Sonika Kar, Prabhaker Sangem, Narlagiri Anusha, Balasubramanian Senthilkumaran

Department of Animal Biology, School of Life Sciences, University of Hyderabad, P.O. Central University, Hyderabad, 500046, Telangana, India

Keywords:

Endocrine disruptors

Aquatic toxicity

Teleosts

Sex-steroids

Pesticides

Biosensors

ABSTRACT

Endocrine disruptors (EDs) are synthetic or natural chemical molecules occurring in environment that have the potential to impart adverse effects on homeostasis of endocrine axis leading to neurological, developmental,immunological and reproductive disarray at organismal level. A wide range of structurally diverse EDs such as,sex-steroid hormone mimics, pesticides and fertilizers, prevail in the environment originating from waste of industries, pharmaceutics, sewage treatment plants and agriculture. In addition, some metals, such as Cu, Hg and Zn, have endocrine disrupting potency in their metallic as well as synthesized nano-particulate forms. There is an increasing concern in research for the plausible threat posed by EDs that can disrupt the endocrine system in aquatic fauna as these compounds are frequently discharged or run-off into water stream. Fishes are well known bio-indicators to understand toxicity of EDs as they are vulnerable to endocrine disruption. Furthermore, EDs have the potential to affect fish-feeding higher vertebrates including mammals and subsequently human, as they make their way up on the food web pyramid due to biomagnification. In light of this, several observations suggesting adverse effects of EDs and the mechanism contributing to endocrine disruption in fish are discussed extensively in this review. This article highlights the necessity to choose a credible model for assessing the toxic effects exerted by EDs. Furthermore, the toxic effects of EDs will be comprehensively reviewed with reference to sexual plasticity, neuroendocrine mechanisms, thyroid and immune modulation, gonadal development and maturation as well as changes in transcriptome/genome profile using fish models to imply ED-induced aquatic pollution in a larger perspective. For decades now, studies on EDs have challenged traditional concepts in toxicology to develop new molecular markers to improve methodologies and to assess the ecological risks associated with field conditions. In this regard, it is imperative to highlight the development of modern diagnostic tools including biosensors to monitor the inadvertent usage of EDs and the resultant environmental risks.Lastly, current limitations in knowledge along with future research perspectives in the field are also highlighted in this article.

1.Introduction

Endocrine disruptors (EDs) are exogenous substances that alter the functions of endocrine system in living organisms causing adverse health effects to an individual or a population. EDs are known to provoke disarray of hormonally controlled physiological parameters (such as mineral and osmotic balances) or functions (such as growth,development and reproduction). There are different mechanisms through which EDs can act, either by binding to hormone receptor or altering the levels of endogenous hormones, or by modulating gene networks. This leads to stimulation/inhibition of downstream cellular and molecular pathways, thereby, affecting the normal parameters and functions. Lately, a large number of

in vitro

assays have been used extensively by several agencies to identify EDs, out of which a majority are organic molecules with low molecular weights. In the contrary,

in vivo

assays for ED detection are highly sensitive and provide biologically ecorelevant results than

in vitro

assays. Moreover, EDs lay among other complex contaminant families that easily arrive at aquatic system exerting adverse effects to aquatic fauna. EDs include pesticides, hormone mimics, heavy metals, polychlorinated biphenyls (PCBs), phthalates, organic solvents, flame retardants, surfactants, pharmaceuticals,etc. In addition to this, some EDs are naturally occurring, such as phyto-/myco-estrogens, which are synthesized naturally by fungi/plants (Liu,Kanjo, & Mizutani, 2010) whereas some other EDs are synthesized by chemical breaking down in industries such as polycyclic aromatic hydrocarbons (PAH). EDs are known to impart irreversible reproductive outcomes in wildlife fauna (Tubbs & McDonough, 2018). Feminization of males and infecundity are few of such outcomes (Gimeno, Gerritsen,Bowmer, & Komen, 1996). Apart from reproductive anomalies, EDs are also known to cause disarray of other endocrine systems across various axes, for example, hypothalamo-hypophyseal-gonadal (HHG) axis, thyroid and other cellular systems by exerting either antagonistic or agonistic effects upon binding to hormone receptors. Furthermore, some of these unfavourable effects (observed in wildlife fauna/experimental organisms) may also prevail in human if exposed at certain concentration and time causing endocrine dysregulation. Impact of ED exposure has been studied extensively in aquatic animals, particularly, in fishes.Hundreds of EDs have been detected till date, however, a very few have been tested. Nonetheless, the mechanism of action is still unknown for a wide range of chemicals. In addition, only a very few combinatorial studies have been done so far. For example, combination of endosulfan(an insecticide) and flutamide (a non-steroidal anti-androgenic drug)impaired both testicular (Rajakumar et al., 2012) and ovarian growth(Chakrabarty et al., 2012) in cat fish. Another study demonstrated the combinatorial disruptive effects of fenvalerate and triadimefon during embryonic development in the Chinese rare minnow,

Gobiocypris rarus

(Wu, Hu, Zhao, Wang, & Jiang, 2018).Fishes are well known bio-indicators to understand toxicity of EDs as they (Hutchinson, Ankley, Segner, & Tyler, 2006) show sexual plasticity and sensitivity towards sex steroids or xenobiotic compounds (Nagahama, Nakamura, Kitano, & Tokumoto, 2004). Incidentally, many sex steroid analogues or pesticides are known to cause estrogenic effects on variety of physiological systems in fish. In this context, model organisms, such as, zebra fish, medaka and fathead minnow, have been scrutinised. This review highlights valuable inferences of ED toxicity gained using model as well as non-model organisms including food fishes like carps (

Cirrhinus mrigala, Catla

, and

Labeo rohita

), followed by murrels(

Channa marulius

,

C. punctatus

and

C. striatus

) and cat fishes (

Clarias gariepinus

,

C. batrachus

and

Heteropneustes fossilis

). In most of the cases,mere resemblance of certain molecular structures of EDs is sufficient enough to act like native steroid hormones and hence targeting the gonads. However, several EDs often alter HHG axis to impart deleterious effects on gonads eventually. Furthermore, EDs have the potential to affect higher vertebrates owing to biomagnification (Borga, Gabrielsen,& Skaare, 2001; Vives, Moreira, Brienza, Zucchi, & Nascimento Filho,2006; Schäfer et al., 2015). In light of this, several observations suggesting adverse effects of EDs in fish in terms of sexual plasticity,neuroendocrine mechanisms, thyroid and immune modulation, gonadal development and maturation along with transcriptome/genome profile modulations are discussed in this review. Furthermore, advancements in ecological risk assessments of EDs are also highlighted.

1.1.EDs: Broader classification, chemical characteristics and mechanism of action

The mechanisms of ED action involve divergent pathways such as estrogenic, androgenic, thyroid and several other pathways which are highly conserved in wildlife and acts primarily through nuclear receptors of androgen, estrogen, progesterone, thyroid hormone, neurotransmitters and aryl hydrocarbon among others as well as steroidogenic enzymes. EDs represent a broad class of molecules including pesticides,industrial contaminants, plasticizers and other chemicals that are present in the environment or in widespread use. Examples of most potent EDs are represented in Table 1. The chemical empirical structures of these EDs are shown in Fig. 1 and are categorized by considering the mode of action in terms of function.

Table 1Examples of most potent EDs, their chemical characterization, usage andmechanism of action.

Table 1(continued)

Fig. 1.Chemical structures of some potent EDs and their mode of action/function via A) Estrogen receptor; B) Androgen receptor; C) Thyroid hormone receptor; D)AchE receptor/Adrenergic receptor; E) Aryl hydrocarbon receptor; F) Progesterone receptor. Additionally, the alphabets (a, b, c) besides certain EDs indicate their multiple actions via estrogen, androgen, thyroid hormone receptors respectively.

1.2.EDs pose environmental risks

Several studies have focussed on sites with highly persistent ED contamination (e.g., the pacific waters, the Great Lakes, the Baltic Sea area, etc) until recently, wherein, less persistent chemicals are also being evaluated. However, the prevailing data on bioaccumulation of EDs in aquatic ecosystems seem alarming and need attention. Hence, more innovative measures for identification, quantification and development of more robust biological assays for ED detection and their risk assessment are needed to establish dose-response and cause-effect relationships in laboratory as well as in field conditions.

Although, there are a very few reports depicting adverse health effects due to ED exposure in human, sufficient laboratory evidences are well documented to implicate ED toxicity to many wildlife species(Tanabe, 2002; Vos et al., 2000). Hence, the threat posed by EDs cannot be overlooked as it can potentially harm fish-consuming fauna vis-`a-vis human health due to biomagnification. Insights from ED research in various animal models have a positive impact on current practice in toxicological testing and screening. Addressing about underlying environmental causes and assessment of risk factors pertaining to endocrine disruption may aid in neutralizing the negative impacts of EDs on wildlife growth and reproduction. One way of tackling these issues is by regulating the use of these disrupting chemicals through bans and restrictions imposed by the government. This has been proven effective in several cases, for example, bans and restrictions on Pb, pesticides such as endosulfan (Sharma et al., 2011), chlorpyrifos, tributyltin, certain PCBs,dichlorodiphenyltrichloroethane (DDT) and other pollutants, wherein exposure has been drastically reduced by such measures. Furthermore,several validation methods have been formulated till date, however,they capture and address a very narrow range of known disrupting effects. Use of biomarkers and ecological risk assessment through exposure characterization to detect and analyze ED contamination are few of the solutions.

1.2.1.Ecological risk assessment and exposure characterization

Ecological risk assessment of EDs depicts the likelihood of adverse effects due to their exposure. At many instances, EDs are known to cause decrease in mass productivity of aquatic fauna. In this context, several laboratory and field studies (Leatherland, 1993; Soimasuo et al., 1995;Allen, et al., 1999b; Matthiessen et al., 2002; Hashimoto et al., 2000;Lubick, 2011; Gimeno et al., 1996) have established dose-response or cause-and-effect relationship between EDs. Several ecotoxicity tests were performed that included assessment of mortality, reduction in growth, reproductive impairment, number of species affected and determination of concentration of EDs in water that causes 50% lethality(LC) as well as immobilization (EC) in a test population, and bioaccumulation of residues in target and non-target tissues. Short-term reproductive assays (Ayobahan et al., 2020; Brain et al., 2018; Dang,Traas, & Vermeire, 2011; Mihaichi et al., 2015; Wheeler,Valverde-Garcia, & Crane, 2019), embryo toxicity tests using fishes(Braunbeck et al., 2005; Lammer et al., 2009), bioconcentration and biomagnification toxicity tests in aquatic systems (Schäfer et al., 2015)are few of the examples of studies to assess the potential risks involving early life-stage and full life-cycle. Several

in vivo

fish screening assays have been reviewed which proposed to provide comprehensive screening stratergies for several hormonal functions (Hutchinson et al.,2000). Such studies have far reaching ecologically important objectives in characterizing the persistence and transport of EDs in the environment. Furthemore, several scientists and agencies worldwide have evaluated chemicals for their endocrine-disrupting potential (Browne,Noyes, Casey, & Dix, 2017).Exposure characterization is one of the risk analyses step in which ecological interaction with an environmental agent of concern (ED in this context) is evaluated, in other words, the potential exposure of animals, plants and water resources to EDs residues in food, water, air and soil can be estimated by exposure characterization. This includes estimation of the frequency, magnitude and duration of exposure, along with characteristics of the exposed population. In this light, scientific tools such as quantitative structure—activity relationship models and several other integrated comptutaional models have emerged for highthroughput screening of compounds (Cohen et al., 2010). Additionally, field studies have provided a more realistic picture about the fate of parent ED compounds and their breakdown products in the environment. Based on the data from

in silico

simulations,

in vitro

assays,

in vivo

screening assays and field studies, United States Environmental Protection Agency has provided a quantitative environmental fate profile for assessment of ED exposure (https://www.epa.gov/pesticide-scie nce-and-assessing-pesticide-risks/aquatic-life-benchmarks-and-ecolog ical-risk).

2.ED contamination in aquatic ecosystems

Aquatic ecosystems are mostly affected by contamination with EDs in addition to other chemical compounds. EDs, produced during water disinfection and treatment of sewage plants, often reach human due to daily house-hold requirements (Holmes, Smeester, Fry, & Weinberg,2017). Conversely, EDs also have the potential to initiate disinfection by-products (DBPs) formation upon reaction with water disinfectants(Bárány, 2013). Concomitant to this, DBPs binds to steroid hormone receptors resulting in altered gene expression and hence endocrine disruption that can incur some adverse effects on public health (Kar &Senthilkumaran, 2020). As mentioned earlier, EDs easily reach fish-consuming higher vertebrates, mammals and eventually human through marine and terrestrial food webs due to biomagnification. For example, embryonic deformities of herring gulls ( fish-consuming birds)have been observed in the Great Lakes contaminated with PCBs and PAH(Fox, 1993). In this context, effects of organohalogens have been reviewed thoroughly in fish-eating birds with reference to reproductive outcomes, hormone metabolism and circulating steroid levels (Bosveld& Berg, 2002). Research works, in the field, depict human exposure to EDs may aid in increase of non-communicable diseases susceptibility and, time to time, have been associated with contamination outbreaks,for example, the National Geophysical Research Institute in Hyderabad,India and Central Pollution Control Board in Delhi have examined heavy metals and other pollutants in the lakes/ponds around Patancheru,Hyderabad for several years now, which accomodates numerous industries producing harmful chemicals, hence, exerting serious adverse effects on the development of inhabiting organisms including fish species (Lubick, 2011).

Apart from teleosts, several studies depicted ED-induced reproductive failure and thyroidal anomalies in seals, whales, etc due to PCBs and polybrominated diphenyl ethers (PBDEs) contamination (Krahn et al.,2009; Nyman, Koistinen, Fant, Vartiainen, & Helle, 2002; Rayne, Ikonomou, Ross, Ellis, & Barrett-Lennard, 2004; Ross, Ellis, Ikonomou,Barrett-Lennard, & Addison, 2000).

Aquatic ecosystems of several nations like United Kingdom (Jobling,Nolan, Tyler, Brighty, & Sumpter, 1998; Tyler & Jobling, 2008; Jobling,Burn, Thorpe, Williams, & Tyler, 2009; Gross-Sorokin Roast & Brighty,2006), the United States (Blazer et al., 2012; Grieshaber et al., 2018;Yamamoto, Garcia, Kupsco & Oliveira-Ribeiro, 2017), Canada (Fuzzen,Bennett, Tetreault, McMaster, & Servos, 2015; Kidd et al., 2007;Tetreault et al., 2011), certain European countries (Jarque et al., 2015;Müller et al., 2019), the Middle East (Gilannejad et al., 2016), Africa(Chukwuka, Ogbeide, & Uhunamure, 2019; Ibor, Adeogun, Fagbohun, &Arukwe, 2016), India (Gupta, Rai, Pandey, & Sharma, 2009; Lubick,2011) and China (Chen et al., 2016; Zheng, Liu, Liu, Jin, & An, 2015),were widely affected by various categories of EDs and heavy metals contamination leading to phenomenal consequences of intersex conditions, gonadal aberrations and altered reproductive responses in fish species. The global status of knowledge on ED exposure with reference to several endocrine axes in fish is reviewed in the following sections.

3.Impact of EDs on HHG axis

EDs affect HHG axis by targeting genes related to monoaminergic systems, gonadotropin-releasing hormone-gonadotropin (GnRH-GTH)axis and steroidogenesis thereby altering the feedback regulation associated with gonadal function and reproduction (Haider & Inbaraj, 1986;Lam, 1983; Senthilkumaran, 2015). By means of steroid mimicking nature, organochlorine and organophosphate families of pesticides exert differential effects on HHG axis targeting gonadal function (Senthilkumaran, 2015). Organochlorine pesticides include DDT, dichlorodiphienylethylene (DDE), dieldrin and endosulfan, among which endosulfan's toxicity studies are abundant. Several studies in fishes revealed toxic effects of endosulfan on testis and ovary by modulating the expression patterns of several gonad-related transcription factors and steroidogenesis-related enzymes in addition to imparting deleterious effects on brain-pituitary (Chakrabarty et al., 2012; Rajakumar et al., 2012). Endosulfan treated males also showed impaired spermatogenesis progression and testis growth in the juvenile Asian cat fish,

C. batrachus

wherein similar effects were observed for flutamide(another ED) exposure (Rajakumar et al., 2012). On the other hand,females showed precocious ovarian development and alteration of ovary-specific genes after the exposure of endosulfan. Serotonin is an excitatory neurotransmitter for GnRH-GTH release and tryptophan hydroxylase (tph) being the rate limiting enzyme for its synthesis in teleosts (Goos, Senthilkumaran, & Joy, 1999; Raghuveer et al., 2011;Saligaut et al., 1992; Senthilkumaran & Joy, 1996; Sudhakumari et al.,2010), any alteration in this loop affects gonadal differentiation or development. ɤ-aminobutyric acid (GABA) receptor blockade (Jia &Misra, 2007; Trudeau, Sloley, & Peter, 1993) by endosulfan and sex-steroids such as 17β-estradiol (E) and T was also evident, which might have an influence on GnRH release as GABA is known for its stimulatory (Senthilkumaran, Okuzawa, Gen, & Kagawa, 2001) and inhibitory (Khan & Thomas, 1999) role on GnRH-GTH release in teleosts. Sex steroids are known to elicit feedback regulation not only by targeting monoaminergic system (Larson, Norris, Gordon Grau, &Summers, 2003; Larson, Norris, & Summers, 2003; Senthilkumaran &Joy, 1996; Trudeau, Peter, & Sloley, 1991) but also neurosteroids(Whitlock, Wolf, & Boyce, 2003). In addition to this, changes in

cyp19a1b

, is responsible for aromatization of androgens, led to increased availability of estrogens (Andersen et al., 1999; Chang et al., 2005; Lee et al., 2001; Rasheeda, Kagawa, Kirubagaran, Dutta-Gupta, & Senthilkumaran, 2010; Zhou et al., 2005), and influenced GnRH neurons in fishes (Amano et al., 1994; Dubois, Florijn, Zandbergen, Peute, & Goos,1998). Incidentally, Joy (1994) established the presence of cholinergic system in teleosts, yet its role in gonadal function is still unclear. In conclusion, endosulfan exerted sexually dimorphic effects (Pandey,1988; Prathibha et al., 2014; Rajakumar et al., 2012; Senthilkumaran,2015; Stanley, Curtis, Massey-Simonich & Tanguay, 2009) at the level of brain by modulating transcripts like GnRH and disrupting serotonergic system loop in teleosts. However, impact of EDs on cholinergic system of fish is yet to be investigated.Reports on malathion exposure (an organophosphate pesticide)showed differential effects on lipid metabolism and growth retardation by attenuating hormone levels and targeting multiple systems in addition to pituitary and ovary in teleosts (Lal, Sarang, & Kumar, 2013; Lal &Singh, 1987; Patil & David, 2010; Singh, 1992). Dose-related differential effects of malathion on

cfgnrh

,

tph

and tyrosine hydroxylase were evident in male and female brain of the air-breathing cat fish,

C. batrachus

(Prathibha et al., 2014). In addition to this, malathion disrupted gonadal development wherein slow progression of spermatogenesis and high number oil droplet oocytes in ovary were evident(Prathibha et al., 2014). GnRH-GTH release is regulated by serotonin and catecholamines (Goos et al., 1999; Senthilkumaran et al., 2001;Senthilkumaran & Joy, 1996) and hence any direct action of malathion on these correlates might inturn affect the expression of certain genes/factors related to neurosteroidogenesis (Prathibha et al., 2014; Senthilkumaran, 2015). In conclusion, it is also plausible to assume a direct action of malathion on gonads which might affect GnRH-GTH. However,the impact of malathion is least studied at the level of HHG axis as a whole and is yet to be explored more.Taken together, EDs belonging to organochlorine and organophosphate families or any other similar compounds impair HHG axis by selective depletion of various genes/factors resulting in gonadal function disarray (Prathibha et al., 2014; Senthilkumaran, 2015). In accordance to this, azocyclotin (an insecticide) altered transcription of several genes in HHG axis implicating adverse consequences on reproduction such as inhibition of oocyte maturation and retarded spermatogenesis in zebra fish (Ma, Cao, Wang, Gui, & Zhu, 2016). A recent study reported that the exposure of herbicide, glyphosate, in the Japanese medaka,

O. latipes

selectively altered neuroendocrine-related genes like

kiss1

and

kiss2

in female brain in addition to reproductive, developmental, and epigenetic modifications (Smith, Vera, & Bhandari, 2019).

Like pesticides, pharmaceuticals found in aquatic ecosystem as well as in wastewater treatment plants have sex steroid mimic nature which can cause serious effects on non-target organisms including fishes (Fent,Weston, & Caminada, 2006). In this line, recent studies in hybrid striped bass and fathead minnows revealed alterations in brain serotonin levels upon exposure to anti-depressant pharmaceutical drugs like fluoxetine and venlafaxine (Gaworecki & Klaine, 2008; Bisesi, Bridges, & Klaine,2014). PAHs such as benzo[α]pyrene also induced deleterious effects on HHG axis in teleosts (Clotfelter, McNitt, Carpenter, & Summers, 2010;Gesto, Tintos, Soengas, & Míguez, 2006; Hose et al., 1981; Yarahmadi,Movahedinia, Savari, Rassouli & Sahraeian, 2013). In conclusion,several pharmaceuticals and naturally occurring EDs altered the monoaminergic system (Cunha, Rodrigues, Santos, Moradas-Ferreira&Fenske, van Aerle, Brack, Tyler, & Segner, 2001) and serotonergic system (Mennigen, Stroud, Zamora, Moon, & Trudeau, 2011) in fishes.However, many more such compounds are yet to be screened, identified and evaluated in this aspect. Subsequent paragraphs highlight the impacts of several other groups of EDs on HHG axis in teleosts.

Exposure of bisphenol A (BPA) in zebra fish embryos caused deleterious effects on neuroendocrine system along with modulation of several brain-related genes like

kiss1

,

kiss1r

,

gnrh3

,

lhβ

,

fshβ

and

er1

(Qiu et al.,2016). Similarly, sex steroid analogues such as 17α-ethinylestradiol(EE) and methyltestosterone (MT) also imparted adverse effects on HHG axis by influencing the monoaminergic-GnRH system in the air-breathing cat fish,

C. gariepinus

(Swapna & Senthilkumaran, 2009).Coincidentally, synthetic progestin, dydrogesterone affected transcription of GnRH and steroidogenesis pathway genes in female zebra fish(Shi et al., 2019). Estrogen and progestin mimics present in the environment affects brain aromatase,

cyp19a1b

, expression in zebra fish(Cano-Nicolau et al., 2016; Zeilinger et al., 2009), which may in turn target feedback loop modulating GnRH-GTH release. Hinfray et al.(2016) successfully employed

in vivo

and

in vitro

bioassays in zebra fish to analyze expression of the brain specific

er

-regulated

cyp19a1b

gene after exposing to estrogen or progestin mimics alone or in combination. In conclusion, sex steroid analogues altered brain function by modulating GnRH and steroidogenic pathway gene expression. Like brain aromatase,

cyp19a1b

(a potential biomarker), other biomarker genes are yet to be identified in order to identify ED-induced anomalies in HHG axis.Like sex steroids, heavy metals for example, Zn, Cu, Fe, Hg, Mn, As,Cd, Cr and Pb are essential regulators of biological functions (Farkas,Salanki, & Varanka, 2000; Handy, 2003; Lwanga, Kansiime, Denny, &Scullion, 2003; Nawaz, Nagra, Saleem, & Priydarshi, 2010). However,they could be highly toxic to living organisms including fishes in their native/compound or nano-particulate forms based on exposure concentration and developmental stages or age (Deepa, Murugananthkumar, Gupta, Gowda, & Senthilkumaran, 2019; Denier, Hill,Rotchell, & Minier, 2009; Mance, 1987). Few of these reports used either adult or juvenile fishes as indicators of metal contamination at several aquatic ecosystems to evaluate the underlying risks to human health due to biomagnification (Vives et al., 2006). Recent studies reported endocrine disrupting activity of heavy metals such as Cd, Cu, Fe, Mn, Pb, Hg,Zn and methyl mercury (MeHg) on HHG axis in several fish species (Cao et al., 2019; Crump & Trudeau, 2009; Gárriz, Fresno, Carriquiriborde, &Miranda, 2019; Olivares-Rubio et al., 2015; Tilton, Foran, & Benson,2003). A study in zebra fish reported adverse effects of metallic compound, BaClon several brain-related and sex steroid receptor genes(Kwon et al., 2016). A report using the Atlantic croaker,

Micropogonias undulatus

depicted that PbCland aroclor 1254 (a PCB) altered the stimulatory serotonin-GnRH pathway that might further lead to gonadal impairment (Khan & Thomas, 1997, 2000; Thomas & Khan, 1997). In summary, heavy metals and metallic compounds exerted adverse effects on HHG axis by modulating sex steroid levels, altering the expressions of several brain related genes and disrupting the serotonin-GnRH pathway.

In view of these, it is possible to suggest that the EDs have the potential to impart serious effects on HHG axis and exhibit biphasic action based on the sex of teleost species to alter sex-wise population in aquatic ecosystem. In fishes, existence of dual GTH receptors in ovarian follicles has been well demonstrated (Oba et al., 1999), yet EDs affecting these correlates are not identified till date. Hence, a focus on GTH receptors might provide more information to understand the impact of EDs on HHG axis. Taken together, EDs possibly target HHG axis by acting through monoaminergic system vis-`a-vis feedback regulation. However,it is also plausible that the action might also be triggered directly on gonads to affect brain-pituitary leading to complex effects on reproductive function. Considering the direct action of EDs on gonads due to sex-steroid mimicking nature, subsequent section analyzes gonad-specific endocrine disruption.

4.Impact of EDs on gonads

Fishes pose diverse reproductive strategies, most of which are poorly understood (Kime, 1998). Sexual plasticity is one of the reproductive strategies in certain fish species wherein individuals can reversibly change their from one functional sex to the other in response to various environmental and social cues. Fish reproduction has been an ecologically relevant indicator of endocrine disruption since last four decades(Arcand-Hoy & Benson, 1998). Exogenous exposure of sex steroid such as 11-ketotestosterone (11-KT), a potent fish androgen, is known to act agonistically on androgen receptor in adult female gold fish (Kobayashi,Aida, & Stacey, 1991). EDs alter the expression of cytochrome P450(CYP) enzymes which play critical roles in sex differentiation and development in teleosts (Kazeto, Place, & Trant, 2004). EDs in the environment are known to affect and exploit naturally occurring sexual plasticity in fish wherein the underlying mechanisms are largely unknown.

4.1. 1.EDs impair vitellogenesis and gonadal dysfunction

As discussed earlier, the scientific community worldwide has been able to establish the induction of VTG production using juvenile or male fish as one of the most notable and convincing biological response linked to EDs exposure that are mostly estrogenic in nature (Harries et al.,1999; Tyler & Routledge, 1998). Several reports in various fish species depicted similar effects (Flammarion et al., 2000; Nichols,Miles-Richardson, Snyder, & Giesy, 1999; Purdom et al., 1994). Studies on concomitant implications of these responses in terms of gonadal impairment and reproductive success are very crucial. Several such research evidences, to the best of our knowledge, are briefly reviewed below.

Alterations in VTG levels was observed upon exposure of synthetic estrogen such as EEand natural estrogens like estrone and Ealong with EDs such as DDT, BPA, methyl paraben, alkylphenols, its derivatives and several pharmaceutical drugs in teleosts (Celius, Haugen,Grotmol, & Walther, 1999; Chen, Wu, Tsai, Hsien, & Huang, 2019;Dambal, Selvan, Lite, Barathi, & Santosh, 2017; Hylland & Haux, 1997;Ibor et al., 2016; Johnson & Lema, 2017; Kim et al., 2012; Larsson et al.,1999; Leatherland, 1993; Nimrod & Benson, 1997; Jobling, Sumpter,Sheahan, Osborne & Matthiessen, 1996). Induction of VTG production is often associated with developmental disruptions and anomalies in fishes such retarded testicular growth and formation of egg cells in testis (Gray& Metcalfe, 1997; Harries et al., 1996; Lye, Frid, & Gill, 1998; Lye, Frid,Gill, Cooper, & Jones, 1999; Lye, Frid, Gill, & McCormick, 1997).

Apart from exerting estrogenic activities, inducing VTG production and causing testicular abnormalities, a number of reports on EDs elucidated occurrences of intersex, wherein, the riverine fish roach,

Rutilus rutilus

; the sturgeon,

Scaphirhynchus platyorynchus;

the male flounder,

P. flesus;

the Nile tilapia,

O. niloticus

; the common dab,

Limanda limanda

and the gudgeon,

Gobio gobio

were scrutinised (Aerle et al.,2001; Harshbarger, Coffey, & Young, 2000; Jobling et al., 1998, 2002;Kosai, Jiraungkoorskul, Sachamahithinant, & Jiraungkoorskul, 2011;Stentiford & Feist, 2005; Tyler & Routledge, 1998), a few of which were confirmed by histology of gonads. Exposure of DDT in the Japanese medaka,

O. latipes

, led to an intersex condition of gonad in male (Metcalfe et al., 2000) and altered gonadal morphology in

O. mossambicus

(Mlambo, Van Vuren, Barnhoorn, & Bornman, 2009). A report using the African cat fish,

C. gariepinus,

depicted that early exposure of EEand diethylstilbestrol (DES) induced morphological changes and altered ovarian steroidogenic pathway (Sridevi et al., 2015). Upon treatment with EE, reduction in spermatids was observed in males whereas MT treatment led to precocious ovarian development in females which were confirmed by gonadal histology (Swapna & Senthilkumaran, 2009). EEexposure altered the expression of

cyp19a1a

and

cyp19a1b

in zebra fish(Kazeto et al., 2004). Similar effects were reported upon exposure to nonylphenol, octylphenol, BPA and benzo[

a

]pyrene (an ED as well as a known carcinogen) in zebra fish (Alharthy, Albaqami, Thornton, Corrales, & Willett, 2017; Kazeto et al., 2004) which was also reported in other fish species such as

Rivulus marmoratus

(Lee, Seo, Kim, Yoon, &Lee, 2006) and

S. salar

(Meucci & Arukwe, 2006).

Synthetic progestins (such as drospirenone, gestodene, norgestrel)and pharmaceutical drug such as (atrazine, fadrozole and metformin)altered sex steroid levels, induced reproductive behavioural changes,pertured steroid synthesis and affected gonadal morphology in several fish species (Ayobahan et al., 2020; Blanco, Fernandes, Medina,Blázquez, & Porte, 2016; Frankel et al., 2016; Hou et al., 2018; Niemuth& Klaper, 2015; Vasanth, Bupesh, Vijayakumar, Subramanian, & Ramasamy, 2018). Trenbolone exerted androgenic effects in fathead minnow(Ankley et al., 2003) like the effects imparted by pesticides such as linuron, fenitrothion, and vinclozolin (Katsiadaki et al., 2006; Makynen et al., 2000).

Other abnormalities included impaired milt production (Jobling et al., 1998) and altered spermatogenesis (Lye et al., 1998) in European flounder. Reduced gonodopodium (modified anal fin) size was observed in the mosquito fish,

G. affinis

(Batty & Lim, 1999) which is very critical for sperm transfer and is formed under the influence of testosterone (T).On the contrary, earlier reports suggested strong masculinization of the anal fin in mosquito fish upon exposure to phytosterols (such as β-sitosterol, campestrol, and stigmastanol) and other naturally occurring androgenic chemicals (Denton, Howell, Allison, McCollum, & Marks,1985; Howell & Denton, 1989; Howell, Black, & Bortone, 1980).Long-term exposure to cythion (an organophosphorus pesticide)affected reproduction in

C. punctatus

by biochemical changes in gonad and degeneration of immature oocyte (Ram & Sathyanesan, 1987).Recent studies on the delta smelt,

Hypomesus transpacificus

and the Japanese medaka,

O. latipes

, reported alteration in endogenous steroid levels and developmental toxicity upon exposure to herbicides like glyphosate and fluridone (Jin et al., 2018, 2020). Apart from chemical compounds mentioned in the previous sections, heavy metals are also known to exert toxic effects on gonads in their native/compound as well as in nano-particulate forms which are brie fly reviewed further.A report using two fish species,

C. punctatus

and

Aorichthys aor

demonstrated occurrence and bioaccumulation of heavy metals in the river Ganges, Allahabad, India wherein metals accumulation in fish muscles was found to be in the order as follows: Zn

＞

Pb

＞

Cu

＞

Cr

＞

Cd(Gupta et al., 2009). A study using the shark cat fish,

Pangasianodon hypophthalmus

and the Japanese eel,

Anguilla japonica

, suggested that heavy metals such as Pb, Mo, Rb and As might exert inhibitory effects on spermatogenesis (Yamaguchi et al., 2007). Concomitant to the findings in these reports, earlier reports suggested that Zn and its compounds accumulation in fish tissues might result in increased mortality rate,growth retardation and hypoxia conditions (Shaf fi, 1979). Pb(NO)caused hormonal imbalances and disarray in normal follicular steroidogenesis in stinging cat fish,

H. fossilis

(Chaube, Mishra, & Singh,2010). Toxic effects of CdClwere observed on rainbow trout's embryos(Lizardo-Daudt & Kennedy, 2008). Another study using

O. mossambicus

depicted estrogen-like growth-promoting property of Cd in juveniles, yet in adult fish it negatively affected reproduction (Amutha & Subramanian, 2013). MeHg and compounds like HgClexerted androgenic affect and inhibited gonadal development in several fish species (Baldigo et al., 2006; Dey & Bhattacharya, 1989; Friedmann, Costain,MacLatchy, Stansley, & Washuta, 2002; Friedmann, Watzin,Brinck-Johnsen, & Leiter, 1996; Kime, 1998; Kirubagaran & Joy, 1988).In the last two decades, increased production and usage of metallic nanoparticles (NPs) have inevitably led to heavy discharge into different compartments of environment including aquatic ecosystems resulting in accumulation at various levels of the food chain. A study, in this light,depicted morphological disarray in testis of

C. batrachus

upon treatment with Cu/Cu-NPs which was further confirmed by transmission electron microscopy (Gupta et al., 2016; Murugananthkumar, Rajesh, & Senthilkumaran, 2016). ZnO-NPs and ZnSOexposure to common carp exhibited defective testicular lumen which was confirmed through histology (Deepa et al., 2019). Furthermore, ZnO-NPs induced developmental malformations and decreased hatching rates in zebra fish embryos (Han, Zhai, Liu, Hao, & Guo, 2017). In conclusion, various categories of EDs led to intersex condition, caused alterations in sex steroid levels and steroidogenic enzyme gene expression, and induced anomalies in gonadal morphology along with several other cellular dysfunctions wherein hormone biomarkers were used to detect EDs induced aberrations. However, identification of genetic biomarkers will ensure toxicity detection elegantly.

4.2. 2.EDs cause decrease/failure of reproductive success

Apart from morphological anomalies and gonadal dysfunctions,several studies depicted reduced spawning success in females (Johnson,Casillas, Collier, McCain, & Varanasi, 1988; Johnson et al., 1992) and reduced sperm motility, spawning and, fertilization success in teleosts(Wu, Zhou, Randall, Woo, & Lam, 2003). Reduction in zebra fish hatching rate was also observed upon exposure to toxaphene, an insecticide (Ree & Payne, 1997). Exposure to p,p-DDE affected the expression of genes involved in reproduction and histological alterations in the number of mature oocytes in largemouth bass and zebra fish, respectively (Garcia-Reyero, Barber, Gross, & Denslow, 2006; Monteiro et al., 2015). Long term exposure of carbaryl on fathead minnow affected their survival, growth, reproduction and induced larval mortality (Carlson, 1972). Moreover, early-life-stage mortality has been observed in the lake trout,

Salvelinus namaycush

, the rainbow trout,

O. mykiss

, the Atlantic salmon,

S. salar

(Hansson & Hahn, 2008) and the zebra fish,

Danio rerio

(King-Heiden et al., 2012) upon exposure to tetrachlorodibenzo-p-dioxins, polychlorinated dibenzo-p-dioxins, dibenzofurans and PCBs. A report on perchlorate (an industrial chemical)depicted significant reduction of spawning success in zebra fish (Mukhi& Patino, 2007).

5.Impact of EDs on thyroidal function

Since last four decades, several evidences of thyroid hyperplasia have been reported in bony fishes upon natural exposure of EDs such as PCBs, polybrominated biphenyls, DDT, its metabolites and insecticides like mirex in the great lake of Eric, Michigan and Ontario of U.S.A.(Black & Simpson, 1974; Drongowski, Wood, & Bouck, 1975; Leatherland, 1992; Moccia, Leatherland, & Sonstegard, 1981; Sonstegard &Leatherland, 1976). Some of these observations were further confirmed by histology using the coho salmon,

O. kisutch

as a model organism,among all other inhabiting fish species (Leatherland, 1992) and by analysing the effect of biomagnification in the herring gull,

Larusargentatus

wherein enlarged thyroid follicles were evident (Moccia,Fox, & Britton, 1986). This paved a way for researchers to analyze the impact of EDs on several other fish species focusing particularly on thyroid. Moreover, EDs such as PCBs and pharmaceuticals often transfer across the trophic levels accumulating more from one to other, ultimately affecting human due to biomagnification (Borgå et al., 2001)Furthermore, consumption of EDs exposed to fish resulted in thyroid hormone disturbances in human wherein impaired triiodothyronine (T)and thyroxine (T) levels were reported in the great lake fish consumers(Persky et al., 2001). Several reports suggested that EDs tend to target thyroid hormone production (Adams, Cyr, & Eales, 2000; Wang et al.,2013). In this context, atlerations in thyroid hormone levels was observed in several fish species upon exposure to EDs including PBDE(Tomy et al., 2004), perchlorate (Mukhi, Carr, Anderson, & Patinõ,2005), pesticides (Guo & Zhou, 2013; Holzer et al., 2017; Tu et al., 2016;Yadav & Singh, 1986; Yu, Chen, Liu, Gui, & Zhu, 2013; Zhang, Zhang,Du, & Zhao, 2017), alkylphenols (Shirdel & Kalbassi, 2016) and phthalates (Zhai et al., 2014). Furthermore, thyroid hormone production is also known to be affected adversely by heavy metals, their compounds and associated nanoparticles. Studies, in this light, depicted that BPA in combination with TiO2-NP led to reduction in Tlevels in zebra fish (Guo et al., 2019). Like metallic pollutants, chemical EDs alter thyroid morphology and also cause thyroid follicular developmental defects. DES, a synthetic estrogen and ioxynil, a herbicide, altered thyroid morphology and impaired thyrocyte development in zebra fish(Campinho & Power, 2013; Li, Canário, Power, & Campinho, 2019).Parental exposure to EDs for example, PBDEs, resulted in thyroid disruption and developmental neurotoxicity in zebra fish offsprings(Chen et al., 2012). Genistein, a phytohormone, caused disarray in thyroidal gene network by altering the expression of genes involved in thyroid gland development (Schiller et al., 2013). In zebra fish, pentachlorophenol exposure elevated expression of genes encoding thyroid stimulating hormone and its receptors (Guo & Zhou, 2013). Likewise,exposure of mono-(2-ethylhexyl) phthalate up-regulated several genes involved in thyroid hormone synthesis, development and hormone metabolism while

ttr

, involved in thyroid hormone binding, was down-regulated in zebra fish (Zhai et al.,2014). In line with this, zebrafish larvae exposed to acetochlor (a herbicide) altered thyroid hormone levels and modulated thyroid hormone synthesis related gene expression(Yang et al., 2016).

Taken together, EDs exerted deleterious effects on thyroid hormone production and thyroid function-associated genes that might cause developmental defects and reproductive dysfunctions. Some of the EDs might exert adverse effects on gonadal development through brainpituitary-thyroid axis (Sharma & Patiño, 2013). Furthermore, accumulation of EDs and their transfer to next generation might cause many notable complications leading to thyroidal toxicity and altered biochemical pathways.

6.Impact of EDs: Immune systems and diseases

The competence of EDs to interfere with immune function is relatively an unexplored area of toxicity. Several studies depicted that EDs exert toxicological effects on fish immune system by altering its functional components (Milla, Depiereux, & Kestemont, 2011; Pandey,Ghorai, & Rai, 2018; Xu, Yang, Qiu, Pan, & Wu, 2013; Yin et al., 2007).For example, malathion exposure resulted in reduced lymphocyte and erythrocyte cell number (Areechon & Plumb, 1990), decreased B-cell activity, lowered T-cell antibody production and weakened the host cell and humoral mediated immune response (Beaman et al., 1999). Similarly, another organophosphorous pesticide, diazoxon reduced lymphocyte proliferation, phagocytosis and spleen growth in the Nile tilapia,

O. niloticus

(Girón-Pérez, Zaitseva, Casas-Solís, & Santerre,2008). Chlorpyrifos is known to cause acute immunotoxicity by altering the serum immunoglobulin M levels in common carp and Nile tilapia(Girón-Pérez, Barcelos-Garcia, Vidal-Chavez, Romero-Bañuelos, &Robledo-Marenco, 2006; Li et al., 2013).Water bodies are greatly influenced by heavy metal contamination that exert adverse effects on immune function in addition to other physiological parameters as discussed in the earlier sections. Studies depicted toxic effects of As and AsOexposure in the walking cat fish,

C. batrachus

wherein the humoral immune system was affected making it susceptible to pathogen (Ghosh, Datta, Bhattacharya, & Mazumder,2007). Another study using the shark cat fish,

P. hypophthalmus

depicted that As alone and in combination with high temperature led to acute immunotoxicity (Kumar, Gupta, Bhushan, & Singh, 2019). Pb and Cd exposure led to loss of helper, memory cell activity and reduction in antibody production in the brown trout,

S. trutta

(O'Neill, 1981).Industrial chemicals such as BPA and akylphenols also exerted immunotoxic effects in the gold fish,

C. auratus

(Yin et al., 2007). A study in common carp demonstrated chronic toxicity and reduction in total number of leukocytes upon nonylphenol and EEexposure which was evident by histology (Schwaiger et al., 2000). Exposure of EE, in Nile tilapia altered the expression of

igfI

and

igfII

in addition to impairment of spleen growth, thereby, increasing disease susceptibility (Shved et al.,2009). Another report in the wolf fish,

Hoplias malabaricus

depicted several morphological defects such as increase in necrosis area and neutrophils cell number in kidney in addition to microautophagy of mitochondria and cell damage in liver upon exposure to tributyltin and Pb (Rabitto et al., 2005).

EDs are also known to target genes encoding immune components such as receptors and antibodies. Several studies depicted that pesticides, DDT derivatives, alkyphenols and sex steroids altered the expression of genes related to innate immune system and immune response in teleosts (Cuesta, Meseguer, & Esteban, 2008; Jin, Chen, Liu,& Fu, 2010; Massart et al., 2015; Willett, Zapata, Hopkins, & Steiner,1997).

In general, EDs affect various components of immune system and associated gene network making fish species susceptible to various diseases.

Since last three decades, documentations on ED-induced fish mortality, impaired fish growth, increase in fish diseases and infections in aquatic ecosystems worldwide are plenty and alarmingly high in number (Blazer et al., 2012; Ripley, Iwanowicz, Blazer, & Foran, 2008;Svensson et al., 1991; Vidal-Martínez et al., 2014). A report using an inorganic metal compound, sodium perchlorate, revealed induction of non-alcoholic fatty liver disease in the stickleback,

G. aculeatus

(Minicozzi, Von Hippel, Furin, & Buck, 2019). In another report, PBDE-47 exposure hampered pathogen resistance in fathead minnow larvae(Thornton, Path, Nystrom, Venables, & Jeffries, 2018). In several cases,the observed effects at population level, such as reproductive failure and outbreaks of diseases, can be linked to ED exposure. Other major effects on aquatic system are due to oil spills (Pérez-del Olmo, Raga, Kostadinova, & Fernández, 2007) or industrialization (Lavado, Thibaut, Raldúa,Martin, & Porte, 2004). In this regard, several studies demonstrated the toxic effects of chemicals such as PAHs, DDT on the Mayan cat fish,

Ariopsis assimilis

and the checkered puffer,

Spheroides testudineus

and their parasites (Khan, 1990; Pech et al. 2009; Vidal-Martínez & Poulin,2003). Furthermore, PAHs and heavy metals could induce carcinogenesis and hence, this property was exploited to test the nature of different class of EDs using fish models (Bailey, Selivonchick, & Hendricks, 1987;Hendricks, Meyers, Shelton, Castee & Bailey, 1985; Khudoley, 1984;Law, Hawkins, Overstreet, & Walker, 1994; Ostrander & Rotchell,2005).

Considering all these reports, ED-induced fish diseases might cause severe losses to aquaculture fish farms. Hence, more ecorelevant research works along with field-based assessments are required at population-, community- and food-web levels.

7.Impact of EDs on other endocrine glands or cellular systems

Environmental contaminants including EDs are often related to increased plasma cortisol concentrations and thereby activating of the brain-pituitary-interrenal axis. Exposure to PAHs, PCBs and heavy metals causes stress and alters adrenal (suprarenal) physiology in fish as demonstrated in the northern pike,

Esox lucius

, the yellow perch,

P. flavescens

and the brown trout,

S. trutta

(Girard, Brodeur, & Hontela,1998; Hontela, Dumont, Duclos, & Fortin, 1995; Hontela, Rasmussen,Audet, & Chevalier, 1992; Norris et al., 1999) leading to conditions that are either directly or indirectly involved in impaired energy metabolism in fish. Exposure of pesticides, like atrazine, induced severe hepatic perturbations in the grey mullet,

Liza ramada

(Biagianti-Risbourg &Bastide, 1995). Diethyl phthalate (DEP) exposure in

C. striatus

caused inhibition of cholesterol biosynthesis in liver followed by necrosis in hepatocytes and cytoplasmic vacuolization. Furthermore, DEP also altered kidney morphology and metabolic machinery of haematopoietic system in fish (George, Gokul, & Malini, 2017) concomitant with another study wherein

C. mrigala

was exposed to fenevalerate (Velmurugan, Selvanayagam, Cengiz, & Unlu, 2007). Ioxynil and DES disrupted cardiovascular development upon individual as well as combinatorial exposures in zebra fish,

D. rerio

(Li et al., 2019).Furthermore, studies on pharmaceuticals affecting the functions of thyroid and interrenal tissues are also plenty (Metcalfe, Miao, Hua,Letcher, & Servos, 2004). Pharmaceutical drugs, such as paracetamol and diclofenac, are known to induce hepatotoxicity and haematologic alterations in the silver bagre cat fish,

Rhamdia quelen

(Guiloski, et al.,2017a,b). Disruption of interrenal steroidogenesis was observed in rainbow trout upon exposure to paracetamol, ibuprofen and salicylic acid (Gravel & Vijayan, 2006; Hontela, 2006). A recent study in zebrafish indicated hepatotoxicity in female liver upon exposure to acetaminophen (Ayobahan et al., 2020).

Taken together, different classes of EDs exert adverse effects to various tissues and cellular systems across all the endocrine axes in fish which is why the need to assess their impacts is crucial though highly challenging. Apart from this, ED-induced effect on regulation of homeostasis and spawning migration in fish has been reviewed in the next section in order to assess ED toxicity at organismal level.

8.EDs affecting spawning migration and osmotic homeostasis

Fishes inhabit in various range of aquatic environments that differ in their osmotic properties. In addition to respiration, osmoregulation contributes to increased exposure of EDs to fish in varying environment(marine, freshwater or brackish) differentially. However, irrespective of the environment they live in, fishes are vulnerable to aquatic pollutants which penetrate their bodies by surrounding water uptake either due to drinking (as in case of marine fish) or osmosis via skin (as in freshwater fish). In addition to this, another significant route of ED exposure is via maternal egg transfer of contaminants (Kraak, Hewitt, Lister, McMaster,& Munkittrick, 2001) together with limited excretory activity in early-life-stage fish. Incidentally, levels of exposure by this route may be higher than that accumulated directly from water via osmoregulation affecting critical periods of development. Furthermore, a report in white sucker fish suggested alteration in pituitary function during spawning migration upon exposure to mixture of contaminants including EDs, in addition to this, reduced levels of T, 17α,20β-dihydroxy-4-pregnen-3-one (Kraak, Munkittrick, McMaster, Portt, & Chang, 1992;McMaster, Kraak, & Munkittrick, 1995) and 11-KT (Munkittrick, Portt,Kraak, Smith, & Rokosh, 1991) were documented in other related studies. Similar effect of altered T/Elevels was observed in the common carp,

C. carpio

(Folmar et al. 1996) and the European flounder (Janssen,Lambert, Vethaak, & Goos, 1997). Another report demonstrated impairment of hyposmoregulation in salmon after Eand nonylphenol exposure (Madsen, Mathiesen, & Korsgaard, 1997). However, the underlying mechanisms of these processes are poorly understood.

9.EDs alter transcriptome or genome profile

Apart from endocrine disruption, EDs often alter transcriptome and genome either directly or indirectly as a consequential effect. This can be validated only with those studies which had been done over generation with genetics or epigenetics based experimental approaches. Some of the reports showed genome-wide changes using carp and zebra fish as animal models after exposing to various EDs. In carp, cDNA microarray was developed to identify minimal subset of genes in order to distinguish ED that interferes normal functions like development, physiology, and reproduction (Moens, Ven

,

Remortel

,

Favero

&

Coen

,

2006). Application of massive parallel sequencing in zebra fish to study EDs targeting early organ development and reproductive dysfunction has been well reviewed (Baker & Hardiman, 2014). In line with this, adverse effects of EDs can be studied in zebra fish through microarrays or whole-genome and next-generation sequencing technologies for quantitative analysis(Caballero-Gallardo, Olivero-Verbel, & Freeman, 2016; Mehinto, Martyniuk, Spade, & Denslow, 2012). These approaches provide global gene expression complemented with transcriptome data as a tool to predict toxicogenomic outcomes of particular EDs in a pool of compounds(Caballero-Gallardo et al., 2016; Mehinto et al., 2012).

10.Modern day diagnostic tools: Use of biosensors to monitor EDs at organismic level and water bodies

Considering a variety of endocrine disruptive effects on various physiological systems, it is important to develop tools for contaminant detection even at low concentration. Screening of EDs is an immense task and a mandatory requirement in order to assess the level of damage of a given environmental site including aquatic systems (Hecker &Hollert, 2011). It is also obligatory to develop new restrictions as well as specific rules and regulations for their usage for various applications related to animal or human welfare. Usage of piezoelectric biosensors for detecting organophosphate and carbamate pesticidal residues in the aquatic environment can be implemented to regulate applications(Marrazza, 2014). Development of CYP enzyme based biosensor in pharma industries to test any foreign chemical metabolism in the body has been well reviewed (Schneider & Clark, 2013). In fact, a transgenic line of zebra fish was generated by making use of CYP-green fluorescence protein construct, driven by

cyp1a1

promoter to detect PCBs (Hung et al., 2012). Next level of approach is to record the damage so that serious measures can be taken to ban or restrict the use of such compounds completely as the prevalence of EDs in environment poses serious health hazards (Schug et al., 2016). Ecological studies from various aquatic site validated the need for ED analysis due to the prevalence of xenobiotic pollutants which often exert steroidal mimic action(Scholz & Klüver, 2009) or regulating a gene pertaining to maturation-inducing hormone synthesizing enzyme (Sreenivasulu et al.,2012). In addition to xenobiotics, biological effects of many classes of EDs have been recorded in the Arctic fish and other wildlife species(Dietz et al., 2019). In view of this, it is essential to develop whole organism model with biomarkers to test EDs. The scientific data available,to the best of our knowledge, on effective biomarkers which have been used extensively to screen EDs in fish is provided in Table 2. Incidentally, zebra fish embryo-based multiplex method was developed using gene-based biomarkers to detect estrogenic, androgenic, and thyroid hormone mimics in aquatic ecosystem (Jarque et al., 2018; Jarque,Ibarra, Rubio-Brotons, García-Fernández, & Terriente, 2019). In spite of using whole organism, it is also worthwhile to use both single (Ren et al.,2017) and complex molecules biosensors to detect specific analyte(Jarque et al., 2019). Such protocols can also be modified and extended to aquatic system as well as nutrition industries processing aquatic foods to check the quality of stored whole organism or meat or food originated from aquatic vertebrates including fishes. Meat and fish quality diagnosis can be efficiently done to test the contaminant levels of dioxin,dioxin-like PCBs (Chobtang et al., 2011), hypoxanthine (Mulchandani,Luong, & Male, 1989; Shen, Yang, & Peng, 1996; Watanabe, Ando,Karube, Matsuoka, & Suzuki, 1983; Watanabe, Toyama, Karube, Matsuoka & Suzuki, 1984), octopine and amines (Shin, Yamanaka, Endo, &Watanabe, 1998), histidine, putrescine, cadaverine (Male, Bouvrette,Luong, & Gibbs, 1996), inosine (Watanabe, Toyama, Karube, Matsuoka,& Suzuki, 1984), polyamines (Chemnitius et al. 1992) and sulfites(Mulchandani, Groom, & Luong, 1991). Such an approach not only certifies the quality but also guarantees those as a food source for consumers. In spite of these advances, it is essential develop rapid diagnostic sensors for quick screening of molecules (Vigneshvar & Senthilkumaran,2018; Vigneshvar, Sudhakumari, Senthilkumaran, & Prakash, 2016) in order to provide an opportunity to implement those for ED contaminants testing. With the invention of high-end instruments, bio-accumulation of various protein or metal NPs can be efficiently analyzed. Most of these inventions related to biosensors have lot of scope and hence, more simplified approach with low cost is the need of hour diagnostic methods (Vigneshvar et al., 2016). These kinds of tools serve in a broader perspective and not just restricted to aquatic system. In fact, use of biosensors with novel modern high-tech gadgets to test the sample quality as per the requirement ranging from nutrition to therapeutics across systems has been well reviewed by Vigneshvar and Senthilkumaran (2018). One of the innovations in this field is complex multiple molecules sensing which has greater applications to test multiple EDs from tissue or organism from aquatic system without any prior separation methodology applications (Jarque et al., 2019). Most of those are highly qualitative rather than quantitative which makes it highly suitable for quick applications for testing purposes. At the same time,quantitative measurements are also possible wherein specific procedural change in policy of usage or restrictions can be designed with the knowledge on the extent of damage. Considering these, modern day biosensors are extremely useful for environmental monitoring with remote sensing and internet of things. Developing of such sensors is highly desirable for ED screening at organismal or tissue level, or even the inhabiting aquatic system.

Table 2Various biomarkers for screening and assessment of EDs in fish.

11.Future perspectives

This review attempted to highlight results of the studies pertaining to ED-induced toxicity and their adverse impact of endocrine axes together with data on effective biomarkers to monitor EDs-mediated risks in teleosts. EDs causing disarray of endocrine axes and tissues by hampering different physiological parameters including brain, thyroid,immune, interrenal and gonadal systems in various fish species are depicted in Table 3. Similarly, deleterious effects exerted by metals including nanoparticles in teleosts are depicted in Table 4. These studies have come in a long way to establish specific biomarkers with

in vitro

culture systems to ultimately introduce

in vivo

analysis. In line with this,a few reports indicated the toxic effects of metal nano-particles toxicity in fish gonads

in vivo

which corroborated with studies using primary cells of fish testis or rat testicular cell lines

in vitro

(Deepa et al., 2019;Murugananthkumar et al., 2016). In this context, development of improved and ecorelevant methodologies for ED and identification of more sensitive and specific biomarkers to screen EDs, would provide more information on endocrine-mediated effects in individuals as well as in populations on a global basis. Progress in understanding the effect of EDs used in teleosts model will contribute further on this broader perspective.

Table 3Impact of EDs on the function of fishendocrine system.

Table 3(continued)

Table 3(continued)

Table 3(continued)

Table 3(continued)

Table 3(continued)

Table 4Heavy metals and metal-NPs exerting deleterious impacts on various endocrine axes in fish.

Table 4 (continued)

12.Research trends and innovative applications

The current research trend comprises the use of several biochemical assays for a few biomarkers to screen EDs wherein, measuring VTG in fish blood serum through ELISA assay is an example (Goksøyr, 2006;Kim et al., 2007). Several chromatography techniques have been standardized for the determination of various EDs (Liu et al., 2006; Mottaleb et al., 2009; Nakata, Kannan, Jones, & Giesy, 2005; Petrović, Hernando,Díaz-Cruz, & Barceló, 2005; Vieno, Tuhkanen, & Kronberg, 2006; Wen,Zhou, Xu, Jin, & Feng, 2006), however, all of which involves usage of sophisticated instruments and are expensive as well as time consuming.Furthermore, these techniques cannot be used for multi-analyte detection and may not be suitable for field studies and

in situ

analysis.Alternatively, various assays were used using fish species to screen EDs(Ayobahan et al., 2020; Brain et al., 2018; Braunbeck et al., 2005; Dang et al., 2011; Lammer et al., 2009) and assess their potential interactions with endocrine system using various fish models, for example, the estrogenic property of butylated hydroxyanisole (an antioxidant) was evaluated using a freshwater teleost,

A

.

testudineus

(Paul, Binitha, &Sunny, 2018). Use of biosensors can be an alternative tool for expensive analytical methods. In this regard, a few ED sensors and biosensors have been developed over the last few years which depict the enormous possibilities that sensor and biosensor technologies hold for detecting and quantifying micro-pollutants in the environment and, some of the EDs such as parathion, dioxins, nonylphenol and BPA (Sacks, Eshkenazi,Neufeld, Dosoretz, & Rishpon, 2000; Kurosawa, Aizawa, & Park, 2005;Evtugyn, Eremin, Shaljamova, Ismagilova, & Budnikov, 2006; Rodriguez-Mozaz, de Alda, & Barceló, 2006; Sánchez-Acevedo, Riu, & Rius,2009). Au-NPs based progesterone-specific aptasensor being one of the examples (Du et al., 2016). Furthermore, these can be potentially used for other compounds in various species with necessary modifications. In future, a standalone biosensor may allow on-site techniques to evaluate EDs wherein graphite electrodes may be used as a transducer and carrier(Labib, Sargent, & Kelley, 2016; Rebollar-Pérez, Campos-Terán,Ornelas-Soto, Méndez-Albores, & Torres, 2016; Salimi, Hallaj, Soltanian, & Mamkhezri, 2007).

Overall, existing knowledge about EDs and their underlying mechanism of actions can be expanded by strengthening international collaboration and network research between the scientific communities of different developing nations. This will certainly pave way to classify pertinent, less pertinent and non-pertinent effects of EDs using various animal models, more importantly teleosts.

Declaration of competing interest

The authors declare that there is no conflicts of interest.

Acknowledgements

SK and PS are grateful to the University of Hyderabad for Non-NET fellowships. NA is thankful to the Junior Research Fellowship support by a grant-in-aid (BT/PR15748/AAQ/3/803/2016) from the Department of Biotechnology (DBT), India awarded to BS. BS is also a recipient of TATA innovation fellowship (BT/HRD/35/01/02/2013) from DBT,India (during the years: 2014—2019) which is acknowledged.