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Effects of xenobiotics and steroids on renal and hepatic estrogen metabolism in lake trout
General and Comparative Endocrinology 148 (2006) 273–281 www.elsevier.com/locate/ygcen
EVects of xenobiotics and steroids on renal and hepatic estrogen metabolism in lake trout Gail F. Jurgella a, Ashok Marwah b, JeVrey A. Malison a, Richard Peterson c, Terence P. Barry a,¤ b
a Department of Animal Sciences, University of Wisconsin-Madison, 1675 Observatory Drive, Madison, WI 53706, USA Department of Biochemistry—Enzyme Institute, University of Wisconsin-Madison, 1710 University Avenue, Madison, WI 53726, USA c School of Pharmacy, University of Wisconsin-Madison, 777 Highland Avenue, Madison, WI 53705, USA
Received 15 December 2005; revised 13 March 2006; accepted 23 March 2006 Available online 4 May 2006
Abstract Experiments were conducted to (1) elucidate the biochemical pathways of E2 metabolism in the lake trout (Salvelinus namaycush) kidney and liver, and (2) test the hypothesis that speciWc xenobiotics and endogenous steroids inhibit E2 metabolism by these tissues. Kidney and liver tissue fragments from immature lake trout were incubated in vitro in the presence of radiolabelled E2 plus various xenobiotics or steroids. E2 metabolites were identiWed by liquid chromatography/mass spectroscopy, and quantiWed by liquid scintillation spectroscopy. A major metabolite produced by both tissues was an unidentiWed hydroxylated estrogen metabolite (E2-OH) with a molecular mass of 288 that was not estriol (16-OH-E2), but possibly 7-OH-E2 or 2-OH-E2 (catecholestrogen). Both tissues also produced estradiol-17glucuronide (E2-17-G), estradiol-17-sulfate (E2-17-S), and estradiol-3-glucuronide (E2-3-G). Compared to the kidney, the liver produced half the amount of conjugated metabolites, but twofold more E2-OH. The following xenobiotics (at a concentration of 100 M) inhibited the production of water-soluble (i.e., conjugated) E2 metabolites by both the kidney and liver: 4,4⬘-(OH)2-3,3⬘,5,5⬘- tetrachlorobiphenyl (4,4⬘-OH-TCB), bisphenol A (BPA), tetrabromobisphenol A (TB-BPA), tetrachlorobisphenol A (TC-BPA), tribromophenol (TBP), trichlorophenol (TCP), and pentachlorophenol (PCP). The alkylphenols, 4-n-nonylphenol (NP), and 4-octylphenol (OP), and 2,2⬘,4,4⬘tetrabromodiphenyl ether (TBDE) had no signiWcant eVect on E2 metabolism by either tissue. Testosterone and 17,20-dihydroxy-4pregnen-3-one inhibited the production of conjugated E2 metabolites by both the kidney and liver. Cortisol and 11-ketotestosterone inhibited E2 metabolism by the liver only. The median inhibitory concentrations (IC50) for 4,4⬘-OH-TCB ranged from 7–32 M in the kidney and 0.6–1.6 M in the liver. For BPA, IC50’s ranged from 40–108 M in the kidney and 11–18 M in the liver. Low doses (0.1 M) of 4,4⬘-OH-TCB and BPA signiWcantly increased estrogen metabolism in the kidney. The results suggest that certain estrogenic xenobiotics and endogenous steroids may inhibit the phase II conjugation of E2 by the kidney and liver of lake trout, and some of the known biological eVects of these compounds are likely mediated, at least partially, by this mechanism of action. © 2006 Elsevier Inc. All rights reserved. Keywords: Lake trout; Endocrine disruption; Steroidogenic enzymes; Estrogen; Xenobiotic; Steroid hormone; Estradiol-17; Steroid; Steroid metabolism
1. Introduction The lake trout (Salvelinus namaycush) is the top natural predator in the Great Lakes food web. During the twentieth century, lake trout populations dropped dramatically in Lake Ontario, and it was thought that this decline was the *
Corresponding author. Fax: +1 608 262 0454. E-mail address: [email protected] (T.P. Barry).
0016-6480/$ - see front matter © 2006 Elsevier Inc. All rights reserved. doi:10.1016/j.ygcen.2006.03.011
result of sea lamprey predation, declining water quality, and over harvesting. Attempts to reestablish naturally reproducing lake trout populations focused primarily on improving these problems, largely without success (Gilbertson, 1992). Later studies suggested that xenobiotic chemicals present in the water might be the principal cause of lake trout recruitment failure. For example, Walker and Peterson (1994) demonstrated that lake trout larvae are highly sensitive to 2,3,7,8-tetrachlorodibenzo-p-dioxin
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(TCDD), and even very low-level exposure to TCDD causes early life stage mortality. Further studies demonstrated that the concentrations of TCDD and related chemicals in Lake Ontario during the 1960s were high enough to explain the observed decline in lake trout populations (Cook et al., 2003). Despite bans on the use of TCDD and other xenobiotics such as polychlorinated biphenyls (PCBs), these lipophilic compounds continue to persist in lake sedimentation basins, and in the tissues of contaminated Wsh, particularly the top trophic-level predators (Humphrey et al., 2000; Safe, 1994). Moreover, the levels of other xenobiotics that are potentially harmful to lake trout, including polybrominated diphenyl ethers (PBDEs), are currently increasing in the Great Lakes due to their continued production and use by industry (deWit, 2002; Hale et al., 2003). Although the levels of most xenobiotics in the Great Lakes and in Great Lakes Wshes are currently too low to cause outright early life stage mortality, they may still reduce lake trout recruitment by acting as endocrine disruptors. Certain PCBs and hydroxylated PCB metabolites are known to modify physiological functions in estrogen-responsive targets such as the reproductive organs, liver, and neuroendocrine centers (Jobling et al., 1995; McLachland and Arnold, 1996; White et al., 1994). Even weakly active compounds may disturb the diVerentiation or function of estrogen-responsive tissues if exposure occurs during critical developmental periods (Colborn et al., 1993). Developing Wsh embryos and larvae are particularly at risk because many xenobiotics can be transferred from the mother to the developing eggs (Peterson et al., 1993; Walker and Peterson, 1994). The mechanisms by which xenoestrogens exert their endocrine-disrupting eVects are not well established. Many potent hydroxylated PCB metabolites (PCB-OHs), for example, show little or no aYnity for the - or -estrogen receptor (ER) subtypes, and thus may not act directly as estrogen receptor agonists (Korach et al., 1988; Kuiper et al., 1998). In humans, Kester et al. (2000) showed that numerous PCB-OHs are powerful inhibitors of the phase II conjugating enzyme, estrogen sulfotransferase (SULT1E1), and indirectly induce estrogenic activity by reducing the metabolism of endogenous estrogen. Indeed, the inhibition of SULT1E1 is the most potent biological eVect described to date regarding the endocrine-disrupting activity of PCBs or their metabolites (Kester et al., 2000). The most potent PCB-OHs, with IC50 values in the subnanomolar range, were compounds with a 4-OH-3,5-dichloro substitution pattern (Kester et al., 2000). In subsequent studies, other xenobiotic compounds with known estrogenic activity, including halogenated bisphenol A (BPA) metabolites, were also shown to be potent inhibitors of SULT1E1 (Kester et al., 2002). Xenobiotic inhibition of phase II conjugating enzymes may also be an important mechanism of action of endocrine disruption in Wsh. For example, Ohkimoto et al. (2003) reported that BPA, 4-n-nonylphenol (NP), and 4-n-octylphenol (OP) inhibited the sulfonation of estradiol-17 (E2) in
zebraWsh (Brachydanio rerio), and van den Hurk et al. (2002) found that several PCB-OHs inhibited both the sulfonation and glucuronidation of benzo[a] pyrene in channel catWsh (Ictalurus punctatus), indicating that both sulfotransferase (SULT), and glucuronosyltransferase (UGT) are targets for inhibition by xenobiotics in Wsh. The goals of the present investigation were to (1) begin characterizing the biochemical pathways of E2 metabolism in the kidney and liver of lake trout, and (2) evaluate the eVects of various estrogenic xenobiotics, as well as several key Wsh steroid hormones, on the production of conjugated, water-soluble estrogen metabolites normally produced by the lake trout kidney and liver. 2. Materials and methods 2.1. Chemicals Steroids were obtained from Sigma Chemical Co. (St. Louis, MO) or Steraloids (Newport, RI). [6,7-3H]-E2 (40 to 60 Ci/mmol) was purchased from Perkin-Elmer Life Sciences (Boston, MA). 4,4⬘-OH-TCB was obtained from Ultra ScientiWc (North Kingstown, RI). BPA, OP, and NP were obtained from Sigma (St. Louis, MO). Tetrachlorobisphenol A (TC-BPA), tetrabromobisphenol A (TB-BPA), PCP, 2,4,6-trichlorophenol (TCP), 2,4,6tribromophenol (TBP), and 2,2⬘,4,4⬘-tetrabromodiphenyl ether (TBDE) were gifts from AccuStandard (New Haven, CT). All other chemicals were obtained from Sigma Chemical Co. (St. Louis, MO) and Fisher ScientiWc (Atlanta, GA). HPLC grade methanol, ethanol, hexane, and acetonitrile were purchased from Sigma–Aldrich (Milwaukee, WI, USA) and used as such. HPLC grade acetic acid was purchased from Fisher ScientiWc (Pittsburgh, PA, USA). Formic acid was purchased from Fluka (Milwaukee, WI, USA). Distilled water was deionized and puriWed (18.25 § 0.05 M-cm) using a Nanopure water system from Barnstead International (Dubuque, IA, USA). Solid phase extraction (SPE) cartridges (Oasis-HLB, 3 cc) were obtained from Waters Corporation (Milford, MA, USA).
2.2. Fish Lake trout were obtained from the Minnesota Department of Natural Resources, Crystal Springs State Fish Hatchery (Altura, MN). Three-yearold, sexually immature Wsh, 30–50 cm in total length, were used in all experiments. The Wsh were maintained in 750 L circular Wberglass tanks supplied with Xow-through carbon-Wltered city water at 10–14 °C. Photoperiod was kept constant at 12 h light/12 h dark. The Wsh were fed once daily to satiation using Silvercup trout feed (Murray, UT). Food was withheld 24 h prior to sacriWce. Care and treatment of the Wsh were in accordance with the guidelines of the University of Wisconsin Research Animal Care Committee.
2.3. In vitro cultures Physiological saline and Hepes-buVered culture medium, developed for culturing salmonid tissues, were prepared and sterilized by ultraWltration at 0.2 m (Barry et al., 1997). Estrogen metabolism was quantiWed by measuring the formation of E2 metabolites after incubation of tissue fragments in 1 ml of culture medium in the presence of [3H]-E2, according to the method of Barry et al. (1997). In brief, the kidney and liver were uniformly fragmented using a Polytron (Brinkmann Instruments, Westbury, NY). The tissue fragments were washed 2–3 times in physiological saline, and 50 l aliquots of tissue (1.16 § 0.03 mg protein) were added to 150 £ 20 mm test tubes and incubated at 15 °C for 1–24 h. Controls were no-tissue and time 0 incubates. The cultures were terminated and extracted as described below. Experiments used tissues from one Wsh cultured in duplicate incubations, and each experiment was conducted two to four times (N D 2 to 4).
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2.4. IdentiWcation of estrogen metabolites Kidney and liver tissue fragments were incubated with 106 dpm [3H]E2 and 3.7 M radioinert E2 for 0, 1, 4, and 24 h. Estradiol-17 and its metabolites were extracted according to the method of Marwah et al. (2002). BrieXy, water (1 £ vol) and acetic acid (0.05 £ vol) were added to the cultures (medium plus tissue). The samples were centrifuged at 1500g for 5 min to pellet the tissue. The supernatants were applied to pre-conditioned (2 ml methanol followed by 2 ml water), 3 ml, solid-phase, extraction cartridges (Oasis-HLB Waters Corp., Milford, MA), and washed twice with 2 ml of 5% methanol. The metabolites were eluted with 3 ml of 100% methanol, and the methanol was evaporated under a stream of nitrogen at 40 °C. The residues were dissolved in 0.2 ml water, and analyzed directly by LC–MS as described below. No radioactivity remained associated with the tissue pellet. Chromatography of E2 and its neutral metabolites was performed on C18 analytical columns (Zorbax-SB, 3.0 £ 150 mm, 3.5 m, Mac-Mod Analytical, Inc., Chadds Ford, PA, and Novapack, 3.9 £ 75 mm, 4.0 m, Waters Corp., Milford, MA) protected with matching C18 guard columns at a Xow rate of 0.4 ml/min for Zorbax C18 column, and at a Xow rate of 0.8 ml/min for Novapack column and a temperature of 40 °C. The solvent was acetonitrile–water with a linear gradient to reach 20–45% acetonitrile at 25 min, 94% at 32 min and back to 20% at 34 min, followed by a 10 min post run time. Water-soluble conjugates were resolved on a C18 analytical column (Zorbax-SB, 3.0 £ 150 mm, 3.5 m, Mac-Mod Analytical, Inc., Chadds Ford, PA) at a Xow-rate of 0.4 ml/min using acetonitrile–water and acetonitrile containing 3% acetic acid and 3% acetic acid gradients. The gradients started with 10% acetonitrile to reach 40% at 30 min, and 96% acetonitrile at 38 min and back to 10% at 40 min. Radioactivity was quantiWed using an inline radioactivity detector. The neutral compounds were analyzed in the mass detector using electro-spray ionization in positive mode. Operating conditions were optimized by Xow injection analysis of estriol (E1), E2 and estrone (E3). Conjugated estrogen metabolites were analyzed in negative ion mode using electro-spray. Operating conditions were: drying gas (N2), 8.0 L/min; drying gas temperature, 350 °C; nebulizer pressure, 40 psi; fragmentor, 80 V; and capillary voltage, 3000 V. In some cases, unknown metabolites were positively identiWed by comparing their mass spectra to puriWed commercial standards.
2.5. EVects of xenobiotics and steroids on estrogen metabolism Fifteen compounds were tested to determine their eVects on the production of conjugated E2 metabolites, including representatives of Wve classes of estrogenic xenobiotics (three BPAs, two alkylphenols, three halogenated phenols, a non-hydroxylated PBDE, and a hydroxylated PCB), and Wve key Wsh steroid hormones. The eVect of these compounds on the production of conjugated, water-soluble E2 metabolites was determined by incubating kidney and liver fragments with 70,000 dpm [3H]-E2 and 0.15 M radioinert E2, for 1 h in the presence of the test compounds. Controls were incubates without tissue (blanks), and without added test compound. All test chemicals were dissolved in 100% ethanol. Ethanol concentrations were kept constant in all cultures at 1%. Most compounds were tested at a single high dose (100 M), although two structurally distinct xenoestrogens were chosen for further characterization: 4,4⬘-(OH)23,3⬘,5,5⬘- tetrachlorobiphenyl (4,4⬘-OH-TCB), the most potent inhibitor of human SULT1E1 (Kester et al., 2000), and BPA, an important xenoestrogen in the Great Lakes, and a known inhibitor of E2 metabolism in Wsh (Ohkimoto et al., 2003). Complete dose–response experiments (0.1– 100 M) were conducted with these compounds to determine their 50% inhibition concentrations (IC50). Water-soluble metabolites were isolated from neutral steroid metabolites by extracting the cultures (medium plus tissue) twice with diethyl ether (5 £ volume). The fractions were separated by freezing the aqueous layer in a dry ice/ethanol bath, and pouring oV the ether. After extraction, the tissue was dissolved by adding 100 l of 1 N NaOH and incubating at 70 °C for 1 h. The samples were neutralized with 100 l of 1 N HCl, and the production of total water-soluble E2 metabolites (conjugates) was quanti-
275
Wed by measuring the radioactivity in a 120 l sub-sample by liquid scintillation counting. The remaining medium and solubilized tissue was diluted tenfold with distilled water, and total protein measured using a commercial Bradford kit (Pierce, Rockford, IL). The fate of E2 in the presence of inhibitory xenobiotics and steroids was not determined by LC–MS.
2.6. Data analysis All inhibition data were corrected for background radioactivity estimated in the blanks. Inhibition data are expressed as percent control after the data were normalized for protein concentrations. Results are expressed as means § SEM. Dose–response data were analyzed by ANOVA, and diVerences in E2 metabolism between control and treatment groups were compared using least squares means analysis at P < 0.01. The IC50 values were calculated using a curve-Wtting program (GraphPad Prism 4.0, San Diego, CA).
3. Results 3.1. IdentiWcation of estrogen metabolites The quantitatively most important E2 metabolite produced by both the lake trout kidney and liver was a singlehydroxylated compound with a molecular mass of 288 (Figs. 1 and 2, Table 1, peak 1, note that peak 3, which appears larger, consists of two co-eluting compounds). This compound, designated E2-OH, was not positively identiWed, although LC–MS analysis with known standards showed that it was not estriol (16-OH-E2). In both tissues, this metabolite appeared after 1 h of incubation, and its levels steadily increased with time (Figs. 1–3, Table 1). The production of E2-OH was higher in the liver than the kidney (Figs. 1–3, Table 1). The kidney and liver produced several major water-soluble E2 metabolites that were positively identiWed by LC–MS (Figs. 1 and 2, Table 1). Estradiol-3-glucuronide (E2-3-G) eluted with a retention time of approximately 13.8 min (Figs. 1 and 2, Table 1, peak 2). Estradiol-17-glucuronide (E2-17-G) and estradiol-17sulfate (E2-17-S) co-eluted with a retention time of approximately 14.9 min (Figs. 1 and 2, Table 1, peak 3); mass spectroscopy analysis showed that E2-17-G and E2-17-S were produced in approximately equal concentrations by both tissues (MS data not shown). All three of these compounds (i.e., E2-3-G, E2-17-G, and E2-17-S) were produced rapidly and appeared after 1 h of incubation in both tissues (Figs. 1–3, Table 1, peaks 2 and 3). In the kidney, the levels of these conjugated E2 metabolites all increased with time, whereas in the liver, their levels remained constant after 1 h (Fig. 3). The kidney and liver also produced several other conjugated metabolites, including the three possible sulfates of E2OH each with a molecular mass of 368 (i.e., sulfate substitutions at positions 3, 17, and the position of the unknown hydroxyl), and the three possible glucuronides of E2-OH each with a molecular mass of 464 (i.e., glucuronic acid substitutions at positions 3, 17, and the position of the unknown hydroxyl). The three E2-OH sulfates (E2-OH-S) co-eluted with a retention time of approximately 9.4 min (Figs. 1 and 2, Table 1, peak 4). The three E2-OH glucuronides (E2-OH-G)
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0h
1h
E2
CPMs
E2
1
5
10
15
20
25
30
5
2
10
3
15
20
25
30
5
E2
10
15 20 25 Time (min)
30
3
4h
24 h
CPMs
E2
1
2
1 3
4
2
5
10
15 20 25 Time (min)
30
5
Fig. 1. Estradiol-17 (E2) metabolism by the lake trout kidney. Data shown are the counts per minute (CPMs) of radiochromatograms of the E2 metabolites produced by incubating lake trout kidney tissue for 0, 1, 4, and 24 h in the presence of 106 dpm [3H]-E2 at 15 °C. The E2 metabolites were separated by HPLC, quantiWed by an inline radiation detector and identiWed by mass specroscopy. Peak 1 is an unidentiWed metabolite of E2 that is hydroxylated at an unknown position. This compound, designated E2-OH, is not estriol (16-OH-E2). Peak 2 is E2-3-glucoronide. Peak 3 consists of two co-eluting compounds produced in approximately equal concentrations: E2-17-glucoronide and E2-17-sulfate. Peak 4 consists of three co-eluting compounds produced in approximately equal concentrations: The three possible sulfates of E2-OH (i.e., E2-OH sulfonated on positions 3, 17, and the position of the unknown hydroxyl group). Peak 5 consists of three co-eluting compounds produced in approximately equal concentrations: The three possible glucoronides of E2OH (i.e., E2-OH glucoronidated on positions 3, 17, and the position of the unknown hydroxyl group). See Table 1 for additional information on peaks 1–5.
co-eluted with a retention time of approximately 10.2 min (Figs. 1 and 2, Table 1, peak 5). The liver produced E2-OH-G and E2-OH-S more rapidly than the kidney. 3.2. EVects of xenobiotics and steroids on estrogen metabolism In 1 h control cultures, the kidney and liver converted 39 § 4% and 15 § 5% of the added E2 into water-soluble, conjugated metabolites, respectively. E2 metabolism was signiWcantly inhibited in both kidney and liver by 100 M of BPA, TC-BPA, TB-BPA, PCP, TCP, TBP, and 4,4⬘-OH-TCB (Table 2). The alkylphenol NP, and TBDE, the only nonhydroxylated xenobiotic tested, had no signiWcant eVect on E2 metabolism by either tissue, and the alkylphenol OP inhibited E2 metabolism by the liver, but not the kidney (Table 2). Incubation of kidney and liver with 100 M E2 signiWcantly reduced the conversion of [3H]-E2 to water-soluble, conjugated metabolites, as expected (Table 2). Testosterone and 17,20-dihydroxy-4-pregnen-3-one (17,20P) inhibited the production of water-soluble E2 conjugates by both the kidney and liver. Cortisol and 11-ketotestosterone inhibited
the production of water-soluble E2 conjugates by the liver, but had no signiWcant eVect on the kidney (Table 2). In kidney and liver, both 4,4⬘-OH-TCB and BPA inhibited the production of water-soluble E2 conjugates in a dosedependent manner, although both compounds were more potent in the liver (Fig. 4). In three independent experiments IC50’s for 4,4⬘-OH-TCB were 32, 28, and 7 M in the kidney, and 1.6, 1.5, and 0.6 M in the liver; IC50’s for BPA were 94, 108, and 40 M for the kidney, and 11, 18, and 11 M for the liver. Low doses (0.1 M) of both 4,4⬘-OH-PCB and BPA signiWcantly increased E2 metabolism in the kidney (Fig. 4). The wide range of IC50’s is likely due to the fact that tissue cultures, and not puriWed subcellular preparations, were used to conduct the dose–response experiments. 4. Discussion 4.1. Estrogen metabolites A neutral steroid with a single hydroxyl substitution (OH-E2) was quantitatively the most important E2 metabolite produced by both lake trout kidney and liver tissue.
G.F. Jurgella et al. / General and Comparative Endocrinology 148 (2006) 273–281
0h
1h
E2
CP M s
E2
277
1
5
10
15
20
25
30
5
2
10
3
15
20
25
30
25
30
1
4h
24 h
CPM s
E2
1
3
3 4 5
45
2
2 E2
5
10
15 20 Time (min)
25
30
5
10
15 20 Time (min)
Fig. 2. Estradiol-17 (E2) metabolism by the lake trout liver. Data shown are radiochromatograms of all neutral and conjugated E2 metabolites produced by incubating lake trout liver tissue for 0, 1, 4, and 24 h in the presence of 106 dpm [3H]-E2 at 15 °C. See Table 1 for the identity of peaks 1–5.
Table 1 Retention times, molecular weights, and identiWcation of HPLC peaks shown in Figs. 1 and 2 Peak no.
Retention time (min)
Molecular weighta
Compound name
Notes
1 2 3
12 13.8 14.9 9.4
5
10.2
464
OH-estradiol Estradiol-3-glucuronide Estradiol-17-glucuronide Estradiol-17-sulfate OH-estradiol-3-sulfate OH-estradiol-17-sulfate OH-estradiol-?-sulfateb OH-estradiol-3-glucuronide OH-estradiol-17-glucuronide OH-estradiol-?-glucuronide
Estradiol-17 hydroxylated at an unknown position
4
288 448 448 352 368
OH-estradiol-sulfonated at positions 3, 17, or ?
OH-estradiolglucuronidated at positions 3, 17, or ?
Peak 1 is an unidentiWed, single hydroxylated metabolite of estradiol-17 that is not estriol (i.e., the OH is not on position 16). Mass spectroscopy showed that peaks 3, 4, and 5 actually consist of two or three co-eluting compounds that were all produced in approximately equal concentrations. a Molecular weight was determined by mass spectroscopy. b Compounds designated with a question mark (?) are conjugated at the position of the unknown hydroxyl group of OH-estradiol.
LC–MS analysis indicated that this metabolite was not estriol (16-OH-E2), and it could possibly be 7-OH-E2 or 2-OH-E2 (catecholestrogen), two major E2 metabolites previously reported in Wsh (Hansson and Rafter, 1983; Klotz et al., 1986; Snowberger and Stegman, 1987). In mammals, some hydroxylated E2 metabolites (e.g., OH at positions 6 or 7) are inactivation products, whereas others (e.g., OH at position 2) have important biological functions distinct from E2 (Zhu and Conney, 1998). The biological role of E2-OH may be clariWed once its identity is established.
Other important products of lake trout renal and hepatic E2 metabolism were E2-3-G, E2-17-G, and E2-17S, indicating that both SULT and UGT enzymes are active in the lake trout kidney and liver. Secondary metabolites with molecular masses of 368 (E2-OH plus SO4) and 464 (E2-OH plus glucuronic acid) were produced in longerterm incubations, indicating that E2-OH is a substrate for SULT and UGT, and/or that E2 conjugates, such as E2-3G, may be substrates for hydroxylation. In general, steroid conjugation reactions, such as those observed in the lake trout kidney and liver, biologically
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Table 2 The eVects of xenobiotics and steroids on the formation of water-soluble estrogen metabolites by the liver and kidney of lake trout Compound
Abbreviation
E2 metabolism (% of control) Kidney
Liver
Bisphenol As Bisphenol A Tetrachlorobisphenol A Tetrabromobisphenol A
BPA TC-BPA TB-BPA
52.0 § 2.0¤¤ 20.2 § 2.8¤¤ 31.0 § 9.1¤¤
24.2 § 4.3¤¤ 18.7 § 6.8¤¤ 17.5 § 0.5¤¤
Alkylphenols 4-n-Nonylphenol 4-Octylphenol
NP OP
100 § 4.4 85.5 § 3.8
80.0 § 24.4 69.5 § 5.1¤¤
Halogenated phenols Pentachlorophenol 2,4,6-Trichlorophenol 2,4,6-Tribromophenol
PCP TCP TBP
72.0 § 8.6¤¤ 34.7 § 2.6¤¤ 41.7 § 2.9¤¤
49.0 § 5.8¤¤ 19.2 § 2.1¤¤ 23.8 § 2.4¤¤
Polybrominated diphenyl ether 2,2⬘,4,4⬘-Tetrabromodiphenyl ether
TBDE
86.2 § 9.24
Hydroxylated Polychlorinated biphenyl 4,4⬘-(OH)2-3,3⬘,5,5⬘-tetrachlorobiphenyl
4,4⬘-OH-TCB
21.2 § 3.4¤¤
15.2 § 3.4¤¤
Steroid hormones Testosterone 11-ketotestosterone 17,20-dihydroxy-4-pregnen-3-one Cortisol 17-Estradiol (positive control)
T 11-KT 17,20P Cortisol E2
46.5 § 8.5¤¤ 83.3 § 6.9 53.1 § 3.6¤¤ 78.7 § 7.6 12.4 § 2.0¤¤
61.1 § 8.7¤¤ 68.2 § 6.4¤¤ 55.1 § 3.3¤¤ 59.2 § 2.5¤¤ 46.0 § 7.3¤¤
109.5 § 6.6
Liver and kidney tissues of lake trout were incubated in vitro for 1 h in the absence (control) or presence of 100 M of the indicated xenobiotics or steroids. The culture medium was extracted to remove neutral steroids, and the radioactivity remaining in the medium was quantiWed by LSC. Estrogen metabolism is expressed as percent of control. Data are means § SEM (N D 4). Asterisks indicate signiWcant diVerence from control at P 6 0.01.
inactivate E2, and water-soluble conjugates are normally excreted via the urine (Roy, 1992). Some conjugated E2 metabolites, however, circulate in the blood at high levels, and serve as a reservoir of circulating E2 that is released by the actions of sulfatases and glucuronidases in peripheral target tissues (Kirk et al., 2003; Zhu and Conney, 1998). In mammals, certain hydroxylated and conjugated E2 metabolites, similar to those identiWed in lake trout (i.e., E2-OH-S and E2-OH-G), have important non-hormonal eVects (Dubey and Jackson, 2001). In the human placenta, for example, 2-hydroxyestradiol-17-sulfate (2-OH-E2-17-S) is a major product of E2 metabolism, and a long-lasting and powerful antioxidant essential to the maintenance of healthy pregnancy (Takanashi, 2003). Further research is necessary to determine if the E2 metabolites produced by the lake trout kidney and liver play biological roles besides steroid inactivation and excretion. 4.2. EVects of xenobiotics and steroids E2 metabolism by the lake trout kidney and liver was inhibited by various xenobiotics, indicating that some endocrine-disrupting chemicals present in the Great Lakes may act in Wsh by increasing the bioavailability of endogenous hormones through inhibition of hormone-metabolizing enzymes. The mechanism by which these compounds act to inhibit E2 conjugation is not known, although it is likely that they are either competitive or allosteric inhibitors. Hydroxyl-
ated xenobiotics are known substrates for Phase II conjugation (Clark et al., 1991; Yokota et al., 1999), and in the present study, only hydroxylated chemicals inhibited SULT and UGT supporting the hypothesis that xenoestrogens act as competitive inhibitors of E2 conjugation, as reported in zebraWsh (Ohkimoto et al., 2003). On the other hand, Kester et al. (2000, 2002) showed that various xenobiotics inhibit human SULT1E1 by acting as non-competitive, allosteric inhibitors. Moreover, van den Hurk et al. (2002) could not rule out the possibility that non-competitive inhibition was the mechanism by which various PCB-OHs inhibited SULT and UGT activities in the catWsh intestine. The most potent inhibitors of human SULT1E1 were co-planar PCB-OHs that had hydroxyl groups Xanked by halogens, and Kester et al. (2000) speculated that the coplanar structure mimics the primarily planar structure of E2, and that the adjacent halogen substitutions increase relative potency by increasing the dissociation of the OH group. Similarly, in the present study, all of the inhibitory compounds, except BPA, were halogenated, and the most potent inhibitors had hydroxyl groups Xanked by halogens, suggesting that xenobiotic chemicals may have a similar mechanism of action in humans and lake trout. The hydroxylated PCB metabolite 4,4⬘-OH-TCB inhibited human hepatic SULT1E1 with an IC50 value of approximately 0.1 nM (Kester et al., 2000), much lower than observed in the lake trout liver. There are several possible explanations for the apparent higher sensitivity in
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Kidney 100
3,3',5,5'-PCB-4,4'-OH
Water-Soluble E2 Metabolites (% of control)
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5: E2-OH-G 4: E2-OH-S 3: E2-17-G and E2-17-S 2: E2-3-G 1: E2-OH
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**
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**
**
0
24
**
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Bisphenol A
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40 20 0 0
1
4
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140
24
Time (hrs) Fig. 3. Estradiol-17 (E2) metabolism by the lake trout kidney (upper graph) and liver (lower graph). Data shown are the integrated peaks from Figs. 1 and 2, and depict the percentage of total E2 metabolites represented by each peak at each time point. See Table 1 for the identity of peaks 1–5. E2-17-G, estradiol-17-glucuronide; E2-17-S, estradiol-17-sulfate, etc.
human vs. lake trout. The human studies were conducted using puriWed cytosolic extracts, whereas the lake trout experiments were conducted using primary tissue cultures. Thus, 4,4⬘-OH-TCB, a highly lipophilic compound, may have become bound to cellular membranes eVectively reducing its bioavailability to the lake trout enzymes. In addition, the lake trout liver expresses both SULT and UGT activities, and therefore, higher concentrations of 4,4⬘-OH-TCB may be required to simultaneously inhibit both enzymes. Finally, 4,4⬘-OH-TCB could act by diVerent mechanisms of action in the two species—allosterically in human (Kester et al., 2000), and competitively in lake trout. An explanation for the observation that low doses of 4,4⬘OH-TCB and BPA increased E2 metabolism in the lake trout kidney is that low doses may allosterically potentiate SULT and UGT activities, whereas higher doses may act as competitive inhibitors. In contrast to 4,4⬘-OH-TCB, BPA inhibited both human hepatic SULT1E1 and lake trout hepatic E2 conjugation with similar IC50 values of approximately 10 M (Kester et al., 2000). Human hepatic SULT1E1 was inhibited by the halogenated BPAs, 3,3⬘,5,5⬘-tetrachlorobisphenol A (TC-BPA) and 3,3⬘,5,5⬘tetrabromobisphenol A (TB-BPA), via a non-competitive mechanism with IC50 values of 29–53 nM and 12–33 nM, respectively, illustrating the importance of halogen substitutions to overall xenobiotic potency in humans (Kester
Water-Soluble E2 Metabolites (% of control)
100
**
120 100 80 60
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[uM] Fig. 4. Dose–response eVects of 3,3⬘,5,5⬘-tetrachloro-4,4⬘-biphenyl-diol (4,4⬘-OH-TCB, upper graph) and bisphenyl A (BPA, lower graph) on the production of water-soluble (conjugated) E2 metabolites produced by the kidney and liver of lake trout incubated for 1 h in the presence of 106 dpm [3H]-E2 at 15 °C. E2 and all neutral metabolites were extracted with diethyl ether, and the radioactivity remaining in the incubation medium was quantiWed. Data shown are mean § SEM (N D 3), and are expressed as percentage of E2 metabolism in the control incubations without added inhibitor. Asterisks indicate values signiWcantly diVerent from control at P 6 0.01.
et al., 2002). Although IC50 values were not determined for TC-BPA and TB-BPA in lake trout, 100 M doses of both compounds inhibited E2 metabolism more than 100 M BPA, suggesting that halogen substitutions may also increase xenobiotic potency in Wsh. Both 4,4⬘-OH-TCB and BPA were more potent inhibitors of E2 conjugation in the liver than the kidney. This may be because the liver has approximately 50% of the E2 conjugating activity as the kidney, and therefore less inhibitor is required to block the limited enzyme activity in the liver. The steroids, T and 17,20P both inhibited the production of water-soluble E2 metabolites by both the kidney and liver. The biological signiWcance of steroid regulation of E2 metabolism is unclear, but suggests that circulating E2 levels and/or E2 accessibility to estrogen target tissues could possibly be regulated by the reproductive state of the Wsh. For example, elevated T levels associated with reproductive maturation may indirectly increase the biopotency of estrogen by inhibiting E2 metabolism by the kidney, liver, and perhaps other estrogen targets. It is unclear why
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11-ketotestosterone and cortisol were only eVective in the liver but not the kidney, but this may also be explained by the observed tissue diVerences in E2 metabolic activity. 5. Concluding remarks This study reports that various xenobiotics found in the aquatic environment have the ability to inhibit E2 metabolism by the lake trout kidney and liver. These results add to the growing literature indicating that many xenoestrogen endocrine disruptors act by inhibiting E2 metabolism in Wsh (Kirk et al., 2003; Ohkimoto et al., 2003; van den Hurk et al., 2002). The signiWcance of this regulation is not fully understood, but may be key to understanding many biological phenomena thought to be regulated by environmental estrogens. Xenobiotics, for example, could increase circulating endogenous E2 concentrations by disrupting normal inactivating pathways. Alternatively, xenobiotics may act directly at the level of the target tissue by inhibiting metabolizing enzymes that normally protect the ER from binding circulating E2. Inhibitory xenobiotics may prevent the formation of important biologically active E2 metabolites, such as antioxidants or biologically active catecholestrogens (Zhu and Conney, 1998). Finally, inhibitory xenobiotics may switch normal E2 metabolic pathways in alternative directions. It has been reported in mammals, for example, that certain phytoestrogens can increase the activity of 2-hydroxylase and decrease the activity of 16-hydroxylase, an alteration associated with a reduced risk of estrogen-mediated cancers (Dubey and Jackson, 2001). Acknowledgments This work was funded by a grant to TB and RP from the University of Wisconsin Sea Grant College Program, National Oceanic and Atmospheric Administration, US Department of Commerce (Federal Grant NA16RG2257; Project No. R/BT-19). Support also came from the University of Wisconsin-Madison College of Agricultural and Life Sciences and School of Natural Resources, and the Wisconsin Department of Natural Resources. The Minnesota Department of Natural Resources donated the lake trout used in this study. Ultra ScientiWc (North Kingstown, RI) donated xenobiotics. We acknowledge the assistance of Colin Jefcoate, Milo Wiltbank, James A. Held, Luciene Lima, Yuliana Ng, Robert Moore, Henry Lardy, and Padma Marwah. References Barry, T.P., Riebe, J., Parrish, J.J., Malison, J.A., 1997. EVects of 17,20dihydroxy-4-pregnen-3-one on cortisol production by rainbow trout interrenal tissue in vitro. Gen. Comp. Endocrinol. 107, 172–181. Clark, D.J., George, S.G., Burchell, B., 1991. Glucuronidation in Wsh. Aquat. Toxicol. 20, 35–56. Colborn, T., von Saal, F.S., Soto, A.M., 1993. Developmental eVects of endocrine-disrupting chemicals in wildlife and humans. Environ. Health Perspect. 101, 378–384.
Cook, P.M., Robbins, J., Endicott, D.D., Lodge, K., Guiney, P.D., Walker, M.K., Zabel, E.W., Peterson, R.W., 2003. EVects of aryl hydrocarbon receptor mediated early life stage toxicity on lake trout populations in Lake Ontario during the 20th century. Environ. Sci. Technol. 37, 3864– 3877. deWit, C.A., 2002. An overview of brominated Xame retardants in the environment. Chemosphere 46, 583–624. Dubey, R.K., Jackson, E.K., 2001. Cardiovascular protective eVects of 17 beta-estradiol metabolites. J. Appl. Physiol. 91, 1868–1883. Gilbertson, M., 1992. PCB and dioxin research and implications for Wsheries research and resource management. Fisheries 17, 26–27. Hale, R.C., Alaee, M., Manchester-Neesvig, J.B., Stapleton, H.M., Ikonomou, M.G., 2003. Polybrominated diphenyl ether Xame retardants in the North American environment. Environ. Int. 29, 771–779. Hansson, T., Rafter, J., 1983. In vitro metabolism of estradiol-17 by liver microsomes from juvenile rainbow trout, Salmo gairdneri. Gen. Comp. Endocrinol. 49, 490–495. Humphrey, H.E.B., Gardiner, J.C., Pandya, J.R., Sweeney, A.M., Gasior, D.M., McCaVrey, R.J., Schantz, S.L., 2000. PCB congener proWle in the serum of humans consuming Great Lakes Wsh. Environ. Health Perspect. 108, 167–172. Jobling, S., Reynolds, T., White, R., Parker, M.G., Sumpter, J.P., 1995. A variety of environmentally persistent chemicals, including some phthalate plasticizers, are weakly estrogenic. Environ. Health Perspect. 103, 582–587. Kester, M.H.A., Bulduk, S., Tibboil, D., Meinl, W., Glatt, H., Falany, C.N., Coughtrie, M.W.H., Bergman, A., Safe, S.H., Kuiper, G.G.J.M., Schuur, A.G., Brouwer, A., Visser, T.J., 2000. Potent inhibition of estrogen sulfotransferase by hydroxolated PCB metabolites: a novel pathway explaining the estrogenic activity of PCBs. Endocrinology 141, 1897–1900. Kester, M.H.A., Bulduk, S., van Toor, H., Tibboel, D., Meinl, W., Glatt, H., Falany, C.N., Coughtrie, M.W.H., Schuur, A.G., Brouwer, A., Visser, T.J., 2002. Potent inhibition of estrogen sulfotransferase by hydroxylated metabolites of polyhalogenated aromatic hydrocarbons reveals alternative mechanism for estrogenic activity of endocrine distrupters. Clin. Endocrinol. 87, 1142–1150. Kirk, C.J., Bottomley, L., Minican, N., Carpenter, H., Shaw, S., Kohli, N., Winter, M., Taylor, E.W., Waring, R.H., Michelangeli, F., Harris, R.M., 2003. Environmental endocrine disrupters dysregulate estrogen metabolism and Ca2+ homeostasis in Wsh and mammals via receptor-independent mechanisms. Comp. Biochem. Phys. (Part A) 135, 1–8. Klotz, A.V., Stegeman, J.J., Woodin, B.R., Snowberger, E.A., Thomas, P.E., Walsh, C., 1986. Cytochrome p-450 isozymes from the marine teleost Stenotomus chrysops: their role in steroid hydroxylation and the inXuence of cytochrome b5. Arch. Biochem. Biophys. 249, 326–338. Korach, K.S., Sarver, P., Chae, K., McLacklan, J.A., McKinney, J.D., 1988. Estrogen receptor-binding activity of polychlorinated hydroxybiphenyls: conformationally restricted structural probes. Mol. Pharmacol. 33, 120–126. Kuiper, G.G.J.M., Lemmen, J.G., Carlsson, B., Corton, J.C., Safe, S.H., Van der Saag, P.T., Van Der Burg, B., Gustafsson, J.A., 1998. Interaction of estrogenic chemicals and phytoestrogens with estrogen receptor . Endocrinology 139, 4252–4263. Marwah, A., Marwah, P., Lardy, H., 2002. Metabolism of dehydroepiandrosterone by rat liver in vitro: an LC–MS study. J. Chromatogr. B. 76, 285–299. McLachland, J.A., Arnold, S.F., 1996. Environmental estrogens. Am. Sci. 84, 452–461. Ohkimoto, K., Sugahara, T., Sakakibara, Y., Suiko, M., Liu, M.Y., Carter, G., Liu, M.C., 2003. Sulfonation of environmental estrogens by zebraWsh cytosolic sulfotransferases. Biochem. Biophys. Res. Commun. 309, 7–11. Peterson, R.E., Theobald, H.M., Kimmel, G.L., 1993. Developmental and reproductive toxicity of dioxins and related compounds: cross-species comparisons. Crit. Rev. Toxicol. 23, 283–335. Roy, A.K., 1992. Regulation of steroid hormone action in target cells by speciWc hormone-inactivating enzymes. Proc. Soc. Exp. Biol. Med. 199, 265–272.
G.F. Jurgella et al. / General and Comparative Endocrinology 148 (2006) 273–281 Safe, S.H., 1994. Polychlorinated biphenyls (PCBs): environmental impact, biochemical and toxic responses, and implications for risk assessment. Crit. Rev. Toxicol. 24, 87–149. Snowberger, E.A., Stegman, J.J., 1987. Patterns and regulation of estradiol metabolism by hepatic microsomes from two species of marine telosts. Gen. Comp. Endocrinol. 66, 256–265. Takanashi, K., 2003. Studies on 2-hydroxyestradiol 17-sulfate derived from fetoplacental unit: the antioxidant as a potential defense substance against preeclampsia. Yakugaku Zasshi. 123, 399–411. van den Hurk, P., Kubiczak, G.A., Lehmler, H.-J., James, M.O., 2002. Hydroxylated polychlorinated biphenyls as inhibitors of the sulfation and glucuronidation of 3-hydroxy-benzo[a]pyrene. Environ. Health Perspect. 110, 343–348.
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EVects of xenobiotics and steroids on renal and hepatic estrogen metabolism in lake trout Gail F. Jurgella a, Ashok Marwah b, JeVrey A. Malison a, Richard Peterson c, Terence P. Barry a,¤ b
a Department of Animal Sciences, University of Wisconsin-Madison, 1675 Observatory Drive, Madison, WI 53706, USA Department of Biochemistry—Enzyme Institute, University of Wisconsin-Madison, 1710 University Avenue, Madison, WI 53726, USA c School of Pharmacy, University of Wisconsin-Madison, 777 Highland Avenue, Madison, WI 53705, USA
Received 15 December 2005; revised 13 March 2006; accepted 23 March 2006 Available online 4 May 2006
Abstract Experiments were conducted to (1) elucidate the biochemical pathways of E2 metabolism in the lake trout (Salvelinus namaycush) kidney and liver, and (2) test the hypothesis that speciWc xenobiotics and endogenous steroids inhibit E2 metabolism by these tissues. Kidney and liver tissue fragments from immature lake trout were incubated in vitro in the presence of radiolabelled E2 plus various xenobiotics or steroids. E2 metabolites were identiWed by liquid chromatography/mass spectroscopy, and quantiWed by liquid scintillation spectroscopy. A major metabolite produced by both tissues was an unidentiWed hydroxylated estrogen metabolite (E2-OH) with a molecular mass of 288 that was not estriol (16-OH-E2), but possibly 7-OH-E2 or 2-OH-E2 (catecholestrogen). Both tissues also produced estradiol-17glucuronide (E2-17-G), estradiol-17-sulfate (E2-17-S), and estradiol-3-glucuronide (E2-3-G). Compared to the kidney, the liver produced half the amount of conjugated metabolites, but twofold more E2-OH. The following xenobiotics (at a concentration of 100 M) inhibited the production of water-soluble (i.e., conjugated) E2 metabolites by both the kidney and liver: 4,4⬘-(OH)2-3,3⬘,5,5⬘- tetrachlorobiphenyl (4,4⬘-OH-TCB), bisphenol A (BPA), tetrabromobisphenol A (TB-BPA), tetrachlorobisphenol A (TC-BPA), tribromophenol (TBP), trichlorophenol (TCP), and pentachlorophenol (PCP). The alkylphenols, 4-n-nonylphenol (NP), and 4-octylphenol (OP), and 2,2⬘,4,4⬘tetrabromodiphenyl ether (TBDE) had no signiWcant eVect on E2 metabolism by either tissue. Testosterone and 17,20-dihydroxy-4pregnen-3-one inhibited the production of conjugated E2 metabolites by both the kidney and liver. Cortisol and 11-ketotestosterone inhibited E2 metabolism by the liver only. The median inhibitory concentrations (IC50) for 4,4⬘-OH-TCB ranged from 7–32 M in the kidney and 0.6–1.6 M in the liver. For BPA, IC50’s ranged from 40–108 M in the kidney and 11–18 M in the liver. Low doses (0.1 M) of 4,4⬘-OH-TCB and BPA signiWcantly increased estrogen metabolism in the kidney. The results suggest that certain estrogenic xenobiotics and endogenous steroids may inhibit the phase II conjugation of E2 by the kidney and liver of lake trout, and some of the known biological eVects of these compounds are likely mediated, at least partially, by this mechanism of action. © 2006 Elsevier Inc. All rights reserved. Keywords: Lake trout; Endocrine disruption; Steroidogenic enzymes; Estrogen; Xenobiotic; Steroid hormone; Estradiol-17; Steroid; Steroid metabolism
1. Introduction The lake trout (Salvelinus namaycush) is the top natural predator in the Great Lakes food web. During the twentieth century, lake trout populations dropped dramatically in Lake Ontario, and it was thought that this decline was the *
Corresponding author. Fax: +1 608 262 0454. E-mail address: [email protected] (T.P. Barry).
0016-6480/$ - see front matter © 2006 Elsevier Inc. All rights reserved. doi:10.1016/j.ygcen.2006.03.011
result of sea lamprey predation, declining water quality, and over harvesting. Attempts to reestablish naturally reproducing lake trout populations focused primarily on improving these problems, largely without success (Gilbertson, 1992). Later studies suggested that xenobiotic chemicals present in the water might be the principal cause of lake trout recruitment failure. For example, Walker and Peterson (1994) demonstrated that lake trout larvae are highly sensitive to 2,3,7,8-tetrachlorodibenzo-p-dioxin
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(TCDD), and even very low-level exposure to TCDD causes early life stage mortality. Further studies demonstrated that the concentrations of TCDD and related chemicals in Lake Ontario during the 1960s were high enough to explain the observed decline in lake trout populations (Cook et al., 2003). Despite bans on the use of TCDD and other xenobiotics such as polychlorinated biphenyls (PCBs), these lipophilic compounds continue to persist in lake sedimentation basins, and in the tissues of contaminated Wsh, particularly the top trophic-level predators (Humphrey et al., 2000; Safe, 1994). Moreover, the levels of other xenobiotics that are potentially harmful to lake trout, including polybrominated diphenyl ethers (PBDEs), are currently increasing in the Great Lakes due to their continued production and use by industry (deWit, 2002; Hale et al., 2003). Although the levels of most xenobiotics in the Great Lakes and in Great Lakes Wshes are currently too low to cause outright early life stage mortality, they may still reduce lake trout recruitment by acting as endocrine disruptors. Certain PCBs and hydroxylated PCB metabolites are known to modify physiological functions in estrogen-responsive targets such as the reproductive organs, liver, and neuroendocrine centers (Jobling et al., 1995; McLachland and Arnold, 1996; White et al., 1994). Even weakly active compounds may disturb the diVerentiation or function of estrogen-responsive tissues if exposure occurs during critical developmental periods (Colborn et al., 1993). Developing Wsh embryos and larvae are particularly at risk because many xenobiotics can be transferred from the mother to the developing eggs (Peterson et al., 1993; Walker and Peterson, 1994). The mechanisms by which xenoestrogens exert their endocrine-disrupting eVects are not well established. Many potent hydroxylated PCB metabolites (PCB-OHs), for example, show little or no aYnity for the - or -estrogen receptor (ER) subtypes, and thus may not act directly as estrogen receptor agonists (Korach et al., 1988; Kuiper et al., 1998). In humans, Kester et al. (2000) showed that numerous PCB-OHs are powerful inhibitors of the phase II conjugating enzyme, estrogen sulfotransferase (SULT1E1), and indirectly induce estrogenic activity by reducing the metabolism of endogenous estrogen. Indeed, the inhibition of SULT1E1 is the most potent biological eVect described to date regarding the endocrine-disrupting activity of PCBs or their metabolites (Kester et al., 2000). The most potent PCB-OHs, with IC50 values in the subnanomolar range, were compounds with a 4-OH-3,5-dichloro substitution pattern (Kester et al., 2000). In subsequent studies, other xenobiotic compounds with known estrogenic activity, including halogenated bisphenol A (BPA) metabolites, were also shown to be potent inhibitors of SULT1E1 (Kester et al., 2002). Xenobiotic inhibition of phase II conjugating enzymes may also be an important mechanism of action of endocrine disruption in Wsh. For example, Ohkimoto et al. (2003) reported that BPA, 4-n-nonylphenol (NP), and 4-n-octylphenol (OP) inhibited the sulfonation of estradiol-17 (E2) in
zebraWsh (Brachydanio rerio), and van den Hurk et al. (2002) found that several PCB-OHs inhibited both the sulfonation and glucuronidation of benzo[a] pyrene in channel catWsh (Ictalurus punctatus), indicating that both sulfotransferase (SULT), and glucuronosyltransferase (UGT) are targets for inhibition by xenobiotics in Wsh. The goals of the present investigation were to (1) begin characterizing the biochemical pathways of E2 metabolism in the kidney and liver of lake trout, and (2) evaluate the eVects of various estrogenic xenobiotics, as well as several key Wsh steroid hormones, on the production of conjugated, water-soluble estrogen metabolites normally produced by the lake trout kidney and liver. 2. Materials and methods 2.1. Chemicals Steroids were obtained from Sigma Chemical Co. (St. Louis, MO) or Steraloids (Newport, RI). [6,7-3H]-E2 (40 to 60 Ci/mmol) was purchased from Perkin-Elmer Life Sciences (Boston, MA). 4,4⬘-OH-TCB was obtained from Ultra ScientiWc (North Kingstown, RI). BPA, OP, and NP were obtained from Sigma (St. Louis, MO). Tetrachlorobisphenol A (TC-BPA), tetrabromobisphenol A (TB-BPA), PCP, 2,4,6-trichlorophenol (TCP), 2,4,6tribromophenol (TBP), and 2,2⬘,4,4⬘-tetrabromodiphenyl ether (TBDE) were gifts from AccuStandard (New Haven, CT). All other chemicals were obtained from Sigma Chemical Co. (St. Louis, MO) and Fisher ScientiWc (Atlanta, GA). HPLC grade methanol, ethanol, hexane, and acetonitrile were purchased from Sigma–Aldrich (Milwaukee, WI, USA) and used as such. HPLC grade acetic acid was purchased from Fisher ScientiWc (Pittsburgh, PA, USA). Formic acid was purchased from Fluka (Milwaukee, WI, USA). Distilled water was deionized and puriWed (18.25 § 0.05 M-cm) using a Nanopure water system from Barnstead International (Dubuque, IA, USA). Solid phase extraction (SPE) cartridges (Oasis-HLB, 3 cc) were obtained from Waters Corporation (Milford, MA, USA).
2.2. Fish Lake trout were obtained from the Minnesota Department of Natural Resources, Crystal Springs State Fish Hatchery (Altura, MN). Three-yearold, sexually immature Wsh, 30–50 cm in total length, were used in all experiments. The Wsh were maintained in 750 L circular Wberglass tanks supplied with Xow-through carbon-Wltered city water at 10–14 °C. Photoperiod was kept constant at 12 h light/12 h dark. The Wsh were fed once daily to satiation using Silvercup trout feed (Murray, UT). Food was withheld 24 h prior to sacriWce. Care and treatment of the Wsh were in accordance with the guidelines of the University of Wisconsin Research Animal Care Committee.
2.3. In vitro cultures Physiological saline and Hepes-buVered culture medium, developed for culturing salmonid tissues, were prepared and sterilized by ultraWltration at 0.2 m (Barry et al., 1997). Estrogen metabolism was quantiWed by measuring the formation of E2 metabolites after incubation of tissue fragments in 1 ml of culture medium in the presence of [3H]-E2, according to the method of Barry et al. (1997). In brief, the kidney and liver were uniformly fragmented using a Polytron (Brinkmann Instruments, Westbury, NY). The tissue fragments were washed 2–3 times in physiological saline, and 50 l aliquots of tissue (1.16 § 0.03 mg protein) were added to 150 £ 20 mm test tubes and incubated at 15 °C for 1–24 h. Controls were no-tissue and time 0 incubates. The cultures were terminated and extracted as described below. Experiments used tissues from one Wsh cultured in duplicate incubations, and each experiment was conducted two to four times (N D 2 to 4).
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2.4. IdentiWcation of estrogen metabolites Kidney and liver tissue fragments were incubated with 106 dpm [3H]E2 and 3.7 M radioinert E2 for 0, 1, 4, and 24 h. Estradiol-17 and its metabolites were extracted according to the method of Marwah et al. (2002). BrieXy, water (1 £ vol) and acetic acid (0.05 £ vol) were added to the cultures (medium plus tissue). The samples were centrifuged at 1500g for 5 min to pellet the tissue. The supernatants were applied to pre-conditioned (2 ml methanol followed by 2 ml water), 3 ml, solid-phase, extraction cartridges (Oasis-HLB Waters Corp., Milford, MA), and washed twice with 2 ml of 5% methanol. The metabolites were eluted with 3 ml of 100% methanol, and the methanol was evaporated under a stream of nitrogen at 40 °C. The residues were dissolved in 0.2 ml water, and analyzed directly by LC–MS as described below. No radioactivity remained associated with the tissue pellet. Chromatography of E2 and its neutral metabolites was performed on C18 analytical columns (Zorbax-SB, 3.0 £ 150 mm, 3.5 m, Mac-Mod Analytical, Inc., Chadds Ford, PA, and Novapack, 3.9 £ 75 mm, 4.0 m, Waters Corp., Milford, MA) protected with matching C18 guard columns at a Xow rate of 0.4 ml/min for Zorbax C18 column, and at a Xow rate of 0.8 ml/min for Novapack column and a temperature of 40 °C. The solvent was acetonitrile–water with a linear gradient to reach 20–45% acetonitrile at 25 min, 94% at 32 min and back to 20% at 34 min, followed by a 10 min post run time. Water-soluble conjugates were resolved on a C18 analytical column (Zorbax-SB, 3.0 £ 150 mm, 3.5 m, Mac-Mod Analytical, Inc., Chadds Ford, PA) at a Xow-rate of 0.4 ml/min using acetonitrile–water and acetonitrile containing 3% acetic acid and 3% acetic acid gradients. The gradients started with 10% acetonitrile to reach 40% at 30 min, and 96% acetonitrile at 38 min and back to 10% at 40 min. Radioactivity was quantiWed using an inline radioactivity detector. The neutral compounds were analyzed in the mass detector using electro-spray ionization in positive mode. Operating conditions were optimized by Xow injection analysis of estriol (E1), E2 and estrone (E3). Conjugated estrogen metabolites were analyzed in negative ion mode using electro-spray. Operating conditions were: drying gas (N2), 8.0 L/min; drying gas temperature, 350 °C; nebulizer pressure, 40 psi; fragmentor, 80 V; and capillary voltage, 3000 V. In some cases, unknown metabolites were positively identiWed by comparing their mass spectra to puriWed commercial standards.
2.5. EVects of xenobiotics and steroids on estrogen metabolism Fifteen compounds were tested to determine their eVects on the production of conjugated E2 metabolites, including representatives of Wve classes of estrogenic xenobiotics (three BPAs, two alkylphenols, three halogenated phenols, a non-hydroxylated PBDE, and a hydroxylated PCB), and Wve key Wsh steroid hormones. The eVect of these compounds on the production of conjugated, water-soluble E2 metabolites was determined by incubating kidney and liver fragments with 70,000 dpm [3H]-E2 and 0.15 M radioinert E2, for 1 h in the presence of the test compounds. Controls were incubates without tissue (blanks), and without added test compound. All test chemicals were dissolved in 100% ethanol. Ethanol concentrations were kept constant in all cultures at 1%. Most compounds were tested at a single high dose (100 M), although two structurally distinct xenoestrogens were chosen for further characterization: 4,4⬘-(OH)23,3⬘,5,5⬘- tetrachlorobiphenyl (4,4⬘-OH-TCB), the most potent inhibitor of human SULT1E1 (Kester et al., 2000), and BPA, an important xenoestrogen in the Great Lakes, and a known inhibitor of E2 metabolism in Wsh (Ohkimoto et al., 2003). Complete dose–response experiments (0.1– 100 M) were conducted with these compounds to determine their 50% inhibition concentrations (IC50). Water-soluble metabolites were isolated from neutral steroid metabolites by extracting the cultures (medium plus tissue) twice with diethyl ether (5 £ volume). The fractions were separated by freezing the aqueous layer in a dry ice/ethanol bath, and pouring oV the ether. After extraction, the tissue was dissolved by adding 100 l of 1 N NaOH and incubating at 70 °C for 1 h. The samples were neutralized with 100 l of 1 N HCl, and the production of total water-soluble E2 metabolites (conjugates) was quanti-
275
Wed by measuring the radioactivity in a 120 l sub-sample by liquid scintillation counting. The remaining medium and solubilized tissue was diluted tenfold with distilled water, and total protein measured using a commercial Bradford kit (Pierce, Rockford, IL). The fate of E2 in the presence of inhibitory xenobiotics and steroids was not determined by LC–MS.
2.6. Data analysis All inhibition data were corrected for background radioactivity estimated in the blanks. Inhibition data are expressed as percent control after the data were normalized for protein concentrations. Results are expressed as means § SEM. Dose–response data were analyzed by ANOVA, and diVerences in E2 metabolism between control and treatment groups were compared using least squares means analysis at P < 0.01. The IC50 values were calculated using a curve-Wtting program (GraphPad Prism 4.0, San Diego, CA).
3. Results 3.1. IdentiWcation of estrogen metabolites The quantitatively most important E2 metabolite produced by both the lake trout kidney and liver was a singlehydroxylated compound with a molecular mass of 288 (Figs. 1 and 2, Table 1, peak 1, note that peak 3, which appears larger, consists of two co-eluting compounds). This compound, designated E2-OH, was not positively identiWed, although LC–MS analysis with known standards showed that it was not estriol (16-OH-E2). In both tissues, this metabolite appeared after 1 h of incubation, and its levels steadily increased with time (Figs. 1–3, Table 1). The production of E2-OH was higher in the liver than the kidney (Figs. 1–3, Table 1). The kidney and liver produced several major water-soluble E2 metabolites that were positively identiWed by LC–MS (Figs. 1 and 2, Table 1). Estradiol-3-glucuronide (E2-3-G) eluted with a retention time of approximately 13.8 min (Figs. 1 and 2, Table 1, peak 2). Estradiol-17-glucuronide (E2-17-G) and estradiol-17sulfate (E2-17-S) co-eluted with a retention time of approximately 14.9 min (Figs. 1 and 2, Table 1, peak 3); mass spectroscopy analysis showed that E2-17-G and E2-17-S were produced in approximately equal concentrations by both tissues (MS data not shown). All three of these compounds (i.e., E2-3-G, E2-17-G, and E2-17-S) were produced rapidly and appeared after 1 h of incubation in both tissues (Figs. 1–3, Table 1, peaks 2 and 3). In the kidney, the levels of these conjugated E2 metabolites all increased with time, whereas in the liver, their levels remained constant after 1 h (Fig. 3). The kidney and liver also produced several other conjugated metabolites, including the three possible sulfates of E2OH each with a molecular mass of 368 (i.e., sulfate substitutions at positions 3, 17, and the position of the unknown hydroxyl), and the three possible glucuronides of E2-OH each with a molecular mass of 464 (i.e., glucuronic acid substitutions at positions 3, 17, and the position of the unknown hydroxyl). The three E2-OH sulfates (E2-OH-S) co-eluted with a retention time of approximately 9.4 min (Figs. 1 and 2, Table 1, peak 4). The three E2-OH glucuronides (E2-OH-G)
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0h
1h
E2
CPMs
E2
1
5
10
15
20
25
30
5
2
10
3
15
20
25
30
5
E2
10
15 20 25 Time (min)
30
3
4h
24 h
CPMs
E2
1
2
1 3
4
2
5
10
15 20 25 Time (min)
30
5
Fig. 1. Estradiol-17 (E2) metabolism by the lake trout kidney. Data shown are the counts per minute (CPMs) of radiochromatograms of the E2 metabolites produced by incubating lake trout kidney tissue for 0, 1, 4, and 24 h in the presence of 106 dpm [3H]-E2 at 15 °C. The E2 metabolites were separated by HPLC, quantiWed by an inline radiation detector and identiWed by mass specroscopy. Peak 1 is an unidentiWed metabolite of E2 that is hydroxylated at an unknown position. This compound, designated E2-OH, is not estriol (16-OH-E2). Peak 2 is E2-3-glucoronide. Peak 3 consists of two co-eluting compounds produced in approximately equal concentrations: E2-17-glucoronide and E2-17-sulfate. Peak 4 consists of three co-eluting compounds produced in approximately equal concentrations: The three possible sulfates of E2-OH (i.e., E2-OH sulfonated on positions 3, 17, and the position of the unknown hydroxyl group). Peak 5 consists of three co-eluting compounds produced in approximately equal concentrations: The three possible glucoronides of E2OH (i.e., E2-OH glucoronidated on positions 3, 17, and the position of the unknown hydroxyl group). See Table 1 for additional information on peaks 1–5.
co-eluted with a retention time of approximately 10.2 min (Figs. 1 and 2, Table 1, peak 5). The liver produced E2-OH-G and E2-OH-S more rapidly than the kidney. 3.2. EVects of xenobiotics and steroids on estrogen metabolism In 1 h control cultures, the kidney and liver converted 39 § 4% and 15 § 5% of the added E2 into water-soluble, conjugated metabolites, respectively. E2 metabolism was signiWcantly inhibited in both kidney and liver by 100 M of BPA, TC-BPA, TB-BPA, PCP, TCP, TBP, and 4,4⬘-OH-TCB (Table 2). The alkylphenol NP, and TBDE, the only nonhydroxylated xenobiotic tested, had no signiWcant eVect on E2 metabolism by either tissue, and the alkylphenol OP inhibited E2 metabolism by the liver, but not the kidney (Table 2). Incubation of kidney and liver with 100 M E2 signiWcantly reduced the conversion of [3H]-E2 to water-soluble, conjugated metabolites, as expected (Table 2). Testosterone and 17,20-dihydroxy-4-pregnen-3-one (17,20P) inhibited the production of water-soluble E2 conjugates by both the kidney and liver. Cortisol and 11-ketotestosterone inhibited
the production of water-soluble E2 conjugates by the liver, but had no signiWcant eVect on the kidney (Table 2). In kidney and liver, both 4,4⬘-OH-TCB and BPA inhibited the production of water-soluble E2 conjugates in a dosedependent manner, although both compounds were more potent in the liver (Fig. 4). In three independent experiments IC50’s for 4,4⬘-OH-TCB were 32, 28, and 7 M in the kidney, and 1.6, 1.5, and 0.6 M in the liver; IC50’s for BPA were 94, 108, and 40 M for the kidney, and 11, 18, and 11 M for the liver. Low doses (0.1 M) of both 4,4⬘-OH-PCB and BPA signiWcantly increased E2 metabolism in the kidney (Fig. 4). The wide range of IC50’s is likely due to the fact that tissue cultures, and not puriWed subcellular preparations, were used to conduct the dose–response experiments. 4. Discussion 4.1. Estrogen metabolites A neutral steroid with a single hydroxyl substitution (OH-E2) was quantitatively the most important E2 metabolite produced by both lake trout kidney and liver tissue.
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0h
1h
E2
CP M s
E2
277
1
5
10
15
20
25
30
5
2
10
3
15
20
25
30
25
30
1
4h
24 h
CPM s
E2
1
3
3 4 5
45
2
2 E2
5
10
15 20 Time (min)
25
30
5
10
15 20 Time (min)
Fig. 2. Estradiol-17 (E2) metabolism by the lake trout liver. Data shown are radiochromatograms of all neutral and conjugated E2 metabolites produced by incubating lake trout liver tissue for 0, 1, 4, and 24 h in the presence of 106 dpm [3H]-E2 at 15 °C. See Table 1 for the identity of peaks 1–5.
Table 1 Retention times, molecular weights, and identiWcation of HPLC peaks shown in Figs. 1 and 2 Peak no.
Retention time (min)
Molecular weighta
Compound name
Notes
1 2 3
12 13.8 14.9 9.4
5
10.2
464
OH-estradiol Estradiol-3-glucuronide Estradiol-17-glucuronide Estradiol-17-sulfate OH-estradiol-3-sulfate OH-estradiol-17-sulfate OH-estradiol-?-sulfateb OH-estradiol-3-glucuronide OH-estradiol-17-glucuronide OH-estradiol-?-glucuronide
Estradiol-17 hydroxylated at an unknown position
4
288 448 448 352 368
OH-estradiol-sulfonated at positions 3, 17, or ?
OH-estradiolglucuronidated at positions 3, 17, or ?
Peak 1 is an unidentiWed, single hydroxylated metabolite of estradiol-17 that is not estriol (i.e., the OH is not on position 16). Mass spectroscopy showed that peaks 3, 4, and 5 actually consist of two or three co-eluting compounds that were all produced in approximately equal concentrations. a Molecular weight was determined by mass spectroscopy. b Compounds designated with a question mark (?) are conjugated at the position of the unknown hydroxyl group of OH-estradiol.
LC–MS analysis indicated that this metabolite was not estriol (16-OH-E2), and it could possibly be 7-OH-E2 or 2-OH-E2 (catecholestrogen), two major E2 metabolites previously reported in Wsh (Hansson and Rafter, 1983; Klotz et al., 1986; Snowberger and Stegman, 1987). In mammals, some hydroxylated E2 metabolites (e.g., OH at positions 6 or 7) are inactivation products, whereas others (e.g., OH at position 2) have important biological functions distinct from E2 (Zhu and Conney, 1998). The biological role of E2-OH may be clariWed once its identity is established.
Other important products of lake trout renal and hepatic E2 metabolism were E2-3-G, E2-17-G, and E2-17S, indicating that both SULT and UGT enzymes are active in the lake trout kidney and liver. Secondary metabolites with molecular masses of 368 (E2-OH plus SO4) and 464 (E2-OH plus glucuronic acid) were produced in longerterm incubations, indicating that E2-OH is a substrate for SULT and UGT, and/or that E2 conjugates, such as E2-3G, may be substrates for hydroxylation. In general, steroid conjugation reactions, such as those observed in the lake trout kidney and liver, biologically
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Table 2 The eVects of xenobiotics and steroids on the formation of water-soluble estrogen metabolites by the liver and kidney of lake trout Compound
Abbreviation
E2 metabolism (% of control) Kidney
Liver
Bisphenol As Bisphenol A Tetrachlorobisphenol A Tetrabromobisphenol A
BPA TC-BPA TB-BPA
52.0 § 2.0¤¤ 20.2 § 2.8¤¤ 31.0 § 9.1¤¤
24.2 § 4.3¤¤ 18.7 § 6.8¤¤ 17.5 § 0.5¤¤
Alkylphenols 4-n-Nonylphenol 4-Octylphenol
NP OP
100 § 4.4 85.5 § 3.8
80.0 § 24.4 69.5 § 5.1¤¤
Halogenated phenols Pentachlorophenol 2,4,6-Trichlorophenol 2,4,6-Tribromophenol
PCP TCP TBP
72.0 § 8.6¤¤ 34.7 § 2.6¤¤ 41.7 § 2.9¤¤
49.0 § 5.8¤¤ 19.2 § 2.1¤¤ 23.8 § 2.4¤¤
Polybrominated diphenyl ether 2,2⬘,4,4⬘-Tetrabromodiphenyl ether
TBDE
86.2 § 9.24
Hydroxylated Polychlorinated biphenyl 4,4⬘-(OH)2-3,3⬘,5,5⬘-tetrachlorobiphenyl
4,4⬘-OH-TCB
21.2 § 3.4¤¤
15.2 § 3.4¤¤
Steroid hormones Testosterone 11-ketotestosterone 17,20-dihydroxy-4-pregnen-3-one Cortisol 17-Estradiol (positive control)
T 11-KT 17,20P Cortisol E2
46.5 § 8.5¤¤ 83.3 § 6.9 53.1 § 3.6¤¤ 78.7 § 7.6 12.4 § 2.0¤¤
61.1 § 8.7¤¤ 68.2 § 6.4¤¤ 55.1 § 3.3¤¤ 59.2 § 2.5¤¤ 46.0 § 7.3¤¤
109.5 § 6.6
Liver and kidney tissues of lake trout were incubated in vitro for 1 h in the absence (control) or presence of 100 M of the indicated xenobiotics or steroids. The culture medium was extracted to remove neutral steroids, and the radioactivity remaining in the medium was quantiWed by LSC. Estrogen metabolism is expressed as percent of control. Data are means § SEM (N D 4). Asterisks indicate signiWcant diVerence from control at P 6 0.01.
inactivate E2, and water-soluble conjugates are normally excreted via the urine (Roy, 1992). Some conjugated E2 metabolites, however, circulate in the blood at high levels, and serve as a reservoir of circulating E2 that is released by the actions of sulfatases and glucuronidases in peripheral target tissues (Kirk et al., 2003; Zhu and Conney, 1998). In mammals, certain hydroxylated and conjugated E2 metabolites, similar to those identiWed in lake trout (i.e., E2-OH-S and E2-OH-G), have important non-hormonal eVects (Dubey and Jackson, 2001). In the human placenta, for example, 2-hydroxyestradiol-17-sulfate (2-OH-E2-17-S) is a major product of E2 metabolism, and a long-lasting and powerful antioxidant essential to the maintenance of healthy pregnancy (Takanashi, 2003). Further research is necessary to determine if the E2 metabolites produced by the lake trout kidney and liver play biological roles besides steroid inactivation and excretion. 4.2. EVects of xenobiotics and steroids E2 metabolism by the lake trout kidney and liver was inhibited by various xenobiotics, indicating that some endocrine-disrupting chemicals present in the Great Lakes may act in Wsh by increasing the bioavailability of endogenous hormones through inhibition of hormone-metabolizing enzymes. The mechanism by which these compounds act to inhibit E2 conjugation is not known, although it is likely that they are either competitive or allosteric inhibitors. Hydroxyl-
ated xenobiotics are known substrates for Phase II conjugation (Clark et al., 1991; Yokota et al., 1999), and in the present study, only hydroxylated chemicals inhibited SULT and UGT supporting the hypothesis that xenoestrogens act as competitive inhibitors of E2 conjugation, as reported in zebraWsh (Ohkimoto et al., 2003). On the other hand, Kester et al. (2000, 2002) showed that various xenobiotics inhibit human SULT1E1 by acting as non-competitive, allosteric inhibitors. Moreover, van den Hurk et al. (2002) could not rule out the possibility that non-competitive inhibition was the mechanism by which various PCB-OHs inhibited SULT and UGT activities in the catWsh intestine. The most potent inhibitors of human SULT1E1 were co-planar PCB-OHs that had hydroxyl groups Xanked by halogens, and Kester et al. (2000) speculated that the coplanar structure mimics the primarily planar structure of E2, and that the adjacent halogen substitutions increase relative potency by increasing the dissociation of the OH group. Similarly, in the present study, all of the inhibitory compounds, except BPA, were halogenated, and the most potent inhibitors had hydroxyl groups Xanked by halogens, suggesting that xenobiotic chemicals may have a similar mechanism of action in humans and lake trout. The hydroxylated PCB metabolite 4,4⬘-OH-TCB inhibited human hepatic SULT1E1 with an IC50 value of approximately 0.1 nM (Kester et al., 2000), much lower than observed in the lake trout liver. There are several possible explanations for the apparent higher sensitivity in
G.F. Jurgella et al. / General and Comparative Endocrinology 148 (2006) 273–281
Kidney 100
3,3',5,5'-PCB-4,4'-OH
Water-Soluble E2 Metabolites (% of control)
60 40 20 0 0
1 4 Time (hrs)
Kidney Liver
140
5: E2-OH-G 4: E2-OH-S 3: E2-17-G and E2-17-S 2: E2-3-G 1: E2-OH
80
279
120
**
100 80 60 **
40
**
20
**
**
0
24
**
0.1
1
10
100
[uM]
Liver
80 60
Bisphenol A
5: E2-OH-G 4: E2-OH-S 3: E2-17-G and E2-17-S 2: E2-3-G 1: E2-OH
40 20 0 0
1
4
Kidney Liver
140
24
Time (hrs) Fig. 3. Estradiol-17 (E2) metabolism by the lake trout kidney (upper graph) and liver (lower graph). Data shown are the integrated peaks from Figs. 1 and 2, and depict the percentage of total E2 metabolites represented by each peak at each time point. See Table 1 for the identity of peaks 1–5. E2-17-G, estradiol-17-glucuronide; E2-17-S, estradiol-17-sulfate, etc.
human vs. lake trout. The human studies were conducted using puriWed cytosolic extracts, whereas the lake trout experiments were conducted using primary tissue cultures. Thus, 4,4⬘-OH-TCB, a highly lipophilic compound, may have become bound to cellular membranes eVectively reducing its bioavailability to the lake trout enzymes. In addition, the lake trout liver expresses both SULT and UGT activities, and therefore, higher concentrations of 4,4⬘-OH-TCB may be required to simultaneously inhibit both enzymes. Finally, 4,4⬘-OH-TCB could act by diVerent mechanisms of action in the two species—allosterically in human (Kester et al., 2000), and competitively in lake trout. An explanation for the observation that low doses of 4,4⬘OH-TCB and BPA increased E2 metabolism in the lake trout kidney is that low doses may allosterically potentiate SULT and UGT activities, whereas higher doses may act as competitive inhibitors. In contrast to 4,4⬘-OH-TCB, BPA inhibited both human hepatic SULT1E1 and lake trout hepatic E2 conjugation with similar IC50 values of approximately 10 M (Kester et al., 2000). Human hepatic SULT1E1 was inhibited by the halogenated BPAs, 3,3⬘,5,5⬘-tetrachlorobisphenol A (TC-BPA) and 3,3⬘,5,5⬘tetrabromobisphenol A (TB-BPA), via a non-competitive mechanism with IC50 values of 29–53 nM and 12–33 nM, respectively, illustrating the importance of halogen substitutions to overall xenobiotic potency in humans (Kester
Water-Soluble E2 Metabolites (% of control)
100
**
120 100 80 60
**
40 20
** **
0 0.1
1
10
100
[uM] Fig. 4. Dose–response eVects of 3,3⬘,5,5⬘-tetrachloro-4,4⬘-biphenyl-diol (4,4⬘-OH-TCB, upper graph) and bisphenyl A (BPA, lower graph) on the production of water-soluble (conjugated) E2 metabolites produced by the kidney and liver of lake trout incubated for 1 h in the presence of 106 dpm [3H]-E2 at 15 °C. E2 and all neutral metabolites were extracted with diethyl ether, and the radioactivity remaining in the incubation medium was quantiWed. Data shown are mean § SEM (N D 3), and are expressed as percentage of E2 metabolism in the control incubations without added inhibitor. Asterisks indicate values signiWcantly diVerent from control at P 6 0.01.
et al., 2002). Although IC50 values were not determined for TC-BPA and TB-BPA in lake trout, 100 M doses of both compounds inhibited E2 metabolism more than 100 M BPA, suggesting that halogen substitutions may also increase xenobiotic potency in Wsh. Both 4,4⬘-OH-TCB and BPA were more potent inhibitors of E2 conjugation in the liver than the kidney. This may be because the liver has approximately 50% of the E2 conjugating activity as the kidney, and therefore less inhibitor is required to block the limited enzyme activity in the liver. The steroids, T and 17,20P both inhibited the production of water-soluble E2 metabolites by both the kidney and liver. The biological signiWcance of steroid regulation of E2 metabolism is unclear, but suggests that circulating E2 levels and/or E2 accessibility to estrogen target tissues could possibly be regulated by the reproductive state of the Wsh. For example, elevated T levels associated with reproductive maturation may indirectly increase the biopotency of estrogen by inhibiting E2 metabolism by the kidney, liver, and perhaps other estrogen targets. It is unclear why
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11-ketotestosterone and cortisol were only eVective in the liver but not the kidney, but this may also be explained by the observed tissue diVerences in E2 metabolic activity. 5. Concluding remarks This study reports that various xenobiotics found in the aquatic environment have the ability to inhibit E2 metabolism by the lake trout kidney and liver. These results add to the growing literature indicating that many xenoestrogen endocrine disruptors act by inhibiting E2 metabolism in Wsh (Kirk et al., 2003; Ohkimoto et al., 2003; van den Hurk et al., 2002). The signiWcance of this regulation is not fully understood, but may be key to understanding many biological phenomena thought to be regulated by environmental estrogens. Xenobiotics, for example, could increase circulating endogenous E2 concentrations by disrupting normal inactivating pathways. Alternatively, xenobiotics may act directly at the level of the target tissue by inhibiting metabolizing enzymes that normally protect the ER from binding circulating E2. Inhibitory xenobiotics may prevent the formation of important biologically active E2 metabolites, such as antioxidants or biologically active catecholestrogens (Zhu and Conney, 1998). Finally, inhibitory xenobiotics may switch normal E2 metabolic pathways in alternative directions. It has been reported in mammals, for example, that certain phytoestrogens can increase the activity of 2-hydroxylase and decrease the activity of 16-hydroxylase, an alteration associated with a reduced risk of estrogen-mediated cancers (Dubey and Jackson, 2001). Acknowledgments This work was funded by a grant to TB and RP from the University of Wisconsin Sea Grant College Program, National Oceanic and Atmospheric Administration, US Department of Commerce (Federal Grant NA16RG2257; Project No. R/BT-19). Support also came from the University of Wisconsin-Madison College of Agricultural and Life Sciences and School of Natural Resources, and the Wisconsin Department of Natural Resources. The Minnesota Department of Natural Resources donated the lake trout used in this study. Ultra ScientiWc (North Kingstown, RI) donated xenobiotics. We acknowledge the assistance of Colin Jefcoate, Milo Wiltbank, James A. Held, Luciene Lima, Yuliana Ng, Robert Moore, Henry Lardy, and Padma Marwah. References Barry, T.P., Riebe, J., Parrish, J.J., Malison, J.A., 1997. EVects of 17,20dihydroxy-4-pregnen-3-one on cortisol production by rainbow trout interrenal tissue in vitro. Gen. Comp. Endocrinol. 107, 172–181. Clark, D.J., George, S.G., Burchell, B., 1991. Glucuronidation in Wsh. Aquat. Toxicol. 20, 35–56. Colborn, T., von Saal, F.S., Soto, A.M., 1993. Developmental eVects of endocrine-disrupting chemicals in wildlife and humans. Environ. Health Perspect. 101, 378–384.
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