Developmental abnormalities and differential expression of genes induced in oil and dispersant exposed Menidia beryllina embryos

Aquatic Toxicology 168 (2015) 60–71

Contents lists available at ScienceDirect

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

Developmental abnormalities and differential expression of genes induced in oil and dispersant exposed Menidia beryllina embryos Olanike K. Adeyemo, Kevin J. Kroll, Nancy D. Denslow ∗ Department of Physiological Sciences and Center for Environmental and Human Toxicology, University of Florida, Gainesville, FL 32611, USA

a r t i c l e

i n f o

Article history: Received 15 July 2015 Received in revised form 19 September 2015 Accepted 21 September 2015 Available online 26 September 2015 Keywords: Menidia beryllina Fish development Oil spills Cardiotoxicity Dispersant

a b s t r a c t Exposure of fish embryos to relatively low concentrations of oil has been implicated in sub-lethal toxicity. The objective of this study was to determine the effects of the exposure of Menidia beryllina embryos at 30–48 h post-fertilization to the water accommodated fractions of oil (WAF, 200 ppm, v/v), dispersants (20 ppm, v/v, Corexit 9500 or 9527), and mixtures of oil and each of the dispersants to produce chemically enhanced water accommodated fractions (CEWAFs) over a 72-hour period. The polyaromatic hydrocarbon (PAH) and benzene, toluene, ethylene and xylene (BTEX) constituents of the 5X concentrated exposure solutions (control, WAF, dispersants and CEWAFs) were determined and those of the 1× exposures were derived using a dilution factor. PAH, BTEX and low molecular weight PAH constituents greater than 1 ppb were observed in WAF and the dispersants, but at much higher levels in CEWAFs. The WAF and CEWAFs post-weathering were diluted at 1:5 (200 ml WAF/CEWAF: 800 ml 25 ppt saltwater) for embryo exposures. Mortality, heartbeat, embryo normalcy, abnormality types and severities were recorded. The qPCR assay was used to quantify abundances of transcripts of target genes for sexual differentiation and sex determination (StAR, dmrt-1, amh, cyp19b, vtg and chg-L,), growth regulation (ghr) and stress response (cyp1a and Hsp90); and gapdh served as the housekeeping gene. Temperature was 21 ± 1.5 ◦ C throughout the experimental period, while mortality was low and not significantly different (p = 0.68) among treatments. Heartbeat was significantly different (0.0034) with the lowest heartbeats recorded in Corexit 9500 (67.5 beats/min) and 9527 (67.1 beats/min) exposed embryos compared with controls (82.7 beats/min). Significantly more treated embryos were in a state of deterioration, with significantly more embryos presenting arrested tissue differentiation compared with controls (p = 0.021). Exposure to WAF, dispersants and CEWAF induced aberrant expression of all the genes, with star, dmrt-1, ghr and hsp90 being significantly down-regulated in CEWAF and cyp19b in Corexit 9527. The cyp1a and cyp19b were significantly up-regulated in CEWAFs and WAF, respectively. The molecular endpoints were most sensitive, especially the expression of star, cyp19b, cyp1a, hsp90 and could therefore be used as early indicators of long term effects of Corexit 9500 and 9527 usage in oil spill management on M. beryllina, a valid sentinel for oil pollution events. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Oil dispersing agents have been employed in spill response for decades. Although the exact constituents of Corexit 9500 and 9527 in common use today is proprietary; they have been reported to consist of non-ionic and/or anionic surfactants in a solvent base designed to enhance oil miscibility under varying temperature and salinity conditions (Major et al., 2012; Place et al., 2010; Singer et al., 1996). When an oil pollution event and the subsequent cleanup

∗ Corresponding author. Fax: +1 352 392 4707. E-mail address: ndenslow@ufl.edu (N.D. Denslow). http://dx.doi.org/10.1016/j.aquatox.2015.09.012 0166-445X/© 2015 Elsevier B.V. All rights reserved.

occur, polycyclic aromatic hydrocarbons (PAHs) and other components of oil and the dispersant used in the cleanup process may persist in the marine environment for a long time thereby creating pathways for lingering biological exposure and associated adverse effects. The toxicity, carcinogenic, mutagenic, and teratogenic properties of PAHs are well documented (Boström et al., 2002; Stegeman et al., 1991). Studies have reported extreme impacts of crude oil and PAHs on the heart of exposed fish embryos including what was described as cardiogenic fluid accumulation syndrome (Brette et al., 2014; Carls et al., 1999; Heintz et al., 1999), and craniofacial and body axis defects (Incardona et al., 2005). GeorgeAres and Clark (2000) reported that oil spill dispersants, Corexit® 9500 and Corexit® 9527 induced low to moderate toxicity in most

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

aquatic species in laboratory tests; while Berninger et al. (2011) studied the effect of oil and oil/9500 mixtures on fish and shrimp species and concluded that the toxicity of oil increases when mixed with dispersant 9500. A similar study by Khan and Payne (2005) looked at the toxicity of 9527 to four different fish species and revealed results similar to Berninger et al., (2011) with an exception of Cunner fish (Tautogolabrus adspersus), where mortality was higher in fish exposed to oil alone. Thus they concluded that toxicity of oil and/or dispersants might be species specific. Adams et al., (2014) also provided evidence in support of the contention that dispersants increase the concentration of oil in test solutions without affecting the toxicity of the dispersed oil. Although aquatic pollutants do impact an extensive range of fish species, more studies use standard laboratory models such as medaka, fathead minnow and zebrafish to determine the impact of pollution (Ankley et al., 2008; Segner, 2009; Scholz and Mayer, 2008). Most environmental risk assumptions about sensitivity and long-term effects of pollutants on fish are based on these few species, which actually may not be natural inhabitants of the aquatic environment under consideration. Hence, Menidia beryllina, an estuarine Atherinid fish commonly known as the inland silverside has been proposed as a more widely distributed fish model, especially for North America because they are found in estuarine and brackish habitats throughout the coast (Brander et al., 2013, 2012; Middaugh and Hemmer, 1992). Atherinids are also listed by USEPA (2002) as model fish species for the Whole Effluent Toxicity Testing Program, while being reported to be more sensitive to toxicants when compared with other species (Clark et al., 1985). Moreover, estuaries are utilized by many species of fish for at least part of their lives and are therefore subject to a wide array of pervasive environmental contaminants. Gene expression profiles from the liver (Whitehead et al., 2012) and gill (Dubansky et al., 2013) tissues of the non-Artherinid Gulf killifish, Fundulus grandis have been used to identify exposure of fish to the toxic components of oil. These have been shown, especially in early life stages, to reflect the types of responses that are expected to precede long-term population-level effects. These include compensatory responses in genes associated with regulation of transcription, cell cycle progression, RNA processing, DNA damage, and apoptosis (Pilcher et al., 2014). Oil pollution from exploration and production processes, natural seeps, atmospheric contribution, freight accidents, industrial discharge, and urban run-off is a significant hazard for the marine environment (Wilson and LeBlanc, 2000). Dispersants are generally used as a frontline means of mitigating the impact of oil spills; nevertheless, the ecological implication of their use is still ambiguous. Dispersants increase the rates of natural hydrocarbon degradation by breaking up oil slicks, which increases the surface area for access to the oil by hydrocarbon-degrading bacteria. While this accelerates biodegradation, dispersion may also increase the amounts of PAH present in the water column resulting in a large increase (5–50 times) in the amount of aromatics and PAHs in the water column (Fingas and Banta, 2008). Exposure of fish embryos to relatively low concentrations of oil has been implicated in sub-lethal toxicity. However, the effect of oil and dispersants; singly and in combination should be more thoroughly evaluated to better understand and anticipate the ecological impacts. The present study therefore looked into the effect of oil and commonly used dispersants (Corexit 9500 and 9527), singly and in combination on the M. beryllina embryos. We also determined the PAHs benzene, toluene, ethylbenzene and xylenes (BTEX) and low molecular weight PAH constituents of the water-accommodated fraction of oil (WAF), dispersants and chemically enhanced water-accommodated fractions of dispersants and oil (CEWAFs). We hypothesized that exposure of Menidia embryos to WAF, CEWAFs, and each dispersant used alone would (1) disrupt tissue

61

differentiation and hence hatching; (2) result in abnormal development of hatched embryos; and (3) disrupt the expression of transcripts of genes associated with sexual differentiation, growth and the stress response.

2. Materials and methods 2.1. Preparation of exposure solutions Crude oil from the Deep Water Horizon used in this study was collected from the leaking well riser by BP (using Entrix sampler) and provided to us by AECOM Environment Toxicology Laboratory. The Corexit 9500 and 9527 were also obtained as a gift. Solutions of BP’s Deep Water Horizon crude oil collected from the leaking well riser, Corexit 9500, Corexit 9527, oil/9500, and oil/9527 mixed in artificial seawater (25 ppt Instant Ocean) were weathered as described by Hemmer et al. (2011). Each solution was prepared in 2.0 L pyrex glass containers and mixed for 7 days by vortexing with a stir bar, and allowed to degas under a fume hood (25 ◦ C). Mixing intensity was maintained so that a vortex extended 2–3 inches under the water surface. Oil (1000 ppm, v/v, 1 ml/L) and dispersants (100 ppm, v/v, 0.1 ml/L) were weathered singly and in combination at a ratio of 10:1 (oil: dispersant, v/v) by adding oil to the mixing seawater, followed by the dispersants according to the manufacturer’s recommended application rate. After 7 days, the solutions were allowed to settle overnight, and the WAF and CEWAF were separated from the oil using separatory funnels. The WAF and CEWAF solutions were further diluted 1:5 into 1 L pyrex glass holding containers with artificial seawater (25 ppt) for embryo exposures. The holding containers were sealed with a lid and opened temporarily to refresh each treatment daily. The final exposure solutions for WAF were the dissolved fraction from 200 ppm oil, 20 ppm for each of the dispersants alone, and the dissolved fraction from a mixture of 200 ppm oil/20 ppm dispersant for the two CEWAFs. This diluted solution was used for fish exposures. One liter subsamples from each undiluted preparation (control water, WAF (oil), dispersants (Corexit 9500 and 9527) and CEWAF (oil/9500, oil/9527)) were collected and shipped to Columbia Analytical Services (Kelso, WA) for analyses of PAH, BTEX and low molecular weight PAHs. As a preservative, 1 mL hydrochloric acid was added to each sample and the samples were stored at 4 ◦ C and shipped on ice. The PAH and BTEX constituents of the 1× exposure waters were derived based on a dilution factor of 5. Values of PAH or BTEX greater than 1 ppb in at least one of control, WAF, dispersants or CEWAFs are presented in the results. Analyses were conducted using gas chromatography/mass spectrometry (GC/MS) and GC MS/MS according to EPA 8270D standard operating method (USEPA, 2007). Oil fingerprinting was achieved using an HPLC/MS/MS following the EPA 3580A method (USEPA, 1992) and volatile organic constituents were determined using the EPA 5030B method with GC/PID (USEPA, 1996).

2.2. Embryo exposures M. beryllina (inland silversides) embryos at 30–48 h postfertilization were exposed in groups of 35–40 embryos per container in 100 ml of either control water or exposure medium that had been diluted 1:5 from the original weathered solutions (WAF, Corexit 9500 or 9527, and CEWAFs (oil/9500, oil/9527)). Exposures were performed in quadruplicate in uncovered glass containers for 72 h. The exposure solutions were renewed with a 50% change on a daily basis. All exposure vessels and measurement devices were made of glass throughout this experiment. Embryos were obtained from Aquatic Biosystems (Fort Collins, CO).

62

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

2.3. Sample collection Daily, embryos were observed for hatching success and mortality; temperature was taken and 50% water changes were performed. The experiments were terminated after 72 h and embryos in each treatment were viewed under the microscope. Video recordings of four embryos per replicate were taken for heart rate (as beats per min) for a total of N = 16 per exposure. Photographic images of abnormalities were also recorded. Using a modified graduated severity index (GSI) method according to Hose et al. (1996) to determine the severity of differentiation abnormality, scores were obtained as follows: 0 = normal embryo, 1 = arrested differentiation with heartbeat, 2 = arrested differentiation with twitching tissue without heart structure, 3 = arrested differentiation, no heartbeat and 4 = undifferentiated egg. Similarly, the number of abnormal embryos observed was recorded and normalcy scores were conferred as follows: a score of 0, if ≥5 abnormal embryos were seen in a microscopic field, 1, if 4 abnormal embryos were observed, 2, if 3 abnormal embryos, were observed, 3, if 2 abnormal embryos were observed, 4, if 1 abnormal embryo was observed, and 5, when no abnormal embryos were observed. Percentage normalcy was then calculated for each treatment as the number obtained per beaker for the four beakers per treatment. The maximum score obtainable was 20. For the gene expression study, 35–40 embryos were pooled per replicate exposure per treatment for a total of n = 4 pools, flash frozen and stored at −80 ◦ C for subsequent RNA extraction, purification and qPCR assay. 2.4. RNA extraction and purification RNA was extracted from pooled embryos using RNA STAT-60 reagent (Tel-TestTM Inc., Friendswood, TX, USA) and DNase treatment was carried out with Turbo DNA-free reagent (AmbionTM Austin, TX, USA) to remove any remaining genomic DNA, following procedures detailed by the manufacturers. The quality and quantity of RNA was first measured using the NanoDropTM ND1000 (NanoDropTM Technologies, Wilmington, DE, USA), and RNA integrity was confirmed using the AgilentTM 2100 BioAnalyzer with the RNA 6000 nanochip. The A260 /A280 values of all samples ranged from 1.82 to 2.04, while RNA integrity values were 8.6–10 for all samples used in the analysis. 2.5. Quantitative real-time PCR (qRT-PCR) In total, 9 target genes and GAPDH as the reference gene were quantified with RT-qPCR. cDNA was synthesized using SuperScriptTM II reverse transcriptase (InvitrogenTM , Carlsbad, CA). Vitellogenin (vtg), choriogenin L (chg-L), and glyceraldehyde-3phosphate dehydrogenase (gapdh), which was the housekeeping gene, had been previously developed for Menidia (Brander et al., 2013). As reported by Jayasinghe et al. (2014), assays for the following genes were also previously developed, validated and optimized: brain aromatase (cyp19b), steroidogenic acute regulatory protein (StAR), anti-mullerian hormone (amh), and doublesex and mab-3 related transcription factor (dmrt-1), and growth hormone receptor (ghr). We also developed, validated and optimized assays for an additional two genes: heat shock protein (hsp90), and cytochrome P450 A1 (cyp1a). The efficiency of amplification ranged between 98 and 105.7%. All primers used for the transcripts are contained in Table S1 (see Supplemental materials for sequences of primers). Real-time quantitative polymerase chain reaction (qPCR) analysis was performed using the AB 7900HT qPCR machine and analyzed by the RQ Manager both from Applied BiosystemsTM (Foster City, CA). The cDNAs (200 ng) from embryos were amplified in 384-well plates with SYBR green supermix (BioRad, Hercules, USA) using

qPCR primers (described in Table S1). Negative template controls (water only) were analyzed for each gene to determine if nonspecific products formed during PCR amplification. The samples and RNA (minus RT controls) were analyzed in duplicate using a two-step protocol with an initial denaturation at 95 ◦ C for 3 min and 40 cycles of 95 ◦ C for 10 s followed by a 1 min annealing step at 58 ◦ C ending with a melting curve with a gradually increasing temperature gradient for 30 s intervals from 55 to 95 ◦ C to verify that single PCR products were amplified. 2.6. Statistical analyses Statistical tests were performed using Graph-Pad Prism version 5.0 (GraphPad Software Inc., La Jolla, CA, USA). Data were expressed as mean ± SEM. One-way analysis of variance (ANOVA) was performed to test overall group differences. Because of the small sample size, data were not normally distributed. The nonparametric Kruskal–Wallis test followed by Dunn’s multiple comparisons test was applied. P < 0.05 was considered significant between control embryos and those exposed to oil and dispersants for mortality, heartbeat and differentiation data. Quantitative PCR data was transformed and analyzed using the relative quantification log 2 (−DCt) method (Livak and Schmittgen, 2001). Data are reported as the log 2 gene transcription relative to gapdh and normalized to the mean transcription of each gene corresponding to the experimental controls to allow for direct comparison. Differences in gene expression (i.e., mRNA levels) between exposures were assessed on normalized data using one-way ANOVA and post-hoc tests as described above. 3. Results 3.1. Quantitation of constituents of oil and dispersants in exposure solutions PAH and BTEX in control, WAF, dispersants or CEWAFs were analyzed by GC MS and chemicals with amounts greater than 1 ppb, present in at least one of these analyses, are presented. The PAH and BTEX constituents of the different solutions are shown in Tables 1 and 2. PAHs found in the oil were cis–trans-decalin (C1–C4), naphthalene (C1–C4), biphenyl, fluorene (C1–C4), phenanthrene (C1–C4), dibenzothiophene (C1–C4), fluoranthenes/pyrenes (C1–C4), naphthobenzothiophenes (C1–C3), chrysene (C1–C4), C30hopane, 4-methyldibenzothiophene, 3-methylphenanthrene, 2-methylphenanthrene, 9-methylphenanthrene, 1-methylphenanthrene, 2-methylnaphthalene, 12,6-dimethylnaphthalene and methylnaphthalene, 2,3,5-trimethylnaphthalene. It is also remarkable that Corexit 9500 contained cis–trans-decalin (C1–C4), naphthalene (C2), fluorene (C2) and phenanthrene in addition to more hydrophilic components not measured in this analysis; while Corexit 9527 contained cis/trans-decalin, naphthalene (C1–C3), biphenyl, fluorene, phenanthrene (C1–C3), 1-methylnaphthalene and 2methylnaphthalene in addition to the hydrophilic components, not analyzed herein. It is however noteworthy that the PAH constituents of the CEWAF were similar to WAF but at much higher quantities. For BTEX determinations, benzene-124 and naphthalene were isolated in WAF and Corexit 9500 at levels greater than 1 ppb. In addition to these, isopropylbenzene and n-propylbenzene were isolated from Corexit 9527. Similar to what was observed with PAHs, it appears that the dispersants had an additive effect on the levels of BTEX and low molecular weight PAH constituents in CEWAF (Table 2).

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

63

Table 1 PAH constituents in WAF, dispersants and CEWAF above 1 ppb. Concentrations are in ppb (␮L). Chemical name

Control

cis/trans-Decalin C1-decalinsa C2-decalins C3-decalins C4-decalins Naphthalene C1-naphthalenes C2-naphthalenes C3-naphthalenes C4-naphthalenes Biphenyl Fluorene C1-fluorenes C2-fluorenes C3-fluorenes Phenanthrene C1-phenanthrenes/Anthracenes C2-phenanthrenes/Anthracenes C3-phenanthrenes/Anthracenes C4-phenanthrenes/Anthracenes Dibenzothiophene C1-dibenzothiophenes C2-dibenzothiophenes C3-dibenzothiophenes C4-dibenzothiophenes C1-fluoranthenes/Pyrenes C2-fluoranthenes/Pyrenes C3-fluoranthenes/ Pyrenes C4-fluoranthenes/Pyrenes C1-naphthobenzothiophenes C2-naphthobenzothiophenes C3-naphthobenzothiophenes Chrysene C1-chrysenes C2-chrysenes C3-chrysenes C4-chrysenes C30-hopane 4-Methyldibenzothiophene 3-Methylphenanthrene 2-Methylphenanthrene 9-Methylphenanthrene 1-Methylphenanthrene 2-Methylnaphthalene 1-Methylnaphthalene 2,6-Dimethylnaphthalene 2,3,5-Trimethylnaphthalene SUM of PAHs a

Oil

0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00

0.01 0.02 0.06 0.16 0.26 5.08 3.18 1.76 0.81 0.41 0.29 0.22 0.26 0.31 0.27 0.34 0.58 0.54 0.34 0.26 0.06 0.16 0.20 0.16 0.09 0.06 0.12 0.13 0.10 0.05 0.06 0.05 0.03 0.08 0.10 0.07 0.06 0.05 0.07 0.13 0.14 0.16 0.12 3.09 2.54 0.75 0.26 24.08

Corexit 9500 0.02 0.02 0.70 5.08 10.08 1.13 0.00 0.49 0.00 0.00 0.00 0.00 0.00 0.01 0.00 0.01 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0 17.55

Corexit 9527 0.02 0.00 0.00 0.00 0.00 0.01 0.01 0.01 0.01 0.00 0.01 0.01 0.00 0.00 0.00 0.01 0.02 0.01 0.08 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.01 0.01 0.00 0.0025 0.24

Oil/9500 0.07 0.37 0.64 1.22 1.46 5.08 5.76 4.60 2.25 1.09 0.58 0.29 0.43 0.76 0.59 0.50 0.90 0.88 0.56 0.39 0.10 0.27 0.43 0.26 0.16 0.10 0.19 0.22 0.17 0.09 0.12 0.08 0.06 0.14 0.16 0.12 0.09 0.08 0.11 0.21 0.21 0.26 0.19 5.00 4.04 2.09 0.80 44.18

Oil/9527 1.62 5.04 10.32 11.64 10.00 9.96 20.64 24.00 14.40 7.64 2.54 1.23 2.53 3.50 3.19 2.40 5.40 5.56 3.53 2.42 0.41 1.30 1.87 1.40 0.92 0.62 1.22 1.40 1.02 0.69 0.82 0.70 0.53 1.17 1.58 1.12 0.88 0.72 0.66 1.20 1.35 1.54 1.13 17.40 13.00 10.72 4.32 217.26

The C# refers to the number of carbons in the chain.

Table 2 BTEX and some low molecular weight PAH constituents in WAF, dispersants and CEWAF above 1 ppb. Concentrations are in ppb (␮g/L). Chemical

Control

Oil

Corexit 9500

Corexit 9527

Oil/9500

Oil/9527

Benzene Toluene Ethylbenzene m,p-Xylenes o-Xylene Isopropylbenzene n-Propylbenzene Trimethylbenzene-135 Trimethylbenzene-124 Naphthalene SUM

0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

2.8 1.6 0.3 1.7 0.8 0.0 0.0 0.1 0.3 0.3 8.1

2.5 3.6 0.3 1.5 0.7 0.0 0.0 0.1 0.3 0.3 9.4

11.6 8.0 1.4 7.6 3.5 0.1 0.2 0.5 1.5 1.4 35.8

23.2 37.2 2.9 14.8 6.8 0.3 0.4 0.9 2.8 2.1 91.5

44.0 72.0 5.2 27.6 12.4 0.6 0.8 1.8 5.2 3.8 173

3.2. Heart rate and tissue differentiation Average temperature (± standard deviation) throughout the experimental period was 21 ± 1.5 ◦ C. Mortality was low and not significantly different (p = 0.68) for controls and embryos exposed to WAF, dispersants and CEWAFs. Heartbeat per minute

(Fig. 1) was significantly different (p = 0.0034) with the lowest beats (bradycardia) recorded in Corexit 9500 (67.5 beats/min) and 9527 (67.1 beats/min) exposed embryos compared with controls (82.7 beats/min). Significantly (p = 0.021) more embryos exposed to oil and dispersants were in a state of arrested differentiation (p = 0.021)

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

5

*

4

*

*

*

3 2 1

00 O il/ 95 27

il/ 95

O

ex

it

95 00 or C

C

or

ex

it

C

95 27

O il

0

on tr ol

Abnormality severity score

64

Exposures Fig. 1. Heart rate in Menidia beryllina embryos exposed to WAF, dispersants and CEWAFs. Heart rate was measured after 72 h of exposure. N = 16 per group. Results are depicted as box and whisker plots with the horizontal line in the box representing the median, and the maximum and minimum values as the whiskers. Significantly different groups were determined using the one-way ANOVA, P < 0.05.

Mean normalcy score

6

*

4

* *

2

Fig. 3. Abnormal differentiation severity score in Menidia beryllina embryos exposed to WAF, dispersants and CEWAFs. N = 16. Scores were developed using the algorithm of Hose et al. (1996) with 0 = normal embryo, 1 = arrested differentiation with heartbeat, 2 = arrested differentiation with twitching tissue without heart structure, 3 = arrested differentiation, no heartbeat and 4 = undifferentiated egg. Results are depicted as box and whisker plots with the horizontal line in the box representing the median, and the maximum and minimum values as the whiskers. Significantly different groups were determined using one-way ANOVA.

Skeletal abnormalities (Fig. 4B) were observed in hatchlings of WAF, dispersant and CEWAF exposed embryos. Additionally lower hatching incidences were observed in CEWAF (9527) exposed embryos. Other abnormalities were head malformations, pericardial and abdominal edema, which were predominant in embryos exposed to Corexit 9500 and CEWAF (oil/9500) (Fig. 5). 3.4. Gene expression

O il/ 95 27

00 95 il/

O

95 27 it

ex or C

it

95 00

O il

ex C

or

C on tr ol

0

Exposures Fig. 2. Normalcy score for Menidia beryllina embryos exposed WAF, dispersants and CEWAFs. Embryos were scored based on the number of abnormal embryos seen per group of 5 individuals per replicate with 4 replicates per condition. N = 20. Scores refer to the following: 0 = 5 abnormal embryos in a microscopic field, 1 = 4 abnormal embryos, 2 = 3 abnormal embryos, 3 = 2 abnormal embryos, 4 = 1 abnormal embryo and 5 = no abnormal embryos. Results are depicted as box and whisker plots with the horizontal line in the box representing the median, and the maximum and minimum values as the whiskers. Significantly different groups were determined using the one-way ANOVA.

compared with control (Fig. 2), except for oil and the oil/9527 combination where differences were not statistically significant. Exposure to dispersants and CEWAFs resulted in significantly (p = 0.017) more severe hatchling abnormalities than in the controls (Fig. 3). 3.3. Morphological abnormalities Fig. 6 demonstrates morphological abnormalities, characterized by deterioration of unhatched embryos (Fig. 4A) and the presence of more cellular debris (Fig. 4A and B) in dispersantand CEWAFs-exposed embryos when compared with control.

A differential expression of transcripts (mRNA levels) was observed for StAR, dmrt-1, amh, cyp19b, vtg and chg-L associated with steroidogenesis and sexual differentiation (Fig. 6a–f); ghr associated with growth regulation (Fig. 6g), and cyp1a, hsp90 associated with stress response (Fig. 6h and i). However, only the expressions of StAR, cyp19b, cyp1a and hsp90 were significantly different (p = 0.021, 0.039, 0.0028 and 0.02, respectively) compared with the control. Dmrt-1 was also significantly down-regulated in WAF, Corexit 9500 and CEWAF. 4. Discussion Crude oils are quite complex and contain lots of different chemical components like aliphatic and aromatic hydrocarbons (PAH), phenols, and other compounds, such as hydroxylated PAHs, among others (Du et al., 2011; Hong et al., 2012). While dispersants have been used to mitigate the effects of oil by making the oil more bioavailable to oil-eating bacteria (Chakraborty et al., 2012; Couillard et al., 2005; Mason et al., 2012; Redmond and Valentine, 2012), their effects on in situ marine flora and fauna, both in short and long term studies have not been fully explained (Almeda et al., 2013; White et al., 2012). There has been mixed opinion about the impact of combinations of oil and dispersants (Adams et al., 2014; Camilli et al., 2010; Lichtenthaler and Daling, 1985; Rico-Martínez et al., 2013). In a laboratory trial by Kuhl et al. (2013), juvenile F. grandis were exposed to WAF, Corexit 9500, and CEWAF treatments in microcosms at a dispersant:oil ratio of 1:10, as recommended by the manufacturer. They observed that Corexit 9500 and CEWAF

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

65

Fig. 4. Panel “A” is showing deterioration of Menidia beryllina embryos exposed to WAF, dispersants and both CEWAF’s compared to control with arrows pointing to abnormal embryos. Panel “B” demonstrates some skeletal malformations observed in hatchlings of Menidia beryllina embryos exposed to WAF, dispersants and CEWAFs with arrows pointing to abnormal hatchlings.

Fig. 5. Other abnormalities observed in hatchlings of Menidia embryos exposed to dispersants and CEWAFs were (a) head malformation, (b) pericardial edema, and (c) abdominal edema.

were consistently more lethal than undispersed oil. They therefore concluded that both Corexit 9500 and CEWAF remained toxic to juvenile F. grandis at 4 ppt nominal salinity for at least 4 weeks of the 16 experimental weeks. Rico-Martínez et al. (2013) also measured the acute toxicity of Corexit 9500 and Corexit dispersed Macondo crude oil to rotifers using the manufacturer’s recommended ratios for deploying Corexit (1:10–1:50) and the actual dispersant:oil ratio (1:130) used in the Deep Water Horizon spill. They submitted

that CEWAFs were 47–52 times more toxic than WAF, suggestive of synergism. The data presented by Hemmer et al. (2011) demonstrated that dispersants alone were less toxic than undispersed or dispersed oil; they rated Corexit 9500 alone as being practically non-toxic and CEWAF as moderately toxic to 11–14 day old M. beryllina, based on their results of 96 h LC50 s (130 ␮l/L and 7.6 mg total petroleum hydrocarbons/L, respectively) both derived from standard short term acute tests at definitive concentrations ranging

66

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

Fig. 6. Q-PCR evaluation of transcript levels for genes of interest in 72 h old embryos. (a–f) shows differential expression of transcripts for genes associated with sex and sexual differentiation, g, differential expression of transcripts for genes associated with growth and (h and i), differential expression of transcripts for genes associated with stress response in WAF, dispersants and CEWAFs exposed Menidia embryos. Significantly different groups were determined using ANOVA.

from 32 to 320 ␮l/L for the dispersant alone and 3.1–100% CEWAF (USEPA, 2002). In estuarine ecosystems, fish are reliable indicator organisms due to their size and mid to high trophic position (Brewton et al., 2013). Menidia are a good sentinel species to assess the effects of the oil spill because they are widespread along the Atlantic coast from Maine to Florida, and along the Gulf of Mexico. According to Whitehead (2013), monitoring of genomic endpoints (e.g., transcriptomics, proteomics) that inform mechanism, and integrative measurements of physiological status, in relevant sentinel species, could enable characterization of cause-and-effect in significant and beneficial ways. In this study, exposure of embryos of M. beryllina to WAF and dispersants and CEWAF induced apical and genomic responses. 4.1. Constituents of oil and dispersants Crude oil is a complex mixture of paraffinic, naphthenic, aliphatic, alicyclic and aromatic hydrocarbons ranging in carbon

number from C1 to >C60; however, its precise composition differs depending on the source (Castellano et al., 2011; OSPAR, 2004; Zhang et al., 2010). The potential transport of oil components in the water column during spill episodes presents an environmental risk factor because crude oils contain a large number of PAHs some of which are known to be human carcinogens such as benzo(a) anthracene, chrysene, benzo(b) fluoranthene, benzo(a) pyrene and benzo(ghi) perylene) (Nikolaou et al., 2009). They have also been said to affect marine larvae, for example, Couillard et al. (2005) reported that EROD activity was significantly increased in mummichog (Fundulus heteroclitus), while a compromise of the immune system of filter feeders, like copepods and oysters was observed by several authors (Auffret et al., 2004; Baumard et al., 1999; Wootton et al., 2003). Some studies have also demonstrated that PAH exposure in polluted environments produces reproductive and developmental impairment in fish (Brette et al., 2014; Carls et al., 1999; Incardona et al., 2005; Monteiro et al., 2000). In the present study, the PAH concentrations ranged from 1 to 24 ppb with the higher values observed in CEWAFs (Table 1).

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

Naphthalenes, phenanthrenes, dibenzothiophenes and biphenyls were generally higher in WAF and CEWAFs, while chrysene and cis–trans decalin were found in levels greater than 1 ppb in CEWAFs alone. Chrysene, phenanthrene and naphthalene belong to the group of homocyclic PAHs reported as being carcinogenic (Incardona et al., 2004; van Lipzig et al., 2005) and are also on the original USEPA list of 32 PAH compounds designated as priority pollutants (USEPA, 2000; Yan et al., 2009). In addition to their carcinogenic properties, some homocyclic PAHs and their metabolites, as well as some heterocyclic PAHs (like dibenzothiophenes and cis-decalin) have been found to exhibit endocrine disrupting effects, such as having estrogenic, anti-androgenic and anti-estrogenic activity (Kizu et al., 2003; Santodonato, 1997). According to the Water Framework Directive 2000/60/EC, PAHs are regarded as priority substances because of their environmental behavior, lipophilicity, and toxicity and are characterized as highly persistent in the environment. Similar to the increase in the PAH levels in the presence of dispersants observed in this study, Zuijdgeest and Huettel (2012) also reported that the difference between PAH release from oil and PAH release from oil in the presence of Corexit was statistically significant. However, in our study all PAHs identified in CEWAF were originally present in oil. After the Deepwater Horizon Accident in June 2010, Allan et al. (2012) reported 170 ± 14 ng/L PAHs mean maximum concentration in water samples collected from four sites: Grand Isle, Louisiana, Gulfport, Mississippi, Gulf Shores, Alabama, and Gulf Breeze, Florida in the Gulf of Mexico, which was significantly higher than the mean baseline concentration (3.8 ± 0.64 ng/L) observed before the accident. Our data showing marked increases in the PAHs in CEWAFs compared to WAF can therefore be used to support the hypothesis that the toxicity attributed to CEWAFs is two to three times greater than that of WAF because of the higher levels of the PAHs in the CEWAFs. Agreement with these ratios was also evident for BTEX and low molecular weight PAH concentrations (Table 2). Several researchers (Adams et al., 2014; Couillard et al., 2005; Ramachandran et al., 2004; Yamada et al., 2003) concluded that this could be because dispersants reduce oil droplet size, thereby increasing the mobility and bioavailability of hydrophobic compounds into the water phase. Increased access to PAHs has significant implications for indwelling flora and fauna as evidenced in the present study by the more deleterious effects observed in dispersants and CEWAF exposed embryos compared to control and WAF exposed embryos. 4.2. Heart rate and tissue differentiation In the current study (Fig. 1), heartbeat was significantly lowest in Corexit 9500 (67.5 beats/min) and 9527 (67.1 beats/min) exposed embryos compared with controls (82.7 beats/min). The heart of embryonic fish has been reported to be a sensitive target organ for the toxic effects of the PAHs in crude oil (Incardona et al., 2011). We observed alterations in the heart beat and severe abnormalities in Menidia embryos exposed to the two Corexits alone and with Corexit and oil mixtures. Similar changes in heart beat and heart malformations such as cardiac dysfunction, edema, spinal curvature, and reduction in the size of the jaw and other craniofacial structures were detected by Incardona et al. (2004) in zebrafish embryos using seven non-alkylated PAHs from petrogenic sources, including five two- to four-ring compounds, suggesting that it is the presence of these PAHs in our CEWAFs that are the culprits. The dibenzothiophenes and phenanthrenes found in CEWAFs at levels higher than 1 ppb in the present study were part of those reported to directly disrupt cardiac function in fish, thus affecting the associated processes of circulation and heart chamber formation (Incardona et al., 2004, 2005). Incardona

67

et al. (2014) also confirmed that the high-energy water accommodated fractions (HEWAF) of Mississippi Canyon 252 oil (MC252) and Alaska North Slope crude oil (ANSCO) caused similar cardiotoxicity in zebrafish embryos. But, in our experiment the dispersants alone also had these effects, possibly by disrupting cell membranes or because of other components in their mixtures. In the current study, abnormal embryos were found in all conditions except in the control exposures. There was a trend towards significance in exposures with dispersants and CEWAF (oil/9500) (p < 0.017). Additionally, a large number of embryos exposed to oil and dispersants were in a state of arrested differentiation (p = 0.021) (Fig. 3). Early life stages of fish and other vertebrates are particularly vulnerable to contaminants because disruption of development affects the fish throughout their lives. Studies have shown that transient and sublethal effects of PAHs on the embryonic heartbeat can result in permanent secondary changes in heart shape and cardiac output and these are detrimental to fish (Hicken et al., 2011; Incardona et al., 2004). In particular, Incardona et al. (2004) showed that PAH exposure resulted in a slower swimming speed and therefore the inability of exposed fish to elude predators. A study by Fucik et al. (1995) considered the effects of generic western and central Gulf of Mexico oil and Corexit dispersant on inland silversides (M. beryllina), Atlantic menhaden (Brevoortia tyrannus), spot (Leiostomus xanthurus), and red drum (Sciaenops ocellatus) and concluded that for the four species and irrespective of the oil source, CEWAF exposures significantly decreased hatching success, as well as early survival (less-than-96 h) of all species except red drum. In contrast, Hemmer et al. (2011) rated Corexit 9500 as slightly toxic to mysids and moderately toxic to inland silversides when they were exposed to CEWAF at a ratio of 1:10 (dispersant: oil) at definitive exposure concentrations ranging from 10 to 100 ␮l/L for mysids and 32–320 ␮l/L for Menidia. 4.3. Morphological abnormalities Abnormalities from the present study included head malformation, abdominal/yolk sac and pericardial edema, and various curvatures of the body axis (Figs. 4 and 5). The evidence of increased levels of PAHs in WAF and CEWAFs observed in the current study lends credence to Carls et al. (1999) who reported that high levels of tricyclic PAHs and their alkylated homologs increased the frequency of malformations. Similar to the deleterious effect on heartbeat, significantly more embryos deteriorated and failed to hatch or had abnormalities in dispersants and CEWAF exposures in the present study. CEWAFs have been reported to induce abnormal tissue differentiation and developmental abnormalities (Fucik et al., 1995). Field and laboratory studies in other species have similarly demonstrated oil-induced embryo-larval toxicity syndrome (Carls et al., 1999; Couillard 2002; Incardona et al., 2004, 2011). 4.4. Gene expression 4.4.1. Expression of genes associated with sexual differentiation and sex determination The differential expression of transcripts for genes associated with sex and sexual differentiation in WAF, dispersants and CEWAF exposed Menidia embryos are presented in Fig. 6. The significant down-regulation of StAR reported in the present study is suggestive of potential consequences for steroidogenesis in teleosts exposed to oil and dispersants. StAR is a protein that controls the rate-limiting step for the initiation of steroidogenesis as it shuttles cholesterol into mitochondria for transformation into sex steroids (Kallen et al., 1998; Chen et al., 2014). Modulation of this process will eventually affect testicular, ovarian and adrenocortical functions (Vang et al., 2007). Arukwe (2008) in his review of the implication of endocrine disrupting chemicals on StAR concluded that “the transcriptional

68

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

modulation of genes associated with steroidogenesis and early sex differentiation would cause alterations of reproduction, sexual differentiation, growth, development, and metabolism because these genes are very important for the development of healthy embryos,” which is also valid for the results presented from our study. Transcription of the testis-specific gene dmrt-1 was similarly significantly down regulated in WAF, Corexit 9500 and CEWAF exposed embryos; while a statistically non-significant down regulation was observed in the Corexit 9527 exposed cohort. The expression of dmrt-1 genes has been reported to be essential for postnatal testicular differentiation in almost all vertebrates (Yamaguchi et al., 2006). The primary function of dmrt-1 genes in the gonad is to promote male-specific and repress female-specific differentiation. Despite the disparities in sexual differentiation in fish, Herpin and Schartl (2011) observed that the expression of dmrt-1 genes correlates with testis development in all teleosts. The amh gene was differentially up regulated by WAF, dispersants and CEWAF exposures compared to control but not at a statistically significant level. In higher vertebrates, amh is responsible for inducing the regression of the mullerian ducts, primordial structures that otherwise would develop into the internal reproductive organs of females (reviewed in Josso et al., 2005). Although fish lack mullerian ducts, other amh functions may be important for gonad development (von Hofsten and Olsson, 2005). cDNAs with homologies to mammalian, bird and reptile amh have been cloned in a variety of fish species, including Atlantic salmon (GenBank accession number AY722411), Japanese flounder (Yoshinaga et al., 2004), zebrafish (Rodríguez-Marí et al., 2005), and Japanese medaka (GenBank accession number AY899282), and all of them have been named after their mammalian orthologue. In all these studies, amh was first expressed at low levels in the undifferentiated gonads of both sexes and then at higher levels in the male gonads compared with those of females during the process of gonadal sex differentiation. A study by Fernandino et al. (2008) in Pejerrey, Odontesthes bonariensis, a teleost fish presenting a strong temperature-dependent sex determination revealed that amh displayed higher expression at masculinizing than at feminizing temperatures. The overexpression of amh has been shown to negatively modulate the differentiation and function of Leydig cells in transgenic mice (Racine et al., 1998) by down regulating several enzymes involved in the steroidogenic pathway with moderate overexpression producing externally masculinized mice with significantly decreased serum testosterone in adulthood (Lyet et al., 1995). It has also been implicated in the inhibition of ovarian cell growth in vitro (Ha et al., 2000). Chg-L, a gene that is exquisitely sensitive to estrogens, trended towards down-regulation in all exposed groups (P = 0.69, 0.39, 0.09, 0.08 and 0.05 for WAF, Corexit 9500, Corexit 9527 and the CEWAFsoil/9500, oil/9527, respectively), suggesting that 17 ␤-estradiol levels were low in the embryos. Some homocyclic PAHs and their metabolites, as well as some heterocyclic PAHs like dibenzothiophenes, cis-decalin reported in this study, have been shown to exhibit effects on hormone-linked processes (Santodonato, 1997). However, the exhibition of estrogen or anti-estrogen activities will be greatly influenced by complex environmental and physiological mechanisms. Exposure to WAF up-regulated (P = 0.01) the expression of the transcript for cyp19b, while Corexit 9527 down-regulated its expression (P = 0.004). No significant differences were observed in Corexit 9500 (P = 0.65) and CEWAFs (P = 0.12 and 0.07 for oil/9500 and oil/9527, respectively) exposed embryos compared to controls. Aromatase, the cyp19 gene product converts testosterone into 17␤estradiol, and its inhibition has been reported to result in female to male sex reversal (von Hofsten and Olsson, 2005). Moreover, cyp19b has been shown to have estrogen response elements in its promoter in several teleost species (Callard et al., 2001; Chang et al., 2005; Le

Page et al., 2008). Furthermore, the up-regulation of cyp19b expression has been claimed to be a reliable biomarker for the presence of environmental estrogens (Wang et al., 2011). Taken together these results suggest that the exposures may have had a direct negative effect on Star, which would result in reducing the amount of steroid hormones (both estrogens and androgens) in the embryos, resulting in the down-regulation of genes such as drmt1 and chg-L. Other genes that we expected to be similarly affected, cyb19b and vtg, were not and maybe this was because these genes are not expressed much in this life stage. While there are many reports of cyp19b sensitivity to estrogens in early life stage fish (Callard et al., 2001 and Cheshenko et al., 2008), in our hands the Menidia cyp19b appears not to be sensitive to estrogens in early life stage fish (unpublished observations). 4.4.2. Expression of genes responsible for stress response and growth regulation Significant up-regulations in cyp1a were observed in fish exposed to WAF (P = 0.002) and CEWAFs (P = 0.03 and 0.004 for oil/9500 and oil/9527, respectively), compared with controls (Fig. 6). Induction of cytochrome P4501A (cyp1a) is a powerful biomarker of exposure to PAHs and polyhalogenated aromatic hydrocarbons (PHAHs, i.e., PCBs and dioxins) and has been used for bio-monitoring assessments in a number of aquatic organisms. Increased expression of cyp1a is indicative of the metabolism of planar hydrocarbons; fish collected after oil spills have been reported to show increases in the transcripts of cyp1a that are correlated with exposure to PAHs in the oil (Lee and Anderson, 2005). In this study, WAF and both CEWAFs potentiated the up regulation of cyp1a relative to 9500 and 9527 alone suggesting that PAH exposure is coming from the oil and is also indicative of increased bioavailability of the PAHs as reported in other studies (Couillard et al., 2005; Ramachandran et al., 2004; Yamada et al., 2003). Cyp1a has been reported to play a major role in the biotransformation of a variety of endogenous substances such as lipids, steroids, and vitamins (Schlenk et al., 2008) and also environmental toxicants, in particular halogenated aromatic hydrocarbons, polycyclic aromatic hydrocarbons, and polychlorinated biphenyls (Stegeman and Hahn, 1994). Garcia et al. (2012) showed that cytochrome P450 genes were up-regulated in Gulf killifish collected at oil-affected sites in Louisiana, and Dubansky et al. (2013) reported that Gulf killifish embryos exposed for 21 days to oiled sediments collected more than a year after a spill were characterized by up-regulation of cyp1a, as well as reduced hatch rates, smaller size at hatch, reduced heart rates, and poor vigor. Hsp90 transcript levels were down regulated in embryos exposed to WAF, dispersants and CEWAF, but the down-regulation was significant for Corexit 9527 (P = 0.01) and CEWAFs (P = 0.02 and 0.01 for oil/9500 and oil/9527, respectively). The gene, hsp90, is not a developmental gene, but encodes a heat shock protein which helps organisms cope with stress, thereby protecting against cellular damage. This is because heat shock proteins bind to other proteins and act as a molecular chaperone, helping them stay in the right shape in the face of stressful conditions or it binds to various proteins that are inherently unstable (Csermely et al., 1998; Iwama et al., 1999). Hsp90 is also a chaperone for nuclear steroid receptors such as ERs and ARs and is required for their stabilization and activity (Osborne et al., 2007). Differential expression of genes coding for detoxifying and other cytoprotective proteins, such as Hsp90 are therefore considered as good environmental stress biomarkers (Yamashita et al., 2010). Additionally, heat shock proteins have been reported to be up-regulated in cells that are exposed to a wide variety of stressors, particularly those that denature proteins (Freeman et al., 1999; Welch, 1993), demonstrating a correlation between increased levels of heat shock proteins and exposure to stressors.

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

The differential expressions of ghr observed in this study can also be associated with stress response. Mean transcriptional levels of ghr tended towards down-regulation, but were only significant in Corexit 9527 (p = 0.04) and CEWAF- exposed embryos (p = 0.02 and 0.016 for oil/9500 and oil/9527, respectively). Growth is a multifactorial characteristic resulting from complex genetic and molecular interactions; the ghr has been reported to mediate somatic growth (Filby et al., 2007). The ghr gene encodes a member of the type I cytokine receptor family, which is a transmembrane receptor for growth hormone. Binding of growth hormone to the receptor leads to receptor dimerization and the activation of an intra- and intercellular signal transduction pathway leading to growth (Brown et al., 2005; Gonzalez et al., 2007; Waters et al., 2006). Although oxygen levels were optimal during our study, low oxygen stress has been shown to similarly down-regulate the levels of ghr in Atlantic salmon (Olsvik et al., 2013; Quinn et al., 2011b). Numerous studies have also profiled global gene expression changes in fishes exposed to other stressors, including elevated temperature (Liu et al., 2013; Quinn et al., 2011a). We can therefore conclude that exposures to WAF, dispersants and CEWAFs served as stressors eliciting the down regulation of ghr observed in this study.

5. Conclusions In summary, our study showed developmental toxicity in Menidia embryos exposed to WAF, dispersants and CEWAFs, indicated by increased skeletal malformation of fry, reduced hatching success, lowered heart rate and heart malformation, and impaired survivorship of fish fry. Yamashita, (2003) noted that stress-induced apoptosis is thought to contribute to abnormal development during embryogenesis. Our data also supports previous assertions that early life stages of fish and other vertebrates are particularly vulnerable to oil pollution. The synergistic toxic effects of oil and dispersant might be due to the dispersants’ ability to solubilize oil into the water column, thereby making it more bioavailable. The evidence of Corexit 9500 and 9527’s toxicity to M. beryllina embryos necessitates considering the types of organisms present at the site of a spill and the developmental stages of those organisms before remedial action can be taken. Genes associated with steroidogenesis and sexual differentiation were differentially expressed in WAF, dispersants and CEWAFs exposed embryos. Some reviews suggest that PAHs can have a significant anti-estrogenic effects by binding to the aryl-hydrocarbon receptor (Navas and Segner, 1998; Palumbo et al., 2009), or a mild estrogenic effect by binding to the estrogen receptor (Nicolas, 1999), although with affinities that are several hundred times lower than that of estradiol (Nimrod and Benson, 1996). The up-regulation of transcription of cyp1a, a powerful biomarker of exposure to PAH and transcriptional target of the pollutant-activated aryl hydrocarbon receptor (AHR) signaling pathway was also observed in this study for WAF and CEWAFs. In embryos, inappropriate AHR activation has been reported to cause developmental abnormalities (Antkiewicz et al., 2006; Clark et al., 2010), as observed in the present study. The down-regulation of Hsp90 observed in this study implies that the ability of the embryos to cope with stress was compromised and as such ties in with the developmental abnormalities including heart rate, growth and differential transcription of genes responsible for sex differentiation and growth. Similarly, the trend towards down regulation of ghr induced by Corexit 9527 and CEWAFs is in agreement with the apical endpoint of abnormal development observed in the hatchlings in this study. This is also in consonance with the results of a study on the effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in developing red seabream (Pagrus major) embryos where Yamauchi et al. (2006) concluded that when cardiac function is impacted resulting

69

in abnormal heartbeat and circulation failure; this subsequently results in body growth retardation via insufficient nutrients. Both the apical and molecular endpoints established in the present study could therefore be used as early indicators of long-term effects of Corexit 9500 and 9527 usage in oil spill management on M. beryllina. Acknowledgements This study was supported by a faculty seed grant from the University of Florida College of Veterinary Medicine. Additional support was provided by a fellowship from the Schlumberger Foundation (OA). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox.2015.09. 012. References Adams, J., Sweezey, M., Hodson, P.V., 2014. Oil and oil dispersant do not cause synergistic toxicity to fish embryos. Environ. Toxicol. Chem. 33 (1), 107–114. Allan, S.E., Smith, B.W., Anderson, K.A., 2012. Impact of the Deepwater Horizon oil spill on bioavailable polycyclic aromatic hydrocarbons in Gulf of Mexico coastal waters. Environ. Sci. Technol. 46 (4), 2033–2039. Almeda, R., Wambaugh, Z., Wang, Z., Hyatt, C., Liu, Z., Buskey, E.J., 2013. Interactions between zooplankton and crude oil: toxic effects and bioaccumulation of polycyclic aromatic hydrocarbons. PLoS One 8 (6), e67212. Ankley, G.T., Miller, D.H., Jensen, K.M., Villeneuve, D.L., Martinovic, D., 2008. Relationship of plasma sex steroid concentrations in female fathead minnows to reproductive success and population status. Aquat. Toxicol. 88, 69–74. Antkiewicz, D.S., Peterson, R.E., Heideman, W., 2006. Blocking expression of AHR2 and ARNT1 in zebrafish larvae protects against cardiac toxicity of 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin. Toxicol. Sci. 94 (1), 175–182. Arukwe, A., 2008. Steroidogenic acute regulatory (StAR) protein and cholesterol side-chain cleavage (P450scc)-regulated steroidogenesis as an organ-specific molecular and cellular target for endocrine disrupting chemicals in fish. Cell Biol. Toxicol. 24 (6), 527–540. Auffret, M., Duchemein, M., Rousseau, S., Boutet, I., Tanguy, A., Moraga, D., Marhic, A., 2004. Monitoring of immunotoxic responses in oysters reared in areas contaminated by the Erika oil spill. Aquat. Living Resour. 17, 297–303. Baumard, P., Budzinski, H., Garrigues, P., Narbonne, J.F., Burgeot, T., Michel, X., Bellocq, J., 1999. Polycyclic aromatic hydrocarbon (PAH) burden of mussels (Mytilus sp.) in different marine environments in relation with sediment PAH contamination, and bioavailability. Mar. Environ. Res. 47 (5), 415–439. Berninger, J.P., Williams, E.S., Brooks, B.W., 2011. An initial probabilistic hazard assessment of oil dispersants approved by the United States National Contingency Plan. Environ. Toxicol. Chem. 30 (7), 1704–1708. Boström, C.E., Gerde, P., Hanberg, A., Jernström, B., Johansson, C., Kyrklund, T., Rannug, A., Törnqvist, M., Victorin, K., Westerholm, R., 2002. Cancer risk assessment, indicators, and guidelines for polycyclic aromatic hydrocarbons in the ambient air. Environ. Health Perspect. 110 (Suppl. (3)), 451–489. Brander, S.M., Cole, B.J., Cherr, G.N., 2012. An approach to detecting estrogenic endocrine disruption via choriogenin expression in an estuarine model fish species. Ecotoxicology 21 (4), 1272–1280. Brander, S.M., Connon, R.E., He, G., Hobbs, J.A., Smalling, K.L., Teh, S.J., White, J.W., Werner, I., Denison, M.S., Cherr, G.N., 2013. From ‘omics to otoliths: responses of an estuarine fish to endocrine disrupting compounds across biological scales. PLoS One 8, e74251, http://dx.doi.org/10.1371/journal.pone.0074251. Brette, F., Machado, B., Cros, C., Incardona, J.P., Scholz, N.L., Block, B.A., 2014. Crude oil impairs cardiac excitation–contraction coupling in fish. Science 343 (6172), 772–776. Brewton, R.A., Fulford, R., Griffitt, R.J., 2013. Gene expression and growth as indicators of effects of the BP Deepwater Horizon oil spill on spotted seatrout (Cynoscion nebulosus). J. Toxicol. Environ. Health A 76 (21), 1198–1209. Brown, R.J., Adams, J.J., Pelekanos, R.A., Wan, Y., McKinstry, W.J., Palethorpe, K., Waters, M.J., 2005. Model for growth hormone receptor activation based on subunit rotation within a receptor dimer. Nat. Struct. Mol. Biol. 12 (9), 814–821. Callard, G.V., Tchoudakova, A.V., Kishida, M., Wood, E., 2001. Differential tissue distribution, developmental programming, estrogen regulation and promoter characteristics of cyp19 genes in teleost fish. J. Steroid Biochem. Mol. Biol. 79 (1), 305–314. Camilli, R., Reddy, C.M., Yoerger, D.R., Van Mooy, B.A., Jakuba, M.V., Kinsey, J.C.M., cIntyre, C.P., Sylva, S.P., Maloney, J.V., 2010. Tracking hydrocarbon plume transport and biodegradation at Deepwater Horizon. Science 330 (6001), 201–204.

70

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71

Carls, M.G., Rice, S.D., Hose, J.E., 1999. Sensitivity of fish embryos to weathered crude oil: Part I. Low-level exposure during incubation causes malformations, genetic damage, and mortality in larval pacific herring (Clupea pallasi). Environ. Toxicol. Chem. 18 (3), 481–493. Castellano, O., Gimon, R., Soscun, H., 2011. Theoretical study of the –␲ and ␲–␲ interactions in heteroaromatic monocyclic molecular complexes of benzene, pyridine, and thiophene dimers: implications on the resin-asphaltene stability in crude oil. Energy Fuels 25 (6), 2526–2541. Chakraborty, R., Borglin, S.E., Dubinsky, E.A., Andersen, G.L., Hazen, T.C., 2012. Microbial response to the MC-252 oil and Corexit 9500 in the Gulf of Mexico. Front. Microbiol. 3 (357), 1–6. Chang, X., Kobayashi, T., Senthilkumaran, B., Kobayashi-Kajura, H., Sudhakumari, C.C., Nagahama, Y., 2005. Two types of aromatase with different encoding genes, tissue distribution and developmental expression in Nile tilapia (Oreochromis niloticus). Gen. Comp. Endocrinol. 141 (2), 101–115. Chen, C., Kuo, J., Wong, A., Micevych, P., 2014. Estradiol modulates translocator protein (TSPO) and steroid acute regulatory protein (StAR) via protein kinase A (PKA) signaling in hypothalamic astrocytes. Endocrinology 155 (8), 2976–2985. Cheshenko, K., Pakdel, F., Segner, H., Kah, O., Eggen, R.I., 2008. Interference of endocrine disrupting chemicals with aromatase CYP19 expression or activity, and consequences for reproduction of teleost fish. Gen. Comp. Endocrinol. 155 (1), 31–62. Clark, J.R., Patrick, J.M., Middaugh, D.P., Moore, J.C., 1985. Relative sensitivity of six estuarine fishes to carbophenothion, chlorpyrifos, and fenvalerate. Ecotoxicol. Environ. Saf. 10 (3), 382–390. Clark, B.W., Matson, C.W., Jung, D., Di Giulio, R.T., 2010. AHR2 mediates cardiac teratogenesis of polycyclic aromatic hydrocarbons and PCB-126 in Atlantic killifish (Fundulus heteroclitus). Aquat. Toxicol. 99 (2), 232–240. Couillard, C.M., 2002. A microscale test to measure petroleum oil toxicity to mummichog embryos. Environ. Toxicol. 17, 195–202. Couillard, C.M., Lee, K., Légaré, B., King, T.L., 2005. Effect of dispersant on the composition of the water-accommodated fraction of crude oil and its toxicity to larval marine fish. Environ. Toxicol. Chem. 24 (6), 1496–1504. Csermely, P., Schnaider, T., So, C., Prohászka, Z., Nardai, G., 1998. The 90-kDa molecular chaperone family: structure, function, and clinical applications. A comprehensive review. Pharmacol. Ther. 79 (2), 129–168. Du, W., Wan, Y., Zhong, N., Fei, J., Zhang, Z., Chen, L., Hao, J., 2011. Status quo of soil petroleum contamination and evolution of bioremediation. Pet. Sci. 8 (4), 502–514. Dubansky, B., Whitehead, A., Miller, J.T., Rice, C.D., Galvez, F., 2013. Multitissue molecular, genomic, and developmental effects of the Deepwater Horizon oil spill on resident Gulf killifish (Fundulus grandis). Environ. Sci. Technol. 47 (10), 5074–5082. Fernandino, J.I., Hattori, R.S., Kimura, H., Strüssmann, C.A., Somoza, G.M., 2008. Expression profile and estrogenic regulation of anti-Müllerian hormone during gonadal development in pejerrey, Odontesthes bonariensis, a teleost fish with strong temperature-dependent sex determination. Dev. Dyn. 237 (11), 3192–3199. Filby, A.L., Thorpe, K.L., Maack, G., Tyler, C.R., 2007. Gene expression profiles revealing the mechanisms of anti-androgen and estrogen-induced feminization in fish. Aquat. Toxicol. 81 (2), 219–231. Fingas, M., Banta, J., 2008. A Review of Literature Related to Oil Spill Dispersants (1997-2008) for Prince William Sound Regional Citizens’ Advisory Council (PWSRCAC). PWSRCAC, Anchorage, Alaska, pp. 168 (contract number-955.08.03). Freeman, M.L., Borrelli, M.J., Meredith, M.J., Lepock, J.R., 1999. On the path to the heat shock response: destabilization and formation of partially folded protein intermediates, a consequence of protein thiol modification. Free Radical Biol. Med. 26 (5), 737–745. Fucik, K.W., Carr, K.A., Balcom, B.J., 1995. Toxicity of oil and dispersed oil to the eggs and larvae of seven marine fish and invertebrates from the Gulf of Mexico. In: Lane, P. (Ed.), Chemicals in Oil Spill Response Philadelphia, vol. 1252. ASTM Special Technical Publication, Philadelphia, PA, pp. 135–171. Garcia, T.I., Shen, Y., Crawford, D., Oleksiak, M.F., Whitehead, A., Walter, R.B., 2012. RNA-Seq reveals complex genetic response to Deepwater Horizon oil release in Fundulus grandis. BMC Genomics 13 (1), 474. George-Ares, A., Clark, J.R., 2000. Aquatic toxicity of two Corexit® dispersants. Chemosphere 40 (8), 897–906. Gonzalez, L., Curto, L.M., Miquet, J.G., Bartke, A., Turyn, D., Sotelo, A.I., 2007. Differential regulation of membrane-associated growth hormone binding protein (MA-GHBP) and growth hormone receptor (GHR) expression by growth hormone (GH) in mouse liver. Growth Horm. IGF Res. 17 (2), 104–112. Ha, T.U., Segev, D.L., Barbie, D., Masiakos, P.T., Tran, T.T., Dombkowski, D., Glander, M., Clarke, T.R., Lorenzo, H.K., Donahoe, P.K., Maheswaran, S., 2000. Müllerian inhibiting substance inhibits ovarian cell growth through an Rb-independent mechanism. J. Biol. Chem. 275 (47), 37101–37109. Heintz, R.A., Short, J.W., Rice, S.D., 1999. Sensitivity of fish embryos to weathered crude oil: Part II. Increased mortality of pink salmon (Oncorhynchus gorbuscha) embryos incubating downstream from weathered Exxon Valdez crude oil. Environ. Toxicol. Chem. 18 (3), 494–503. Hemmer, M.J., Barron, M.G., Greene, R.M., 2011. Comparative toxicity of eight oil dispersants, Louisiana sweet crude oil (LSC), and chemically dispersed LSC to two aquatic test species. Environ. Toxicol. Chem. 30 (10), 2244–2252. Herpin, A., Schartl, M., 2011. Dmrt-11 genes at the crossroads: a widespread and central class of sexual development factors in fish. FEBS J. 278 (7), 1010–1019.

Hicken, C.E., Linbo, T.L., Baldwin, D.H., Willis, M.L., Myers, M.S., Holland, L., Larsen, M., Stekoll, M.S., Rice, S.D., Collier, T.K., Scholz, N.L., Incardona, J.P., 2011. Sublethal exposure to crude oil during embryonic development alters cardiac morphology and reduces aerobic capacity in adult fish. Proc. Natl. Acad. Sci. U. S. A. 108 (17), 7086–7090. Hong, S., Khim, J.S., Ryu, J., Park, J., Song, S.J., Kwon, B.O., Choi, K., Ji, K., Seo, J., Lee, S., Park, J., Lee, W., Choi, Y., Lee, K.T., Kim, C., Shim, W.J., Naile, J.E., Giesy, J.P., 2012. Two years after the Hebei Spirit oil spill: residual crude-derived hydrocarbons and potential AhR-mediated activities in coastal sediments. Environ. Sci. Technol. 46 (3), 1406–1414. Hose, J.E., McGurk, M.D., Marty, G.D., Hinton, D.E., Brown, E.D., Baker, T.T., 1996. Sublethal effects of the (Exxon Valdez) oil spill on herring embryos and larvae: morphological, cytogenetic, and histopathological assessments, 1989–1991. Can. J. Fish. Aquat. Sci. 53 (10), 2355–2365. Incardona, J.P., Collier, T.K., Scholz, N.L., 2004. Defects in cardiac function precede morphological abnormalities in fish embryos exposed to polycyclic aromatic hydrocarbons. Toxicol. Appl. Pharmacol. 196 (2), 191–205. Incardona, J.P., Collier, T.K., Scholz, N.L., 2011. Oil spills and fish health: exposing the heart of the matter. J. Expo. Sci. Environ. Epidemiol. 21 (1), 3–4. Incardona, J.P., Carls, M.G., Teraoka, H., Sloan, C.A., Collier, T.K., Scholz, N.L., 2005. Aryl hydrocarbon receptor-independent toxicity of weathered crude oil during fish development. Environ. Health Perspect., 1755–1762. Incardona, J.P., Gardner, L.D., Linbo, T.L., Brown, T.L., Esbaugh, A.J., Mager, E.M., Stieglitz, J.D., French, B.L., Labenia, J.S., Laetz, C.A., Tagal, M., Sloan, C.A., Elizur, A., Benetti, D.D., Grosell, M., Block, B.A., Scholz, N.L., 2014. Deepwater Horizon crude oil impacts the developing hearts of large predatory pelagic fish. Proc. Natl. Acad. Sci. U. S. A. 111 (15), E1510–E1518. Iwama, G.K., Vijayan, M.M., Forsyth, R.B., Ackerman, P.A., 1999. Heat shock proteins and physiological stress in fish. Am. Zool. 39 (6), 901–909. Jayasinghe, S., Kroll, K., Adeyemo, O., Lavelle, C., Denslow, N., Mehinto, A., Bay, S., Maruya, K., 2014. Linkage of In Vitro Assay Results With In Vivo End Points, Final Report—Phase 1. San Francisco Estuary Institute Richmond, CA, contribution # 734. Josso, N., Belville, C., di Clemente, N., Picard, J.Y., 2005. AMH and AMH receptor defects in persistent Müllerian duct syndrome. Hum. Reprod. Update 11 (4), 351–356. Kallen, C.B., Billheimer, J.T., Summers, S.A., Stayrook, S.E., Lewis, M., Strauss, J.F., 1998. Steroidogenic acute regulatory protein (StAR) is a sterol transfer protein. J. Biol. Chem. 273 (41), 26285–26288. Khan, R.A., Payne, J.F., 2005. Influence of a crude oil dispersant, Corexit 9527, and dispersed oil on capelin (Mallotus villosus), Atlantic cod (Gadus morhua), longhorn sculpin (Myoxocephalus octodecemspinosus), and cunner (Tautogolabrus adspersus). Bull. Environ. Contam. Toxicol. 75 (1), 50–56. Kizu, R., Okamura, K., Toriba, A., Kakishima, H., Mizokami, A., Burnstein, K.L., Hayakawa, K., 2003. A role of aryl hydrocarbon receptor in the antiandrogenic effects of polycyclic aromatic hydrocarbons in LNCaP human prostate carcinoma cells. Arch. Toxicol. 77 (6), 335–343. Kuhl, A.J., Nyman, J.A., Kaller, M.D., Green, C.C., 2013. Dispersant and salinity effects on weathering and acute toxicity of South Louisiana crude oil. Environ. Toxicol. Chem. 32 (11), 2611–2620. Le Page, Y., Menuet, A., Kah, O., Pakdel, F., 2008. Characterization of a cis-acting element involved in cell-specific expression of the zebrafish brain aromatase gene. Mol. Reprod. Dev. 75 (10), 1549–1557. Lee, R.F., Anderson, J.W., 2005. Significance of cytochrome P450 system responses and levels of bile fluorescent aromatic compounds in marine wildlife following oil spills. Mar. Pollut. Bull. 50 (7), 705–723. Lichtenthaler, R.G., Daling, P.S., 1985. Aerial application of dispersants—comparison of slick behavior of chemically treated versus non-treated slicks. International Oil Spill Conference, vol. 1985. American Petroleum Institute, pp. 471–478. Liu, S.K., Wang, X.L., Sun, F.Y., Zhang, J.R., Feng, J.B., Liu, H., Rajendran, K.V., Sun, L.Y., Zhang, Y., Jiang, Y.L., Peatman, E., Kaltenboeck, L., Kucuktas, H., Liu, Z.J., 2013. RNA-Seq reveals expression signatures of genes involved in oxygen transport, protein synthesis, folding, and degradation in response to heat stress in catfish. Physiol. Genomics 45 (12), 462–476. Livak, K.J., Schmittgen, T.D., 2001. Analysis of relative gene expression data using real-time quantitative PCR and the 2−CT method. Methods 25 (4), 402–408. Lyet, L., Louis, F., Forest, M.G., Josso, N., Behringer, R.R., Vigier, B., 1995. Ontogeny of reproductive abnormalities induced by deregulation of anti-müllerian hormone expression in transgenic mice. Biol. Reprod. 52 (2), 444–454. Major, D., Zhang, Q., Wang, G., Wang, H., 2012. Oil-dispersant mixtures: understanding chemical composition and its relation to human toxicity. Toxicol. Environ. Chem. 94 (9), 1832–1845. Mason, O.U., Hazen, T.C., Borglin, S., Chain, P.S., Dubinsky, E.A., Fortney, J.L., Han, J., Holman, H.Y., Hultman, J., Lamendella, R., Mackelprang, R., Malfatti, S., Tom, L.M., Tringe, S.G., Woyke, T., Zhou, J., Rubin, E.M., Jansson, J.K., 2012. Metagenome, Metatranscriptome and single-cell sequencing reveal microbial response to Deepwater Horizon oil spill. ISME J. 6 (9), 1715–1727. Middaugh, D.P., Hemmer, M.J., 1992. Reproductive ecology of the inland silverside, Menidia beryllina, (Pisces: Atherinidae) from Blackwater Bay, Florida. Copeia, 53–61. Monteiro, P.R.R., Reis-Henriques, M.A., Coimbra, J., 2000. Polycyclic aromatic hydrocarbons inhibit in vitro ovarian steroidogenesis in the flounder (Platichthys flesus L.). Aquat. Toxicol. 48 (4), 549–559. Navas, J.M., Segner, H., 1998. Antiestrogenic activity of anthropogenic and natural chemicals. Environ. Sci. Pollut. Res. 5 (2), 75–82.

O.K. Adeyemo et al. / Aquatic Toxicology 168 (2015) 60–71 Nicolas, J.M., 1999. Vitellogenesis in fish and the effects of polycyclic aromatic hydrocarbon contaminants. Aquat. Toxicol. 45 (2), 77–90. Nikolaou, A., Kostopoulou, M., Lofrano, G., Meric, S., 2009. Determination of PAHs in marine sediments: analytical methods and environmental concerns. Global nest. Int. J. 11 (4), 391–405. Nimrod, A.C., Benson, W.H., 1996. Environmental estrogenic effects of alkylphenol ethoxylates. Crit. Rev. Toxicol. 26 (3), 335–364. Olsvik, P.A., Vikeså, V., Lie, K.K., Hevrøy, E.M., 2013. Transcriptional responses to temperature and low oxygen stress in Atlantic salmon studied with next-generation sequencing technology. BMC Genomics 14 (1), 817. Osborne, N., Sherry, J., Rendell, J.L., Currie, S., 2007. The role of hsp90 in 17 alpha-ethynylestradiol-induced endocrine disruption in rainbow trout hepatocytes. Ecotoxicol. Environ. Saf. 68, 13–19. OSPAR, 2004. Guidelines for Monitoring the Environmental Impact of Offshore Oil and Gas Activities. Oslo and Paris Commissions, London, Ref. 2004-11E. Palumbo, A.J., Denison, M.S., Doroshov, S.I., Tjeerdema, R.S., 2009. Reduction of vitellogenin synthesis by an aryl hydrocarbon receptor agonist in the white sturgeon (Acipenser transmontamus). Environ. Toxicol. Chem. 28, 1749–1755. Pilcher, W., Miles, S., Tang, S., Mayer, G., Whitehead, A., 2014. Genomic and genotoxic responses to controlled weathered-oil exposures confirm and extend field studies on impacts of the Deepwater Horizon oil spill on native killifish. PLoS One 9 (9), e106351, http://dx.doi.org/10.1371/journal.pone. 0106351. Place, B., Anderson, B., Mekebri, A., Furlong, E.T., Gray, J.L., Tjeerdema, R., Field, J., 2010. A role for analytical chemistry in advancing our understanding of the occurrence, fate, and effects of Corexit oil dispersants. Environ. Sci. Technol. 44 (16), 6016–6018. Quinn, N.L., McGowan, C.R., Cooper, G.A., Koop, B.F., Davidson, W.S., 2011a. Identification of genes associated with heat tolerance in Arctic charr exposed to acute thermal stress. Physiol. Genomics 43 (11), 685–696. Quinn, N.L., McGowan, C.R., Cooper, G.A., Koop, B.F., Davidson, W.S., 2011b. Ribosomal genes and heat shock proteins as putative markers for chronic, sublethal heat stress in Arctic charr: applications for aquaculture and wild fish. Physiol. Genomics 43 (18), 1056–1064. Racine, C., Rey, R., Forest, M.G., Louis, F., Ferré, A., Huhtaniemi, I., Josso, N., di Clemente, N., 1998. Receptors for anti-Müllerian hormone on Leydig cells are responsible for its effects on steroidogenesis and cell differentiation. Proc. Natl. Acad. Sci. U. S. A. 95 (2), 594–599. Ramachandran, S.D., Hodson, P.V., Khan, C.W., Lee, K., 2004. Oil dispersant increases PAH uptake by fish exposed to crude oil. Ecotoxicol. Environ. Saf. 59 (3), 300–308. Redmond, M.C., Valentine, D.L., 2012. Natural gas and temperature structured a microbial community response to the Deepwater Horizon oil spill. Proc. Natl. Acad. Sci. U. S. A. 109 (50), 20292–20297. Rico-Martínez, R., Snell, T.W., Shearer, T.L., 2013. Synergistic toxicity of Macondo crude oil and dispersant Corexit 9500A® to the Brachionus plicatilis species complex (Rotifera). Environ. Pollut. 173, 5–10. ˜ Rodríguez-Marí, A., Yan, Y.L., BreMiller, R.A., Wilson, C., Canestro, C., Postlethwait, J.H., 2005. Characterization and expression pattern of zebrafish Anti-Müllerian hormone (Amh) relative to sox9a, sox9b, and cyp19a1a, during gonad development. Gene Expr. Patterns 5 (5), 655–667. Santodonato, J., 1997. Review of the estrogenic and antiestrogenic activity of polycyclic aromatic hydrocarbons: relationship to carcinogenicity. Chemosphere 34 (4), 835–848. Scholz, S., Mayer, I., 2008. Molecular biomarkers of endocrine disruption in small model fish. Mol. Cell. Endocrinol. 293, 57–70. Segner, H., 2009. Zebrafish (Danio rerio) as a model organism for investigating endocrine disruption. Comp. Biochem. Physiol. C 149, 187–195. Schlenk, D., Celander, M., Gallagher, E.P., George, S., James, M., Kullman, S.W., Van der Hurk, P., Willett, K., 2008. Biotransformation in fishes. In: Di Giulio, R.T., Hinton, D.E. (Eds.), The Toxicology of Fishes. CRC Press Taylor & Francis Group, pp. 153–234. Singer, M.M., George, S., Jacobson, S., Lee, I., Weetman, L.L., Tjeerdma, R.S., Sowby, M.L., 1996. Comparison of acute aquatic effects of the oil dispersant Corexit 95000 with those of other Corexit series dispersants. Ecotoxicol. Environ. Saf. 35, 183–189. Stegeman, J.J., Hahn, M.E., 1994. Biochemistry and molecular biology of monooxygenases: current perspectives on forms, functions, and regulation of cytochrome P450 in aquatic species. In: Aquatic Toxicology: Molecular, Biochemical & Cellular Perspectives. CRC Press, Boca Raton, FL, USA, pp. 87–206. Stegeman, J.J., Smolowitz, R.M., Hahn, M.E., 1991. Immunohistochemical localization of environmentally induced cytochrome P450IA1 in multiple organs of the marine teleost Stenotomus chrysops (Scup). Toxicol. Appl. Pharmacol. 110 (3), 486–504. USEPA, 1992. Method 3580, Solvent dilution of a non-aqueous waste sample prior to cleanup and/or analysis, Revision 1, Washington, D.C. http://www.epa.gov/ solidwaste/hazard/testmethods/sw846/pdfs/3580a.pdf.

71

USEPA, 1996. Method 5030, Purge-and-Trap for aqueous samples, pp. 5030-1-21, Washington, D.C. http://www.epa.gov/osw/hazard/testmethods/sw846/pdfs/ 5030b.pdf. USEPA, 2000. Toxic Release Inventory Public Data Release. Office of Environmental Information, United States Environmental Protection Agency, Washington, D.C http://www.epa.gov/triinter/tridata/index.htm. USEPA, 2002. Methods for Measuring the Acute Toxicity of Effluent and Receiving Waters to Freshwater and Marine Organisms, 5th edition. U.S. Environmental Protection Agency, Washington, D.C, EPA-821-R-02-012. USEPA, 2007. Method 8270D (SW-846): Semivolatile organic compounds by gas chromatography/mass spectrometry (GC/MS), Revision 4. Washington, D.C. http://www.epa.gov/osw/hazard/testmethods/sw846/pdfs/8270d.pdf. Vang, S.H., Kortner, T.M., Arukwe, A., 2007. Steroidogenic acute regulatory (StAR) protein and cholesterol side-chain cleavage (p450 scc) as molecular and cellular targets for 17␣-ethynylestradiol in salmon previtellogenic oocytes. Chem. Res. Toxicol. 20 (12), 1811–1819. van Lipzig, M.M., Vermeulen, N.P., Gusinu, R., Legler, J., Frank, H., Seidel, A., Meerman, J.H., 2005. Formation of estrogenic metabolites of benzo [a] pyrene and chrysene by cytochrome P450 activity and their combined and supra-maximal estrogenic activity. Environ. Toxicol. Pharmacol. 19 (1), 41–55. von Hofsten, J., Olsson, P., 2005. Zebrafish sex determination and differentiation: involvement of FTZ-F1 genes. Reprod. Biol. Endocrinol. 3, 63–74, http://dx.doi. org/10.1186/1477-7827-3-63. Wang, J., Shi, X., Du, Y., Zhou, B., 2011. Effects of xenoestrogens on the expression of vitellogenin (vtg) and cytochrome P450 aromatase (cyp19a and b) genes in zebrafish (Danio rerio) larvae. J. Environ. Sci. Health 46 (9), 960–967. Waters, M.J., Hoang, H.N., Fairlie, D.P., Pelekanos, R.A., Brown, R.J., 2006. New insights into growth hormone action. J. Mol. Endocrinol. 36 (1), 1–7. Welch, W.J., 1993. How cells respond to stress. Sci. Am. 268 (5), 56–64. White, H.K., Hsing, P.Y., Cho, W., Shank, T.M., Cordes, E.E., Quattrini, A.M., Nelson, R.K., Camilli, R., Demopoulos, A.W.J., German, C.R., Brooks, J.M., Roberts, H.H., Shedd, W., Reddy, C.M., Fisher, C.R., 2012. Impact of the Deepwater Horizon oil spill on a deep-water coral community in the Gulf of Mexico. Proc. Natl. Acad. Sci. U. S. A. 109 (50), 20303–20308. Whitehead, A., 2013. Interactions between oil-spill pollutants and natural stressors can compound ecotoxicological effects. Integr. Comp. Biol. 53 (4), 635–647. Whitehead, A., Dubansky, B., Bodinier, C., Garcia, T.I., Miles, S., Pilley, C., Raghunathan, V., Roach, J.L., Walker, N., Walter, R.B., Rice, C.D., Galvez, F., 2012. Genomic and physiological footprint of the Deepwater Horizon oil spill on resident marsh fishes. Proc. Natl. Acad. Sci. U. S. A. 109 (50), 20298–20302. Wilson, V.S., LeBlanc, G.A., 2000. The contribution of hepatic inactivation of testosterone to the lowering of serum testosterone levels by ketoconazole. Toxicol. Sci. 54 (1), 128–137. Wootton, E., Dyrynda, E., Pipe, R., Raccliffe, N., 2003. Comparisons of PAH-induced immuno-modulation in three bivalve molluscs. Aquat. Toxicol. 65, 13–25. Yamada, M., Takada, H., Toyoda, K., Yoshida, A., Shibata, A., Nomura, H., Wada, M., Nishimura, M., Okamoto, K., Ohwada, K., 2003. Study on the fate of petroleum-derived polycyclic aromatic hydrocarbons (PAHs) and the effect of chemical dispersant using an enclosed ecosystem, mesocosm. Mar. Pollut. Bull. 47 (1), 105–113. Yamaguchi, A., Lee, K.H., Fujimoto, H., Kadomura, K., Yasumoto, S., Matsuyama, M., 2006. Expression of the DMRT-1 gene and its roles in early gonadal development of the Japanese pufferfish Takifugu rubripes. Comp. Biochem. Physiol. D Genomics Proteomics 1 (1), 59–68. Yamashita, M., 2003. Apoptosis in zebrafish development. Comp. Biochem. Physiol. B Biochem. Mol. Biol. 136 (4), 731–742. Yamashita, M., Yabu, T., Ojima, N., 2010. Stress protein HSP70 in fish. Aqua-BioSci. Monogr. 3 (4), 111–141. Yamauchi, M., Kim, E.Y., Iwata, H., Shima, Y., Tanabe, S., 2006. Toxic effects of 2, 3, 7, 8-tetrachlorodibenzo-p-dioxin (TCDD) in developing red seabream (Pagrus major) embryo: an association of morphological deformities with AHR1, AHR2 and CYP1A expressions. Aquat. Toxicol. 80 (2), 166–179. Yan, J., Wang, L., Fu, P.P., Yu, H., 2009. Photomutagenicity of 16 polycyclic aromatic hydrocarbons from the US EPA priority pollutant list. Mutat. Res. 557 (1), 99–108. Yoshinaga, N., Shiraishi, E., Yamamoto, T., Iguchi, T., Abe, S.I., Kitano, T., 2004. Sexually dimorphic expression of a teleost homologue of Müllerian inhibiting substance during gonadal sex differentiation in Japanese flounder, Paralichthys olivaceus. Biochem. Biophys. Res. Commun. 322 (2), 508–513. Zhang, Z., Gai, L., Hou, Z., Yang, C., Ma, C., Wang, Z., Sun, B., He, X., Tang, H., Xu, P., 2010. Characterization and biotechnological potential of petroleum-degrading bacteria isolated from oil-contaminated soils. Bioresour. Technol. 101 (21), 8452–8456. Zuijdgeest, A., Huettel, M., 2012. Dispersants as used in response to the MC252-spill lead to higher mobility of polycyclic aromatic hydrocarbons in oil-contaminated Gulf of Mexico sand. PLoS One, e50549, http://dx.doi.org/10. 1371/journal.pone.0050549.