Responses to various exposure durations of levonorgestrel during early-life stages of fathead minnows (Pimephales promelas)

Aquatic Toxicology 161 (2015) 33–40

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Aquatic Toxicology journal homepage: www.elsevier.com/locate/aquatox

Responses to various exposure durations of levonorgestrel during early-life stages of fathead minnows (Pimephales promelas) Matthew D. Overturf ∗ , Duane B. Huggett Department of Biology, University of North Texas, Denton, TX 76203, USA

a r t i c l e

i n f o

Article history: Received 1 October 2014 Received in revised form 26 January 2015 Accepted 28 January 2015 Available online 29 January 2015 Keywords: Levonorgestrel Early-life stage mRNA expression Fathead minnow

a b s t r a c t Pharmaceuticals are routinely detected in the environment; and several of these compounds have been extensively researched due to their potential impacts to the endocrine system of aquatic organisms. The negative reproductive consequences of synthetic progestins in teleost species have only recently been investigated. The current study examined different exposure periods that may be most sensitive for levonorgestrel (LNG) in early-life stages of fathead minnow larvae. Larvae were exposed to a single concentration of LNG (125 ng/L) for different durations from fertilized egg through 28 days post hatch (dph) with growth and mRNA expression of FSH, 3␤-HSD, 20␤-HSD, and CYP19a1 measured. Regardless of the duration of exposure, LNG significantly decreased growth in the fathead minnow larvae at day 28. For both 20␤-HSD and CYP19a1, mRNA expression was decreased following LNG exposure durations ≥7 days. 3␤-HSD and FSH showed similar trends after exposure to LNG with later stages of development exhibiting decreased expression. 20␤-HSD and 3␤-HSD were the only transcripts to remain down regulated once larvae were moved to clean water after the 7–14 dph LNG exposure. This study is the first to investigate the effects of exposure time to a synthetic progestin on developing fish. Future research is needed to understand what impacts these changes have on adult stages of development. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Pharmaceuticals and their metabolites are routinely found in surface waters. These compounds have received considerable attention recently, especially for their potential to disrupt the endocrine system of many aquatic species (Ankley et al., 2007). Much research has been centered on synthetic estrogens and their endocrine disruption potential (Corcoran et al., 2010; Kidd et al., 2007), whereas, the synthetic progestins have only recently been examined. Both synthetic estrogens and progestins are commonly used in oral contraceptives and hormone replacement therapies; therefore, these compounds can find their way into surface waters through the inefficient removal of sewage treatment facilities. Progestins have been detected in the environment at concentrations ranging from 1 to 199 ng/L (Al-Odaini et al., 2010; Andersson et al., 2005; Chang et al., 2011; Kolpin et al., 2002; Lopez de Alda et al., 2002; Petrovic et al., 2002; Viglino et al., 2008; Vulliet et al., 2008) and are shown to negatively impact aquatic animals at comparable

∗ Corresponding author. Present address: Faculty of Science, University of Ontario, Institute of Technology, 2000 Simcoe Street, North Oshawa, ON L1H 7K4, USA. Tel.: +1 905 721 8668x2941. E-mail address: [email protected] (M.D. Overturf). http://dx.doi.org/10.1016/j.aquatox.2015.01.029 0166-445X/© 2015 Elsevier B.V. All rights reserved.

concentrations (DeQuattro et al., 2012; Paulos et al., 2010; Runnals et al., 2013; Zeilinger et al., 2009). In mammals, synthetic progestins inhibit ovulation and proliferation of the endometrium by binding to the progesterone receptor (Croxatto et al., 2004; Durand et al., 2001; Marions et al., 2002, 2004). These progestins exhibit binding affinities to the other steroid hormone receptors (e.g., androgen and mineralocorticoid receptor) depending on the parent molecule they are derived (Sitruk-Ware, 2004). The derivatives of 19-nortestosterone, levonorgestrel (LNG; gonane) and norethindrone (NOR; estrane), have androgenic activity (Africander et al., 2011). These principles have been demonstrated in teleost species where masculinization of females occur after exposure to LNG and NOR; additionally, ovulation (egg laying) is also significantly reduced by these synthetic progestins (Paulos et al., 2010; Runnals et al., 2013; Zeilinger et al., 2009). These effects have only been demonstrated in adult fish; therefore, little is known about the impacts of these synthetic progestins on early-life development of aquatic species. There is limited data that suggests that synthetic progestins have negative impacts on early-life stages of aquatic organisms. Early-life stage studies of NOR and LNG have subsequently been conducted with fathead minnow larvae demonstrating significant reduction in growth of larval fishes (Overturf et al., 2012, 2014). Studies in amphibians have indicated disruption of the thyroid

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M.D. Overturf, D.B. Huggett / Aquatic Toxicology 161 (2015) 33–40

Fig. 1. Experimental design for examining various durations of exposure for levonorgestrel in early-life stages of the fathead minnow. Bolded lines with double arrows indicate duration of LNG exposure for that particular treatment group. Outside of the bolded lines with double arrows signify when larval were in clean water. Endpoints measured were: hatchability – 0 dph; survival and growth – 28 dph; and mRNA expression – 0, 3, 7, 14, and 28 dph.

system by LNG resulting in impaired metamorphosis of Xenopus laevis tadpoles (Lorenz et al., 2011). The reduction in growth of larval fish exposed to progestins may be attributed to the disruption of thyroid responsive pathways. However, mRNA expression during early-life stages of fish has been investigated and has indicated changes in expression of steroidogenic enzymes, steroid receptors, and gonadotropins fathead minnows 28 days post hatch (dph; Overturf et al., 2014) and zebrafish 48–96 h post fertilization (hpf; Zucchi et al., 2012). With gonadal steroids linked to growth in fish (Bhatt et al., 2012), it is critical to understand if the impacts that synthetic progestins have on the steroidogenic pathway are linked to the observed reduction in larval fish growth. Adverse outcome pathways (AOP), conceptual frameworks that link direct molecular events to an adverse outcome relevant to risk assessment (Ankley et al., 2010), allow risk assessors to evaluate chemicals using fewer resources and animals. These pathways are valuable in identifying key events in which high-throughput testing methods can be developed and applied (OECD, 2011; Volz et al., 2011). Measuring mRNA transcripts following an early-life stage toxicity study is one such approach to establish endocrine disruption markers for AOPs; however, there may be critical windows of sensitivity during the typical 28 day early-life stage study in which mRNA transcripts are more greatly influenced. Also, such windows of sensitivity may determine if shorter duration studies could be conducted to reduce testing time and cost. In a previous study, Overturf et al. (2014) demonstrated that a 28 day LNG exposure resulted in reduced larval growth in fathead minnows along with the down-regulation of mRNA transcripts for follicle stimulating hormone (FSH␤) and a few steroidogenic enzymes: hydroxysteroid dehydrogenases (3␤-HSD and 20␤-HSD) and aromatase (CYP19a1). Therefore, the purpose of this study was to investigate if critical windows of sensitivity for growth exist in the exposure of LNG to larval fathead minnows and to verify that the aforementioned gene transcripts are responsible for any observed growth effects. Fathead minnow larvae were exposed during dif-

ferent timeframes and then allowed to grow out in clean water for the duration of the test (Fig. 1). Mean dry weight, as a measure of growth, was evaluated at the end of the study; whereas, mRNA expression of the above mentioned genes was analyzed at the end of LNG exposure as well as the end of the test. This is the first study to investigate different exposure durations of a pharmaceutical throughout larval development within the framework of the OECD (1992), 210 Fish Early-Life Stage Toxicity Test. 2. Materials and methods 2.1. Chemicals Levonorgestrel (LNG), dimethylformamide (DMF), TRI reagent, Tris–EDTA, isopropanol, ethanol, methanol, hexane, and ethyl acetate were all purchased from Sigma–Aldrich (St. Louis, MO). RNAlater and RNAseZap were both purchased from Ambion (Life Technologies, Grand Island, NY). Rotor-Gene SYBR Green RT-PCR Kit was purchased from Qiagen (Valencia, CA). 2.2. Experimental protocol The study was conducted in accordance to the OECD 210 Guidelines (1992) with modifications (Fig. 1) described below. Newly fertilized (∼24 h old) fathead minnow embryos (Aquatic Biosystems, Ft. Collins, CO) were divided among seven groups (100 embryos/group) consisting of a dilution water control, solvent vehicle control (DMF), and five treatment groups. Embryos were divided equally into four replicates for each control and treatment groups. All treatment groups received a nominal dose of LNG at 125 ng/L, however, for different exposure periods. The nominal concentration of 125 ng/L LNG was the lowest observed effect concentration for growth in a 28 day fathead minnow early-life stage toxicity study (Overturf et al., 2014). Fig. 1 demonstrates the experimental design with exposure durations as follows: fertilized egg through 3,

M.D. Overturf, D.B. Huggett / Aquatic Toxicology 161 (2015) 33–40

35

Table 1 Primers used for real-time PCR analysis. Provided in the table are NCBI accession number, FW and RV primer sequences, the goodness of fit (R2 ) for the standard curve, and efficiency of the primer set. Gene

NCBI accession

FW primer (5 –3 )

RV primer (5 3 )

R2

Efficiency (%)

FSH 3␤-HSD 20␤-HSD CYP19a RPL8

DQ242616 DT361291 DT259130 AF288755 AY919670

TCGGCTTTCCAATATCTCCATT ATGAGATGCCCTACCCAAAGAC TGTCATGCTCTTCTGCCGATA TGCTGACACATGCAGAAAAACTC CATACCACAAGTACAAGGCCAAGA

ATGCAGTTGTGTCGAGATGTGAT CCCTTTACCTTTGTGCCATTG CTTGCTAACAAAGCTGGACACATT CAGCTCTCCGTGGCTCTGA ACCGAAGGGATGCTCAACAG

0.998 0.993 0.995 0.996 0.995

110 103 97 103 97

3–7, 7–14, 14–28 days post hatch (dph) and fertilized egg through 28 dph. A static renewal system was utilized with a 60% water change daily with each replicate. A stock solution of LNG was prepared with 100% DMF; however, the final solvent volume in each treatment was 0.001%. After each exposure duration, larvae were transferred to clean dilution water for the remainder of the test. Test vessels were placed in a walk-in environmental chamber where temperature and photoperiod were maintained at 23 ± 1 ◦ C and 16 h light:8 h dark, respectively. Larvae were fed daily Artemia nauplii ad libitum daily. Survival was monitored daily with percent survival calculated at 28 dph. Hatching success was determined at 0 dph. Six larvae were collected from dilution water controls at 0, 3, 7, 14, and 28 dph, whereas, six larvae were only collected at 3, 7, 14, and 28 dph for the solvent control group. For the treatments groups, six larvae were collected at the end of duration for that treatment and at 28 dph (i.e., six larvae were collected at 3 dph for the egg through 3 dph treatment group). All larvae collected were placed in RNAlater and stored at −20 ◦ C until further processing. Growth was monitored at 28 dph with remaining larvae and represented by mean dry weight for each treatment. For dry weight measurements, 10 larvae from each replicate were pooled and measured using a Mettler H51AR analytical balance following drying for 24 h at 60 ◦ C. 2.3. RNA isolation RNA was isolated from individual whole fathead minnow larvae that were placed in RNA later. All supplies and surfaces were RNase free and/or treated with 75% ethanol and RNAseZap. Samples were then homogenized in TRI reagent, vortexed, and centrifuged. The resulting middle transparent layer was then transferred into chloroform, vortexed, and then centrifuged leading to a separation into 3 phases: bottom layer (protein), middle layer (DNA), and a top transparent layer (RNA). The RNA was removed and added to isopropanol and centrifuged. The resulting pellet was then washed with 75% ethanol, centrifuged, and followed by the removal of ethanol. At this point, the pellet was re-suspended in 30 ␮l Tris–EDTA buffer and stored at −80 ◦ C until processing by qRTPCR. Purity and concentration of isolated RNA were analyzed using a Synergy 2 multi-mode microplate reader, Gen5 software, and a Take3 micro-volume plate (BioTek Winooski, VT). RNA quality was assessed with the Bio-Rad ExperionTM Automated Electrophoresis System (Hercules, CA). The RNA Quality Indicator value for all samples was greater than 9. All RNA samples were diluted with Tris–EDTA buffer to 40 ng/␮L for qRT-PCR processing. 2.4. qRT-PCR method mRNA expression of FSH, 3␤-HSD, 20␤-HSD, and CYP19a (aromatase) were analyzed by using the Rotor-Gene SYBR Green RT-PCR kit. Primers for each transcript analyzed are found in Table 1. Reaction samples were prepared with 40 ng/␮L RNA plus 200 nmol forward primer and 200 nmol reverse primer. Final concentrations in each reaction vial were as follows: SYBR Green 1×, forward primer 1 ␮M, reverse primer 1 ␮M, reverse transcriptase 0.25 ␮L, RNA 4 ng/␮L, and 19% RNase-free water. qRT-PCR was per-

formed using a Qiagen Rotor-Gene 6000 cycler following the Qiagen Rotor-Gene SYBR Green handbook procedure for Real-Time OneStep RT-PCR. Reverse transcription took place at 55 ◦ C for 10 min followed by a 5 min incubation period at 95 ◦ C. Two-step cycling was carried out for 5 s and 10 s at 95 ◦ C and 60 ◦ C, respectively. Real-time fluorescence and cycle times (Ct value) were compared to the reference gene RPL8. Ct values were then analyzed using the 2−CT method (Livak and Schmittgan, 2001). RPL8 expression did not change between the control and treatment groups (ANOVA, p > 0.05). 2.5. Liquid chromatography–tandem mass spectrometry (LC–MS/MS) At two periods during the experiment, mean measured LNG concentrations were determined over the 24 h period using liquid chromatography–tandem mass spectrometry (LC–MS/MS). Specifically, water samples were taken immediately and then 24 h after re-dosing. d6-norethindrone (d6-NOR; Toronto Research Chemicals, Toronto, ON, Canada) was utilized as an internal standard for quantification of LNG. Levonorgestrel was extracted from water samples using a liquid–liquid extraction method (1:1 mixture of hexane and ethyl acetate). Both LNG and d6-NOR required a derivatization process using a mixture of hydrazinopyridine and trifluoroacetic acid (HP/TFA) for analysis (Hala et al., 2011). Chromatographic separation of derivatized LNG was conducted using a Waters 2695 separations module coupled to a Waters 2998 UV–vis detector and a Waters Sunfire C18 column (2.1 × 50 mm, 3.5 ␮m particle size; Milford, MA, USA). Electrospray ionization (positive ion mode) and mass-spectrometric analysis was conducted using a quadrupole–hexapole–quadrupole instrument (Micromass Quattro UltimaTM mass detector, Manchester, UK). Samples were quantified using a nine point standard curve ranging from 0.39–100 ␮g/L. Efficiency of the LC–MS/MS was verified by analyzing check standards (i.e., 12.5 ␮g/L) before and after each run. 2.6. Statistical analysis Differences in survival, dry weight, and mRNA expression were determined by testing for statistical significance of treatment levels relative to controls. A Shapiro-Wilk’s and Bartlett’s test was used to test for normality and homogeneity of variances, respectively. A one-way analysis of variance (ANOVA) followed by a Tukey’s HSD (dry weight) and Dunnett’s (mRNA expression) post-hoc test was used to determine significance relative to the controls. Significance was reported at ˛ ≤ 0.05. GraphPad (La Jolla, CA) Prism 5 was utilized for statistical and graphical analysis of data. 3. Results 3.1. Water quality and nominal concentration verification Water quality was monitored weekly throughout the duration of the study. Temperature and pH were maintained between 21.8–22.5 ◦ C and 7.6–8.4 ◦ C, respectively. Alkalinity

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Fig. 2. Growth (average dry weight) of fathead minnow larvae at 28 dph. Letters above bars indicate significant differences between treatment groups (Tukey’s HSD, p ≤ 0.05). Values are means ± standard error of the mean, n = 4.

a)

b)

FSH

3β-HSD 1.5

Fold Change (log2)

Fold Change (log2)

1.5 1.0 0.5 0.0 -0.5 0

3

7

14

***

1.0

*

0.5 0.0 -0.5

28

0

Days post hatch

c)

7

14

28

Days post hatch

d)

20β-HSD

CYP19a1 1.5

1.0

*

0.5

*

***

0.0 -0.5 0

3

7

14

28

Days post hatch

Fold Change (log2)

1.5

Fold Change (log2)

3

*

*

14

28

1.0 0.5 0.0 -0.5 0

3

7

Days post hatch

Fig. 3. mRNA expression of fathead minnow larvae not exposed to LNG at different time points throughout the study. (a) FSH, (b) 3␤-HSD, (c) 20␤-HSD, and (d) CYP19a. *(p ≤ 0.05) and ***(p ≤ 0.001) indicate significant differences between 0 days post hatch and subsequent time points (Dunnett’s post hoc test). Values are means ± standard error of the mean, n = 6.

and hardness were maintained between 110–115 mg/L and 160–174 mg/L CaCO3 , respectively. Conductivity ranged from 522 to 528 ␮mhos/cm, while dissolved oxygen ranged from 7.8 to 8.2 mg/L. Levonorgestrel was not detected (MDL = 0.390 ng/L) in dilution water controls. The nominal concentration for all test solutions was 125 ng/L. Renewed solutions of LNG ranged were on average 99.1% of the nominal concentration, while solutions prior to renewal (24 h old) were on average 45.6% of the nominal concentration. Therefore, the mean measured concentration of LNG over the 24 h period

was calculated to be 72.4% of the nominal concentration resulting in an exposure concentration for LNG of 90.5 ng/L. 3.2. Survival and growth Hatching success among all treatments was greater than 90%, while survival for all treatments was greater than 85% (data not shown). Both hatching success and percent survival meet testing criteria for the OECD 210 Guidelines (1992). Growth was significantly reduced (F(6,21) = 10.43, p < 0.0001) in all treatment groups

M.D. Overturf, D.B. Huggett / Aquatic Toxicology 161 (2015) 33–40

a)

b)

FSH

3β-HSD

1.0

1.0

0.5 0.0 -0.5 -1.0

***

-1.5

***

Fold Change (log2)

Fold Change (log2)

37

-2.0

0.5 0.0 -0.5 -1.0

*

-1.5

**

**

Levonorgestrel Exposure Duration

So l

Levonorgestrel Exposure Duration

c)

20β-HSD

d)

CYP19a1

ve

nt C on tr ol Eg g -3 d 3d -7 d 7d -1 4d 14 d -2 8d Eg g -2 8d

So lv en tC on tr ol Eg g -3 d 3d -7 d 7d -1 4d 14 d -2 8d Eg g -2 8 d

-2.0

1.0

0.0 -0.5 -1.0

*

**

*

-1.5

nt C on tr ol Eg g -3 d 3d -7 d 7d -1 4d 14 d -2 8d Eg g -2 8d ve

0.5 0.0 -0.5 -1.0 -1.5

**

**

***

-2.0

-2.0

So l

Fold Change (log2)

0.5

Levonorgestrel Exposure Duration

So lv en tC on tr ol Eg g -3 d 3d -7 d 7d -1 4d 14 d -2 8d Eg g -2 8d

Fold Change (log2)

1.0

Levonorgestrel Exposure Duration

Fig. 4. mRNA expression following each treatment group’s exposure to LNG. Different treatment groups had been exposed to LNG at different times during development (Fig. 1). (a) FSH, (b) 3␤-HSD, (c) 20␤-HSD, and (d) CYP19a. *(p ≤ 0.05), **(p ≤ 0.01), and ***(p ≤ 0.001) indicate significant differences between treatment groups and control (Dunnett’s post hoc test). Values are means ± standard error of the mean, n = 6.

relative to control and solvent control (Tukey’s HSD, p < 0.05; Fig. 2).

3.3. mRNA expression Fathead minnow larvae were collected at periods throughout the study based on the critical periods of sensitivity (Fig. 1). For controls, six larvae were collected at 0, 3, 7, 14, and 28 dph, whereas, larvae were only collected at 3, 7, 14, and 28 dph for solvent controls. Larvae were also collected from LNG treatments following exposure duration and at 28 dph. mRNA expression for the four genes analyzed was detected in all samples. Solvent controls resulted in no significant differences in mRNA expression compared to controls at their respective time points (data not shown). All data presented below are in relation to solvent controls collected at the sampling time point with the exception to the results provided in Fig. 3 in which water only controls at each sampling time point were compared to larvae collected at 0 dph. For graphical purposes, only one control was used for Fig. 4 as no significant differences were found between the fold change (log 2) values among the controls from each sampling time point (data not shown). Expression for FSH resulted in no significant changes (F(4,24) = 2.26, p = 0.09) throughout development among unexposed larvae (Fig. 3a). FSH transcripts were only significantly down-regulated following exposure to LNG for durations of 14–28 dph and egg to 28 dph (F(5,28) = 15.24, p < 0.0001; Dunnett’s, p < 0.001; Fig. 4a).

3␤-HSD expression continuously increased throughout development among larvae not exposed to LNG resulting in a significant increase at 14 and 28 dph (F(4,24) = 7.53, p = 0.0004; Dunnett’s, p < 0.05; Fig. 3b). Similar results were seen in 3␤-HSD expression as was in FSH expression following LNG exposure. Both the 14–28 dph and egg to 28 dph exposures resulted in significantly down-regulation of 3␤-HSD (F(5,28) = 10.65, p < 0.0001; Dunnett’s, p < 0.01; Fig. 4b). In contrast to FSH, 3␤-HSD expression was downregulated after the 7–14 dph exposure (F(5,28) = 10.65, p < 0.0001; Dunnett’s, p = 0.031; Fig. 4b) and once LNG exposure was removed following the 7–14 dph exposure (F(3,25) = 7.64, p = 0.0009; Dunnett’s, p = 0.0003; Fig. 5b). For 20␤-HSD, expression steadily increased throughout the duration of the test resulting in a significant increase at 7, 14, and 28 dph as compared to time-point 0 (F(4,24) = 8.24, p = 0.0002; Dunnett’s, p ≤ 0.03; Fig. 3c). 20␤-HSD expression was down-regulated following exposure to LNG in exposures with durations ≥7 days (F(5,29) = 4.32, p = 0.0047; Dunnett’s, p ≤ 0.02; Fig. 4c). Furthermore, once larvae were moved to clean water, 20␤-HSD expression in larvae remained down-regulated at 28 dph in the 7–14 dph treatment (F(3,19) = 7.66, p = 0.0015; Dunnett’s, p = 0.013; Fig. 5c). Similar results are demonstrated with CYP19a expression as is with 20␤-HSD. Throughout the 28 dph, CYP19a expression was continuously increased with significant changes at 14 and 28 dph (F(4,24) = 4.66, p = 0.0063; Dunnett’s, p < 0.05; Fig. 3d). LNG exposure resulted in reductions in CYP19a expression in all exposure durations, in which significant down-regulation occurred in the

38

M.D. Overturf, D.B. Huggett / Aquatic Toxicology 161 (2015) 33–40

a)

b)

FSH

3β-HSD 1.0

Fold Change (log2)

Fold Change (log2)

1.0 0.5 0.0 -0.5 -1.0 -1.5 -2.0

0.5 0.0 -0.5 -1.0

-2.0 Solvent Control

Egg - 3d (25)

3d - 7d (21)

7d - 14d (14)

Solvent Control

Levonorgestrel Exposure Duration

c)

d)

20β-HSD

Egg - 3d (25)

3d - 7d (21)

7d - 14d (14)

Levonorgestrel Exposure Duration

CYP19a1 1.0

0.5 0.0 -0.5

*

-1.0 -1.5

Fold Change (log2)

1.0

Fold Change (log2)

***

-1.5

0.5 0.0 -0.5 -1.0 -1.5 -2.0

-2.0 Solvent Control

Egg - 3d (25)

3d - 7d (21)

7d - 14d (14)

Levonorgestrel Exposure Duration

Solvent Control

Egg - 3d (25)

3d - 7d (21)

7d - 14d (14)

Levonorgestrel Exposure Duration

Fig. 5. mRNA expression of fathead minnows at 28 dph for those treatment groups that were placed in water for remainder of study (see Fig. 1). (a) FSH, (b) 3␤-HSD, (c) 20␤-HSD, and (d) CYP19a. Number in parentheses under each x-axis label indicates the number of days of depuration for each treatment. *(p ≤ 0.05) and ***(p ≤ 0.001) indicate significant differences between treatment groups (Dunnett’s post hoc test). Values are means ± standard error of the mean, n = 6.

7–14 dph, 14–28 dph, and egg to 28 dph (F(5,28) = 5.97, p = 0.0007; Dunnett’s, p ≤ 0.01; Fig. 4d). In contrast with 20␤-HSD, CYP19a recovered to baseline levels once LNG exposure was removed (Fig. 5d). 4. Discussion This is the first study to investigate if critical windows of sensitivity to a synthetic progestin exist in early-life stages of a teleost fish. A similar study was conducted in fathead minnow larvae in which exposure of early-life stages to the synthetic estrogen, ethinylestradiol, led to alterations in gonadal development in adult stages (van Aerle et al., 2002). In contrast, this study was terminated at 28 dph with only growth and alterations in mRNA expression being measured. Standard early-life stage study (OECD, 1992) endpoints include survival and growth assessments at 28 dph. The reduction of growth, as mean dry weight, in this study is comparable to previous studies with synthetic progestins. Both NOR and LNG reduce growth in fathead minnow larvae at 740 and 86.9 ng/L, respectively (Overturf et al., 2012, 2014). Contrarily, X. laevis tadpoles increased weight after exposure to 3124 ng/L LNG which can be attributed to the overstimulation of thyroid stimulating hormone ␤ (TSH␤; Lorenz et al., 2011). The thyroid system of larval fathead minnows has not been thoroughly investigated to explain the changes in growth observed in previous studies; although changes in expression of steroid hormone receptors, steroidogenic enzymes, and gonadotropins have been examined (Overturf et al., 2014). The expression of follicle stimulating hormone (FSH) and

three steroidogenic enzymes (3␤-HSD, 20␤-HSD, and CYP19a) was down-regulated at 86.9 ng/L LNG in which growth was also reduced in fathead minnow larvae (Overturf et al., 2014). Therefore, these transcripts were investigated in the present study. In fish, the natural progestin, 17␣,20␤-dihydroxy-4-pregnen3-one (17,20-DP), is converted from 17␣-OH-progesterone by the enzyme 20␤-HSD. In the present study, 20␤-HSD was downregulated after exposure to LNG for periods ≥7 days. Adult female fathead minnows showed similar decreases in 17,20-DP after 7 days of exposure to LNG (Overturf et al., 2014). These data potentially suggest that synthetic progestins induce a negative feedback to decrease 17,20-DP production in fish. In mammals, progesterone production is stimulated by the gonadotropin LH; subsequently, progesterone inhibits the release of LH through gonadotropin releasing hormone (GnRH) modulation involving the binding of progesterone to the nuclear progesterone receptor (Skinner et al., 1998). With LNG having high affinity for mammalian nuclear progesterone receptors (Africander et al., 2011), this potential mechanism may also occur in teleost species; however, additional research is needed to verify these claims. CYP19a is responsible for catalyzing the aromatization of androgens to estrogens; thereby, converting testosterone to 17␤-estradiol, in addition, to the formation of estrone from androstenedione. Estrone is further converted to 17␤-estradiol through 17␤-HSD. Therefore, CYP19a is important in regulating estrogen concentrations in both mammals and fish. Gonadal steroids are linked to growth in piscine species. Bhatt et al.

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(2012) demonstrated that removal of gonads led to a decrease in body weight of Nile tilapia (Oreochromis mossambicus) and consequently steroid hormone levels including 17␤-estradiol were also significantly reduced. Exposure of adult female fathead minnows to LNG for 7 days resulted in a non-significant trend of reduced 17␤-estradiol (Overturf et al., 2014). In the present study, CYP19a1 was significantly reduced after exposure to LNG following exposure durations ≥7 days. Therefore, the down-regulation of CYP19a1 may lead to a reduction in 17␤-estradiol which could decrease the mean dry weight of fathead minnow larvae. However, removal of LNG exposure resulted in CYP19a1 expression returning to baseline levels in the present study. Therefore, a longer study design may be needed to determine if the change in CYP19a1 expression can lead to a recovery in body weight of fathead minnows. In addition, 17␤-estradiol is crucial for sexual differentiation. Larval fishes exposed to CYP19a inhibitors resulted in the sex-reversal of genotypic females to phenotypic males, similar to that observed in fish larvae exposed to androgenic compounds (reviewed in Leet et al., 2011). In fathead minnows, ovaries begin developing 10 dph with pre-meiotic cells emerging between 10 dph and 25 dph and primary oocytes developing 25–120 dph (van Aerle et al., 2004). With the down-regulation of CYP19a1 observed in the present study during the same windows of ovarian differentiation, changes in sexual differentiation may be seen in later developmental stage of fathead minnows, further supporting the androgenic potential of LNG. However, longer studies would be needed to investigate the role LNG may play in sexual differentiation, especially with the observed removal of LNG exposure leading to a recovery in CYP19a expression. In Coho salmon (Onchorhynchus kisutch), an increase concentration of FSH␤ was responsible for increases in 3␤-HSD (Luckenbach et al., 2011). Therefore, the decrease in FSH␤ expression seen in later stages of development in the present study may be responsible for down-regulation of 3␤-HSD in the same developmental stages. Natural androgens are also known to down-regulate 3␤-HSD in female rainbow trout (Onchorhynchus mykiss) at 65 days post fertilization (Govoroun et al., 2001). As LNG has a relative binding affinity of 58% to the mammalian androgen receptor (Sitruk-Ware, 2004), the down-regulation of 3␤-HSD in the present study may also be attributed to the androgenic potential of LNG as previously reported (Runnals et al., 2013; Zeilinger et al., 2009). The 7–14 dph exposure period appears to be a critical window of sensitivity for 3␤-HSD as it remained down regulated after LNG was removed. However, additional research may be needed to completely understand the potential impacts of the reduction in 3␤-HSD expression. Overall, these data suggest that while shorter-term tests (i.e., shorter exposure periods) may yield similar changes in growth, the changes in mRNA expression produced during these shorter exposure periods do not translate to long term changes associated with typical 28 day early life-stage toxicity studies. With the movement to shift toxicity testing from historical in vivo responses (i.e., changes in growth) to short term in vitro testing and predictive toxicology (NRC, 2007), there are concerns about relying on these alternative testing protocols for regulatory decisions (reviewed in Bus and Becker, 2009). As previously mentioned, the short term changes in mRNA expression in early life-stages presented in this study and in the literature (Zucchi et al., 2012) do not explain future implications of endocrine disruption throughout development. However, these data do contribute to understanding potential endocrine activity in teleost species. Therefore, future studies should take into consideration high-throughput testing (e.g., next-generation sequencing) throughout all life stages to

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further define mechanisms of disruption to endocrine active compounds to aid in refining testing strategies to align NRC’s goals. 5. Conclusions This is the first study to investigate the critical windows of sensitivity for a synthetic progestin in larval fathead minnows. Larval growth was decreased despite larvae being exposed to LNG. Of the mRNA transcripts investigated in this study, 20␤-HSD and 3␤-HSD remained down-regulated in the 7–14 dph exposure group once removed to clean water. Further testing would be needed to elicit the significance of these findings. Also, other endocrine related systems (i.e., thyroid system) need to be investigated to determine if the reduction in growth could be related to these systems. Furthermore, determining the critical windows of sensitivity for endocrine active compounds would aid in refining regulatory testing strategies by reducing testing duration, which in turn, would reduce both the number of animals needed and the cost of testing. References Africander, D., Verhoog, N., Hapgood, J.P., 2011. Molecular mechanisms of steroid receptor-mediated actions by synthetic progestins used in HRT and contraception. Steroids 76, 636–652. Al-Odaini, N.A., Zakaria, M.P., Yaziz, M.I., Surif, S., 2010. Multi-residue analytical method for human pharmaceuticals and synthetic hormones in river water and sewage effluents by solid-phase extraction and liquid chromatography–tandem mass spectrometry. J. Chromatogr. A 1217, 6791–6806. Andersson, J., Ekheden, Y., Kaj, L., Remberger, M., Woldegeorgis, A., 2005. Results from the Swedish National Screening Programme, 2005. In: Subreport 1: Antibiotics, Anti-Inflammatory Substances, and Hormones. IVL Report, pp. 1–98. Ankley, G.T., Brooks, B.W., Huggett, D.B., Sumpter, J.P., 2007. Repeating history: pharmaceuticals in the environment. Environ. Sci. Technol. 41, 8211–8217. Ankley, G.T., Bennett, R.S., Erickson, R.J., Hoff, D.J., Hornung, M.W., Johnson, R.D., Mount, D.R., Nichols, J.W., Russom, C.L., Schmieder, P.K., Serrrano, J.A., Tietge, J.E., Villeneuve, D.L., 2010. Adverse outcome pathways: a conceptual framework to support ecotoxicology research and risk assessment. Environ. Toxicol. Chem. 29, 730–741. Bhatt, S., Iwai, T., Miura, C., Higuchi, M., Shimizu-Yamaguchi, S., Fukada, H., Miura, T., 2012. Gonads directly regulate growth in teleosts. PNAS 109, 11408–11412. Bus, J.S., Becker, R.A., 2009. Toxicity testing in the 21st century: a view from the chemical industry. Toxicol. Sci. 112, 297–302. Chang, H., Wan, Y., Wu, S., Fan, Z., Hu, J., 2011. Occurrence of androgens and progestogens in wastewater treatment plants and receiving river waters: comparison to estrogens. Water Res. 45, 732–740. Corcoran, J., Winter, M.J., Tyler, C.R., 2010. Pharmaceuticals in the aquatic environment: a critical review of the evidence for health effects in fish. Crit. Rev. Toxicol. 40, 287–304. Croxatto, H.B., Brache, V., Pavez, M., Cochon, L., Forcelledo, M.L., Alvarez, F., Massai, R., Faundes, A., Salvatierra, A.M., 2004. Pituitary–ovarian function following the standard levonorgestrel emergency contraceptive dose or a single 0.75-mg dose given on the days preceding ovulation. Contraception 70, 442–450. DeQuattro, Z.A., Peissig, E.J., Antkiewicz, D.S., Lundgren, E.J., Hedman, C.J., Hemming, J.D.C., Barry, T.P., 2012. Effects of progesterone on reproduction and embryonic development in the fathead minnow (Pimephales promelas). Environ. Toxicol. Chem. 31, 851–856. Durand, M., del Carmen Cravioto, M., Raymond, E.G., Duran-Sanchez, O., De la Luz Cruz- Hinojosa, M., Castell-Rodriguez, A., Schiavon, R., Larrea, F., 2001. On the mechanisms of action of short-term levonorgestrel administration in emergency contraception. Contraception 64, 227–234. Govoroun, M., McMeel, O.M., D’Cotta, H., Ricordel, M.-J., Smith, T., Fostier, A., Guiguen, Y., 2001. Steroid enzyme gene expression during natural and androgen-induced gonadal differentiation in the rainbow trout, Oncorhynchus mykiss. J. Exp. Zool. 290, 558–566. Hala, D., Overturf, M.D., Petersen, L.H., Huggett, D.B., 2011. Quantification of 2-hydrazinopyridine derivatized steroid hormones in fathead minnow (Pimephales promelas) blood plasma using LC-ESI+/MS/MS. J. Chromatogr. B 879, 591–598. Kidd, K.A., Blanchfield, P.J., Mills, K.H., Palace, V.P., Evans, R.E., Lazorchak, J.M., Flick, R.W., Kolpin, D.W., 2007. Collapse of a fish population after exposure to a synthetic estrogen. PNAS 104, 8897–8901. Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber, L.B., Buxton, H.T., 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in US streams: 1999–2000: a national reconnaissance. Environ. Sci. Technol. 36, 1202–1211. Leet, J.K., Gall, H.E., Sepúlveda, M.S., 2011. A review of studies on androgen and estrogen exposure in fish early life stages: effects on gene and hormonal

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