Re-evaluating the Significance of Estrone as an Environmental

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Reevaluating the Significance of Estrone as an Environmental Estrogen Gerald T. Ankley, David Feifarek, Brett Blackwell, Jenna E. Cavallin, Kathleen M. Jensen, Michael D. Kahl, Shane Poole, Eric Randolph, Travis Saari, and Daniel L. Villeneuve Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b00606 • Publication Date (Web): 22 Mar 2017 Downloaded from http://pubs.acs.org on March 24, 2017

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Environmental Science & Technology

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Reevaluating the Significance of Estrone as an Environmental Estrogen

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Gerald T. Ankley1, *, David Feifarek1, Brett Blackwell1, Jenna E. Cavallin2, Kathleen M.

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Jensen1, Michael D. Kahl1, Shane Poole1, Eric Randolph3, Travis Saari1, Daniel L. Villeneuve1

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US Environmental Protection Agency

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Badger Technical Services

Oak Ridge Institute of Science Education

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6201 Congdon Boulevard

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Duluth, Minnesota 55804 USA

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*Corresponding Author: [email protected]; T: (218) 529-5147; F: (218) 529-5003

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ACS Paragon Plus Environment


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Abstract

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Studies worldwide have demonstrated the occurrence of feminized male fish at sites impacted by

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human and animal wastes. A variety of chemicals could contribute to this phenomenon, but those

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receiving the greatest attention in terms of research and monitoring have been 17β-estradiol (β-

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E2) and 17α-ethinylestradiol, due both to their prevalence in the environment and strong

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estrogenic potency. A third steroid, estrone (E1), also can occur at high concentrations in surface

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waters, but generally has been of lesser concern due to its relatively lower affinity for vertebrate

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estrogen receptors. In an initial experiment, male fathead minnow (Pimephales promelas) adults

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were exposed for 4-d to environmentally-relevant levels of waterborne E1, which resulted in

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plasma β-E2 concentrations similar to those found in reproductively-active females. In a second

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exposure we used 13C-labeled E1, together with liquid chromatography-tandem mass

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spectrometry, to demonstrate that elevated β-E2 measured in the plasma of the male fish was

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indeed derived from the external environment, most likely via a conversion catalyzed by one or

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more 17β-hydroxysteroid dehydrogenases. The results of our studies suggest that the potential

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impact of E1 as an environmental estrogen currently is underestimated.

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Introduction

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More than two decades ago a team of scientists documented the extensive occurrence of

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feminized male fish in effluent-dominated streams and rivers in the UK, and proposed that the

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observations were due to estrogenic chemicals [1-3]. The presence of feminized fish at these and

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many other sites throughout the world has been a critical contributor to widespread (and

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ongoing) concerns about the possible effects of endocrine-disrupting chemicals on human health

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and the environment. Hundreds, if not thousands of papers have been published concerning

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occurrence in the environment and effects of environmental estrogens on fish. Based in part on

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analyses at sites in the UK, two chemicals of particular concern relative to feminized fish have

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been the natural and synthetic steroids 17β-estradiol (β-E2) and 17α-ethinylestradiol (EE2) [4,5].

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Not only are both commonly detected in wastewater treatment plant (WWTP) effluents, but they

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are among the most potent chemicals known in terms of binding to and activating estrogen

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receptor-alpha (ER-α), the primary ER isoform associated with many key reproductive and

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developmental pathways in vertebrates. However, a number of other compounds also have been

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proposed/evaluated as contributing to estrogenicity in environmental samples, including

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industrial chemicals such as bisphenol A and alkylphenols, the natural steroid estrone (E1), and

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some phytoestrogens [e.g., 6-12].

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Estrogens are a frequent target of environmental monitoring programs, which employ a variety

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of instrumental and/or biological methods to determine, respectively, presence of specific

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chemicals of interest or integrated measures of estrogenic activity. For example, over the past 5

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yr our laboratory has been involved in studies at a number of sites around the Great Lakes

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focused on chemicals of emerging concern, including those that disrupt endocrine function, using

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an integrated suite of targeted analytical techniques and in vitro and in vivo bioassays [13]. In



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these studies we employ: (a) an analytical approach that encompasses about 130 chemicals

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including a variety of known estrogens (e.g., E1, β-E2, EE2, bisphenol A), (b) in vitro assays for

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bioactivities such as estrogen and androgen receptor activation, and (c) analysis of different types

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of pathway-specific biomarkers in samples from caged fathead minnows (Pimephales promelas),

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including indicators of reproductive endocrine function (e.g., plasma sex steroid concentrations,

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gonadal expression of genes involved in steroid synthesis, etc.). There have been a number of

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recent papers describing facets of the application of this integrated approach at different Great

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Lakes sites [14-19]. Many of these results have been consistent with expectations based on other

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studies (e.g., occurrence of in vitro estrogenicity associated with WWTP discharges [16, 17]),

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but there also have been some unexpected observations. One involved the occurrence of elevated

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concentrations of β-E2 in the plasma of caged male fathead minnows (deployed from controlled

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laboratory cultures) in the vicinity of a WWTP discharging to the Duluth-Superior Harbor, MN.

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This observation was reproducible both over time (within a given study), and across multiple

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study years (SI Figure S1).

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Plasma concentrations of β-E2 in untreated male fathead minnows reared in our laboratory are

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historically quite low, typically close to method detection limits [20], so this observation from

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the field was of interest. We initially speculated that the elevated β-E2 may be derived from

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upregulation of cytochrome P450 aromatase (CYP19), which converts testosterone to β-E2 in

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vertebrates, but there was little evidence of increased expression of CYP19 transcripts or ex vivo

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E2 production in testes of the caged fish (data not shown). Interestingly, in analyzing samples

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from a caged fathead minnow study conducted at a site in the western US we noted an even more

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pronounced elevation of plasma β-E2 in male fathead minnows. A notable chemistry observation

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from that deployment site, which received a WWTP discharge, was high concentrations (ca. 150


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ng/L) of E1 in surface waters collected during the fish exposure (Marc Mills, USEPA,

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Cincinnati, OH; personal communication). Significantly, data from studies conducted with the

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WWTP affecting the Duluth-Superior Harbor caged fish site also indicated that relatively high

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concentrations (> 40 ng/L) of E1 could occur in the effluent [17]. This suggested to us that the

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elevated β-E2 in males might result from exposure to E1. Specifically, it is relatively well

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established in rodents and humans that E1 can be converted to β-E2 by multiple17β-

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hydroxysteroid dehydrogenase (17βHSD) isoforms [21-25]. Although not as well characterized,

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there is evidence for both structural and functional conservation of several of these 17βHSDs in

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teleost fish [26-30]. Consequently, it seemed plausible that elevated β-E2 observed in male

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fathead minnows from the caging studies may be enzymatically derived from E1. If this were

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true, it would have implications for assessing the ecological risk of estrogens in the environment

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because, depending upon the test system used, E1 and β-E2 can differ quite substantially in their

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predicted potency.

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The aims of the present study were two-fold. First, we sought to determine whether exposure of

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male fathead minnows to E1 concentrations similar to those present in the environment would

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result in elevated β-E2 in the fish. Secondly, to establish whether observed elevations in β-E2

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actually occurred as a result of (presumably metabolic) conversion of external E1 by the fish, we

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utilized an approach involving exposure of the fish to 13C-labeled E1.

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Materials and Methods

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Experimental Design and Biological Measurements

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Two exposures were conducted with sexually-mature (6-7 month old) male fathead minnows

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from an on-site culture. The exposures were 4-d in duration to mimic our typical caged-fish



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deployment regime [16]. In Experiment 1, three replicate tanks containing six males were used

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for each treatment. Target water concentrations were 0 (Lake Superior water control); 1.8, 5.5,

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16.7, 50 and 150 ng E1/L; and (as a positive control) 1 and 10 ng β-E2/L. The E1 and β-E2

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(>99% and >98% purity, respectively; Sigma-Aldrich, St. Louis, MO, USA) were dissolved

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directly (without solvent) in Lake Superior water, which served as a stock solution for

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subsequent dilution and delivery to the test tanks at a flow rate of approximately 45 mL/min. In

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Experiment 2, fish were exposed using the same experimental design and conditions to 2,3,4-

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13

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250 and 500 ng/L. All experiments were conducted at 25+1ºC under a 16:8 light:dark

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photoperiod. Fish were fed thawed adult brine shrimp ad libitum twice daily. Procedures

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involving animals were performed in accordance with an externally-approved Animal Care and

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Use Protocol.

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Water samples were collected from the test tanks daily during both experiments and analyzed for

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E1 and β-E2 using the procedures described below. After 4 d fish were anaesthetized with a

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buffered solution of MS-222 (Sigma-Aldrich), weighed, and blood samples collected from

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caudal vasculature with a heparinized glass capillary tubule. Plasma was immediately isolated

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using centrifugation and stored at -80ºC until analyzed for steroids using either

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radioimmunoassay (RIA; Experiment 1) or liquid chromatography-tandem mass spectrometry

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(LC-MS/MS; Experiment 2). Additional tissues collected from the males included liver and

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testis, which were flash-frozen in liquid nitrogen and stored at -80ºC until used for gene

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expression measurements.

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Concentrations of β-E2 in plasma samples from Experiment 1 were measured using a RIA

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technique adapted to small volume samples [20]. To assess the degree to which E1 or 17α-

C-labeled E1 (13C-E1; Sigma-Aldrich, >98 % purity), at nominal water concentrations of 125,


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estradiol (α-E2, a possible E1 metabolite [31]) might bias our determination of β-E2

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concentrations, we conducted an experiment to assess possible cross-reactivity of the two

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steroids with the β-E2 RIA antibody. Briefly, the normal RIA for β-E2 was conducted with RIA

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buffer that had been spiked either with E1 or α-E2 at concentrations ranging from 50 to 5000

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ng/L. No discernable cross-reactivity was noted for α-E2, while E1 bound to the antibody with

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slightly less than 5% of the affinity of β-E2 (data not shown).

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The abnormal production of vitellogenin (VTG; egg yolk protein precursor) in male fish is

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widely accepted as a marker of exposure to exogenous estrogens [32]. While multiple ER

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isoforms can be involved in VTG production to some degree [e.g., 33], induction of the

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lipoprotein in males is regarded primarily as an indicator of activation of ER-α [34]. In both

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experiments, hepatic expression of VTG mRNA was measured using quantitative reverse-

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transcription polymerase chain reaction (QPCR), as described elsewhere (SI Table S1 [17]). In

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Experiment 1 we also measured VTG protein in plasma of the fish using an enzyme-linked

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immunosorbent assay with a fathead minnow polyclonal antibody [35].

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QPCR also was used to determine expression of transcripts for cyp19a1a (aromatase) in testis of

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the fish from Experiment 1 (SI Table S1 [36]). In Experiment 2, transcripts of four 17βHSDs

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were measured both in liver and testis samples using QPCR. Mindnich and Adamski [29]

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summarize comparative structural and functional data for 17βHSDs in humans, rodents and

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zebrafish, noting that four forms, 1, 3, 7 and 12 (12a in zebrafish), have been reported in one or

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more of these taxa as capable of converting E1 to β-E2. To measure transcripts of these four

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17βHSDs, gene-specific primers were designed using sequences from a fathead minnow genome

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assembly (SI Table S1 [37]). Total RNA was extracted from all samples using RNeasy Mini kits

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(Qiagen, Hilden, Germany) according to the manufacturer’s protocol, and subsequently diluted



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to 10 ng/µL prior to QPCR analysis. Quality and quantity of RNA was determined using a

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Nanodrop ND-1000 spectrophotometer (Nanodrop Technologies, Thermo Fisher Scientific,

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Wilmington, DE). All RNA samples had A260 nm/A280 nm ratios > 1.86. Relative transcript

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abundance of the four 17βHSDs was evaluated using Power SYBR Green RNA-to-CT 1-step kits

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(Applied Biosystems, Foster City, CA, USA). Each 20 µL reaction included 20 ng total RNA,

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215 nM forward primer and 215 nM reverse primer (see SI Table S1 for primer sequences). The

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thermocycling program included an initial reverse transcription step with a 30 min cycle at 48°C,

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followed by PCR amplification at 95°C for 10 min and 40 cycles of 95°C for 15 s and 60°C for 1

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min, and a final dissociation step of 95°C for 15 s, 60°C for 1 min, and 95°C for 15 s. Relative

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abundance of the gene transcripts was quantified using gene-specific cDNA standard curves with

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six concentrations following a 10-fold dilution series. The cDNA standards were prepared

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through a series of PCR amplifications using JumpStart Taq DNA Polymerase (Sigma-Aldrich,

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St. Louis, MO, USA) according to the manufacturer’s protocol. For all QPCRs, amplification

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efficiencies ranged from 89.9 to 107.8%. At least 10% of the samples were run in duplicate

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within each QPCR plate, and the mean (SD) coefficient of variation between the duplicate

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samples was 12.2 (17.4).

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Instrumental Analytical Measurements

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Water concentrations of E1 and β-E2 in Experiment 1 were determined daily using LC-MS/MS

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with isotope dilution quantification. Water was collected from each tank and concentrated to

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achieve a target concentration between 1–5 µg/L in final extracts. Prior to extraction, each

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sample was spiked with 0.3 ng of stable labeled 13,14,15,16,17,18-13C E2 (13C6-E2; Cambridge

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Isotope Laboratories Inc., Andover, MD, USA) to serve as the isotope dilution standard. Water

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samples were extracted using Strata-X solid phase extraction cartridges (60 mg, 3 mL;


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Phenomenex, Torrance, CA, USA), which were subsequently eluted with methanol. The eluate

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was evaporated to dryness under a gentle stream of nitrogen. Extracts were reconstituted in 25:75

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water:methanol and stored at -20ºC until analysis.

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Both E1 and β-E2 were determined by LC-MS/MS with atmospheric pressure photoionization

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(APPI) using an Agilent 6410 system. A 40 µL aliquot of extract was injected and separated

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using a Kinetex PFP column (2.1 x 100 mm, 2.6µm; Phenomenex) using a gradient elution at

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0.35 mL/min. Under the given conditions, β-E2 and E1 eluted at 7.1 and 7.3 min, respectively.

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Two ion transitions were monitored for each target analyte and one transition monitored for the

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isotopic standard (SI Table S2). The limit of detection (LOD) for both E1 and β-E2 was 0.125

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ng/L. Analysis of matrix spikes and duplicates was performed with each sample set to verify

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accuracy and precision. Mean (SD) percent matrix spike recoveries for E1 (n=5) and β-E2 (n=7)

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were 110.5 (12.8) and 104.1 (8.1), respectively. Percentage duplicate agreement values for E1

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(n=7) and β-E2 (n=1) were 90.4 (8.2) and 92.5, respectively.

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In Experiment 2, water and plasma concentrations of 13C-labeled steroids were measured using

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LC-MS/MS. In addition to E1, concentrations of both E2 stereoisomers (α, β) were determined in

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the test samples. Water from the exposure tank samples was collected daily. Prior to extraction

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or analysis, samples were spiked with 13C6-E2 to achieve a final concentration of 3.0 µg/L.

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Water from the control and 125 ng/L nominal treatment tanks was extracted/processed as for

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Experiment 1, while water samples from the 250 and 500 ng/L treatments were directly injected

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for LC-MS/MS analysis. Both stable isotope-labeled and unlabeled E1 and E2 were determined

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using an Agilent 6410 LC-MS with APPI. A 100 µl aliquot was separated using a Zorbax RRHD

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StableBond C18 column (2.1 x 50 mm, 1.8 µm; Agilent Technologies, Santa Clara, CA, USA)

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under gradient conditions at 0.4 mL/min to fully resolve α- and β- isomers of E2 (Fig. 1). Two



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ion transitions were monitored for each target analyte and one transition monitored for the

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isotopic standard (SI Table 2). Since both the number and position of 13C labels differed between

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surrogate 13C6-E2 and potential 13C-E1 derived 13C3-E2, the compounds could be readily

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distinguished from one another. The LOD of all compounds in water was 2.9 ng/L. The

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percentage matrix spike recovery of 13C-E1 in Lake Superior water (n=10) was 102.4 (11.2) and

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percent duplicate agreement (n=12) was 87.9 (7.5).

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Plasma was extracted using Novum supported liquid extraction cartridges (Phenomenex).

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Briefly, 15 µL of plasma was spiked with 0.25 ng 13C6-E2, diluted to 200 µL, and loaded onto

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the cartridges. Cartridges were eluted with methylene chloride which was evaporated to dryness

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under a gentle stream of nitrogen. Extracts were reconstituted in 100 µL of 25:75 water:methanol

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and stored at -20ºC until analysis. Analysis of plasma was identical to exposure water with the

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exception of use of a 40 µL injection volume. The slopes of the E1 and 13C-E1 calibration curves

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were parallel when determining exposure tank concentrations (data not shown), thus,

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quantifications of 13C-E1, 13C3-α-E2, and 13C3-β-E2 in plasma were calculated using the

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standard curves of the corresponding unlabeled compounds. The LOD of all compounds in

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plasma was 0.42 µg/L. Percent matrix spike recoveries of E1, α-E2, and β-E2 in charcoal

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stripped serum (n=6) were 100.7 (6.0), 92.3 (7.6), and 96.6 (4.3), respectively. Mean percent

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agreement values for 13C3-labeled E1, α-E2, and β-E2 in duplicate plasma samples (n=4) from

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exposed individuals were 92.4 (4.7), 81.9 (14.5), and 88.3 (9.0), respectively.

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Statistical Analysis

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Data were assessed for the assumptions of normality and homogeneity of variance using

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Kolmogorov-Smirnov and Levene’s tests, respectively. When data conformed to parametric

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assumptions, one-way analysis of variance (ANOVA) was used to test for differences across


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treatment groups, followed by Duncan’s multiple range test to determine differences among

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chemical treatments. When data did not meet parametric assumptions, they were transformed

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(log10) and analyzed as described above, or with a Kruskal-Wallis test followed by Dunn’s post

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hoc test. Differences were considered significant at p < 0.05. Statistical analyses were conducted

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using Statistica 8 (StatSoft Inc., Tulsa, OK, USA). Results

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Experiment 1

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Mean (SD, n=4) measured concentrations of the estrogens in the first experiment were consistent

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over the course of the 4-d assay. Concentrations of E1 were close to the target values at 1.66

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(0.33), 4.97 (1.12), 15.6 (4.43), 41.9 (8.83) and 144 (49.6) ng/L. No β-E2 was detected in tank

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water from the E1 treatments. The high β-E2 treatment group was similarly close to the 10 ng/L

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target concentration at 7.73 (1.44) ng/L, while the measured concentration in the low treatment

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was about 40% of nominal at 0.43 (0.08) ng/L. There was detectable E1 in water from both the

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low and high β-E2 treatments, at 0.38 (0.12) and 1.82 (0.91) ng/L, respectively. Neither steroid

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was detected in water from the control tanks.

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Plasma concentrations of β-E2 (determined via RIA) were elevated in a concentration-dependent

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manner in males exposed to E1 (Fig. 2). Increases in β-E2 were statistically significant in

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(measured) E1 treatments of 15.6 ng/L and greater. Exposure of the males to the high, but not

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low, concentration of β-E2 also significantly increased plasma β-E2 in the fish, to a level roughly

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comparable to that observed in males from the 15.6 ng E1/L treatment (Fig. 2).

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Hepatic expression of vtg was increased in a dose-dependent manner in male fathead minnows

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exposed both to E1 and β-E2, with the greatest degree of induction observed in the 41.9 and 144



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ng E1/L treatment groups (Fig. 3A). Induction of VTG in plasma of the fish closely paralleled

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hepatic mRNA expression (Fig. 3B).

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Expression of transcripts for aromatase (cyp19a1a) did not differ significantly from controls in

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males exposed either to E1 or β-E2 (data not shown).

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Experiment 2

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Mean (SD, n=4) measured concentrations of 13C-E1 were 30 to 50% lower than the target values

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at 60.7 (4.57), 161 (14.0), and 360 (38.2) ng/L, but were relatively stable over the course of the

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4-d exposure. Neither α- nor β-13C-E2 were detected in water from the treatment tanks, and none

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of the steroid isotopes were detected in the control tanks.

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Water-borne exposure of male fathead minnows to 13C-E1 elevated plasma levels of 13C-E1 and

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both E2 stereoisomers in a concentration-dependent manner (Fig. 4 A-C). Plasma concentrations

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of 13C-E1 were the highest, followed by 13C-β-E2, and then 13C-α-E2. None of the labeled

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steroids were detected in plasma from the control males.

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Expression of hepatic vtg was non-detectable in control males, and was induced in a dose-

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dependent manner by the 13C-E1 treatments (Fig. 5). Transcripts for four different 17βHSDs (1,

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3, 7, 12a) were detected in testis of the fish (Fig. 6 A-D). Genes coding for two 17βHSDs (3, 7)

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also were detected in the liver, albeit at much lower basal expression levels (ca. 10- and 3-fold,

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respectively) than in the testis. Exposure to 13C-E1 significantly down-regulated transcripts for

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two of the four 17βHSDs (3, 7) present in the testis (Fig. 6 B, C).

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Discussion


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A variety of chemical and non-chemical stressors can impact reproductive endocrine function in

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vertebrates. Consequently, when some of our recent field studies with caged fathead minnows

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showed an abnormal elevation of E2 in males, there was concern both for causes and possible

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biological implications of the response. Common to the sites where detectable β-E2 was noted in

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the males was the presence of relatively high concentrations of E1. For example, at one location

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where about 3 µg β-E2 /L was observed in male plasma (determined using the RIA technique

282

described herein), almost 150 ng E1/L was measured in a corresponding 4-d composite sample

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of surface water. This observation corresponds favorably with data from the present lab studies

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in which a 4-d exposure of male fathead minnows to 144 ng E1/L (Experiment 1) or 161 ng (13C-

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) E1/L (Experiment 2) both resulted in about 5 µg β-E2/L plasma. Data from the second study

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with labeled E1 also clearly demonstrate that β-E2 observed in the males was derived from

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external E1, as opposed to a product of perturbed steroid metabolism (e.g., increased CYP19

288

activity). The biological/toxicological significance of E1-derived β-E2 in the males is difficult to

289

fully define; for example, although hepatic vtg expression was consistently induced in E1-treated

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fish in both experiments, it is impossible to ascertain the degree to which the different steroids

291

contributed to this response. From a completely pragmatic perspective, however, it is noteworthy

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that plasma concentrations of β-E2 observed in males exposed to about 150 ng E1/L either in the

293

field or lab are comparable to those in reproductively-active female fathead minnows [20]. This

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indicates a strong potential for feminization of the fish by the converted β-E2, irrespective of any

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contributions associated with direct activation of ER-α by E1. Overall, we believe that the ability

296

of fish to convert E1 to β-E2 necessitates reconsideration of the potential ecological risks of E1.

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There are widely-varying estimates in the literature of the relative potency of E1 versus β-E2 in

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different bioassays relevant to assessing potential ecological effects. In general, data from in



299

vitro systems suggest β-E2 is far more potent than E1. For example, Denny et al. [38] reported

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that the binding affinity of E1 relative to β-E2 to fathead minnow and rainbow trout ERs in

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ovarian cytosol preparations was 28 and 5%, respectively. Peterson and Tollefsen [39] similarly

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reported that β-E2 was about 100-times more potent than E1 in inducing VTG in isolated

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rainbow trout hepatocytes. In the most comprehensive comparative study available, Lange et al.

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[40] employed transcriptional activation assays with ER-α—reporter gene constructs transfected

305

into HEK293 cells, and found that β-E2 was more potent than E1 by factors of about 19-, 26-,

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36-, 30-, 55- and 65-fold, respectively, using cloned receptors from zebrafish, medaka, roach,

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fathead minnow, three-spined stickleback and common carp. There are some notable examples,

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however, in which the differences in potency between E1 and β-E2 in an in vitro system are

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much less pronounced. The T47D-KBluc assay [41], a human ER-α —based system that has

310

been commonly used to screen complex environmental samples for estrogenicity [e.g., 17, 42], is

311

slightly more sensitive to E1 than β-E2. This quite probably is due, however, to the presence of

312

functional 17βHSDs capable of converting E1 into β-E2 in this human breast cancer cell line

313

[43].

314

In vivo assessments of E1 versus β-E2 potency in fish species for which relevant comparative in

315

vitro data also exist tend to indicate a more equivalent potency for the two steroids, again likely

316

due to the presence of 17βHSDs. For example, Thorpe et al. [44] reported that β-E2 was only

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about 2- to 3-fold more potent than E1 in terms of inducing VTG in juvenile rainbow trout, as

318

opposed to the two-order of magnitude difference suggested by ER binding/activation studies

319

[38, 39]. Analogously, Van den Belt et al. [45] reported similar effects concentrations for

320

induction of VTG in zebrafish by E1 and β-E2, in contrast to the 19-fold difference in potency of

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the two steroids suggested by Lange et al. [40]. Finally, both Panter et al. [46] and Dammann et


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al. [47] conducted 21-d tests with the fathead minnow that examined a variety of endpoints

323

(including VTG induction), and did not observe greater potency for β-E2 than E1 which, again,

324

is much different than expectations based on comparative ER-α transactivation studies with the

325

two steroids [40]. Due to a variety of pharmacokinetic factors, including metabolism, in vivo

326

potency values for test chemicals typically are expected to be more accurate than those derived

327

from in vitro systems. This clearly is true for E1.

328

The occurrence and functional roles of 17βHSDs in mammals have been well studied, especially

329

from a biomedical perspective [21-25]. Far less is known concerning this family of enzymes in

330

fish. Mindnich and Adamski [29] summarized comparative structural and catalytic attributes of

331

17βHSDs in humans, rodents (mice and rats), and zebrafish. Of the 14 known 17βHSDs in

332

mammals, nine sequence homologs occur in zebrafish. While some forms are most highly

333

expressed in gonads and liver, expression of other 17βHSDs occurs in a wide variety of tissues

334

[29]. Four of the homologs shared across humans, rodents, and zebrafish, 17βHSDs 1, 3, 7 and

335

12a, can convert E1 to β-E2 in one or more of the taxonomic groups. While the present study

336

was not designed to demonstrate exactly which 17βHSD isoform(s) might be involved in the

337

conversion of E1 to β-E2, it is noteworthy that transcripts were detected in the testis for all four

338

of the 17βHSDs reported to catalyze this reaction, and genes coding for two of the 17βHSDs also

339

were present in the liver. One line of evidence suggesting that at least some of the testicular

340

17βHSDs (3, 7) have a direct functional role in endocrine function of the fish is the fact that

341

exposure to E1 caused a down-regulation in expression of their transcripts, a response consistent

342

with feed-back inhibition in the hypothalamic-pituitary-gonadal axis. Additional studies are

343

needed, however, to achieve a deeper understanding of the catalytic nature and tissue distribution



344

of these 17βHSDs in the fathead minnow (and other fish species), including better definition of

345

their role in the conversion of E1 to β-E2.

346

An unanticipated observation from Experiment 2 was the occurrence of appreciable

347

concentrations not only of β-E2, but 13C-α-E2 in the 13C-E1-exposed males. The possible

348

biological significance of the presence of α-E2 is uncertain, as there is comparatively little

349

information concerning the in vivo effects or estrogenic potency of this stereoisomer in fish. In

350

the only in vivo study we were able to identify on the topic, Shappell et al. [48] reported that α-

351

E2 was 8 to 9-fold less potent than β-E2 in inducing plasma VTG in a 21-d test with adult

352

fathead minnows. This lesser potency of α-E2 than β-E2 relative to activation of ER-α is

353

consistent with observations from mammalian systems evaluating comparative estrogenic effects

354

of the two stereoisomers [e.g., 49-51]. Therefore, it is quite possible that α-E2 contributed

355

comparatively little to cumulative estrogenic “activity” (as measured by VTG induction) in the

356

fish in the present studies. However, recent studies in mammalian systems suggest that the

357

primary transcription factor associated with α-E2 may not actually be ER-α, but a less-well

358

characterized ER-X, whose biological function(s) are unknown [52]. Similarly, the synthetic

359

pathway through which α-E2 might have arisen from external E1 in the present study is

360

uncertain. It is noteworthy, however, that there have been reports of the occurrence of elevated

361

α-E2 in human females dosed with E1-sulfate [31]. Overall, additional work is needed to better

362

document the occurrence of α-E2 in fish as a consequence of environmental exposure to other

363

steroidal estrogens, as well as to better understand the possible impacts of α-E2 on normal

364

estrogen-signaling pathways.

365

In summary, the extensive occurrence of feminized male fish in field settings impacted by

366

human and animal wastes has been important in fueling broader concerns for the potential effects


Page 16 of 34

Page 17 of 34


367

of endocrine disrupting chemicals on both human health and the environment. Due to their early

368

association with feminized fish, ubiquitous nature in WWTP discharges, and demonstrated

369

potency relative to activation of vertebrate ERs, including ER-α, the two chemicals that have

370

received the most attention as environmental estrogens have been β-E2 and EE2. There has been

371

less concern about the potential ecological risks of E1 despite the fact it often is detected in

372

environmental samples, frequently at concentrations one to two orders-of-magnitude higher than

373

either β-E2 or EE2 [17, 42, 53-57] and, in some instances, has been correlated with the

374

occurrence of feminized fish [e.g., 56]. This is at least in part because E1 is far less potent than

375

β-E2 or EE2 in in vitro systems focused, for example, on binding to or activation of fish ERs

376

[e.g., 38-40]. However, demonstration that E1 can be readily converted to β-E2 by fish indicates

377

that the two steroids perhaps should be considered equivalent in terms of their potential

378

ecological hazard.

379

Supporting Information

380

Figure S1. Plasma concentrations of 17β-estradiol in adult male fathead minnows caged at a field

381

site near a wastewater treatment plant in the Duluth-Superior Harbor.

382

Table S1. Fathead minnow gene-specific primers for quantitative polymerase chain reaction

383

analyses for target genes.

384

Table S2. Ion transitions (m/z) of monitored estrogens from Experiments 1 and 2.

385

386

Acknowledgements



387

We thank Vickie Wilson for providing comments on this paper, and L. Earl Gray and Glen Van

388

Der Kraak for helpful discussions. We also acknowledge the contributions of Ted Smith through

389

activities associated with the Great Lakes Restoration Initiative. This manuscript has been

390

reviewed in accordance with the requirements of the US Environmental Protection Agency

391

(EPA) Office of Research and Development. The views expressed in this work are those of the

392

authors and do not necessarily reflect the views or policies of the US EPA, nor does the mention

393

of trade names or commercial products constitute endorsement or recommendation for use. References

394

395

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Figure Legends

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581

Figure 1. Example chromatogram of 17β-estradiol and 17α-estradiol (A, C) and 13C3-labeled

582

analogs (B, D) in calibration standards at 1.25 µg/L (A, B), and in extracted plasma (C, D) at

583

quantified at 10.1 and 10.7 µg/L of 13C3-17β-estradiol and 13C3-17α-estradiol, respectively.

584

Figure 2. Mean (standard error of the mean, n=18) 17β-estradiol (β-E2) concentrations

585

determined via radioimmunoassay in plasma from male fathead minnows exposed to estrone

586

(E1) or β-E2 for 4-d. Different letters indicate significant (p

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