Assessment of estrogenic endocrine-disrupting chemical actions in the brain using in vivo somatic gene transfer.
In both female and male vertebrates, estrogens affect many aspects of development, growth, sexual differentiation, and reproductive behavior. Estrogens also exert positive and negative feedback effects on the hypothalamopituitary axis to regulate the secretion of gonadotropic and other pituitary hormones (Hess 2003; Korach et al. 2003; McLachlan 2001; Trudeau 1997). Estrogens, notably 17[beta]-estradiol ([E.sub.2]), are also involved in reproductive disorders such as breast and endometrial cancers (Feigelson and Henderson 1996; Graham et al. 2000). It is now recognized that there is worldwide contamination of water systems with chemicals and pharmaceuticals that mimic or inhibit estrogen action (Kolpin et al. 2002; Metcalfe et al. 2003; Ternes et al. 1999). The contraceptive steroid ethinylestradiol (EE2) and the natural hormone [E.sub.2] are among the most commonly detected hormones in surface waters and effluents from sewage treatment plants (Ternes et al. 1999). [E.sub.2] and EE2 were detected in effluents of sewage treatment plants in different countries at concentrations ranging up to 64 ng/L and 42 ng/L, respectively (Yin et al. 2002). The presence of these estrogens in Canadian sewage treatment plants has been documented with median concentrations of 9 ng/L for EE2 and 6 ng/L for [E.sub.2] (Ternes et al. 1999). A recent study of 139 U.S. rivers reported maximum concentrations of 830 ng/L (~ 2.8 nM) for EE2 and 200 ng/L (~ 0.7 nM) for [E.sub.2] (Kolpin et al. 2002).
The xenoestrogen bisphenol A (BPA) is primarily used in the production of polycarbonate and epoxy resins and is found in many plastic products, including food can linings and dental sealants. The widespread industrial and household use, economic importance, and near ubiquitous presence of BPA in the environment (Lee and Peart 2000; Staples et al. 1998) emphasize its risk as an endocrine disruptor. Concentrations of BPA in surface waters have been reported to be, in the most severe cases, as high as 17,200 [micro]g/L (~ 47 gM) in leachates from hazardous waste landfill sites (Yamamoto et al. 2001), but usually concentrations have been around or below 1 [micro]g/L (~ 2.7 nM) (Belfroid et al. 2002). However, the concentration of BPA in many polluted lakes and rivers is not known.
A host of developmental and reproductive abnormalities in many species, including humans (Guillette et al. 1995; McLachlan 2001; Tyler et al. 1998), result from exposure to estrogenic endocrine-disrupting chemicals (EDCs). For example, octylphenol, BPA, and EE2 all stimulate abnormal production of the egg yolk protein vitellogenin in male fish (Arukwe 2001; Sumpter and Jobling 1995). Moreover, BPA induced testis-ova in medaka exposed to a concentration of 10 [micro]g/L (~ 27 nM) (Metcalfe et al. 2001). Other studies showed that estrogenic EDCs cause sex reversal in frogs and feminization of secondary sex characteristics in fish (Arcand-Hoy and Benson 1998; Bogi et al. 2002; Mackenzie et al. 2003).
The diversity of structure and origin of the multitude of compounds currently known to bind to estrogen receptors (ER)-0t and ER-[beta] make it difficult to predict activities in vivo in vertebrate animals (Sanchez et al. 2002; Segner et al. 2003; Yoon et al. 2001). Large-scale screening for estrogenic activities by traditional physiologic and toxicologic methods is time-consuming and costly. A variety of effective in vitro ER binding assays and estrogen-responsive reporter systems in bacterial, yeast, and vertebrate cell systems have defined much of our understanding of estrogen and EDC actions (Ackermann et al. 2002; Matthews et al. 2002; Metivier et al. 2001, 2003; Petit et al. 1997; Zacharewski 1997). However, results derived in vitro for ER binding, hepatocyte vitellogenin induction, or ER reporter gene assays often do not always accurately reflect results obtained in vivo (Andersen et al. 1999; Segner et al. 2003). When [E.sub.2] or estrogenic mimics bind to ERs, receptor dimerization and recruitment of transcriptional comodulators are initiated, and the hormone-receptor complex binds to the estrogen response element (ERE) and subsequently regulates transcription in an ordered and cyclic manner (Metivier et al. 2003; RobinsonRechavi et al. 2003). Some of the discrepancies between in vitro assays and in vivo physiologic experiments may reflect the observations that ER[alpha] and ER[beta] differ dramatically in tissue and cellular distributions, biologic function (Abraham et al. 2004; Hess 2003; Korach et al. 2003), and their affinities for estrogenic chemicals (Le Guevel and Pakdel 2001; Yoon et al. 2001). Moreover, the likelihood that the availability of transcriptional comodulators of the ERs in vitro and in vivo is similar is highly unlikely (Graham et al. 2000), and thus, in vitro models cannot mimic the complexities of whole animal systems with respect to estrogen-dependent processes and responses to EDCs.
To begin to overcome some of the challenges of in vivo assessment of EDC modulation of gene transcription, we have validated polyethylenimine (PEI)-mediated somatic gene transfer (Lemkine and Demeneix 2001; Ouatas et al. 1998) to introduce an estrogen response element-thymidine kinase-luciferase (ERE-TK-LUC) construct into the intact brain. The effects of environmentally relevant concentrations of estrogenic pollutants on the expression of an established ERE reporter system characterized in vitro have been studied in several cell lines (Ackermann et al. 2002; Metivier et al. 2001). We have adapted somatic gene transfer procedures previously used for the Xenopus laevis tadpole (Ouatas et al. 1998) to demonstrate that waterborne estrogenic pollutants regulate transcription in vivo, both in X. laevis tadpoles and in the adult goldfish, Carassius auratus.
Materials and Methods
Plasmid constructs. We used a consensus ERE with a minimal thymidine kinase promoter driving firefly luciferase activity (ERE-TK-LUC) as described previously (Metivier et al. 2001). This ERE reporter system is well characterized in vitro in several cell lines (Ackermann et al. 2002; Metivier et al. 2001) and responds to both zebrafish (Menuet et al. 2002) and goldfish ER-[alpha] and ER-[beta] subtypes (Marlatt V, Trudeau VL, Moon TW, unpublished data). cytomegalovirus (CMV)-luciferase (CMV-LUC) and CMV-green fluorescent protein (CMV-GFP) were from Vical Inc. (San Diego, CA, USA) and Invitrogen (Carlsbad, CA, USA), respectively.
Luciferase activity. Brains from luciferase-transfected X. laevis tadpoles or goldfish were dissected and frozen in liquid nitrogen and stored at -80[degrees]C until assayed for luciferase activity [relative light units (RLUs)] according to the manufacturer's instructions (Promega, Charbonnieres, France). Frozen brains were sonicated in ice-cold luciferase lysis buffer (200 [micro]L for tadpoles, 500 pL for goldfish) and then centrifuged 10 min at 12,000g (4[degrees]C) to precipitate nonsoluble particles and proteins. Twenty microliters of the supernatant was mixed by vortexing with 100 [micro]L luciferase substrate and counted immediately (10 sec) using a single-well luminometer as previously reported (Ouatas et al. 1998).
Assessment of ERE- TK-LUC activity in the brains of X. laevis tadpoles. Previous data have demonstrated that somatic gene transfer is an effective method to study thyroid hormone (TH) responses in the X. laevis tadpole (Ouatas et al. 1998). To avoid possible TH-[E.sub.2] interactions in the brain (Dellovade et al. 1999), we used Nieuwkoop-Faber (NF) stage 54 X. laevis tadpoles (Nieuwkoop and Faber 1967) in which TH synthesis was inhibited by 1 g/L sodium perchlorate to determine whether waterborne estrogenic chemicals activate ERE-TK-LUC injected into the larval brain. In all cases, we report nominal water concentrations of estrogenic chemicals. In experiment 1, tadpoles were preexposed for 48 hr to 0.5 nM EE2, 5 nM [E.sub.2], 50 nM BPA (bisphenol A methylactylate; Sigma, St. Louis, MO, USA), or ethanol vehicle (0.4 mL in 4 L water in 10-L glass tanks; 20-22[degrees]C). In experiment 2, tadpoles were similarly preexposed to 200 nM tamoxifen (TAM; Sigma), a mixed ER antagonist/agonist. After the preexposure period, tadpoles were injected with ERE-TK-LUC (200 ng in 1 [micro]L) complexed with 6 equivalents (eq) of 22 kDa polyethylenimine (PEI; Euromedex, Souffelweyersheim, France) in a 5% glucose solution into the brain as previously described (Ouatas et al. 1998) and returned to clean water freshly treated with estrogenic chemicals and exposed a further 48 hr. Animals were then sacrificed and whole brains dissected for determination of total luciferase activities.
Development of a somatic gene transfer method for the goldfish brain. All fish were purchased from a local supplier (Paris, France) and maintained at 20-22[degrees]C. Adult male and female goldfish were used to optimize in vivo transfer methods and to determine if waterborne estrogenic chemicals activate ERE-TK-LUC injected into the adult brain. First, we established the least intrusive method for injection into the forebrain region. Stereotaxic methods have been established for brain third ventricular injections of medium- to large-sized goldfish (25-35 g), which involved surgical opening of the cranium (Peter and Gill 1975). A modification of this method was used to inject CMV-GFP (800 ng in 2 [micro]L; 6 eq of PEI) to determine the regions transfected by ventricular injections in adults. The skull was opened with fine scissors, and, rather than using a Hamilton syringe as originally reported (Peter and Gill 1975), we used a fine glass capillary held in a micromanipulator as reported for tadpoles (Ouatas et al. 1998). Animals were sacrificed 6 days after brain injections. Whole brain was dissected and first examined directly without fixation using epifluorescence microscopy (Olympus, Hamburg, Germany) to determine if GFP was being expressed. Some brains were fixed in 2% paraformaldehyde in phosphate buffer and processed for standard cryostat sectioning as reported previously (Coen et al. 1999).
This surgical approach permits precise injection into the regions of interest but is slow, highly invasive, and not amenable to the treatment of large numbers of animals. We developed an alternative approach that involves only minor surgery and is more rapidly completed. Animals were anesthetized in 0.05% MS-222 and placed in a sponge holder. Under a dissection microscope and using a modeler's drill apparatus (model 28-515; Proxxon, Niersbach, Germany) with a 0.5-mm bit attached, a small hole was made in the cranium at the midline 1-2 mm posterior to the posterior margins of the eye. In small goldfish (3-10 g), preliminary trials using 0.1% fast green dye (Sigma) established that an injection of 4 [micro]L at an angle of approximately 45-50[degrees] relative to the top of the head and at a depth of 3-4 mm would partially fill the brain ventricle and expose cells in the forebrain and optic tectum to the injected solution. In a trial using CMV-LUC (4 [micro]L of 500 ng DNA/pL; n = 5), approximately 80-90% of the total brain luciferase activity was found in the telencephalon and optic tectum, whereas the hypothalamus and cerebellum plus hindbrain had very low levels of transfection (data not shown).
To establish the concentration of PEI necessary for optimal transfection, small goldfish (3-10 g) were injected with 1 g CMV-LUC in 4 [micro]L complexed with 0, 3, 6, and 9 eq of PEI in a 5% glucose solution. Briefly, as previously described for tadpoles (Ouatas et al. 1998), the required amount of PEI is calculated based on the fact that 1 lag DNA contains 3 nmol phosphate and that 1 [micro]L 0.1 M PEI is equivalent to 100 nmol of amine nitrogen. Therefore, to condense 10 pg DNA with 6 eq of PEI, 180 nmol PEI (i.e., 1.8 [micro]L of 0.1 M PEI) is required. We also performed a time-course study in which animals were injected with 1 pg CMV-LUC in 4 [micro]L complexed with 6 eq of PEI, and whole brains dissected at 2, 12, 24, 48, and 96 hr. After dissection, whole brains were immediately frozen in liquid nitrogen and processed for luciferase activity as described above for tadpole brain.
Effects of estrogenic chemicals on ERE-TK-LUC in goldfish brain. For this experiment, we used small goldfish of both sexes (in 50-70 L glass tanks). Because these animals were in the early stages of seasonal gonadal redevelopment and could not be distinguished by external secondary sex characteristics, they were randomly assigned to each of the treatment groups. To determine whether waterborne estrogenic chemicals activate ERE-TK-LUC injected into the adult brain, groups of animals were preexposed for 48 hr to 10 nM [E.sub.2], 10 nM EE2, or ethanol vehicle (0.1 mL/L water). After the preexposure period, ERE-TK-LUC was injected as described above, and the fish were returned to water freshly treated with estrogenic chemicals and exposed a further 48 hr, at which time the water was changed again. The injected ERE-TK-LUC (1 [micro]g DNA in 4 [micro]L) was complexed with 6 eq of 22 kDa PEI in a 5% glucose solution. Whole brains were dissected at 96 hr after injection. Injections, exposures, and dissections were randomized over 3 days. At the time of dissection, body weights and sex of the animals were recorded.
Statistical analysis. The levels of luciferase activity (RLU) per whole X. laevis tadpole brain are expressed relative to mean expression levels per Experiment (i.e., for the corrected RLU the mean equals 1). Goldfish injected with the ERE-TK-LUC construct varied in size (3-10 g), and therefore an additional correction was made based on milligrams of brain protein in the extracted luciferase fraction that was measured according to the manufacturer's instructions (BioRad, Marnes-la-Coquette, France). Data were analyzed by one-way or two-way analysis of variance (ANOVA) or Student's t-test as appropriate (SigmaStat, version 2.03; SPSS Inc., Chicago, IL, USA).
Effects of estrogenic chemicals on ERE-TK-LUC activity in the brains of X. laevis tadpoles.
Figure 1A shows the effects of exposure to [E.sub.2] (5 nM), E[E.sub.2] (0.5 nM), and BPA (50 nM) on luciferase expression in the brains of ERE-TK-LUC-injected tadpoles. In this experiment the average activity (1 corrected RLU) represents approximately 73,000 RLU/brain. All data are expressed relative to this average. There was an effect of treatment (p < 0.004, one-way ANOVA) on luciferase activity. In the group treated with [E.sub.2], mean levels were approximately 1.4-fold higher than in controls; however, this difference did not achieve statistical significance (p > 0.05). In contrast, E[E.sub.2] induced a 1.9-fold increase (p < 0.002) in luciferase activity. Similarly, BPA also induced a 1.5-fold increase (p < 0.03) in luciferase activity measured in the whole brain. Figure 1B shows the effects of exposure to TAM (200 nM) on luciferase expression in the brains of ERE-TK-LUC-injected tadpoles. In this experiment the average activity (1 corrected RLU unit) represents approximately 38,000 RLU. All data are expressed relative to this average value. TAM induced a 2.3-fold increase (p < 0.003, t-test) in luciferase activity.
[FIGURE 1 OMITTED]
Somatic gene transfer in the goldfish brain. When we injected directly into the brain third ventricle of medium sized fish, cells in the telencephalon, optic tectum, and occasionally in the hypothalamus (not shown) were transfected with CMV-GFP (800 ng in 2 [micro]L). Figure 2A shows the general distribution of GFP-expressing cells in a freshly dissected whole brain. Cells in the telencephalon close to the midline and brain third ventricle, as well as some cells in the optic tectum, were visualized with epifluorescence microscopy. Examples of two neurons expressing GFP are shown in Figure 2B and C. GFP was expressed in the cell body and also extensively in neuronal processes extending laterally away from the ventricular wall (represented by the border between Figure 2B,C). Note also that synaptic boutons and dendrites are also labeled with GFP. Cells in the nucleus preopticus periventricularis and nucleus preopticus (Peter and Gill 1975) also expressed GFP (not shown).
[FIGURE 2 OMITTED]
Figure 3A illustrates the effect of PEI concentrations on transfection efficiency in the goldfish brain. Whereas 3 eq of PEI was only minimally effective, 6 eq of PEI produced maximal luciferase expression at 48 hr after brain injections. There was no further enhancement of transfection using 9 eq of PEI. Using 6 eq of PEI to complex CMV-LUC, a time-course analysis (Figure 3B) was performed. The highest luciferase expression was 96 hr after brain injection.
[FIGURE 3 OMITTED]
Effects of estrogenic chemicals on ERE-TK-LUC in goldfish brain. After having established a method for injection of DNA into the goldfish brain (Figures 2 and 3), we examined the effects of [E.sub.2], EE2, and BPA in small female and male goldfish. Figure 4 shows the effects of exposure to [E.sub.2] (10 nM), EE2 (10 nM), and BPA (100 nM) on luciferase expression in the brains of ERE-TK-LUC-injected females and males. In this experiment the average activity (1 corrected unit) represents approximately 15,000 RLU/mg protein. All data are expressed relative to this average value. The effects of the various treatments on luciferase activity was dependent on the sex of the fish (two-way ANOVA: sex x treatment, p < 0.019). Basal luciferase activity was similar in control females and males (p > 0.05). In males treated with [E.sub.2], mean levels were approximately 1.5-fold higher than in controls; however, this change was not statistically significant (p > 0.05). Additionally, [E.sub]2 did not affect (p > 0.05) luciferase activity in females. In contrast, waterborne EE2 induced a 2-fold increase (p > 0.05) in luciferase activity in the male brain but had no effect in females (p > 0.05). Moreover, BPA did not affect (p > 0.05) luciferase activity in either sex.
[FIGURE 4 OMITTED]
Our results indicate that waterborne estrogenic chemicals can modulate brain activity in aquatic vertebrates. Using somatic gene transfer into the brains of tadpoles and adult fish, we showed that estrogenic chemicals activate expression of an introduced ERE-TK-LUC construct. This required adaptation of somatic gene transfer methods previously used in X. laevis (Ouatas et al. 1998) and mice (Guissouma et al. 1998) to study TH-driven gene expression and in Xenopus tropicalis (Rowe et al. 2002) to analyze apoptosis during metamorphosis. To our knowledge, PEI-mediated somatic gene transfer into the fish brain has not been previously reported. Optimal transfections were achieved with 1 big DNA complexed with 18 nmol 22 kDa PEI 4 days after brain injections. However, longer time periods were not analyzed, and it is possible that expression in the adult goldfish brain would increase after 96 hr.
The potent estrogen from female contraceptives, EE2, and the natural estrogen [E.sub.2] are found at picomolar to nanomolar concentrations in both European and North American sewage effluents and surface waters (Kolpin et al. 2002; Metcalfe et al. 2003; Ternes et al. 1999). We showed that short-term exposure to 0.5 nM EE2 in tadpoles and 10 nM EE2 in male goldfish increased the activity of a known estrogen-responsive reporter gene construct by approximately 2-fold. In contrast, female goldfish were not responsive to 10 nM waterborne E[E.sub.2]. The plasticizing agent BPA and the mixed ER antagonist/agonist TAM were both estrogenic in tadpole brain. In both male and female goldfish, high levels (100 nM) of BPA did not activate the estrogen-responsive reporter gene construct injected into the brain.
Although we did not directly compare transfection efficiencies in tadpoles and goldfish, there appears to be an important difference. Based on the number of GFP-positive cells and the basal levels of luciferase expression, transfection appears less efficient in adult goldfish compared with larval tadpole brain (Ouatas et al. 1998). The maximum activity of reporter gene luciferase from a whole brain per milligram of protein showed that transfection is ~ 30-fold more efficient in the tadpole compared with the adult goldfish. The reasons for this are unknown but likely relate to differences in injection methods, ratio of brain volume to injection volume, and/or cellular characteristics of larval versus adult brain. Results in goldfish are, however, similar to those obtained with PEI-mediated transfection of hypothalamic neurons of neonatal mice with the same CMV-LUC construct (Guissouma et al. 1998). Our results showed that cells in the adult goldfish forebrain and optic tectum are transfectable in vivo. Autoradiographic (Kim et al. 1978), immunocytochemical (Navas et al. 1995), and in situ hybridization (Menuet et al. 2002) studies showed that both ER-[alpha] and ER-[beta] are expressed in the telencephalon and hypothalamus and especially in the preoptic area of fish. Using reverse-transcriptase polymerase chain reaction (RT-PCR), Choi and Habibi (2003) also detected both ERs in goldfish brain. Our results showed that in vivo transfection in the goldfish telencephalon can be used to study the regulation of ERE-driven expression by an estrogenic pollutant.
In both animal models, there was a relatively high basal luciferase activity in controls. This is likely due to two interacting factors: high in vivo activity of the minimal thymidine kinase promoter and effects of endogenous neuroestrogen on basal expression of the ERE-TK-LUC gene construct. In X. laevis tadpoles, estrogen production in the brain has not been studied, but at NF stage 54, whole-body [E.sub.2] levels are easily detectable despite having declined relative to very high levels in early stages of development (Bogi et al. 2002). Male and female gonads are distinguishable by gross morphologic characteristics at NF stage 56 (Bogi et al. 2002). Therefore, it is likely that our tadpoles were producing endogenous estrogen. Relatively high basal ERE-TK-LUC activity at this stage of tadpole development suggests that ERs are active and/or that endogenous [E.sub.2] is being produced and delivered to the transfected cells. The goldfish brain has a remarkable capacity to produce [E.sub.2] from testosterone because of very high aromatase activity (Callard et al. 2001; Pasmanik and Callard 1988). The dose of [E.sub.2] we used is within the physiologic range and thus would be unlikely to raise brain [E.sub.2] above endogenous brain [E.sub.2] concentrations, especially in females. It is known that EE2 is more potent that [E.sub.2] in several assay systems using the same ERE-TK-LUC reporter gene (Ackermann et al. 2002; Le Guevel and Pakdel 2001). In female goldfish, 10 nM EE2 did not affect luciferase expression, similar to what was observed with [E.sub.2]. This is in contrast to males where EE2 induced a 2-fold increase in activity. We have previously observed marked sex differences in goldfish neuroendocrine responses to sex steroids (Bosma et al. 2001). For example, whereas testosterone inhibited the expression of glutamic acid decarboylases (GAD65 and GAD67) in the telencephalon of sexually mature males, it was without effect in females (Lariviere K, Trudeau VL, unpublished data).
Our results indicate that short-term exposure to environmentally relevant water levels of BPA (50 nM,--18 [micro]g/L) can activate the ERE-TK-LUC construct in the tadpole brain. In contrast to effects in fish (Metcalfe et al. 2001; Staples et al. 1998), the effects of BPA in amphibians are not well studied. Kloas et al. (1999), using a static renewal exposure protocol, reported that BPA has estrogenic activity at 2.3 [micro]g/L (-6.3 nM) and induces female-biased sex reversal in X. laevis. In a second study, the same researchers found that 100 nM BPA induced female-biased sex reversal in X. laevis (Levy et al. 2004). However, in a flow-through exposure system (Pickford et al. 2003), there were no observable effects of a range of BPA concentrations (0.83-497 [micro]g/L; ~ 2.3 nM-1.4 [micro]M) on larval growth, development, or sexual differentiation of X. laevis tadpoles. High, nonenvironmental levels of BPA (10-25 [micro]M; 3,644-9,110 [micro]g/L) have both teratogenic and antimetamorphic actions in X laevis (Iwamuro et al. 2003), suggesting interference with the thyroid system. It is difficult at present to reconcile the different conclusions concerning the estrogenicity of BPA in frogs. However, given that BPA is continually being added to aquatic ecosystems through industrial and sewage effluent discharges and activates a known ERE-reporter gene construct in tadpole brain, it is a contaminant of environmental concern.
Activation of an ERE-dependent luciferase reporter gene in both tadpole and fish indicates that waterborne estrogens can directly modulate transcription of estrogen-responsive genes in the brain. Previous work from our laboratory demonstrated that environmentally relevant levels of the estrogenic pollutant octylphenol modulates the expression of multiple hypothalamic genes in leopard frog tadpoles (Crump et al. 2002) and in hatchling snapping turtles (Trudeau et al. 2002). In the latter study, differential display PCR was used, and it is not known if the affected transcripts were directly or indirectly regulated by 4-t-octylphenol or [E.sub.2]-As quantified in these latter studies using reverse Northern blotting, changes in several hypothalamic mRNAs induced by waterborne environmentally relevant levels of octylphenol in these studies were approximately 2-fold. This level of gene expression is similar to what we observed with ERE-dependent luciferase induction after EE2, BPA, and TAM exposures.
In this article we provide a method to study the direct regulation of estrogenresponsive genes in vivo in tadpoles and fish. The power of this approach is that it is possible to determine whether an estrogenic chemical is acting on a certain tissue. The ERE-dependent luciferase reporter gene is injected at a specific site and is responsive to the known nuclear ER subtypes (Menuet et al. 2002). Moreover, because the technique is based on the transcriptional mechanism of action of estrogen, a positive effect of a given chemical can be interpreted as activation of the ER. A somatic gene transfer technique may be generalized to other aquatic species, because it is easier and less time-consuming than identifying ER-regulated genes and their promoters in each species. The main limitation of the somatic gene transfer method described is that it is unlikely to detect indirect and nongenomic effects of estrogenic chemicals.
We thank K. Palmier for performing the protein assays; G. Benisti, J.-P. Chaumeil, and E. LeGoff for animal care and technical support; and A. DeLuze for discussions and advice.
Abraham IM, Todman MG, Korach KS, Herbison AE. 2004. Critical in viva roles for classical estrogen receptors in rapid estrogen actions on intracellular signaling in mouse brain. Endocrinology 145(7):3055-061.
Ackermann GE, Brombacher E, Fent K. 2002. Development of a fish reporter gene system for the assessment of estrogenic compounds and sewage treatment plant effluents. Environ Toxicol Chem 21 (9):1864-1875.
Andersen HR, Andersson AM, Arnold SF, Autrup H, Barfoed M, Beresford NA, et al. 1999. Comparison of short-term estrogenicity tests for identification of hormone-disrupting chemicals. Environ Health Perspect 107(suppl 1):89-108.
Arcand-Hoy LD, Benson WH. 1998. Fish reproduction: an ecologically relevant indicator of endocrine disruption. Environ Toxicol Chem 17(1):49-57.
Arukwe A. 2001. Cellular and molecular responses to endocrinemodulators and the impact on fish reproduction. Mar Pollut Bull 42(8):643-655.
Belfroid A, van Velzen M, van der Horst B, Vethaak D. 2002. Occurrence of bisphenol A in surface water and uptake in fish: evaluation of field measurements. Chemosphere 49(1):97-103.
Bogi C, Levy G, Lutz I, Kloas W. 2002. Functional genomics and sexual differentiation in amphibians. Camp Biochem Physiol B Biochem Mol Biol 133(4):559-570.
Bosma PT, Blazquez M, Fraser EJ, Schulz RW, Docherty K, Trudeau VL. 2001. Sex steroid regulation of glutamate decarboxylase mRNA expression in goldfish brain is sexually dimorphic. J Neurochem 76(4):945-956.
Callard GV, Tchoudakova AV, Kishida M, Wood E. 2001. Differential tissue distribution, developmental programming, estrogen regulation and promoter characteristics of cyp19 genes in teleost fish. J Steroid Biochem Mol Biol 79(1-5):305-314.
Choi CY, Habibi HR. 2003. Molecular cloning of estrogen receptor alpha and expression pattern of estrogen receptor subtypes in male and female goldfish. Mol Cell Endocrinol 204(1-2):169-177.
Coen L, Kissa K, Le Mevel S, Brulet P, Demeneix BA. 1999. A somatic gene transfer approach using recombinant fusion proteins to map muscle-motoneuron projections in Xenopus spinal cord. Int J Dev Biol 43(8):823-830.
Crump D, Lean D, Trudeau VL. 2002. 0ctylphenol and UV-B radiation alter larval development and hypothalamic gene expression in the leopard frog (Rana pipiens). Environ Health Perspect 110:277-284.
Dellovade TL, Kia HK, Zhu YS, Pfaff DW. 1999. Thyroid hormone coadministration inhibits the estrogen-stimulated elevation of preproenkephalin mRNA in female rat hypothalamic neurons. Neuraendocrinology 70(3):168-174.
Feigelson HS, Henderson BE. 1996. Estrogens and breast cancer. Carcinogenesis 17(11):2279-2284.
Graham JD, Bain DL, Richer JK, Jackson TA, Tung L, Horwitz KB. 2000. Thoughts on tamoxifen resistant breast cancer. Are coregulators the answer or just a red herring? J Steroid Biochem Mol Biol 74(5):255-259.
Guillette LJ Jr, Crain DA, Rooney AA, Pickford DB. 1995. Organization versus activation: the role of endocrinedisrupting contaminants (EDCs) during embryonic development in wildlife. Environ Health Perspect 103(suppl 7):157-164.
Guissouma H, Ghorbel MT, Seugnet I, Ouatas T, Demeneix BA. 1998. Physiological regulation of hypothalamic TRH transcription in viva is T3 receptor isoform specific. FASEB J 12(15):1755-1764.
Hess RA. 2003. Estrogen in the adult male reproductive tract: A review. Reprod Biol Endocrinol 1(1):52. Available: http://www.rbej.com/content/1/1/52 [accessed 13 January 20051.
Iwamuro S, Sakakibara M, Terao M, Ozawa A, Kurobe C, Shigeura T, et al. 2003. Teratogenic and anti-metamorphic effects of bisphenol A on embryonic and larval Xenopus laevis. Gen Comp Endocr 133(2):189-198.
Kim YS, Stumpf WE, Sar M. 1978. Topography of estrogen target cells in the forebrain of goldfish, Carassius auratus. J Camp Neural 182(4):611-620.
Kloas W, Lutz I, Einspanier R. 1999. Amphibians as a model to study endocrine disruptors: II. Estrogenic activity of environmental chemicals in vitro and in viva. Sei Total Environ 225(1-2):59-68.
Kolpin DW, Furlong ET, Meyer MT, Thurman EM, Zaugg SD, Barber LB, et al. 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams, 1999-2000: a national reconnaissance. Environ Sci Technol 36(6):1202-1211.
Koraeh KS, Emmen JM, Walker VR, Hewitt SC, Yates M, Hall JM, et al. 2003. Update on animal models developed for analyses of estrogen receptor biological activity. J Steroid Biochem Mol Biol 86(3-5):387-391.
Lee HB, Peart TE. 2000. Determination of bisphenol A in sewage effluent and sludge by solid-phase and supercritical fluid extraction and gas chromatography/mass spectrometry. J AOAC Int 83(2):290-297.
Le Guevel R, Pakdel F. 2001. Assessment of oestrogenic potency of chemicals used as growth promoter by in-vitro methods. Hum Reprod 16(5):1030-1036.
Lemkine GF, Demeneix BA. 2001. Polyethylenimines for in viva gene delivery. Curr Opin Mol Ther 3(2):178-182.
Levy G, Lutz I, Kruger A, Kloas W. 2004. Bisphenol A induces feminization in Xenopus laevis tadpoles. Environ Res 94(1):102-111.
Mackenzie CA, Berrill M, Metcalfe C, Pauli BD. 2003. Gonadal differentiation in frogs exposed to estrogenic and antiestrogenic compounds. Environ Toxicol Chem 22(10):2466-2475.
Matthews JB, Fertuck KC, Celius T, Huang YW, Fang CJ, Zacharewski TR. 2002. 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 82(2-3):181-194.
McLachlan JA. 2001. Environmental signaling: what embryos and evolution teach us about endocrine disrupting chemicals. Endocr Rev 22(3):319-341.
Menuet A, Pellegrini E, Anglade I, Blaise O, Laudet V, Kah 0, et al. 2002. Molecular characterization of three estrogen receptor forms in zebrafish: binding characteristics, transactivation properties, and tissue distributions. Biol Reprod 66(6): 1881-1892.
Metcalfe CD, Koenig BG, Bennie DT, Serves M, Ternes TA, Hirsch R. 2003. Occurrence of neutral and acidic drugs in the effluents of Canadian sewage treatment plants. Environ Toxicol Chem 22(12):2872-2880.
Metcalfe CD, Metcalfe TL, Kiparissis Y, Koenig BG, Khan C, Hughes RJ, et al. 2001. Estrogenic potency of chemicals detected in sewage treatment plant effluents as determined by in vivo assays with Japanese medaka (Oryzias latipes). Environ Toxicol Chem 20(2):297-308.
Metivier R, Penot G, Flouriot G, Pakdel F. 2001. Synergism between ERalpha transactivation function 1 (AF-1) and AF-2 mediated by steroid receptor coactivator protein-1: requirement for the AF-1 alpha-helical core and for a direct interaction between the N- and C-terminal domains. Mol Endocrinol 15(11):1953-1970.
Metivier R, Penot G, Hubner MR, Reid G, Brand H, Kos M, et al. 2003. Estrogen receptor-alpha directs ordered, cyclical, and combinatorial recruitment of cofactors on a natural target promoter. Cell 115(6):751-763.
Navas JM, Anglade I, Bailhache T, Pakdel F, Breton B, Jego P, et al. 1995. Do gonadotrophin-releasing hormone neurons express estrogen receptors in the rainbow trout? A double immunohistochemical study. J Comp Neurol 363(3):461-474.
Nieuwkoop P, Faber J. 1967. Normal Table of Xenopus laevis (Daudin). Amsterdam:Elsevier North Holland.
0uatas T, Le Mevel S, Demeneix BA, de Luze A. 1998. T3-dependent physiological regulation of transcription in the Xenopus tadpole brain studied by polyethylenimine based in vivo gene transfer. Int J Dev Biol 42(8):1159-1164.
Pasmanik M, Callard GV. 1988. Changes in brain aromatase and 5 alpha-reductase activities correlate significantly with seasonal reproductive cycles in goldfish (Carassius auratus). Endocrinology 122(4):1349-1356.
Peter RE, Gill VE. 1975. A stereotaxic atlas and technique for forebrain nuclei of the goldfish, Carassius auratus. J Comp Neurol 159(1):69-101.
Petit F, Le Goff P, Cravedi JP, Valotaire Y, Pakdel F. 1997. Two complementary bioassays for screening the estrogenic potency of xenobiotics: recombinant yeast for trout estrogen receptor and trout hepatocyte cultures. J Mol Endocrinol 19(3):321-335.
Pickford DB, Hetheridge M J, Caunter JE, Hall AT, Hutchinson TH. 2003. Assessing chronic toxicity of bisphenol A to larvae of the African clawed frog (Xenopus laevis) in a flow-through exposure system. Chemosphere 53(3):223-235.
Robinson-Rechavi M, Escriva Garcia H, Laudet V. 2003. The nuclear receptor superfamily. J Cell Sci 116(Pt 4):585-586.
Rowe I, Coen L, Le Blay K, Le Mevel S, Demeneix BA. 2002. Autonomous regulation of muscle fibre fate during metamorphosis in Xenopus trapicalis. Dev Dyn 224(4):381-390.
Sanchez R, Nguyen D, Rocha W, White JH, Mader S. 2002. Diversity in the mechanisms of gene regulation by estrogen receptors. Bioessays 24(3):244-254.
Segner H, Navas JM, Schafers C, Wenzel A. 2003. Potencies of estrogenic compounds in in vitro screening assays and in life cycle tests with zebrafish in vivo. Ecotoxicol Environ Saf 54(3):315-322.
Staples CA, Dorn PB, Klecka GM, O'Block ST, Harris LR. 1998. A review of the environmental fate, effects, and exposures of bisphenol A. Chemosphere 36(10):2149-2173.
Sumpter JP, Jobling S. 1995. Vitellogenesis as a biomarker for estrogenic contamination of the aquatic environment. Environ Health Perspect 103(suppl 7):173-178.
Ternes TA, Stumpf M, Mueller J, Haberer K, Wilken RD, Servos M. 1999. Behavior and occurrence of estrogens in municipal sewage treatment plants I. Investigations in Germany, Canada and Brazil. Sci Total Environ 225(1-2):81-90.
Trudeau VL. 1997. Neuroendocrine regulation of gonadotrophin II release and gonadal growth in the goldfish, Carassius auratus. Rev Reprod 2(1):55-68.
Trudeau VL, Chiu S, Kennedy SW, Brooks RJ. 2002. Octylphenol (OP) alters the expression of members of the amyleid protein family in the hypothalamus of the snapping turtle, Chelydra serpentina serpentina. Environ Health Perspect 110:269-275.
Tyler CR, Jobling S, Sumpter JP. 1998. Endocrine disruption in wildlife: a critical review of the evidence. Crit Rev Toxicol 28(4):319-361.
Yamamoto T, Yasuhara A, Shiraishi H, Nakasugi O. 2001. Bisphenol A in hazardous waste landfill leachates. Chemosphere 42(4):415-418.
Yin GG, Kookana RS, Ru YJ. 2002. Occurrence and fate of hormone steroids in the environment. Environ Int 28(6):545-551.
Yoon K, Pallaroni L, Stoner M, Baido K, Safe S. 2001. Differential activation of wild-type and variant forms of estrogen receptor alpha by synthetic and natural estrogenic compounds using a promoter containing three estrogen-responsive elements. J Steroid Biochem Mol Biol 78(1):25-32.
Zacharewski T. 1997. In vitro bioassays for assessing estrogenic substances. Environ Sci Technol 31(3):613-623.
Vance L. Trudeau, (1) Nathalie Turque, (2) Sebastien Le Mevel, (2) Caroline Alliot, (2) Natacha Gallant, (1) Laurent Coen, (2) Farzad Pakdel, (3) and Barbara Demeneix (2)
(1) Centre for Advanced Research in Environmental Genomics, Department of Biology, University of Ottawa, Ottawa, Ontario, Canada; (2) UMR-5166, Evolution des Regulations Endocriniennes, Museum National d'Histoire Naturelle, Centre National de la Recherche Scientifique, Paris, France; (3) UMR-6026, Equipe d'Endocrinologie Moleculaire de la Reproduction, Centre National de la Recherche Scientifique, Universite de Rennes, Rennes, France
Address correspondence to V.L. Trudeau, Department of Biology, Third Floor, MacDonald Hall, University of Ottawa, Ottawa, Ontario K1N 6N5 Canada. Telephone: (613) 562-5800 ext. 6165. Fax: (623) 562-5486. E-mail: email@example.com
V.L.T. was Professeur Invite at the Museum National d'Histoire Naturelle, Paris, when this work was initiated.
This study was supported by the Natural Sciences and Engineering Research Council of Canada, the Canadian Network of Toxicology Centres, the University of Ottawa, Museum National d'Histoire Naturelle, Paris, and Euromedex.
Received 14 July 2004; accepted 2 December 2004.
|Printer friendly Cite/link Email Feedback|
|Publication:||Environmental Health Perspectives|
|Date:||Mar 1, 2005|
|Previous Article:||Acceleration of autoimmunity by organochlorine pesticides in (NZB x NZW)[F.sub.1] mice.|
|Next Article:||Summary of a workshop on the development of health models and scenarios: strategies for the future.|