technical replicates per concentration tested. Data were normalized for responses of the different ERs to the individual chemicals, where zero and one hundred were defined as the smallest and largest responses, respectively, for each data set. Using GraphPad Prism (Graph Pad Software, San Diego, CA), dose−response data were subsequently analyzed by fitting three parametric nonlinear regression (slope = 1) curves onto the normalized data and EC50 values were calculated from these curves. The relative potencies (RP) of individual test compounds were calculated for each species as ((EC50 of E2/ EC50 of the test compound) × 100).

■

RESULTS Transactivation of Fish ERα Exposed to E2, Equilin and DES. The differences in transcriptional activities mediated by E2 were small between the species with only a 3-fold difference in ERα sensitivity between the most (zebrafish, EC50 = 1.20 × 10−10 M) and the least (carp, EC50 = 3.55 × 10−10 M) sensitive species. DES and equilin both induced the transactivation of all fish ERαs examined (Figure S4, SI). The most sensitive species for responses to DES was the medaka and the least sensitive species for DES was bluegill. For equilin, the most sensitive species were guppy and roach and again bluegill was the least sensitive species. Transactivation of Fish ERα Exposed to Alkylphenols, Bisphenol A, and Triclosan. The EC50 values calculated indicated species differences in the sensitivity of ERα to BPA, NP, and 4tOP between the nine fish species examined (Table 1). Compared with E2, they were all relatively weak estrogens for inducing transactivation of ERα. BPA, NP and 4tOP, 5258



Article

Figure 2. Construction of ERα chimeras and transcriptional activities of wild type ERαs and chimeras where the LBD or A/B domains have been substituted for different species. (A) Schematic presentation of medaka, carp and roach wild type ERαs and chimera ERαs consisting of wild type ERαs with either the LBD (Ai) or the A/B domain (Aii) has been substituted by the respective domain of a different species. (B) Transactivation of wild type and LBD chimera ERαs by 17ß-estradiol (E2; Bi) or nonylphenol (NP; Bii). (C) Transactivation of wild type and A/B-domain chimera by E2 (Ci) NP (Cii). Dose−response curves were fitted on data normalized between 0 and 100%, where zero and one hundred were defined as the smallest and largest value in each data set, respectively. Data are presented as mean ± SEM from three independent assays each consisting of three technical replicates per concentration tested.

Investigating Molecular Mechanisms of Species Differences in ERα SensitivityTransactivation of ERα Chimeras. When comparing the transactivation of medaka, carp and roach wild type ERαs induced by the natural steroid E2, differences in the EC50s were between 1.4-fold (medaka compared to roach) and 2.1-fold (carp compared to roach). For BPA, these differences in wild type ERα sensitivities were higher compared to E2 (3.4-fold, medaka compared to roach; 1.7-fold, roach compared to carp and 5.6-fold, medaka compared to carp). Equally for NP, differences in wild type ERα sensitivities were also higher compared to E2 (1.9-fold, roach compared to carp; 20-fold, medaka compared to roach and 38-fold, medaka compared to carp) Figures 2 (NP) and S6 (BPA) and Table S2, SI). Assays using chimera ERαs showed that the sensitivity of chimeras consisting of a wild type ERα of one species in which the LBD has been replaced by that of another species was much more comparable to the sensitivity of the wild type ERα of the species from which the LBD originated (Table S2, SI). As an example, the EC50 values for E2 for wild type medaka and carp

values for E2 varied across a very narrow concentration range (0.12−0.38 nM) and all were more sensitive to DES compared with E2. The differences in EC50 between E2 and DES were within 1 order of magnitude for most species (overall, the EC50 values for DES ranged between 0.008−0.154 nM). The EC50 for equilin were also comparable across all species (in the range between 10−9 and 10−8 M) and were approximately 10-fold higher when compared with E2 (for most species). Differences in sensitivities between the fish species became more apparent for weaker estrogens. While EC50 values for BPA, NP, 4tOP and the o,p′-metabolites of DDT ranged between 10−8 and 10−6 M for guppy, medaka, stickleback, and bluegill, EC50 values for the cyprinid species fathead minnow, roach, carp, goldfish, and zebrafish were around 1 order of magnitude lower. Cyprinid fish were particularly insensitive to the o,p′-metabolites of DDT. Furthermore, the order of responsiveness for all noncyprinid fish species was more or less NP≈4tOP>o,p′DDT > BPA>o,p′-DDE> o,p′-DDD, whereas this trend was less consistent for the cyprinid species. 5259



Article

ERα were 3.50 × 10−4 μM and 5.22 × 10−4 μM, respectively. When replacing the LBD in medaka ERα with the LBD of carp ERα, the EC50 for this chimera was 5.47 × 10−4 μM E2, and similar to the EC50 of the carp wild type ERα. Similarly, when replacing the LBD in the carp ERα with medaka ERα LBD, the EC50 was 1.39 × 10−4 μM E2, therefore, much closer to the medaka wild type ERα than carp ERα. This effect became even more apparent for data for BPA and NP (Figures 2 (NP) and S6 (BPA) and Table S2, SI). In contrast to the LBD, transactivation of chimera ERαs in which the A/B domain of a ERα from one species was replaced by an A/B domain of another species remained similar to the transactivation of the ERα of the species from which the domains C−F (i.e., including the LBD) were derived (Figure 2 and Table S2, SI). This was particularly apparent for NP with EC50 values of 0.06 μM, 2.45 μM and 1.29 μM for wild type medaka, carp and roach ERα, respectively. When removing the A/B domain from the medaka ERα and replacing it with the A/ B domain of carp or roach ERα, the EC50s were 0.10 μM NP and 0.02 μM NP, respectively, for those chimeras, being similar to those of the wild type medaka ERα (Table S2, SI). Similarly for BPA, when removing the A/B domain from the roach ERα and replacing it with the A/B domain of medaka ERα, an EC50 value of 1.20 μM BPA was determined for the chimera which was most similar to the roach wild type ERα (EC50 = 1.29 μM BPA) rather than the medaka ERα (EC50 = 0.06 μM BPA) (Table S2, SI). Together, these results suggest that the LBD determines the sensitivity of different fish species to the different environmental estrogenic chemicals as assessed using in vitro reporter gene assays. Comparison of the LBD. Among the teleost ERαs investigated in this study, the LBD is relatively well conserved with homologies between 81 and 98%, contrasting with only 64−66% homologies with the LBD of the human ERα. Comparison between the amino acid residues involved in the binding of E2 to the LBD of the human ERα and the corresponding residues in the teleost sequences revealed that all but two residues were conserved between species (Figure S5, SI). The residues involved in E2-binding and not fully conserved were only different in one out of the 10 sequences compared: L349 of the human ERα aligned with a methionine residue in ERα from all nine fish species and methionine in position 421 of the human ERα aligned with a methionine residue in all fish species except from the bluegill (phenylalanine). The ligand-inducible activation function AF-2 (LLLEML, see16) located in the LBD was completely conserved between human ERα and all fish ERαs.

nine different species studied and this is perhaps not surprising given the natural role of E2 and its functional conservation among vertebrates in many physiological processes (for review see ref 17). The small differences in transcriptional activity between the fish species for E2 are consistent with the high level conservancy of the amino acids which constitute the ligand-binding pocket in the LBD. Contrasting with these finding for E2, responses of the different fish ERαs to other, weaker environmental estrogens varied, and for some widely so. Alkylphenols and BPA were less potent compared with E2 (between 201- and 4545-fold, 181- and 2778-fold and 833- and 7692-fold for NP, 4tOP and BPA, respectively). The stickleback ERα was the most sensitive to NP and 4tOP, whereas zebrafish (NP) and roach and carp (4tOP) ERαs were the least sensitive species. The difference in relative potencies compared to E2 between the most and sensitive species were 22.5-fold for NP (only 4.7-fold among cyprinids) and 15.3-fold for 4tOP (only 2.4-fold among cyprinids). For BPA, guppy ERα was the most sensitive whereas carp ERα was the least sensitive with an overall difference of 9.2-fold between relative potencies between these two species. All three compounds are known to only possess very low binding affinities to ERα compared to E2 (summarized by ref 18) which most likely explains the low potencies compared to E2. DES was found to be around 10-times more potent than E2 in the activation of ERα for all the fish species studied. This difference may relate to the ER/ligand interaction kinetics. In humans, kinetic analysis of ER/ligand-interactions for the ERs has shown the dissociation rate can vary widely and for DES the dissociation occurs much more slowly than for E2 and even more so for BPA.19 This observation is consistent also with data showing that DES has a high affinity for human ERs20 and there is a positive correlation between ligand binding and receptor activation. DDT and its metabolites have been established as weak environmental estrogens21 and this is confirmed for all the fish ERs tested. The guppy ERα was the most sensitive to o,p′-DDT and its o,p′-metabolites, whereas carp ERα was the least sensitive to these compounds. The relative potency of these compounds compared with E2 was comparable among the cyprinid species, but lower than that for the other fish species tested, the guppy, bluegill, stickleback and medaka. Interestingly, the carp ERα was the most sensitive to p,p′-DDD, but the least sensitive to all other environmental estrogens. Contrasting with the o,p′-metabolites, there did not appear any clear differences in relative potencies for the p,p′-metabolites compared to E2 across species. This might be caused by the involvement of other amino acid residues in the binding of p,p′metabolites to the LBD compared with the o,p′-metabolites. Triclosan appeared to induce luciferase activity in bluegill, guppy, medaka, and stickleback, but did so only at very high μM concentrations, and possibly bordering on toxicity thresholds, and in cyprinid ERαs, there was no activation whatsoever for triclosan at any of the tested concentrations. The lack of ERα activation is consistent with findings from a study employing an in vitro cell-based ER-mediated bioassay, where triclosan at concentrations between 10−8 and 10−4 M were ineffective.22 We would conclude that although triclosan may act as a thyroid disruptor23,24 and as an antiandrogen,25 it is not an environmental estrogen, at least not acting via ERα. Accounting for the Differences in Potencies of the Environmental Estrogens between Fish Species. Overall, we found that cyprinid ERαs were less sensitive to EDCs than

■

DISCUSSION The relative importance of various estrogens in the aquatic environment and their effects on reproduction is dependent on their ability to disrupt the normal pattern of ER-mediated signaling in the body of exposed organisms. Fish species may vary in their responsiveness and thus their susceptibility to the effects of environmental estrogens but the reasons for this are little understood. We assessed the interactions of a range of estrogenic EDCs with ERαs for nine different fish species using reporter gene assays, and we provide clear evidence of differential responsiveness for different estrogenic EDCs between fish species that might be explained by ligand-binding in the ERα. Responses to Natural Steroid and Xenobiotic Estrogens. The responses of the ERαs to E2 were similar for the 5260



Article

those of other fish groups. Comparing the fish ERα sequences, there appears to be a number of amino acid residues (22 out of 235) that are conserved between stickleback, guppy, medaka, and bluegill. At these 22 positions, amino acid residues are also identical in the five cyprinid species, but differ from these of the other four fish species. These amino acid substitutions might, at least partially, explain the lower sensitivity of cyprinids to the estrogenic EDCs tested. Two conserved amino acid substitutions in the binding pocket for E2 have been shown to differ between human and rainbow trout ERα (human ERα: L349 and M528 and rainbow trout ERα M317 and I496).26 Comparing these residues with the sequences of the nine species investigated in this study, L349 of the human ERα aligned with a methionine residue in ERα in all nine fish species, which is also the case in rainbow trout (M317).26 L349 has previously been established as a residue involved in the binding of E2 to the LBD of the human ERα.27 As this amino acid is conserved between all nine teleost species, it is unlikely to account for the differences in sensitivity between cyprinid and noncyprinid species. On the other hand, while differing between the human and rainbow trout sequences,26 M528 in the human sequence is conserved across the nine species investigated in the current study. With the exception of L4349, M421 is the only other residue known to be involved in the binding of E2 to the LBD of the human ERα27 that does not align fully across the studied teleosts. Rather it aligns with a methionine residue in all fish species, with the exception of the bluegill which contains a phenylalanine in that position. Given the predominant conservation of this residue it is unlikely that this residue accounts for the differences in sensitivity between cyprinid and noncyprinid species. Of potentially greater relevance for the differences in sensitivity between cyprinid and noncyprinid species might be N348 in the human ERα LDB due to its location in-between two residues that are involved in E2 binding.27 This residue is conserved between the human ERα and the cyprinid ERαs, but different in the noncyprinid species which all contain a serine in this position. Differences in binding affinities of xenoestrogens to the ER may relate to differences in amino acids that are different from those involved in E2 binding, as for instance has been shown for musk compounds.28 The residues in the human ERα LBD previously shown to be different binding sites of, for example, musk compounds compared to E2 (i.e., E353 and H524)28 are conserved across all 10 species compared here. Overall, we identified 22 amino acids that are identical in the cyprinid species, but different from the remaining species among which they were also identical. Mutation assays for residues that differ between species would be helpful to further understanding of the differential activities in future work. Interestingly, even though stickleback are known to be relatively tolerant of contaminated environments,29 the stickleback ERα response was more sensitive to most of the compounds tested than for any of the cyprinid ERαs. The mechanisms behind these differences in relative sensitivities are less clear. The DBD is highly conserved among the fish ERαs studied and is perhaps unlikely to be the source of variation in the ER activation between the fish species for the different environmental estrogens. The LBD influences both ligand binding and cofactor interaction and several coactivators are reported to interact with the AF-2 region in the LBD of ERs.30 These interactions at the LBD, therefore, are a more likely to account for some of the differences in the ERα activation seen across the different fish species. However, given that the LBD is

still relatively highly conserved among teleosts (81−98%) and all amino acid residues of the human ERα known to interact with E2 are 100% conserved among the used test species, differences in the coactivators for ER interaction in fish may be the source of the response variation. Having said this, due to the structural differences of the xenoestrogens compared to the natural ligand E2, it is also possible that different amino acids are involved in the binding of ligands compared to E2. These differences in binding of xenoestrogens to ERα and resulting differences in receptor activation might further be associated with differences in the recruitment of coactivators that are necessary for the ligand-induced transcriptional activation. The existence of ligand-dependent differences in the ability of ERs to bind coactivator proteins in vitro has previously been shown for E2, DES, 4tOP and BPA.31 By constructing and applying chimera ERs, we were in fact able to show that the different responsiveness mainly depends on the LBD (and likely the AF-2 located within) and swapping the A/B domains (containing AF-1) did not lead to differences in ER activation. This complies with a study in which different domains of the rainbow trout ER were replaced by the corresponding domains of the human ER.32 Comparing Findings on ER Activation Across Different In Vitro Systems. Several comparative in vitro studies have been performed to assess the differences in responsiveness of ERs to estrogens among fish species, e.g. refs 15 and 33; amphibians, e.g., ref 34; and reptiles, e.g., ref 35, and to assess for differences between fish and other vertebrates.36 The assays employed, however, vary and comparisons between ER responses between different in vitro assays is complicated by the use of different cell systems, which will differ, among other things, in the availability of required cofactors, and different reporter genes. It has been shown that particularly the number of ERE repeats in the reporter gene plays a crucial role in the sensitivity of different assays.37 Matthews et al.38 reported similar responsiveness of human, mouse, chicken, green anole, Xenopus, and rainbow trout ERα to E2, but clear differences in response magnitude to xenoestrogens, just as we find for our studies. In contrast, Sumida et al.36 found no significant species differences in the transactivation in their reporter gene assay by chemicals between human, rat, chicken, alligator (Caiman), whiptail lizard and African clawed frog ER. Interestingly, both studies reported a thermo-dependent alteration in the susceptibility of rainbow trout ER to E2.36,38 Temperaturedependent ligand affinity has also been reported in vitro for blue tilapia (Oreochromis aureus) ER showing a reduced affinity of the ER to estrogens at elevated temperatures, which could be related to sequence differences in the LBD.39 A comparison of the sensitivities of human and rainbow trout ER for estrogens in the same cell systems and using the same reporter gene has also revealed temperature-dependent effects as, independent of the cell system (teleost or mammalian) used for the reporter assay, ten times higher estrogen concentrations were found to be required in rainbow trout to induce a response comparable to the human ER.37,40 Given that the different species in our study span both temperate and tropic habitats and the reporter gene assays were all performed at the same temperature which is optimal for the mammalian host cell, this might also have played a role in species-different sensitivities. We did not, however, see an obvious pattern to support this being the case (i.e., responses for temperate species were not clearly collectively separated from the responses for tropical species). 5261



Article

Author Contributions

Investigations on species differences in responsiveness to estrogens in vivo are complicated by factors relating to the duration of reproductive development, metabolism, seasonal effects, water temperatures, and also by the presence of several ER subtypes in tissues. The receptor-dependent reporter gene assay system, in contrast allows for greater control in the development for understanding on ER-ligand across different species. Here, we report the ligand- and species-differences for ERα only, but recently, stable reporter gene assays have been developed for other ER subtypes in fish and application of these assays demonstrated selective activation of ER subtypes (in zebrafish) for different chemical ligands.41 These zebrafish assays have the advantage of using stably transfected cells of a zebrafish cell line and, if recruitment of transcriptional cofactors differs between species, especially the use of species-specific cells suggests recruitment of relevant cofactors for the activation of ERs. Having said this, and even with the limitations of the reporter assays system employed in these studies, we were able to show species differences in ERα activation and evolutionary linkages in sensitivity. Together with the fact that xenoestrogens appear to have different affinities for different ER subtypes, we suggest that the use of only a few species followed by cross-species extrapolation is difficult and not enough to protect species as a whole in the risk assessment process for exposure to environmental estrogens. An analysis of the literature for effects in vivo supports the findings that different species of fish show differences in their responsiveness and thus potentially, their susceptibility to the effects of environmental estrogens.14,15 It should also be kept in mind that even though many of these environmental chemicals might only be weak ER agonists and might not appear to cause any adverse effects individually, they generally occur as complex mixtures in the aquatic environment with a higher likelihood for activation of ERs and physiological effects. In summary, we have established an effective in vitro screening tool for environmental estrogens that provides insights into their likely potencies. They further allow studies on the comparative function of ERs thus providing insights into potential risks from environmental contaminants. The differences seen in responsiveness to different estrogenic chemicals between species indicate that the risk of endocrine disruption cannot necessarily be predicted for all wildlife based on simply examining receptor activation from a few selected test species. A limited number of fish species are used for environmental risk assessment, and risk of endocrine disruption cannot necessarily be predicted for all wildlife studies by simply examining receptor activation for a few model fish species.

■

ASSOCIATED CONTENT

■

AUTHOR INFORMATION

◆

These authors contributed equally.

Notes

The authors declare no competing financial interest.

■

ACKNOWLEDGMENTS We thank Drs. T. Hirai, T. Kobayashi, M. Nagae, N. Mitsui, M. Miyahara, and Y. Ohnishi for their kind support. This work was supported by UK-Japan Research Collaboration Grants from the Ministry of the Environment (Japan) and the Department for Environment, Food and Rural Affairs (UK). Further funding was obtained from Grants-in-Aid for Scientific Research (23254003, 24370029) from the Japan Society for the Promotion of Science, Core Research for Evolutional Science and Technology (CREST) from Japan Science and Technology Agency. A.L. was supported by grants from the UK Natural Environmental Research Council (NE/D002818/1 and NE/ E016634/1) awarded to C.R.T.

■

REFERENCES

(1) Krust, A.; Green, S.; Argos, P.; Kumar, V.; Walter, P.; Bornert, J. M.; Chambon, P. The chicken estrogen-receptor sequence - homology with v-erbA and the human estrogen and glucocorticoid receptors. EMBO J. 1986, 5 (5), 891−897. (2) Klinge, C. M. Estrogen receptor interaction with estrogen response elements. Nucleic Acids Res. 2001, 29 (14), 2905−2919. (3) Kraus, W. L.; McInerney, E. M.; Katzenellenbogen, B. S. Liganddependent, transcriptionally productive association of the amino- and carboxyl-terminal regions of a steroid hormone nuclear receptor. Proc. Natl. Acad. Sci. U.S.A. 1995, 92 (26), 12314−12318. (4) Tora, L.; White, J.; Brou, C.; Tasset, D.; Webster, N.; Scheer, E.; Chambon, P. The human estrogen receptor has two independent nonacidic transcriptional activation functions. Cell 1989, 59 (3), 477− 487. (5) Hawkins, M. B.; Thornton, J. W.; Crews, D.; Skipper, J. K.; Dotte, A.; Thomas, P. Identification of a third distinct estrogen receptor and reclassification of estrogen receptors in teleosts. Proc. Natl. Acad. Sci. U.S.A. 2000, 97 (20), 10751−10756. (6) Couse, J. F.; Korach, K. S. Estrogen receptor null mice: What have we learned and where will they lead us? Endocr. Rev. 1999, 20 (3), 358−417. (7) Mills, L. J.; Chichester, C. Review of evidence: Are endocrinedisrupting chemicals in the aquatic environment impacting fish populations? Sci. Total Environ. 2005, 343 (1−3), 1−34. (8) Scholz, S.; Klüver, N. Effects of endocrine disrupters on sexual, gonadal development in fish. Sex. Dev. 2009, 3 (2−3), 136−151. (9) Donohoe, R. M.; Curtis, L. R. Estrogenic activity of chlordecone, o,p′-DDT and o,p′-DDE in juvenile rainbow trout: Induction of vitellogenesis and interaction with hepatic estrogen binding sites. Aquat. Toxicol. 1996, 36 (1−2), 31−52. (10) Jobling, S.; Sumpter, J. P.; Sheahan, D.; Osborne, J. A.; Matthiessen, P. Inhibition of testicular growth in rainbow trout (Oncorhynchus mykiss) exposed to estrogenic alkylphenolic chemicals. Environ. Toxicol. Chem. 1996, 15 (2), 194−202. (11) Lange, A.; Paull, G. C.; Coe, T. S.; Katsu, Y.; Urushitani, H.; Iguchi, T.; Tyler, C. R. Sexual reprogramming and estrogenic sensitization in wild fish exposed to ethinylestradiol. Environ. Sci. Technol. 2009, 43 (4), 1219−1225. (12) Metcalfe, C. D.; Metcalfe, T. L.; Kiparissis, Y.; Koenig, B. G.; Khan, C.; Hughes, R. J.; Croley, T. R.; March, R. E.; Potter, T. Estrogenic potency of chemicals detected in sewage treatment plant effluents as determined by in vivo assays with Japanese medaka (Oryzias latipes). Environ. Toxicol. Chem. 2001, 20 (2), 297−308. (13) Metcalfe, T. L.; Metcalfe, C. D.; Kiparissis, Y.; Niimi, A. J.; Foran, C. M.; Benson, W. H. Gonadal development and endocrine responses in Japanese medaka (Oryzias latipes) exposed to o,p′-DDT

* Supporting Information S

Supplemental experimental section, results and figures for comparison of the domain structures, sequence homologies and evolutionary relationships between fish and the human ERαs, dose−response profiles of fish ERαs and gene transcriptional activities for wild type and chimera ERαs. This material is available free of charge via the Internet at http://pubs.acs.org. Corresponding Authors

*Phone: +44 1392 724450; fax; +44 01392 724000; e-mail: c.r. [email protected]. *Phone: +81-564-59-5235; fax: +81-564-59-5236; e-mail: [email protected]. 5262



Article

in water or through maternal transfer. Environ. Toxicol. Chem. 2000, 19 (7), 1893−1900. (14) Jobling, S.; Casey, D.; Rodgers-Gray, T.; Oehlmann, J.; SchulteOehlmann, U.; Pawlowski, S.; Baunbeck, T.; Turner, A. P.; Tyler, C. R. Comparative responses of molluscs and fish to environmental estrogens and an estrogenic effluent. Aquat. Toxicol. 2004, 66 (2), 207−222. (15) Lange, A.; Katsu, Y.; Miyagawa, S.; Ogino, Y.; Urushitani, H.; Kobayashi, T.; Hirai, T.; Shears, J. A.; Nagae, M.; Yamamoto, J.; Ohnishi, Y.; Oka, T.; Tatarazako, N.; Ohta, Y.; Tyler, C. R.; Iguchi, T. Comparative responsiveness to natural and synthetic estrogens of fish species commonly used in the laboratory and field monitoring. Aquat. Toxicol. 2012, 109, 250−258. (16) Tanenbaum, D. M.; Wang, Y.; Williams, S. P.; Sigler, P. B. Crystallographic comparison of the estrogen and progesterone receptor’s ligand binding domains. Proc. Natl. Acad. Sci. U.S.A. 1998, 95 (11), 5998−6003. (17) Segner, H.; Casanova-Nakayama, A.; Kase, R.; Tyler, C. R. Impact of environmental estrogens on fish considering the diversity of estrogen signaling. Gen. Comp. Endrocrinol. 2013, 191, 190−201. (18) Dang, Z. Comparison of relative binding affinities to fish and mammalian estrogen receptors: The regulatory implications. Toxicol. Lett. 2010, 192 (3), 298−315. (19) Rich, R. L.; Hoth, L. R.; Geoghegan, K. F.; Brown, T. A.; LeMotte, P. K.; Simons, S. P.; Hensley, P.; Myszka, D. G. Kinetic analysis of estrogen receptor/ligand interactions. Proc. Natl. Acad. Sci. U.S.A. 2002, 99 (13), 8562−8567. (20) Metzger, D. A.; Curtis, S.; Korach, K. S. Diethylstilbestrol metabolites and analogs - differential ligand effects on estrogenreceptor interactions with nuclear matrix sites. Endocrinology 1991, 128 (4), 1785−1791. (21) Oien, C. W.; Hurd, C.; Vorojeikina, D. P.; Amold, S. F.; Notides, A. C. Transcriptional activation of the human estrogen receptor by DDT isomers and metabolites in yeast and MCF-7 cells. Biochem. Pharmacol. 1997, 53 (8), 1161−1172. (22) Ahn, K. C.; Zhao, B.; Chen, J.; Cherednichenko, G.; Sanmarti, E.; Denison, M. S.; Lasley, B.; Pessah, I. N.; Kultz, D.; Chang, D. P. Y.; Gee, S. J.; Hammock, B. D. In vitro biologic activities of the antimicrobials triclocarban, its analogs, and triclosan in bioassay screens: Receptor-based bioassay screens. Environ. Health Perspect 2008, 116 (9), 1203−1210. (23) Pinto, P. I. S.; Guerreiro, E. M.; Power, D. M. Triclosan interferes with the thyroid axis in the zebrafish (Danio rerio). Toxicol. Res. 2013, 2 (1), 60−69. (24) Veldhoen, N.; Skirrow, R. C.; Osachoff, H.; Wigmore, H.; Clapson, D. J.; Gunderson, M. P.; Van Aggelen, G.; Helbing, C. C. The bactericidal agent triclosan modulates thyroid hormone-associated gene expression and disrupts postembryonic anuran development. Aquat. Toxicol. 2006, 80 (3), 217−227. (25) Rostkowski, P.; Horwood, J.; Shears, J. A.; Lange, A.; Oladapo, F. O.; Besselink, H. T.; Tyler, C. R.; Hill, E. M. Bioassay-directed identification of novel antiandrogenic compounds in bile of fish exposed to wastewater effluents. Environ. Sci. Technol. 2011, 45 (24), 10660−10667. (26) Matthews, J. B.; Clemons, J. H.; Zacharewski, T. R. Reciprocal mutagenesis between human α(L349, M528) and rainbow trout (M317, I496) estrogen receptor residues demonstrates their importance in ligand binding and gene expression at different temperatures. Mol. Cell. Endocrinol. 2001, 183 (1−2), 127−139. (27) Brzozowski, A. M.; Pike, A. C. W.; Dauter, Z.; Hubbard, R. E.; Bonn, T.; Engstrom, O.; Ohman, L.; Greene, G. L.; Gustafsson, J. A.; Carlquist, M. Molecular basis of agonism and antagonism in the oestrogen receptor. Nature 1997, 389 (6652), 753−758. (28) Schreurs, R. H. M. M.; Quaedackers, M. E.; Seinen, W.; van der Burg, B. Transcriptional activation of estrogen receptor ERα and ERβ by polycyclic musks is cell type dependent. Toxicol. Appl. Pharmacol. 2002, 183 (1), 1−9. (29) Sanchez, W.; Aït-Aïssa, S.; Palluel, O.; Ditche, J.-M.; Porcher, J.M. Preliminary investigation of multi-biomarker responses in three-

spined stickleback (Gasterosteus aculeatus L.) sampled in contaminated streams. Ecotoxicology 2007, 16 (2), 279−287. (30) Danielian, P. S.; White, R.; Lees, J. A.; Parker, M. G. Identification of a conserved region required for hormone dependent transcriptional activation by steroid-hormone receptors. EMBO J. 1992, 11 (3), 1025−1033. (31) Routledge, E. J.; White, R.; Parker, M. G.; Sumpter, J. P. Differential effects of xenoestrogens on coactivator recruitment by estrogen receptor (ER) alpha and ER beta. J. Biol. Chem. 2000, 275 (46), 35986−35993. (32) Petit, F. G.; Valotaire, Y.; Pakdel, F. 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. 2000, 28 (14), 2634−2642. (33) Katsu, Y.; Kohno, S.; Hyodo, S.; Ijiri, S.; Adachi, S.; Hara, A.; Guillette, L. J.; Iguchi, T. Molecular cloning, characterization, and evolutionary analysis of estrogen receptors from phylogenetically ancient fish. Endocrinology 2008, 149 (12), 6300−6310. (34) Katsu, Y.; Taniguchi, E.; Urushitani, H.; Miyagawa, S.; Takase, M.; Kubokawa, K.; Tooi, O.; Oka, T.; Santo, N.; Myburgh, J.; Matsuno, A.; Iguchi, T. Molecular cloning and characterization of ligand- and species-specificity of amphibian estrogen receptors. Gen. Comp. Endrocrinol. 2010, 168 (2), 220−230. (35) Katsu, Y.; Matsubara, K.; Kohno, S.; Matsuda, Y.; Toriba, M.; Oka, K.; Guillette, L. J.; Ohta, Y.; Iguchi, T. Molecular cloning, characterization, and chromosome mapping of reptilian estrogen receptors. Endocrinology 2010, 151 (12), 5710−5720. (36) Sumida, K.; Ooe, N.; Saito, K.; Kaneko, H. Limited species differences in estrogen receptor alpha-medicated reporter gene transactivation by xenoestrogens. J. Steroid Biochem. Mol. Biol. 2003, 84 (1), 33−40. (37) Petit, F.; Valotaire, Y.; Pakdel, F. Differential functional activities of rainbow trout and human estrogen receptors expressed in the yeast Saccharomyces cerevisiae. Eur. J. Biochem. 1995, 233 (2), 584−592. (38) Matthews, J. B.; Fertuck, K. C.; Celius, T.; Huang, Y. W.; Fong, C. J.; Zacharewski, T. R. Ability of structurally diverse natural products and synthetic chemicals to induce gene expression mediated by estrogen receptors from various species. J. Steroid Biochem. Mol. Biol. 2002, 82 (2−3), 181−194. (39) Tan, N. S.; Frecer, V.; Lam, T. J.; Ding, J. L. Temperature dependence of estrogen binding: Importance of a subzone in the ligand binding domain of a novel piscine estrogen receptor. Biochimica et Biophysica Acta (BBA) - Molecular Cell Research 1999, 1452 (2), 103−120. (40) Le Dréan, Y.; Kern, L.; Pakdel, F.; Valotaire, Y. Rainbow trout estrogen receptor presents an equal specificity but a differential sensitivity for estrogens than human estrogen receptor. Mol. Cell. Endocrinol. 1995, 109 (1), 27−35. (41) Cosnefroy, A.; Brion, F.; Maillot-Maréchal, E.; Porcher, J.-M.; Pakdel, F.; Balaguer, P.; Aït-Aïssa, S. Selective activation of zebrafish estrogen receptor subtypes by chemicals by using stable reporter gene assay developed in a zebrafish liver cell line. Toxicol. Sci. 2012, 125 (2), 439−449.

5263


Differing Species Responsiveness of Estrogenic Contaminants in Fish

Recommend Documents