Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?

General and Comparative Endocrinology xxx (2015) xxx–xxx

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Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance? Adam R. Schwindt Colorado State University, Cooperative Fish and Wildlife Research Unit, Department of Fish, Wildlife, and Conservation Biology, 201 Wagar Hall, Campus Delivery 1484, Fort Collins, CO 80523-1484, United States

a r t i c l e

i n f o

Article history: Available online xxxx Keywords: Epigenetics Vertebrate Invertebrate Freshwater Marine Pollution

a b s t r a c t The effects of endocrine disrupting compounds (EDCs) on aquatic wildlife are increasingly being recognized for their complexity. Investigators have detected alterations at multiple levels of biological organization in offspring exposed to EDCs through the blood or germ line of the parents, suggesting that generational consequences of EDCs are evident. Exposure to EDCs through the parents is concerning because if the resulting phenotype of the offspring is heritable and affects fitness, then evolutionary consequences may be evident. This review summarizes the evidence for transgenerational effects of EDCs in aquatic wildlife and illustrates cases where alterations appear to be transmitted maternally, paternally, or parentally. The literature indicates that EDC exposure to the parents induces developmental, physiological, endocrinological, and behavioral changes as well as increased mortality of offspring raised in clean environments. What is lacking, however, is a clear demonstration of heritable transgenerational effects in aquatic wildlife. Therefore, it is not known if the parental effects are the result of developmental or phenotypic plasticity or if the altered phenotypes are durably passed to subsequent generations. Epigenetic changes to gene regulation are discussed as a possible mechanism responsible for EDC induced parental effects. Additional research is needed to evaluate if heritable effects of EDCs are evident in aquatic wildlife, as has been demonstrated for terrestrial mammals. Ó 2015 Elsevier Inc. All rights reserved.

1. Introduction The role of endocrine disrupting compounds (EDCs) in altering the physiological performance of aquatic wildlife is a topic that continues to garner attention. In particular, the study of EDCs has focused on chemical-induced alterations to the hypothalamuspituitary-thyroid (Carr and Patiño, 2011) and -gonadal axes (Tyler et al., 1998). As defined by the Endocrine Society, an EDC is ‘‘a compound, either natural or synthetic, which, through environmental or inappropriate developmental exposures, alters the hormonal and homeostatic systems that enable the organism to communicate with and respond to its environment’’ (DiamantiKandarakis et al., 2009). Evidence exists that EDCs induce life-stage dependent physiological responses depending on the developmental state of the organism (Guillette et al., 1995). For example, EDC exposure early in life can induce durable changes that become evident only as an adult (Milston et al., 2003; Schwindt et al., 2014). Of additional concern is the possibility that EDCs induce changes in the parents that can be passed to the offspring and affect their physiological performance (Skinner et al., 2011). This is despite E-mail address: [email protected]

the offspring never having experienced the chemical directly in the environment. Changes observed in subsequent generations following exposure to the parents suggest that those altered phenotypes may be heritable (Skinner et al., 2011). If the changes affect fitness, as has been demonstrated in mammals (Anway et al., 2005), transgenerational inheritance of effects resulting from EDC exposures could influence evolutionary processes (Bossdorf et al., 2008). Despite the extensive use of aquatic organisms in ecotoxicological research, transgenerational effects of EDCs in aquatic wildlife are poorly understood. It has been known for some time that larval stages are more sensitive to EDC exposure than adults (e.g. McKim, 1977), and that female oviparous animals transfer EDCs to their eggs (e.g. Brooks et al., 1997; Miller, 1993). Transgenerational inheritance of EDC effects has been convincingly demonstrated in mammals (Anway et al., 2005; Bruner-Tran and Osteen, 2011; Chamorro-García et al., 2013; Crews et al., 2007; Manikkam et al., 2013, 2012; Wolstenholme et al., 2012). Therefore, the possibility exists that transgenerational consequences of parental EDC exposures exist for aquatic organisms as well. The purpose of this review is to synthesize the current findings and suggest critical research needs related to potential transgenerational effects of EDCs in aquatic organisms.

http://dx.doi.org/10.1016/j.ygcen.2015.01.020 0016-6480/Ó 2015 Elsevier Inc. All rights reserved.

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx

Transgenerational effects are alterations in offspring endocrinological, physiological, or developmental performance due to a change in the environment experienced by the parents, or the F0 generation in the case of asexual or parthenogenic organisms. The transgenerational effects may persist over multiple generations even in an unaltered environment (Skinner et al., 2011). This suggests that the effects are passed through the germ line, are thus inherited by the offspring (Skinner et al., 2011), and may have evolutionary consequences (Bossdorf et al., 2008; Jablonka and Raz, 2009). Transgenerational effects can result from a variety of alterations in the parental environment. For example, the nutritional state of zebrafish (Danio rerio) parents (Schwerte et al., 2005), environmental temperature in the sheepshead minnow (Cyprinodon variegatus) (Salinas and Munch, 2012), and parental disease state of humans (Gluckman et al., 2007) can all induce transgenerational effects. In the field of ecotoxicology, the primary variable of interest with regard to transgenerational effects is exposure to EDCs present in the environment. Transgenerational effects in aquatic ecotoxicology are poorly studied but could play an important role in how aquatic organisms respond to contaminants over multiple generations. The review begins by addressing the parental, maternal, and paternal effects of EDCs in aquatic vertebrates and invertebrates and is organized by taxa where published literature exists. In all examples discussed herein, the gametes of the parental generation (F0) just prior to fertilization and subsequent effects on all developmental stages are defined as effects on the F1 generation. Likewise, the F1 gametes pre- and post-fertilization and all subsequent effects on the offspring of the F1 generation are defined as effects on the F2 generation. I discuss epigenetics as a mechanism behind transgenerational inheritance of EDC effects and the possibility of such occurrences in aquatic taxa. I conclude the review with a discussion of potentially promising and efficient experimental designs and suggest animal models for detecting transgenerational effects. 2. Parental effects Parental effects are those which derive from exposure to both parents in the laboratory or parents collected from the same polluted environment. This does not necessarily indicate that both parents contribute equally to the generational consequences. Rather, one cannot distinguish the difference between the maternal or paternal contributions because both sperm and egg were exposed. In some cases the parents are exposed early in life and allowed to recover for an extended period of time prior to reproduction (e.g. Holdway et al., 2008; Schwindt et al., 2014). In other cases reproductively mature adults are exposed and then transferred to clean water just prior to reproduction (e.g. Foran et al., 2002; Matta et al., 2001). The biological responses observed from parental exposures are diverse and include effects on survival, behavior, growth, reproduction, hormone production, enzymes, and gene expression. 2.1. Fish 17a-ethinylestradiol (EE2), a common constituent in human contraceptive pills, is a frequently used estrogen for studying parental effects in vertebrates and invertebrates (Nash et al., 2004; Schwindt et al., 2014; Segner et al., 2003). To evaluate the parental effects of EE2 on reproductive parameters in the saltwater sheepshead minnow, Zillioux et al. (2001) exposed mixed sex populations to waterborne EE2 (1.7, 18.1, 117, 723 ng/L) from the subadult stage to sexual maturity (Table 1). Mature F0 were removed from EE2 and allowed to reproduce for 14 days in clean water.

Hatch success in the F1 was reduced at 18.1 and 117 ng/L compared to water and vehicle controls (triethylene glycol) (Zillioux et al., 2001). Generational effects of EE2 on the zebrafish have also been identified. Compared to water controls, reduced fertilization success of the F1 generation parentally exposed to waterborne EE2 (3, 4.5 ng/L) prior to sexual maturation and then raised in clean water has been observed (Nash et al., 2004; Segner et al., 2003). Likewise, Hill and Janz (2003) observed reduced egg viability and hatchability compared to vehicle (acetone) controls in F1 zebrafish whose parents were exposed to waterborne EE2 (10 ng/L) from 2 to 60 days post hatch and then reared to adulthood in clean water (Table 1). In the fathead minnow (Pimephales promelas), reduced F2 embryonic survival and reduced numbers of F2 larvae compared to water controls were observed following EE2 exposure (3.2 ng/L) to the F1 parents early in life (Schwindt et al., 2014) (Table 1). Notably, the effects described above (Hill and Janz, 2003; Nash et al., 2004; Schwindt et al., 2014; Segner et al., 2003) occurred at environmentally relevant EE2 concentrations (Kostich et al., 2013). Effects on endpoints directly relevant to the endocrine system are also evident following parental EE2 exposure. Adult medaka (Oryzias latipes) were exposed to waterborne EE2 (0.2, 5, 500, 2000 ng/L) for 2 weeks and allowed spawn (Foran et al., 2002). Resulting F1 offspring were raised to adulthood (3 months) in clean water and assessed for hepatic estrogen receptor (ER) and vitellogenin (VTG) content (Table 1). VTG is an estrogen responsive protein and used as biomarker of estrogen exposures (Schwindt et al., 2007). Compared to vehicle (ethanol) controls, F1 males parentally exposed to 2000 ng/L showed increased hepatic ER, and hepatic VTG increased in F1 females whose parents were exposed to 0.2 ng/L (Foran et al., 2002). Altogether, results from the above studies suggest that EE2 induces a suite of parental effects in fish ranging from decreased survival to improper activation of estrogen signaling pathways. The detergent additive nonylphenol (NP) binds weakly to ERs (Blair et al., 2000) and induces physiological changes in male fish such as induction of VTG (Pedersen et al., 1999). Parental effects of NP include reduced F1 hatch success (Hill and Janz, 2003; Holdway et al., 2008) but directly relevant endocrine effects can last for years after the original exposure. Schwaiger et al. (2002) exposed F0 rainbow trout (Oncorhynchus mykiss) to NP (1.2, 10.4 lg/L) for 4 months and then induced spawning. In the F1 offspring, there was no effect on male VTG or sex ratios, but female VTG was elevated (10.4 lg/L) compared to water controls (Table 1). Interestingly, NP disrupted steroid hormone biosynthesis because 17b-estradiol (E2) was elevated in F1 males and testosterone (T) was 13 higher in treated females (10.4 lg/L) compared to water controls. The effects on the sex steroids were evident even though F1 fish were reared for three years in clean water (Schwaiger et al., 2002). Compared to the parental effects of EDCs on reproductive endpoints, parental effects on growth and development have received little attention. However, some evidence exists for parental effects of brominated flame retardents on growth and development. Polybrominated diphenyl ethers (PBDE) are currently used as flame retardants, are similar in structure to polychlorinated biphenyls (PCB), and are ubiquitous in the environment (Hooper and McDonald, 2000). Physiological effects following controlled exposures point toward disruption of the thyroid axis (Carr and Patiño, 2011). To assess the parental effects of PBDEs, Yu et al. (2011) exposed F0 zebrafish to a waterborne PBDE mixture (DE71) (1, 3, 10 lg/L) from the embryo stage to sexual maturation (5 months). PBDE concentrations in the eggs transferred maternally ranged from 1689 to 13,701 (ng/g) (Yu et al., 2011) (Table 1). Decreased F1 hatch success and growth were observed following parental exposures to PBDEs (3 and 10 lg/L) compared to vehicle (dimethyl sulfoxide (DMSO)) controls (Yu et al., 2011).

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx Table 1 Parentala Effects of endocrine disrupting compounds on aquatic wildlife. Taxa

Fish Salvelinus namaycush

Chemical (action)

Concentration (route of exposure)

Developmental class(es) exposed

Exposure duration

Comparison group (vehicle)

Endpoints

References

b

Hatchery: 0.1– 0.21 lg/g Lake: 0.4– 4.3 lg/g (multiple) TBT or PCBc1 lg/g or TBT + PCB – 1 lg/g each (dietary) 0.4, 2, 4, 15 lg/ g (dietary) 0, 10, 100 mg/ kg (multiple, injection)

Adult

Lifetime exposure in L. Michigan; eggs collected and raised in hatchery (clean water) for 7 months

Hatchery raised fish

c

Binder and Lech (1984)

Adult

Daily feeding for 3 weeks

Vehicle control (acetone)

f

Decr fertilization (TBT + PCB), hatch (TBT or PCB), and swim-up success (all groups) (F1)

Nirmala et al. (1999)

Adult

Twice daily feeding for 6 weeks Lifetime exposure (mix of g OC) in Newark Bay, NJ, 6 months depuration, then single injection PCB mixture Adult = Lifetime P 1.67 ng/L Fertilization to juvenile P3 ng/L From 2 to 60 days post hatch

Vehicle control (acetone) h Ref site, Flax Pond, NY, Vehicle control (heptane)

Incr weight (F1) Decr larval survival (F2) Decr CYP1A1 (F1)

Matta et al. (2001) Elskus et al. (1999)

Not specified

Decr fertilization success (F1)

Vehicle control (acetone) Water control (ethanol)

Decr egg viability, hatch success, and swim-up success (F1) Decr 11-kTj (F0) Decr fertilization success and male infertility (F1) Decr numbers of offspring (F1) and survival (F2)

Segner et al. (2003) Hill and Janz (2003) Nash et al. (2004)

Water control Vehicle control (triethylene glycol) Vehicle control (ethanol)

PCBs unspecified congeners (multiple)

Oryzias latipes

PCBs and eTBT unspecified congeners

Fundulus heteroclitus Fundulus heteroclitus

PCBs (Arochlor 1268) PCBs (Arochlors 1248 and 1262)

Danio rerio

i

Danio rerio

EE2

Danio rerio

EE2

Pimephales promelas

EE2

0, 3.2, 5.3 ng/L (water)

Adult Juvenile

Adult = lifetime Juvenile 0–75 days post fertilization Adult = lifetime or juvenile 0 6 104 days

Cyprinodon variegatus

EE2

Adult

Sub-adult to adult 59 days

Oryzias latipes

EE2

0, 1.73, 18.1, 117, 328, 723 ng/L (water) 0, 0.2, 5, 500, 2000 ng/L (water)

Adult

2 weeks prior to spawning

0, 50, 100% effluent (water)

Adult

24–36 h post fertilization to adult

1500 lg/L (water)

Adult

0, 1, 5, 10, 22, 33, 50 lg/L (water) 0, 1.16, 10.4 lg/L (water) 0, 10, 30, 100 lg/L (water) 0, 50, 100, 500, 1000, 2250, 5000 lg/L (water) 0, 0.05, 0.1 mg/ kg (dietary – contaminated fish) 0.2, 0.5, 1, 11 lg/g (dietary) 0, 1, 3, 10 lg/L (water)

Fish Pimephales promelas

Danio rerio

Melanotaenia fluviatilis

EE2 (estrogen)

n WW effluent; 1.7 ng/L E2 equivalent (estrogen, multiple) Bis-phenol A (estrogen)

Oncorhynchus mykiss

Endosulfan (estrogen, antiandrogen) Nonylphenol (estrogen)

Danio rerio

Nonylphenol

Melanotaenia fluviatilis

Nonlyphenol

Micropogonias undulatus

Methyl-Hg (multiple)

Fundulus heteroclitus

Methyl-Hg

Danio rerio

r PDBE mixture (thyroid)

P1.67 ng/L P3 ng/L (water) 0, 1, 10, 100 ng/L (water) 0, 0.5, 4.5 ng/L (water)

Adult

Adult Juvenile Juvenile

Adult Juvenile

Water control (methanol)

Incr dCYP1A1 (F1)

Decr hatch success (F1)

Schwindt et al. (2014) Zillioux et al. (2001)

Decr hatch success Incr male kE2, hepatic lER, and female mVTG (F1)

Foran et al. (2002)

Water control (effluent)

Early onset of reproductive behaviors and secondary sex characters (F1)

Sowers et al. (2009)

Lifetime

Not specified

Decr fertilization success (F1)

Adult

4h

Adult

10 days/month for 4 months

Water control Vehicle control (acetone) Water control (ethanol)

Decr oGSI in males, pHSI in females (F1), and hatch success (F2) Incr E2 in males Incr qT in females (F1)

Juvenile

From 2 to 60 days post hatch

Vehicle control (acetone)

Adult

24 h

Water control Vehicle control (acetone)

Decr egg viability and hatch and swim-up success (F1) Decr hatch success (F2)

Segner et al. (2003) Holdway et al. (2008) Schwaiger et al. (2002) Hill and Janz (2003) Holdway et al. (2008)

Adult

1 month

Uncontaminated fish

Decr swimming speed and response to vibration stimulus (F1)

Alvarez et al. (2006)

Adult

Twice daily feeding for 6 weeks

Vehicle control (acetone)

Female biased sex ratio and Incr weight (F1)

Matta et al. (2001)

Adult Larvae

Adult = lifetime Larvae 0–10 days post hatch

Vehicle control (dimethyl sulfoxide)

Incr sT3 and tT4 Decr hatch success, and growth Altered gene expression (F1)

Yu et al. (2011)

(continued on next page)

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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Table 1 (continued) Taxa

Amphibians Anaxyrus americanus

Anaxyrus americanus

Anaxyrus americanus

Invertebrates Crassotrea gigas Hyalella azteca

Americamysis bahia a b c d e f g h i j k l m n o p q r s t

Chemical (action)

Concentration (route of exposure)

Developmental class(es) exposed

Exposure duration

Comparison group (vehicle)

Endpoints

References

Total Hg

Ref site eggs – 21 ng/g Hg site eggs – 149 ng/g (multiple) 0.1, 2.5, 10.1 lg/g (dietary) Ref site eggs – 21 ng/g Hg site eggs – 149 ng/g (multiple) 0.1, 2.5, 10.1 lg/g (dietary) Ref site eggs – 21 ng/g Hg site eggs – 149 ng/g (multiple) 0.1, 2.5, 10.1 lg/g (dietary)

Adult Larvae

Adult = lifetime Larvae fed daily from days 4 to 28 post-hatch

Ref site Control diet (agar-gelatin)

Decr swimming speed, survival and disrupted metamorphosis (F1) relative to dietary exposure

Bergeron et al. (2011)

Adult

Adult = lifetime

Ref site

Todd et al. (2011)

Larvae

Larvae fed daily from days 4 to 28 post-hatch

Control diet (agar-gelatin)

Decr growth and disrupted metamorphosis Incr spinal malformations (F1) relative to ref site

Adult Larvae

Adult = lifetime Larvae fed daily from days 4 to 28 post-hatch then fed crickets

Ref site Control diet (agar-gelatin)

Decr growth 1 year postmetamorphosis (F1) relative to ref site

Todd et al. (2012)

Nonylphenol (estrogen) EE2 (estrogen)

0, 1, 100 lg/L (seawater) 0.1–10 lg/L (water)

Larvae Not specified

48 h during days 7–8 postfertilization 15 weeks

Seawater control (methanol) Not specified

Decr gamete viability and survival (F1) Disrupted male gonad (F2) histology

Fenoxycarb (juvenile horrnone)

0, 1, 6, 43 lg/L (water)

Adult

Lifetime

Water control (not specified)

Decr numbers of offspring Female biased sex ratio (F2)

Nice et al. (2003) Segner et al. (2003) McKenney (2005)

Total Hg

Total Hg

Males and females were exposed. Polychlorinated biphenyls. Increased. Cytochrome P450 1A1. Tributyltin. Decreased. Organochlorines. Reference. 17a-ethinylestradiol. 11keto-testosterone. 17b-estradiol. Estrogen receptor. Vitellogenin. Wastewater. Gonadosomatic index. Hepatosomatic index. Testosterone. Polybrominated diphenyl ethers. 3,5,30 -triiodothyronine. Thyroxine.

The work by Yu et al. (2011) is one of the few examples that evaluated thyroid hormones and gene expression in the F1 generation following exposure to the parents. Parental exposure to 10 lg/L PBDEs led to increased 3,5,30 -triiodothyronine (T3) and thyroxine (T4) in the F1 larvae compared to vehicle controls. Parental exposure to PBDEs (10 lg/L) increased gene expression involved with development and growth of the thyroid gland (hhex, nkx2.1) and thyroid hormone synthesis (thyroglobulin), but reduced thyroid stimulating hormone subunit b (TSHb) expression (Yu et al., 2011). Two deiodinase genes, Dio1 and Dio2, were upregulated or downregulated, respectively, in larvae from parents exposed to (10 lg/L) and uridine diphosphoglucuronosyl transfer-

ase (UGT1) was also upregulated compared to vehicle controls (Yu et al., 2011). The pattern of altered gene expression alongside increased T3 and T4 levels in F1 larvae exposed through the F0 parents is suggestive of hyperthyroidism (Yu et al., 2011). The work by Yu et al. (2011) is the first to suggest that EDC exposure in fish induces thyroid disruption in the offspring, effects that are coincident with reduced somatic growth. While more research is needed, these results provide the basis for the potential construction of a mechanistic model of parentally induced thyroid disruption in fish. Future research should investigate the durability of these changes in the F1 adults and evaluate if the same alterations are present in the F2 larvae.

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx

Rarely are aquatic organisms exposed to EDCs singularly in their natural environments (Kortenkamp, 2007). Rather, aquatic biota are exposed to complex mixtures of EDCs with multiple modes of action (Kortenkamp, 2007); for example, exposure to treated (Jobling et al., 1998; Vajda et al., 2008) and untreated (Orlando et al., 2004) wastewater (WW). Wild fish populations (Rutilus rutilus, Catastomus commersoni) show disrupted reproductive physiology below WW treatment plants (Jobling et al., 1998; Vajda et al., 2008) and naive fathead minnows exposed to WW show signs of reproductive disruption (Vajda et al., 2011). To evaluate the parental effects of WW, Sowers et al. (2009) exposed fathead minnows from larvae to adulthood in 50% and 100% WW with an estimated estrogenicity of 1.7 ng/L E2. Following a 48 h depuration period in clean water, reproduction ensued and the F1 generation was raised to adulthood in clean water. Compared to controls (no WW) secondary sex characters and onset of reproductive behavior were actually increased in the F1 generation (50%, 100% WW) following parental exposure (Sowers et al., 2009) (Table 1). The authors propose that the observed results were potentially due to the F1 generation being derived from naturally resistant F0 parents (Sowers et al., 2009). Polluted bays and estuaries retain complex contaminant mixtures from multiple environmental insults (Oberdörster and Cheek, 2001). In areas with extensive ship building and maintenance, EDCs like de-fouling agents can be present. Tributyltin (TBT) is a de-fouling agent and is associated with disrupted population dynamics and altered reproductive development in invertebrates (Matthiessen and Gibbs, 1998). To evaluate the generational effects of a binary mixture of EDCs, Nirmala et al. (1999) exposed F0 medaka to dietary TBT (1 lg/g) and PCBs (1 lg/g, unspecified congeners) separately or TBT and PCBs (1 lg/g each of TBT and PCBs) together for 3 weeks (Table 1). Fish were fed TBT and PCB contaminated food (10 mg feed/g) daily for 3 weeks (Nirmala et al., 1999). Fertilized eggs were collected during spawning, allowed to develop in clean water, and assessed for contaminant concentrations. Concentrations of TBT in the eggs ranged from 264 to 279 ng/g transferred from females with body-burdens ranging from 2390 to 2610 ng/g (Nirmala et al., 1999) indicating that the eggs retained approximately 10% of the female body burden. Concentrations of PCBs in the females or the eggs were not determined (Nirmala et al., 1999). Parental exposure to the TBT and PCB mixture led to reduced F0 spawning frequency, fewer eggs, and reduced fertilization success of the F1 compared to either chemical alone. Successfully fertilized eggs from TBT and PCB exposed parents survived as well as control fish fed vehicle (acetone) treated food (Nirmala et al., 1999). F1 offspring from parents exposed to TBT showed reduced survival at 12 days compared to vehicle (acetone) controls. Swim-up success of the F1 larvae was reduced in all groups compared to vehicle controls (Nirmala et al., 1999). Contaminant detoxification pathways can also be altered by parental EDC exposures. Cytochrome P450 1A1 (CYP1A1) is a monoxygenase involved in phase I metabolism of many contaminants (Andersson and Förlin, 1992; Stegeman and Kloepper-Sams, 1987). Elskus et al. (1999) evaluated CYP1A1 in wild caught mummichog (Fundulus heteroclitus) from highly contaminated Newark Bay, New Jersey, USA and an uncontaminated reference site. Following capture, the F0 were transferred to clean seawater for 6 months and then subjected to an intraperitoneal injection of Arochlor 1248 and 1262 (10, 100 mg/kg) (Table 1). After 7 weeks, the F0 fish were artificially spawned and F1 larvae fixed for immunohistochemical analysis of CYP1A1 localization. CYP1A1 in F1 fish from the reference site showed a dose-dependent response to increases in the PCB injections. Conversely, F1 larvae from contaminated parents did not show a CYP1A1 induction (Elskus et al., 1999). The results suggest that unresponsive CYP1A1 in the offspring following parental PCB exposure is maladaptive. That is,

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without functional CYP1A1 mummichog larvae are missing a key component in pollutant detoxification pathways. Adult lake trout (Salvelinus namaycush) exposed for their entire life to contaminated water in Lake Michigan, USA were captured and artificially spawned in the laboratory (Binder and Lech, 1984). Induction of monooxygenase systems were compared in embryos from contaminated Lake Michigan parents and embryos from lake trout reared in unpolluted hatchery water. CYP1A1 was induced to a greater degree in embryos and larvae from polluted parents relative to the hatchery controls (Binder and Lech, 1984) (Table 1). That CYP1A1 was differentially regulated in Binder and Lech (1984) and Elskus et al. (1999) could be due to the fact that CYP1A1 responds to a variety of contaminants (Whyte et al., 2000) and there are likely different contaminant profiles in Newark Bay (Elskus et al., 1999), compared to Lake Michigan (Binder and Lech, 1984). This illustrates the point that depending on the specificity of a biomarker to an EDC, or suite of EDCs, organismal responses can be difficult to compare across studies. Methyl mercury (MeHg) is a persistent organic pollutant (Jones and de Voogt, 1999) that has endocrine effects despite the absence of a well characterized mode of action (Iavicoli et al., 2009). Sex steroid biosynthesis and spermatogenesis can be disrupted by MeHg exposures (Iavicoli et al., 2009) and the parental effects of MeHg are evident in fish. Wild caught sexually mature male and female Atlantic Croaker (Micropogonias undulatus) were fed fish contaminated with MeHg (0.05, 0.1 mg/kg/day) for 1 month (Alvarez et al., 2006) (Table 1). Females transferred MeHg to eggs resulting in concentrations ranging from 0.3 to 4.6 ng/g. As indicated by regression analysis, increasing MeHg concentrations induced slower swimming speed and increased time to respond to a vibration stimulus in the F1 larvae (Alvarez et al., 2006). These results suggest that parental MeHg exposures predispose offspring to poor survival skills (Alvarez et al., 2006), even though the offspring themselves were not exposed in the environment.

2.2. Amphibians Among aquatic vertebrates, amphibians appear to be particularly sensitive to parentally-derived Hg. Hg has a high affinity for lipids and proteins, and bioaccumulates in organisms following dietary exposure (Morel et al., 1998). Maternal transfer of Hg to the egg likely poses not only a means of excretion from the mother but also a means of exposing the embryo (Bergeron et al., 2010). Thus, assessing the relative contributions of parentally-derived, versus trophically-derived, Hg on amphibian toxicity should provide clues as to the relative importance of the exposure scenarios. Following parental and dietary exposure of wild-caught American toads (Anaxyrus americanus) to total Hg, (Bergeron et al., 2011) demonstrated that parentally-derived Hg had a greater effect on survival and health of the F1 offspring (Table 1). The toads were captured downstream of a Hg (female Hg = 2250 ng/g, eggs = 149 ng/g) contamination source on the South River, Virginia, USA and from a reference site upstream (female Hg = 160 ng/g, eggs = 21 ng/g) (Bergeron et al., 2011). Toads from each site were fed either a control diet (dry feed with background Hg = 0.01 lg/ g) or dry feed spiked with Hg (2.5, 10.1 lg/g of food) suspended in a agar-gelatin mixture. Parentally exposed (but not dietary exposed) tadpoles re-absorbed their tails slower, exhibited decreased swimming speed, had to be prodded to swim, and survived poorly (Bergeron et al., 2011). That is, parental Hg exposure disrupted the timing and success of metamorphosis to a greater degree than dietary Hg. The effects of parentally derived Hg seem more severe because parental exposures may act at more sensitive developmental timepoints (e.g. gametogenesis) than dietary Hg; although this has not been explicitly addressed.

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx

Using amplexing American toads from the same Hg contamination source and reference site as Bergeron et al. (2011), Todd et al. (2011) transported pairs into the lab where they reproduced overnight. Eggs were collected and larvae were fed a control (parental effect), low, and high Hg dose (same feed and Hg concentrations as (Bergeron et al., 2011)) to evaluate growth of the metamorphosed toads (Todd et al., 2011). Body mass was reduced in the F1 generation parentally exposed relative to the F1 generation from parents captured at the reference site, but survival was not affected (Table 1). Incidence of spinal malformations was increased in F1 toads from contaminated parents relative to F1 toads from reference site parents (Todd et al., 2011). To assess the effects on growth and metamorphosis of parental Hg later in life, Todd et al. (2012) followed the toads used in Todd et al. (2011) for one year after metamorphosis. The F1 toads were placed in outdoor mesocosms with refugia and leaf-litter. Activity resumed after winter and crickets were added as supplemental food. Results from their previous work indicated that the body mass of parentally exposed F1 toads during metamorphosis was 5% less than reference site toads (Todd et al., 2011). After one year, body mass of parentally exposed F1 toads was 7% less suggesting that there were no latent effects from the parental exposure because effects on growth did not increase appreciably through time (Todd et al., 2012) (Table 1). However, the lack of recovery in F1 toads one year later demonstrates the lasting effects of parentally derived Hg on growth and development. 2.3. Invertebrates Compared to the vertebrates relatively little is known of the parental effects of EDCs on invertebrates. Environmentally relevant concentrations of NP (1, 100 lg/L) were used in a 48 h exposure of larval Pacific oysters (Crassotrea gigas) (F0) altering sex ratio and inducing intersex individuals 10 months later (Nice et al., 2003) (Table 1). Effects were also evident in the gametes (F1) (1, 100 lg/L) that showed reduced viability compared to seawater controls. The mechanisms controlling the disruption reported by Nice et al. (2003) are unknown. ERs have been identified in the Pacific oyster (Matsumoto et al., 2007) but there is no known ligand (Castro and Santos, 2014). Most research indicates that the actions of EDCs are exerted through steroid hormone receptors; however, this may be a simplistic view of EDC mode of action (Tabb and Blumberg, 2006). EDCs such as NP can affect other signaling pathways independent of ERs. For example, NP has disrupted aromatase activity and increased aryl-hydrocarbon receptor activity in vitro (Bonefeld-Jørgensen et al., 2007). As such, the parental effects of NP may be exerted through the ER or a combination of alternative modes of action. Fenoxycarb is an insecticide that acts as a juvenile hormone agonist used primarily to control coleopteran and lepidopteran pests (Dhadialla et al., 1998). Exposures to non-target organisms, however, can disrupt reproductive output. Using the saltwater opposum shrimp (Americamysis bahia) as a model, McKenney (2005) found reduced numbers of F1 offspring (6 lg/L) and a female biased sex ratio (1, 43 lg/L) in the F2 generation compared to seawater controls following parental fenoxycarb exposure (0, 1, 6, 43 lg/L) (Table 1). The mechanisms of fenoxycarb action in aquatic crustaceans are poorly understood (Dhadialla et al., 1998) but the above results suggest that developmental and reproductive pathways may be affected. 3. Maternal effects The term maternal effects is borrowed somewhat from the literature using internal fertilizing animals as the experimental model.

That is, maternal effects derive from the physiological and behavioral state of the mother during gestation and subsequent care of the offspring (Fox and Mousseau, 1998). The research on maternal effects is broad and certainly not limited to EDC exposures (Fox and Mousseau, 1998). Maternal effects as they are applied to EDC exposed aquatic oviparous and viviparous organisms require that only the female be exposed. Parthenogenic invertebrates have also been used (Tsui and Wang, 2004) and provide an additional avenue for studying maternal effects of EDCs. Significant research has been conducted on maternal effects of EDCs in birds, reptiles, and mammals (Hamlin and Guillette, 2010). However, as illustrated below, the maternal effects of EDCs in aquatic wildlife have only recently been subject to evaluation. 3.1. Fish Comprehensive analysis of the maternal effects of E2 and octylphenol (OP) has been performed on the marine dwelling viviparous eelpout (Zoarces viviparous). Following fertilization with an unexposed male, Rasmussen et al. (2002) subjected pregnant eelpout to waterborne E2 (0.5 lg/L) or OP (25, 100 lg/L) for 17 and 35 days. Embryos were removed and assessed for a variety of endpoints. F1 embryonic mortality at day 35 from mothers exposed to OP (100 lg/L) was significantly above vehicle (isopropanol) controls. Mass of embryos derived from mothers exposed to E2 and all levels of OP was reduced compared to controls (Table 2). Effects on ER pathways were also observed. VTG cDNA was upregulated in F1 embryos compared to vehicle controls and was localized (immunohistochemistry) in hepatocytes following maternal exposure to OP (100 lg/L). Additionally ER mRNA (in situ hybridization) was found in F1 male and female embryonic gonads in the vehicle control, but not in maternally exposed embryos. Gonads were also affected in embryos whose mothers were exposed to OP (100 lg/L) where 32% could not be identified as male or female. These findings demonstrate the estrogenicity of OP and its ability to act at multiple levels of biological organization, given the effects on gene expression, the gonad, and on embryo growth. The maternal effects of PCBs on zebrafish embryos exhibit a range of severity depending on the congener (Westerlund et al., 2000). Adult females were injected (peanut oil) with 10 (1 lmol/ kg) individual PCB congeners (PCBs 60, 104, 112, 126, 143, 173, 184, 190 and the hydroxylated OH-PCB30 and OH-PCB61) or E2 (1, 1000 lmol/kg) 10 days prior spawning with unexposed males (Westerlund et al., 2000). Mortality of F1 embryos increased up to 75%, becoming significant (relative to control and injection control) on day eight (Table 2). In the following decreasing order of severity, PCBs 184 < 190 < 126 < 112 < 173 < 143 < 60 < 104 induced mortality. Mortality associated with maternal exposure to the hydroxylated (estrogenic) PCBs was compared to the E2 groups, with OH-PCB30 inducing mortality (approximately 75%) similar to both E2 doses. ER mRNA expression in unexposed F1 embryos increased during the blastula and gastrula developmental periods but not during cleavage, segmentation, pharyngial development, or at hatch (Westerlund et al., 2000). The authors conclude that maternal PCB and E2 exposures can have lethal effects possibly mediated through disruption of estrogen signaling pathways (Westerlund et al., 2000). In some cases maternal exposure to EDCs alters the sensitivity of the offspring to a direct environmental exposure as an adult. Six month old adult female medaka were exposed to waterborne ortho, para0 -dichlorodiphenyltrichloroethane (o, p0 -DDT) (2.5 lg/L) for 2 weeks and then spawned with unexposed males (Metcalfe et al., 2000). Embryos were raised to adulthood (approximately 7 months) and then exposed to E2 (12 mg/L) for 4 days. F1 adults maternally exposed to o, p0 -DDT showed a greater induction of

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx Table 2 Maternala effects of endocrine disrupting compounds on aquatic wildlife. Chemical (action)

Concentration (route of exposure)

Developmental class(es) exposed

Exposure duration

Comparison group (vehicle)

Endpoints

References

b

PCBs – 400– 1000 ppb PBDEs – 100– 150 ppb OC pesticides – 5–750 ppb (multiple) 0, 1 lmol/kg injection

Adult

Lifetime

Hatchery raised fish

e Decr egg yolk, whole body volume, and brain volume (F1)

Ostrach et al. (2008)

Adult

Injected 10 days prior to spawning

Decr survival

Westerlund et al. (2000)

11 lg/L blood plasma 0.5 lg/L (water) 0, 1, 1000 nmol/kg (injection)

Adult

Pregnant females 17 days or 35 days

Control Injection control (peanut oil) Vehicle control (isopropanol)

Decr growth g Incr hVTG and iER

Rasmussen et al. (2002)

Adult

Injected 10 days prior to spawning

Decr survival ncr ER

Westerlund et al. (2000)

5, 55 lg/L blood plasma 0, 25, 100 lg/L (water) 0, 2.5 lg/L (water)

Adult

Pregnant females 17 days or 35 day

Rasmussen et al. (2002)

Adult

2 weeks prior to spawning

Vehicle control (acetone)

Decr survival and growth Incr VTG and ER Gonad abnormalities Delayed hatch (F1) Incr VTG (F1 adults)

Methyl-Hg (multiple)

0, 28 nM (dietary – green algae)

3 day old daphnids

6 h/day for 5 days

E2 (estrogen)

0, 0.5, 50 mg/ kg (injection)

Adult Embryo

Adult = 1/week for 8 weeks then 3 week recovery before spawning; Embryos exposed to E2, kOP, lDDD, m TBT for 96 h post fertilization

Underdeveloped eggs and Decr numbers of offspring (F1) Decr E2 sensitivity; Incr DDD and TBT sensitivity; Incr nOR expression (F1)

Tsui and Wang (2004)

Strongylocentrotus purpuratus

Strongylocentrotus purpuratus

Octylphenol (estrogen)

0, 0.5, 50 mg/ kg (injection)

Adult Embryo

Adult = 1/week for 8 weeks then 3 week recovery before spawning; embryos exposed to E2, OP, DDD, TBT for 96 h post fertilization

Decr E2sensitivity; Incr OP, DDD, TBT sensitivity; Incr OR expression (F1)

Roepke et al. (2006)

Strongylocentrotus purpuratus

o

o,p,-DDD (multiple)

0, 0.5, 50 mg/ kg (injection)

Adult Embryo

Adult = 1/week for 8 weeks then 8 week recovery before spawning; Embryos exposed to E2, OP, DDD, TBT for 96 h post fertilization

Decr E2, OP, DDD sensitivity; Incr TBT sensitivity; Decr OR expression (F1)

Roepke et al. (2006)

Strongylocentrotus purpuratus

Tributyltin (multiple)

0, 0.5, 50 lg/ kg (injection)

Adult Embryo

Adult = 1/week for 8 weeks then 3 week recovery before spawning; Embryos exposed to E2, OP, DDD, TBT for 96 h post fertilization

Water control (aqueous Na2CO3) Control Injection control (dimethyl sulfoxide) Control Injection control (dimethyl sulfoxide) Control Injection control (dimethyl sulfoxide) Control Injection control (dimethyl sulfoxide)

Incr OP and TBT sensitivity; Incr OR expression (F1)

Roepke et al. (2006)

Taxa

Fish Morone saxatilis

PCBs, PBDEs, dOC pesticides (multiple)

c

Danio rerio

Zoarces viviparus

Danio rerio

E2

Zoarces viviparus

Octylphenol (estrogen)

Oryzias latipes

j

Invertebrates Daphnia magna

a b c d e f g h i j k l m n o

PCBs multiple congeners (multiple) f E2 (estrogen)

o,p0 -DDT (estrogen)

Control Injection control (peanut oil) Vehicle control (isopropanol)

Metcalfe et al. (2000)

Roepke et al. (2006)

Females were exposed or effects were due to the female parent. Polychlorinated biphenyls. Polybrominated diphenyl ethers. Organochlorines. Decreased. 17b-estradiol. Increased. Vitellogenin. Estrogen receptor. ortho, para-Dichlorodiphenyltrichloroethane. Octylphenol. Dichlorodiphenyldichloroethane. Tributyltin. Orphan receptor. orth, para-Dichlorodiphenyldichloroethane.

VTG following E2 exposure compared to vehicle (acetone) controls (Metcalfe et al., 2000) (Table 2). These results suggest that EDC exposed mothers alter the sensitivity of their offspring to environ-

mental exposures later in life. This phenomenon has also been demonstrated in invertebrates by Roepke et al. (2006) as is discussed later.

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx

The San Francisco Bay, California, USA estuary once retained healthy populations of striped bass (Morone saxatilis) and many stressors have contributed to population declines (Ostrach et al., 2008). In addition to water diversions and invasive species, EDCs also appear to play an important role (Ostrach et al., 2008). Adult male and female striped bass were captured from the San Francisco estuary and returned to the hatchery for spawning and assessment of F1 larval fish health and development (Ostrach et al., 2008). Eggs were analyzed for a suite of EDCs finding that PCBs, PBDEs, and many organochlorine pesticides were significantly higher than hatchery control eggs. The control eggs were produced from hatchery broodstock. Relative to the hatchery controls, F1 larvae from estuary collected parents exhibited reduced whole body volume, egg yolk, and brain volume (Ostrach et al., 2008) (Table 2). In Ostrach et al. (2008) river captured females were spawned with river captured (EDC exposed) males but follow-up experiments indicated that the above effects on the F1 striped bass were of maternal origin. That is, river captured females spawned with hatchery males produced abnormal offspring. And, river captured males spawned with hatchery females produced normal offspring (Dr. David Ostrach, personal communication). 3.2. Invertebrates The maternal effects of EDCs on aquatic invertebrates have also received significant attention in the literature. For example, the parthenogenic freshwater crustacean Daphnia magna transfers 11% of dietary MeHg (28 nM in green algae) to the F1 resulting in significantly underdeveloped eggs and reduced numbers of F1 offspring compared to water controls (Tsui and Wang, 2004) (Table 2). The work by Tsui and Wang (2004) is an unique application of an ideal model for studying maternal effects. The parthenogenic D. magna exhibit extremely fast generation times (Mount and Norberg, 1984) making them ideal organisms for the study of maternal effects. Additionally, new gene expression analysis tools are available for D. magna (Poynton et al., 2007) which should facilitate a more mechanistic understanding of maternally induced EDC effects. Recent work provides evidence that orphan receptors (OR) respond to a variety of EDCs, including estrogens, and may help regulate offspring response to maternal EDC exposures. Roepke et al. (2006) injected (eight weekly) female purple sea urchins (Strongylocentrotus purpuratus) with a variety of EDCs, assessed OR expression in the eggs, spawned the females with unexposed males, and evaluated the sensitivity of F1 embryos re-exposed for 96 h post fertilization to each EDC (Table 2). All comparisons were made relative to maternally exposed F1 embryos that were not re-exposed. Two estrogenic (E2 or OP; 0.5, 50 mg/kg) and non-estrogenic (TBT; 0.5, 50 lg/kg, or ortho, para0 -dichlorodiphenyldichloroethane (o, p0 -DDD); 0.5, 50 mg/kg) EDCs were dissolved in DMSO and administered in seawater. In F1 embryos maternally exposed to E2, the incidence of morphological defects following E2 re-exposure decreased (Roepke et al., 2006), suggesting that maternal E2 prepares the embryo for a potential environmental exposure to E2. In contrast, maternal E2 exposure increased morphological defects of F1 embryos subsequently exposed to TBT and o, p0 -DDD (Roepke et al., 2006). These results suggest that maternal E2 exposure can be protective, but can also hamper development of F1 embryos subsequently exposed to nonestrogenic EDCs. Interestingly, F1 embryos exposed to TBT showed increased morphological abnormalities if they had been exposed maternally to at least one concentration of E2, OP, TBT, or o, p0 DDD (Roepke et al., 2006). Morphological abnormalities following embryonic, but not maternal, exposure to any of the EDCs were no different than unexposed embryos (Roepke et al., 2006), demonstrating the significance of maternal effects. The developmental

effects described above may be regulated in part by OR activity. As described by Roepke et al. (2006), sea urchin eggs collected preand post-injection were assayed for OR (SpSHR2) mRNA expression. OR expression in the eggs increased in response to maternal E2 (0.5, 50 mg/kg), OP (50 mg/kg), and TBT (0.5 lg/kg) exposure and decreased in response to o, p0 -DDD (0.5 mg/kg) relative to seawater controls (Roepke et al., 2006). Taken together these results indicate that maternal EDC exposures can alter the sensitivity of the offspring to EDC exposures later in life. The altered sensitivity in the offspring may be due in part to changes in OR signaling during development. 4. Paternal effects Compared to parental and maternal effects, EDC effects derived solely from the male parent have received less attention. Evaluating the contribution of paternal effects to generational consequences of EDC exposures requires exposing males and breeding with unexposed females. Alternatively, paternal effects can be evaluated by comparing the reproductive success of exposed males and females to unexposed males bred with the same exposed females. 4.1. Fish The paternal effects of EE2 were nicely demonstrated by Nash et al. (2004) using male replacement experiments. Male and female zebrafish were exposed to waterborne EE2 (4.5 ng/L) for their entire life and placed in reproductive trials with clean water. Normal reproductive behavior was observed but fertilization success was significantly reduced relative to water controls. (Nash et al., 2004) (Table 3). When the investigators replaced exposed males with control males, reproductive output was comparable to the unexposed breeders. These results suggest that the EE2 exposed males were responsible for the observed reproductive disruption (Nash et al., 2004). The lack of fertilization in exposed males induced by EE2 (Nash et al., 2004) suggests that the testis may be particularly sensitive to steroid estrogens. To elucidate the generational consequences of EE2 on the testis, Schultz et al. (2003) artificially fertilized unexposed rainbow trout eggs with sperm from adult males exposed to waterborne EE2 (15.6, 131 ng/L) for 62 days just prior to fertilization. A significant reduction in F1 embryo survival (50%) was observed (15.6, 131 ng/L) compared to vehicle (methanol) controls, suggesting that EE2 was acting through the sperm (Schultz et al., 2003) (Table 3). It was unclear however what developmental stage of the F0 was most sensitive to the EE2 exposure. To address this question Brown et al. (2007) exposed clonal male rainbow trout to waterborne EE2 (0.8, 8.3, 65 ng/L) during sexual differentiation (21 consecutive days) and during sexual maturation (56 consecutive days). Semen was collected from both groups and used to artificially fertilize eggs pooled from several unexposed females. Compared to vehicle (methanol) controls, F1 embryonic survival at 19 days post fertilization was lower for males exposed during maturation at all concentrations, but not for males exposed during differentiation (Table 3). The male sex steroid 11-ketotestosterone (11-kT) was reduced and luteinizing hormone was increased in the F0 male parents (65 ng/L exposure level) in the group exposed during sexual maturation (Brown et al., 2007) suggesting that EE2 exerted anti-androgenic effects on the male parents. It is possible that EE2 altered feedback mechanisms controlling androgen production given that increased gonadatropins were observed coincident with reduced 11-kT. The work by Nash et al. (2004), Schultz et al. (2003), and Brown et al. (2007) suggested that the effects of EE2 were passed

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx Table 3 Paternala effects of endocrine disrupting compounds on aquatic wildlife. Chemical (action)

Concentration (route of exposure)

Developmental class(es) exposed

Exposure duration

Comparison group (vehicle)

Endpoints

References

b

EE2 (estrogen)

0, 0.5, 4.5 ng/L (water)

Adult

Lifetime

c

Decr fertilization success (F1)

Nash et al. (2004)

Oncorhynchus mykiss

EE2

0, 15.6, 131, 750 ng/ L (water)

Adult

62 days leading up to spawning

d

Oncorhynchus mykiss

EE2

Pre-testicular differentiation for 21 days or mid-spermatogenesis for 56 days

21 days 109 ng/L or 56 days 0.8, 8.3, 65 ng/L

Incr e17,20-DHP Decr f11-kT (F0) Decr survival (F1) Incr gLH, Decr 11kT (F0) Decr survival (F1) (56 day exposure)

Schultz et al. (2003) Brown et al. (2007)

Oncorhynchus mykiss

EE2

109 ng/L blood plasma 0.8, 8.3, 65 ng/L blood plasma (water) 0, 8.9 ng/L (water)

Water control (ethanol) Vehicle control (methanol) Vehicle control (methanol)

Mid-spermatogenesis

56 days

Oncorhynchus mykiss

EE2

0, 0.8, 8.3, 65 ng/L blood plasma (water)

Mid-spermatogenesis

56 days

Decr survival Anueploid sperm and embryos (F1) Survival (F2 not affected)

Brown et al. (2008) Brown et al. (2009)

Taxa

Fish Danio rerio

a b c d e f g

Vehicle control (methanol) Vehicle control (methanol)

Male parents were exposed. 17a-Ethinylestradiol. Decreased. Increased. 17a, 20b-dihydroxyprogesterone. 11keto-testosterone. Luteinizing hormone.

paternally to the offspring. One question that remained was the durability of the EE2 effects. Using the same rainbow trout in Brown et al. (2007) the investigators followed the F1 generation to adulthood and spawned the males with unexposed females (Brown et al., 2009). They tracked embryo survival of the F2 generation finding no effect of EE2 leading to the conclusion that the paternal effects of EE2 are transient, at least as they pertain to embryonic survival (Brown et al., 2009) (Table 3). The same pattern appears evident in zebrafish given that exposure of the F0 generation (4.5 ng/L EE2) had no effect on F1 survival or on F2 survival compared to water controls (Nash et al., 2004). The research discussed above shows that EE2 can reduce survival of the F1 generation following F0 male exposure, but what mechanism regulates the effect? More work by Brown et al. (2008) indicates that EE2 induces chromosomal abnormalities in the sperm and appears to be one plausible mechanism. Again using clonal trout, Brown et al. (2008) exposed males to waterborne EE2 (9 ng/L) for 56 days or to a vehicle control (methanol). Semen was collected and assessed for aneuploidy. The semen was also used to fertilize unexposed eggs to determine if the F1 embryos were also aneuploid. The results indicated that all sperm and most embryos were aneuploid suggesting that chromosomal abnormalities were responsible for poor embryonic survival (Brown et al., 2008) (Table 3). The mechanism of action for the paternal effects was that EE2 caused genetic abnormalities that prevented proper embryonic development and eventual mortality. 5. EDCS, epigenetics, and transgenerational inheritance A current definition of epigenetics issued by Jablonka and Raz (2009) is ‘‘the study of the processes that underlie developmental plasticity and canalization and that bring about persistent developmental effects in both prokaryotes and eukaryotes.’’ Additionally, epigenetic changes occur in the absence of changes to the DNA sequence and derive from alterations in DNA methylation, histone modification, and chromatin remodeling (Bird, 2002). In aquatic wildlife exposed to EDCs only changes to DNA methylation have

been investigated (Head et al., 2012; Vandegehuchte and Janssen, 2011). The mechanisms, proposed and understood, of epigenetics and epigenetic inheritance have been reviewed elsewhere (Head, 2014; Jablonka and Raz, 2009; Vandegehuchte and Janssen, 2011). This discussion will be limited to the published evidence that EDCs can induce changes consistent with the current understanding of epigenetic mechanisms. One method of identifying changes consistent with epigenetics is measuring the DNA methylation state of C-G repeats (CpG islands) in the promoter regions of genes (Bird, 2002). In a simplistic sense, de-methylation coincides with gene activation and, conversely, methylation of CpG islands is associated with deactivation (Bird, 2002). Methylation of DNA is controlled by methyltransferase enzymes (Bird, 2002). In African clawed frogs (Xenopus laevis), changes in DNA methyltransferase expression from stage three oocytes to mature ova suggests that methylation is important for normal gametogenesis (Kimura et al., 1999). The importance of DNA methyltransferase in gonad development is also evident in fish. Mhanni and McGowan (2002) identified changes in the DNA methyltransferase gene in the zebrafish embryo and the mature ovary. During embryogenesis DNA methlytransferase is visible in the one to two cell stage, is highest at the blastula stage (3–3.5 h post hatch), and declines to an intermediate level between the two-celled embryo and the blastula stage. Methyltransferase in the gonad was highly variable in developing follicles and became undetectable in the mature oocyte (Mhanni and McGowan, 2002). The results indicate that properly coordinated changes in methylation, and thus epigenetic processes, are important for correctly timed embryogenesis and gametogenesis in aquatic organisms. In fish there is evidence that EDCs induce changes consistent with epigenetic mechanisms. For example, zebrafish exposed to high concentrations of EE2 (100 ng/L) show significant de-methylation of the 50 region of the VTG1 gene compared to vehicle (DMSO) controls (Strömqvist et al., 2010). Given that many EDCs disrupt gametogenesis (Kime and Nash, 1999), perhaps through altering methylation state (Strömqvist et al., 2010), it is possible that

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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A.R. Schwindt / General and Comparative Endocrinology xxx (2015) xxx–xxx

epigenetic mechanisms could be at least partly responsible for adverse effects on the gonad. To address this question explicitly, Contractor et al. (2004) exposed adult medaka to high EE2 concentrations (500 ng/L) for 14 days. The DNA methylation state of the estrogen receptor a (ERa) and aromatase (CYP19A1) genes in the brain, gonad, and liver were assessed. Changes in methylation state of ERa were found in the male brain and testes, but not in the liver relative to vehicle (ethanol) controls. No change in ERa methylation was found in the tissues of female fish. Both male and female medaka showed changes in CYP19A1 methylation state. Methylation increased in the liver and brain of males and decreased in the brain of females. Interestingly, no changes in the methylation of CYP19A1 were observed in the gonads (Contractor et al., 2004). The results indicate that steroid estrogens alter methylation state in a tissue and sex specific manner. Future work is needed to empirically link epigenetic mechanisms with alterations in gonad function, behavior, and reproduction in a transgenerational framework. Transgenerational epigenetic inheritance of phenotypes induced by EDCs has not been explicitly addressed in aquatic organisms. However, work on mammals suggest that EDCs can induce heritable changes to phenotypes. For example, altered (relative to vehicle (DMSO) controls) DNA methylation patterns observed in the F0 generation following exposure to the fungicide vinclozolin were found in rats three generations removed from the original exposure (Anway et al., 2005). These results convincingly link EDC exposure with heritable changes in DNA methylation and a maladaptive phenotype (subfertility) that was observed in the F2-F4 generations (Anway et al., 2005). This work (Anway et al., 2005) is one of a growing number of studies on mammals (Bruner-Tran and Osteen, 2011; Chamorro-García et al., 2013; Crews et al., 2007; Manikkam et al., 2013, 2012; Wolstenholme et al., 2012) that meet the definition of transgenerational epigenetic inheritance (Guerrero-Bosagna and Skinner, 2012). Studies of this duration and consequence have not been performed on aquatic organisms but would be of value given the potential evolutionary implications for wild populations (Bossdorf et al., 2008). The key to demonstrating transgenerational inheritance is extending the duration of studies for at least one generation after the initial exposure (Ho and Burggren, 2010). For example, if the F0 are exposed and then reproduce, the germ line forming the F1 was exposed through the mother and father (Ho and Burggren, 2010). Any effects observed in the F2 generation could then be suggestive of transgenerational inheritance because the F2 was never exposed. In the case of internal fertilizing organisms, if the F0 female is already pregnant, the F1 embryo and its germ line (what eventually develops into the F2) are exposed through the mother (Ho and Burggren, 2010). Thus, effects observed in the F3 generation may be indicative of transgenerational inheritance. While it may seem a semantic argument, it is important to distinguish between what looks like transgenerational effects and what is truly transgenerational inheritance. The implications of transgenerational inheritance of EDC effects are different than parental effects that induce transient phenotypic or developmental plasticity. That is, effects of EDC passed to the offspring for generations after the initial exposure suggest that effects of EDCs can be inherited through the germ line of the parents, grandparents, or great grandparents. The gonads are particularly sensitive to EDCs during development. Exposure during sexual differentiation can produce durable changes to germ line in mammals (Skinner et al., 2011). The effects (such as infertility or obesity) are observed in subsequent generations being transmitted by the sperm and eggs, are heritable, and derive from a diverse set of compounds from the insect repellent n,n-diethylmeta-toluamide (DEET) to bisphenol A (Anway et al., 2005; Bruner-Tran and Osteen, 2011; Chamorro-García et al.,

2013; Crews et al., 2007; Manikkam et al., 2013, 2012; Wolstenholme et al., 2012). For detailed examples of transgenerational inheritance of EDC effects in mammals see the excellent reviews by Crews and McLachlan (2006), Skinner et al. (2011), and Walker and Gore (2011). Demonstration of EDC induced transgenerational inheritance has been accomplished using pharmacologic doses in laboratory strains of mice and rats. In aquatic organisms, evidence of transgenerational inheritance of EDC effects remains limited. Future research should be extended into the F3 and F4 generations to provide more convincing evidence of transgenerational inheritance and that the EDC effects are passed through the germ line (Ho and Burggren, 2010; Skinner, 2008). Despite being resource intensive, investigators have much to gain by evaluating transgenerational inheritance of EDC effects in aquatic taxa. That is, if transgenerational inheritance is occurring, a reevaluation of the ecological and evolutionary consequences of EDC exposures will be needed in wild populations of aquatic organisms.

6. Conclusions and recommendations The literature to date demonstrates convincing evidence that some EDCs induce parental effects and in some cases investigators have isolated maternal or paternal side as the culprit. However, the parental effects of EDCs on aquatic organisms reviewed herein cannot be evaluated for their evolutionary significance because the heritability of transgenerational effects has not been demonstrated. The published literature in many cases employs excellent experimental designs aimed at identifying parental, maternal, or paternal effects. However, the studies have not been extended to the F3 or F4 generations, and thus cannot rule out other explanatory factors for the generational effects such as phenotypic plasticity or developmental plasticity (Bossdorf et al., 2008; Ho and Burggren, 2010). Given the significant progress and findings on heritable transgenerational effects of EDCs in mammals (Anway et al., 2005; Bruner-Tran and Osteen, 2011; Chamorro-García et al., 2013; Crews et al., 2007; Manikkam et al., 2013, 2012; Wolstenholme et al., 2012), hopefully similar work is underway in aquatic wildlife. The utility of ‘‘common garden’’ experiments has been demonstrated in several occasions in the literature (e.g. Hopkins et al., 2006; Todd et al., 2012) and is an ideal design for detecting transgenerational effects (Bossdorf et al., 2008). Common garden experiments involve the collection of organisms from different environments, for example a polluted site and reference site. The organisms are then brought to the laboratory where they are raised in a common environment. After a number of generations, an evaluation of physiological performance is conducted. Differences between populations can thus be attributed to the unique characteristics of the polluted site. If isolation of epigenetics as the mechanism behind heritable trangenerational phenotypes is the goal, investigators should use genetically similar organisms exposed for one or more generations experimentally and then raised in a common environment (Bossdorf et al., 2008). The use of genetically similar individuals is needed to ensure that the altered phenotype is not due to genetic differences (Bossdorf et al., 2008). Thus, increased used of parthenogenetic organisms with fast generation times such as Daphnia sp. or other Cladocerans will likely speed progress on evaluating heritability due to epigenetic mechanisms (Harris et al., 2012; Vandegehuchte et al., 2010). While Daphnia sp. and the approximately 90 species of fish, amphibians and lizards that are parthenogenic, they are incompletely parthenogenic and not true clonal populations (Lampert and Schartl, 2010). Still, use of even incomplete parthenogenic organisms should decrease the genetic noise found in out-bred populations. The use of clonal male rainbow trout has been effectively used to demonstrate

Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020

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paternal effects of EE2 (Brown et al., 2009, 2008, 2007) suggesting that these clonal fish are a valuable experimental model. In some instances the goal may be to identify heritable transgenerational effects regardless of the mechanism (i.e. epigenetic or genetic). In this case, standard model organisms used in aquatic toxicity testing should suffice as long as the experiment is conducted long enough (F3–F4) to ensure heritability. The application of gene expression analysis tools is currently underused in studies aimed at identifying transgenerational effects of EDCs in aquatic wildlife. The majority of biological endpoints investigated to date involve overt changes to reproductive output, morphology, behavior, and survival. In a few cases, steroid hormones (e.g. Foran et al., 2002; Schwaiger et al., 2002; Yu et al., 2011) and gene expression (e.g. Roepke et al., 2006; Yu et al., 2011) were evaluated in offspring from exposed parents. But a broader application of sublethal endocrine relevant endpoints is needed for evaluating the health of offspring derived from EDC exposed parents. Microarray technology could be an efficient means of detecting durable changes in gene expression patterns because of the ability to monitor the activity of large numbers of genes over many generations. While much has been accomplished future research must extend into the F3 generation and preferably the F4 generation. Depending on resources, research should also include an evaluation of durable changes to gene regulation, hormone levels, or changes in receptor abundance and localization. Acknowledgments The author is grateful for comments provided by Dr. Alan Vajda, Dr. Dan Villeneuve, and three anonymous reviewers whose input greatly improved the manuscript. References Alvarez, M.D.C., Murphy, C.A., Rose, K.A., McCarthy, I.D., Fuiman, L.D., 2006. Maternal body burdens of methylmercury impair survival skills of offspring in Atlantic croaker (Micropogonias undulatus). Aquat. Toxicol. 80, 329–337. http://dx.doi.org/10.1016/j.aquatox.2006.09.010. Andersson, T., Förlin, L., 1992. Regulation of the cytochrome P450 enzyme system in fish. Aquat. Toxicol. 24, 1–19. http://dx.doi.org/10.1016/0166-445X(92)90014E. Anway, M.D., Cupp, A.S., Uzumcu, M., Skinner, M.K., 2005. Epigenetic transgenerational actions of endocrine disruptors and male fertility. Science 308, 1466–1469. http://dx.doi.org/10.1126/science.1108190. Bergeron, C.M., Bodinof, C.M., Unrine, J.M., Hopkins, W.A., 2010. Bioaccumulation and maternal transfer of mercury and selenium in amphibians. Environ. Toxicol. Chem. 29, 989–997. http://dx.doi.org/10.1002/etc.125. Bergeron, C.M., Hopkins, W., Todd, B.D., Hepner, M.J., Unrine, J.M., 2011. Interactive effects of maternal and dietary mercury exposure have latent and lethal consequences for amphibian larvae. Environ. Sci. Technol. 45, 3781–3787. http://dx.doi.org/10.1021/es104210a. Binder, R.L., Lech, J.J., 1984. Xenobiotics in gametes of Lake Michigan lake trout (Salvelinus namaycush) induce hepatic monooxygenase activity in their offspring. Fundam. Appl. Toxicol. 4, 1042–1054. http://dx.doi.org/10.1016/ 0272-0590(84)90244-6. Bird, A., 2002. DNA methylation patterns and epigenetic memory. Genes Dev. 16, 6– 21. http://dx.doi.org/10.1101/gad.947102. Blair, R.M., Fang, H., Branham, W.S., Hass, B.S., Dial, S.L., Moland, C.L., Tong, W., Shi, L., Perkins, R., Sheehan, D.M., 2000. The estrogen receptor relative binding affinities of 188 natural and xenochemicals: structural diversity of ligands. Toxicol. Sci. 54, 138–153. http://dx.doi.org/10.1093/toxsci/54.1.138. Bonefeld-Jørgensen, E.C., Long, M., Hofmeister, M.V., Vinggaard, A.M., 2007. Endocrine-disrupting potential of bisphenol A, bisphenol A dimethacrylate, 4n-nonylphenol, and 4-n-octylphenol in vitro: new data and a brief review. Environ. Health Perspect. 115 (Suppl.), 69–76. http://dx.doi.org/10.1289/ ehp.9368. Bossdorf, O., Richards, C.L., Pigliucci, M., 2008. Epigenetics for ecologists. Ecol. Lett. 11, 106–115. http://dx.doi.org/10.1111/j.1461-0248.2007.01130.x. Brooks, S., Tyler, C., Sumpter, J., 1997. Egg quality in fish: what makes a good egg? Rev. Fish Biol. Fish. 416, 387–416. http://dx.doi.org/10.1023/A:1018400130692. Brown, K.H., Schultz, I.R., Cloud, J.G., Nagler, J.J., 2008. Aneuploid sperm formation in rainbow trout exposed to the environmental estrogen 17alphaethynylestradiol. Proc. Natl. Acad. Sci. U.S.A. 105, 19786–19791. http:// dx.doi.org/10.1073/pnas.0808333105. Brown, K.H., Schultz, I.R., Nagler, J.J., 2007. Reduced embryonic survival in rainbow trout resulting from paternal exposure to the environmental estrogen 17alpha-

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Please cite this article in press as: Schwindt, A.R. Parental effects of endocrine disrupting compounds in aquatic wildlife: Is there evidence of transgenerational inheritance?. Gen. Comp. Endocrinol. (2015), http://dx.doi.org/10.1016/j.ygcen.2015.01.020