Pyrethroid Pesticides as Endocrine Disruptors: Molecular Mechanisms

Critical Review pubs.acs.org/est

Pyrethroid Pesticides as Endocrine Disruptors: Molecular Mechanisms in Vertebrates with a Focus on Fishes Susanne M. Brander,*,† Molly K. Gabler,† Nicholas L. Fowler,† Richard E. Connon,§ and Daniel Schlenk‡ †

Biology & Marine Biology, University of North Carolina at Wilmington, 601 South College Road, Wilmington, North Carolina, 28403, United States ‡ Environmental Sciences, College of Natural and Agricultural Sciences, University of California at Riverside, 900 University Avenue, Riverside, California 92521, United States § Anatomy, Physiology & Cell Biology, School of Veterinary Medicine, University of California at Davis, One Shields Avenue, Davis, California 95616, United States S Supporting Information *

ABSTRACT: Pyrethroids are now the fourth most used group of insecticides worldwide. Employed in agriculture and in urban areas, they are detected in waterways at concentrations that are lethally and sublethally toxic to aquatic organisms. Highly lipophilic, pyrethroids accumulate in sediments and bioaccumulate in fishes. Additionally, these compounds are demonstrated to act as endocrine disrupting compounds (or EDCs) in mammals and fishes, and therefore interfere with endocrine signaling by blocking, mimicking, or synergizing endogenous hormones through direct receptor interactions, and indirectly via upstream signaling pathways. Pyrethroid metabolites have greater endocrine activity than their parent structures, and this activity is dependent on the enantiomer present, as some pyrethroids are chiral. Many EDCs studied thus far in fish have known estrogenic or antiestrogenic effects, and as such cause the inappropriate or altered expression of genes or proteins (i.e., Vtg−vitellogenin, Chg−choriogenin), often leading to physiological or reproductive effects. Additionally, these compounds can also interfere with other endocrine pathways and immune response. This review highlights studies that focus on the mechanisms of pyrethroid biotransformation and endocrine toxicity to fishes across a broad range of different pyrethroid types, and integrates literature on the in vitro and mammalian responses that inform these mechanisms.

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INTRODUCTION The use of pyrethroid pesticides has increased by an order of magnitude over the past 20 years, as organophosphate pesticides that are acutely neurotoxic to mammals are phased out.1 For example, usage data across 38 countries indicated that less than 100 t (MT) of pyrethroids were applied in the year 2000, but by 2009 this exceeded 1500 MT and a continuing upward trend was evident.2 They are now the fourth major group of insecticides in use worldwide.3 Synthetic pyrethroids are based on pyrethrins, compounds naturally synthesized by chrysanthemum flowers, but are chemically modified to increase persistence and toxicity.4 As such they are lethally toxic to many organisms at environmental concentrations (ng−μg/L).1,5−8 Pyrethroids are also extremely toxic to fishes, 1000 times more so than in mammals.9−11 Used in agriculture and in urban areas, they are found in mosquito control sprays, lice shampoo, and residential yard formulations. Pyrethroids have also become ubiquitous in treated wastewater effluent.12−18 Due to being highly lipophilic, pyrethroids are demonstrated to bioaccumulate in both fishes (12−4938 ng/ L) and marine mammals (mean averages of 7.04 ng/g adults, 68.4 ng/g calves). A recent study conducted in Spain found pyrethroids in 100% of tissue samples collected from riverine fish.19,20 © 2016 American Chemical Society

The intended mechanism of pyrethroids - paralysis via sodium channel overactivation, causes mortality in fish at high concentrations (low μg/L) and swimming abnormalities at lower concentrations (ng/L).21−23 However, recent results from in vitro and in vivo assays reveal that some pyrethroids can act as estrogens (bind to estrogen receptors and/or stimulate estrogen synthesis or signaling pathways), antiestrogens (block estrogen receptors and/or inhibit estrogen synthesis or signaling pathways) and/or antiandrogens (block androgen receptor and/or inhibit androgen synthesis or signaling pathways) via multiple mechanisms.24−29 For example, aqueous exposure to the pyrethroid bifenthrin alters an estrogendependent (choriogenin) protein in juvenile fish (Menidia beryllina) at concentrations as low as 1 ng/L.28,29 Furthermore, pyrethroid metabolites are reported to have even greater estrogenic activity than parent compounds.24,27,30 Pyrethroids are now confirmed to have endocrine disrupting properties, and join the ranks with many other compounds known to mimic or interfere with the endocrine system. Received: Revised: Accepted: Published: 8977

May 6, 2016 July 21, 2016 July 27, 2016 July 27, 2016 DOI: 10.1021/acs.est.6b02253 Environ. Sci. Technol. 2016, 50, 8977−8992

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Environmental Science & Technology

differences in the rate of reduction dependent upon which enantiomer was examined (ie. cis versus trans permethrin). The rate of transformation has been shown to be soil specific, and to also differ between terrestrial and aquatic environments. Furthermore, since many bacteria inhabit the water column instead of sediments, these bacteria may not degrade pyrethroids extensively since the chemicals quickly adsorb to sediment.55 Degradation rates is also affected by the concentration of dissolved organic matter (DOM). Yang et al.56 showed that pesticide hydrophobicity does not solely describe pesticide affinity for DOM nor bioavailability. They reported that the effects of DOM on the toxicity of cyfluthrin were more pronounced than seen with permethrin. Thus, bioavailability is differentially affected by DOM depending on pyrethroid type. Degradation rate in sediments varies widely depending on sediment texture, organic matter content, bacterial species present, whether conditions are aerobic or anaerobic, and other abiotic factors (i.e., temperature). The first order half-life of pyrethroid mixtures can range from 2.9 to >200 days under aerobic conditions (median =18 days) and from 20 to 200 days under anaerobic conditions (median = 70 days), with bifenthrin having a longer half-life than other pyrethroids tested.57 Chen et al.58 reported that the bacterium Brevibacterium aureum was able to use cyfluthrin as its sole carbon source, where normally pesticide-degrading bacteria need a supplemental carbon source (sucrose, glucose) before degrading the pesticide.59,60 With concentrations at 50 mg/L the bacteria were able to degrade 65% in a 12−48 h period and up to 87% after 120 h. While different strains of bacteria can degrade numerous pesticide types, many species, such as B. aureum do so in a unique manner, hence, the wide range of reported half-lives differs among abiotic and biotic, habitat-specific processes. Biotransformation (or lack thereof) has been suggested as one possible mechanism explaining the relative sensitivity of fish to acute pyrethroid toxicity.61 Pyrethroids undergo hydrolytic ester cleavage primarily in the plasma of mammals, but in fish rates of oxidative biotransformation in liver microsomes are greater than esterase activities.9,61−63 Some pyrethroids are metabolized predominantly via NADPHcatalyzed oxidative processes, such as bifenthrin, S-bioallethrin, 1R and 1S-cis-permethrin. Others undergo ester cleavage exclusively, including bioresmethrin and cypermethrin. Betacyfluthrin and cis and trans-permethrin can be converted to metabolites via both oxidative and hydrolytic pathways. Generally, the trans isomers of Type I pyrethroids (i.e., permethrin), which are esters of primary or secondary alcohols, are metabolized more rapidly (via ester cleavage) than cis isomers, which are slowly metabolized via oxidative processes via cytochrome P450 enzymes (CYPs, i.e. CYP3A4 in humans). Type II pyrethroids (i.e., cypermethrin), which consist of an ester of a secondary alcohol with a cyano group at one of the chiral centers, can undergo oxidative reactions as well as ester cleavage, but have a lower rate of biotransformation, and thus are generally more acutely toxic than Type I compounds.46 In a study on the stereoselective metabolism and conversion of four different permethrin enantiomers using human liver microsomes, the formation of the permethrin metabolite 3-(4′hydroxyphenoxy)-benzyl alcohol) (3,4 PBOH), was stereoselective, with the greatest catalytic efficiency observed after incubation of CYPs with 1R trans permethrin. Notably, this study demonstrated that not all catalytic events involving pyrethroids are stereoselective. Formation of the metabolite 3,4

Endocrine-disrupting chemicals (EDCs) act by blocking, mimicking, or synergizing endogenous hormones.31−34 Common sources of EDCs are pharmaceuticals, surfactants in cleaning detergents, and pesticides. Many of the EDCs studied thus far on fish have known estrogenic or antiestrogenic effects and may alter sex ratios or cause the inappropriate or altered production of proteins (Vtg−vitellogenin: egg yolk precursor, Chg−choriogenin: egg envelope precursor) in males and/or females.35,36 Additionally, these compounds also interfere with immune response, growth and development, as well as increasing osmotic stress, protein degradation, and altering numerous metabolic processes.1,34 Furthermore, it has been established that endocrine disrupting effects at the molecular and/or organismal level can result in reduced population size.36,37 As such, the increased usage of pyrethroid pesticides represents a threat to the health and persistence of fish populations globally. This review highlights studies that specifically focus on the many mechanisms of endocrine toxicity of pyrethroids to fishes, with an emphasis on commonly used pyrethroids, and integrates literature on in vitro and mammalian responses that inform these mechanisms.

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BIOTRANSFORMATION Upon reaching ecosystems, pyrethroids are metabolized by bacteria, fish, and other organisms3,38−40 and the metabolites appear to be both estrogenic and antiandrogenic.24,27,28,30,41,42 Therefore, fish are simultaneously exposed to parent pyrethroid compounds, to metabolites generated internally, and to metabolites produced in the environment either via abiotic factors (hydrolysis, photolysis) or by bacteria. Synthetic pyrethroids may contain one to three different chiral centers, comprising between two to eight enantiomers.43−47 Some pyrethroids are formulated to consist of just a single stereoisomer (i.e., deltamethrin), while others contain a blend of cis and trans varieties (i.e., permethrin).48 Each cis and trans oriented molecule also has two enantiomers; for example, 1R cis and 1S cis bifenthrin (Supporting Information (SI) Figure S1). These various isomers can have different physicochemical properties and may have different environmental fates due to differences in toxicity and metabolic processing.49 For example, bacteria showed preferential degradation of trans diastereomer of permethrin over the cis diastereomer, and 1S cis bifenthrin or cis permethrin were selectively degraded over the 1R cis enantiomer (which was shown to be the most ecotoxic).50 As the trans diasteromers are selectively degraded because the reaction Km is lower in comparison to that of cis forms, this leaves the more biologically active cis forms in higher relative concentrations.46,51 As such, toxicity may be disproportional based upon the pesticide concentration.52 Pyrethroids are generally decomposed via cleavage of esters bonds (often hydrolytically) and oxidation of acid and alcohol functional groups.39,43,46,53 They partition out of the water column quickly due to their high octanol−water partition coefficient (Kow) and therefore have relatively high affinity for organic matrices within sediments.49 This makes them subject to metabolic breakdown by mixtures of bacteria inhabiting sediments of fresh, estuarine, and marine waters. Mixed cultures of both neustonic and epiphytic bacteria degrade pesticides more effectively than pure cultures of either type of bacteria.54 Proteobacteria showed the highest number of degrading bacterial isolates for bifenthrin and permethrin.55 S. acidaminiphila reduced the half-lives of pyrethroids, with 8978

DOI: 10.1021/acs.est.6b02253 Environ. Sci. Technol. 2016, 50, 8977−8992

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PBOH, which is primarily generated via carboxylesterase cleavage, did not differ significantly between enantiomers.51 Like permethrin (Type I), cypermethrin’s (Type II) enantiomers can undergo esterase-catalyzed biotransformation. The resulting cyclopropyl acids in both can be further conjugated with glucuronide or glycine (organism dependent) enhancing excretion. The cyanohydrin and aromatic alcohols can also be oxidized and hydroxylated in different positions. The hydroxylated compounds may be conjugated with glucuronide, glycine, taurine, or sulfate groups. This can also occur at carbonyl functionalities. Permethrin metabolites (Figure 1; biotransformations), and possibly the metabolites of other pyrethroids, have also been

Critical Review

BIFENTHRIN

Effects in Fishes. Bifenthrin (Talstar) is used residentially as a granular lawn application, in termite and fire ant liquids, and commercially for similar types of pest control.17,65 Its desired mechanism causes a constant flux of sodium to rapidly trigger nerve impulses, resulting in muscle spasms and neurological disturbances.66 It is now one of the most commonly detected pyrethroids in the U.S.,13 and is also one of the most toxic, with sublethal effects such as altered swimming performance or gene/protein expression in fishes commonly observed at ng/L to low μg/L concentrations,21,23,28,67 and mortality or severe developmental effects observed at higher concentrations.68 Studies that have evaluated gene or protein expression in response to bifenthrin exposure have observed alterations in endocrine-related transcripts or proteins.22,28,29,69,70 For example, Beggel et al.22 observed upregulation of vitellogenin (Vtg) transcripts and down-regulation of CYP3A, a P450 enzyme involved in steroid and pyrethroid metabolism in fathead minnow (P. promelas) after treatment at low μg/L concentrations. Additionally, even low ng/L concentrations of bifenthrin can induce altered expression of estrogen-mediated proteins, as seen in Brander et al.28 Another similarity across studies is the biphasic nature of response to a range of bifenthrin concentrations, with low concentrations inducing an estrogenic response and an indication of negative feedback on estrogen signaling occurring at higher concentrations.22,28,29 For example, Brander et al.28 observed after a 14 day aqueous exposure to three concentrations of bifenthrin, that the lowest concentration used (1 ng/L) induced the highest levels of choriogenin (estrogen-induced egg coat protein) in the whole body homogenate of juvenile M. beryllina. Notably, choriogenin was significantly higher in the homogenate of fish exposed to 1 ng/L bifenthrin than it was in fish exposed to a positive control of 1 ng/L ethinylestradiol (EE2). At higher exposure concentrations of 10 and 100 ng/L bifenthrin, choriogenin levels were still higher than the negative control but were not significantly different from 1 ng/L EE2. The best-fit dose−response curve for bifenthrin was biphasic, or unimodal, indicating that bifenthrin may be acting via different mechanisms depending upon the concentration (Figure 2). Beggel et al.22 observed that gene expression in response to bifenthrin was biphasic at lower concentrations (75−150 ng/L) but linear starting at higher concentrations (>150 ng/L) in P. promelas, and Brander and colleagues29 also saw biphasic responses in juvenile M. beryllina. At 0.5 ng/L bifenthrin generally inhibited the expression of estrogen-responsive genes, but expression increased at higher concentrations (5, 50 ng/L). Such responses have been seen in many EDC studies, and are similar to responses generated by administration of hormones.33 Administration of EE2, a synthetic estrogen, can cause higher expression of estrogen receptor (ESR) genes in fish at low concentrations and/or short exposure durations and lower expression of the same ESRs at higher concentrations and/or longer exposure durations.71 Such biphasic responses can indicate the downregulation of receptors at higher concentrations, as the endocrine system uses feedback to regulate response. Alternatively, responses may be endocrine-mediated at lower concentrations but represent a more general toxic response at higher concentrations, since endogenous hormones tend to activate responses at pg or ng/L concentrations but saturate target receptors or enzymes at higher levels. As such, when responses are biphasic it is not possible to predict

Figure 1. Proposed biotransformation pathway of permethrin in fish ((i) ester hydrolysis; (ii) CYP-catalyzed hydroxylation; (iii) glucuronide/sulfate conjugation); * indicates a chiral center. Reprinted with permission from Nillos, M. G.; Chajkowski, S.; Rimoldi, J. M.; Gan, J.; Lavado, R.; Schlenk, D., Stereoselective Biotransformation of Permethrin to Estrogenic Metabolites in Fish. Chem. Res. Toxicol. 2010, 23, (10), 1568−1575. Copyright 2010 American Chemical Society.

shown to affect membrane fluidity and polarity via lipid and protein oxidation in rats. Some metabolites (i.e., 3 phenoxybenzyl alcohol) induced lipid peroxidation, while others were stronger inducers of protein damage (i.e., 3 phenoxybenzoic acid). These metabolites reduced membrane fluidity potentially impacting sodium and calcium channels limiting ionic transport across the cell membrane.64 In fishes and mammals, it has been noted that some metabolites of pyrethroids (i.e., permethrin, cypermethrin, bifenthrin) have molecular structures that are more likely to interact with estrogen receptors than would the parent compound.24,27,30,51 While the parent pyrethroid compounds show little resemblance to estradiol and thus have limited estrogenicity, their metabolites have much greater structural similarity to estradiol.3 These findings highlight the importance of biotransformation and stereoselectivity in the estrogenic activity of permethrin and bifenthrin. Stereoisomers with the greatest estrogenic activity are also the most persistent compounds in the environment, making them more likely to be transported to an aquatic ecosystem.27 The similarities between the metabolic pathways for bifenthrin and permethrin indicate that other pyrethroids likely follow similar routes of biotransformation. 8979

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of fishes to bifenthrin in in vivo studies may be due to the metabolism of bifenthrin to the 4-hydroxy metabolite (most common metabolite in Blue gill sunfish − Suprenant 1986), in M. beryllina.30 Studies show that when bifenthrin is used at environmentally relevant concentrations in in vitro bioassays, that there is little to no estrogenic activity.28,30,72 However, in combination with alkylphenols and alkylphenol polyethoxylates (surfactants and known EDCs) and the herbicide diuron, in vivo exposed male fish had significantly increased production of the egg protein vitellogenin.74 Additionally, Brander and colleagues28 observed that production of another egg protein (choriogenin) was elevated in M. beryllina after exposure for 14 days to 1 ng/L bifenthrin, but that the same concentration antagonized the estrogen receptor in vitro. The biphasic or nonmonotonic response seen in vivo and in vitro,28 commonly observed with other EDCs,33 may also be due to the activity of metabolites, which are more similar in structure to endogenous estrogen than the parent bifenthrin. Since fish may have a limited ability to continue metabolizing bifenthrin as they are exposed to increasing concentrations (due to enzyme saturation), this may manifest as a decreased estrogenic response as concentration increases.28−30 In vitro, results have been contradictory, in that some cell lines indicate estrogen antagonism,28 whereas others indicate estrogen agonism.70 Results seem to be dependent on the cell line used and whether multiple isoforms of the estrogen receptor are present, which is something that should be addressed in future experiments. In vivo, bifenthrin effects suggest weak estrogen agonism, with higher concentrations or extended exposures resulting in negative feedback. While metabolites of bifenthrin may play a role in ER activation, results across species are conflicting. In a study using two strains of rainbow/steelhead trout (Onchorynchus mykiss), NADPH-catalyzed microsomal clearance of bifenthrin in liver did not correlate with estrogenic activity.73 However, individual metabolites were not identified. Observed activity in vivo but not in vitro, or in mixture but not alone, may point to other molecular targets rather than direct interaction of the parent compound with the ER. Possible mechanisms are 1. Enhanced absorption of the pesticides due to surfactant presence, 2. Induction of biotransformation enzymes that convert the pesticide into a more potent metabolite, 3. Inhibition of estradiol clearance through inhibition of steroid catabolism, 4. Upstream enhancement of steroid biosynthesis.74 With regard to steroid biosynthesis, studies in salmonids indicate that bifenthrin impacts gonadotropin release and subsequent E2 biosynthesis through impairment of dopaminergic control, possibly dependent upon exposure duration75 dopamine receptor 2A (DR2A) expression was associated with an increase in plasma E2 following exposure in juvenile rainbow trout treated with 1.5 μg/L for 96 h and 2 weeks, and a significant increase in the relative expression of vitellogenin mRNA at 2 weeks was observed.75 In addition, an increase in tyrosine hydroxylase transcript levels in brains of bifenthrin treated fish was also observed at 96 h, indicative of dopamine production. Temporal differences were observed with gonadotropin-releasing hormone (GnRH) 2 which was increased at 96 h but decreased after 2 weeks exposure, consistent with feedback loop activation of E2. Similar results were observed in zebrafish and mammals with other pyrethroids.76,77 These data contrast the studies of Liu and colleagues78 which showed that bifenthrin inhibited a number of genes involved in steroidogenesis (rat granulosa cells), such as StAR (steroidogenic acute

Figure 2. Choriogenin concentration−response of juvenile Menidia beryllina to ethinylestradiol, bifenthrin, and permethrin. Most parsimonious curve fit to concentration−response data (sigmoidal or unimodal) using nonlinear regression. Each data point represents the combined whole body homogenate of 6−10 juvenile M. beryllina exposed for 14 days to the corresponding concentration. The horizontal dotted line across the top panel represents the maximum choriogenin response in pyrethroid-exposed fish. Reprinted with permission from Brander, S. M.; He, G.; Smalling, K. L.; Denison, M. S.; Cherr, G. N., The in vivo estrogenic and in vitro antiestrogenic activity of permethrin and bifenthrin. Environ. Toxicol. Chem. 2012, 31, (2), 2848−2855. Copyright 2012 John Wiley and Sons.

responses at lower concentrations from those observed at high concentrations.33,34 Bifenthrin exposure also alters apical end points in fishes, such as gonadosomatic index and ovarian follicle size, concurrent with changes in plasma estradiol levels, and said effects can differ depending on the salinity at which fish are exposed.72,73 As mentioned earlier in the discussion of biotransformation, bifenthrin’s 4-hydroxy metabolite is established to have greater estrogenic activity than bifenthrin, and the estrogenic responses 8980

DOI: 10.1021/acs.est.6b02253 Environ. Sci. Technol. 2016, 50, 8977−8992

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formulations, mRNA levels of StAR, and P450scc were reduced in the testes of male offspring.82 Studies in mammals have expanded beyond estrogenic and antiestrogenic or antiandrogenic effects. A recent study demonstrated that exposure to bifenthrin induced progesterone and human chorionic gonadotropin secretion in human trophoblast cells, caused upregulation of GnRH and its receptor, and also interfered with steroidogenesis. Effects were stereoselective, whereby 1S bifenthrin caused greater disruption than 1R bifenthrin, partly due to ER binding.80 Bifenthrin also affected thyroid hormone signaling, as the Talstar formulation of bifenthrin was found to suppress serum concentrations of T3 and T4 in rats.83 As discussed later in this review, studies in fishes should be expanded to address bifenthrin’s effects on additional aspects of the endocrine system, such as effects on the thyroid and adrenals, as has been done in mammals.

regulatory protein), CYP19a1 (cytochrome P450 aromatase), P450scc (cytochrome P450 side cholesterol chain cleavage enzyme), and PR (progesterone receptor). Low concentrations of bifenthrin (0.5 ng/L) also significantly modulated the expression of a suite of estrogen-influenced genes in M. beryllina, including GPR30 (membrane estrogen receptor), ESR1 (ERα), TRa (thyroid receptor alpha), and MAPK14 (mitogen activated protein kinase). M. beryllina exposed to this same concentration had significantly reduced output of fertilized eggs.29 These finding indicate that bifenthrin and its metabolites may influence estrogen signaling via a myriad of pathways in addition to operating as weak ER antagonists or agonists. Because the parent compound and metabolite are likely exerting their effects simultaneously and may be competing with each other, if the parent is acting as a weak agonist and the metabolite(s) as weak antagonist(s), responses are often difficult to interpret. In addition, the accumulation and likely elimination of bifenthrin is stereoseletive (see above) as the 1Renantiomer was higher in livers of zebrafish than the corresponding 1S enantiomer which was the most estrogenic.70 This and variation in sensitivity between species may account for the results of studies in which an estrogenic response to bifenthrin alone was not observed (e.g., ref 74). Effects in Vitro. Bifenthrin has both estrogenic and antiestrogenic effects in cell lines. Both Zhao et al.79 and Wang et al.70 showed that bifenthrin increases the proliferation of MCF-7 cells in the classical E-SCREEN assay, and that it caused increased expression of the estrogen-responsive transcription factor ps2. Interestingly, while the estrogenic response to bifenthrin can be blocked using ICI, an estrogen antagonist,70 bifenthrin also acts as an antiestrogen in the ER CALUX cell line.28 Whether bifenthrin causes an estrogenic or antiestrogenic response appears to depend on which enantiomer is employed. The 1S cis bifenthrin enantiomer has greater estrogenicity than 1R cis79 and stimulates greater expression of estrogen-responsive genes such as steroidogenic enzymes (i.e., CYP17, CYP19), other steroid receptors (i.e., PR), or hypothalamic genes (i.e., GnRH, GnRHR). However, a racemic mixture of bifenthrin increases the expression of the same genes to less of an extent.80 The same enantiomer (1S cis) was also shown to disrupt luteinizing hormone (LH)-induced ovulatory genes in rat ovarian granulosa cells, resulting in a reduction of prostaglandin E2, which triggers the ovulatory event.80 Notably, docking studies performed by Zhao et al.80 demonstrate that 1S cis bifenthrin is able to form a hydrogen bond with residue Thr347 in the ERα ligand binding domain (LBD), whereas 1R cis bifenthrin cannot. Binding also differed at the ERβ LBD, with 1S cis bifenthrin forming three hydrogen bonds, while 1R cis bifenthrin only formed two. Binding scores for both receptors were higher for 1S cis bifenthrin.80 As a racemic mixture was employed in the Brander et al.28 study in which bifenthrin was determined to be antiestrogenic but not estrogenic in vitro, it may be that less estrogenic or antiestrogenic enantiomers were competing for binding with 1S cis bifenthrin. Furthermore, the ER CALUX contains mRNA for both ER α and β,81 whereas many other cell lines only contain ER α. Effects in Mammals. To date, responses to bifenthrin in mammals are mostly in line with those observed in fishes. In vivo bifenthrin blocked the expression of prostaglandins, which play a key role in ovulation.78 When pregnant rats were exposed to the two cis isomers commonly found in commercial

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PERMETHRIN Effects in Fishes. While permethrin is less acutely toxic than bifenthrin, it also exerts sublethal effects on swimming performance and is toxic both alone and in combination with other pyrethroids via ion channel disruption and induction of hypoxia.84−86 Cis permethrin is more acutely toxic than trans permethrin by an order of magnitude.9,43 Cis permethrin preferentically bioaccumulates because it is metabolized more slowly.77,87−89 Permethrin appears to act as an estrogen in vivo, but at higher concentrations than bifenthrin. Similar to other pyrethroids, estrogenic activity varies depending on the enantiomer being evaluated. For example, 1S cis permethrin showed greater estrogenic activity than 1R cis permethrin in vivo (male Japanese medaka, vitellogenin) and in vitro.69,89 Brander et al.28 showed a similar elevation in choriogenin protein production in M. beryllina with 1 μg/L permethrin. Similarly, Nillos et al.27 showed an increase in vitellogenin mRNA at 10 μg/L permethrin in Japanese medaka (Oryzias latipes). The ability of permethrin to induce egg yolk protein or mRNA expression (vitellogenin) in male or juvenile fish, indirectly indicating ER binding and/or activation of the estrogen response element, has been shown in zebrafish (Danio rerio) as well.89,90 Permethrin appears to be metabolically transformed into a more estrogenic compound. This was confirmed by Nillos et al.,27 who observed that all enantiomers of permethrin were oxidized to a 4′-hydroxy permethrin (4 OH) metabolite and underwent esterase cleavage to 3-phenoxybenzyl alcohol (3 PBOH) and 3-(4′-hydroxyphenoxy)-benzyl alcohol) (3,4′ PBOH) in rainbow trout hepatocytes (Oncorhynchus mykiss). Racemic 4 OH permethrin as well as 3 PBOH, and 3,4 PBOH possessed significant estrogenicity, however 1S trans permethrin underwent esterase cleavage more extensively than the corresponding 1R trans permethrin (Figure 127). Similar results were garnered from a study performed by Jin et al.91 with permethrin and cypermethrin. Both parents showed estrogenic activity, but one of the metabolites (PBOH) had greater estrogenic potential. Recent research suggests that pyrethroid metabolism may be species-specific and may even depend on the salinity at which fish are exposed.73 Metabolism in mammals (rat liver microsomes) is similar in that trans permethrin is metabolized more easily and rapidly than the cis diasteromers.51 However, the diastereomers also have different metabolites, with cis permethrin becoming metabolically activated (increased estrogenicity) but not trans permethrin. Cis and trans permethrin also possess different 8981

DOI: 10.1021/acs.est.6b02253 Environ. Sci. Technol. 2016, 50, 8977−8992

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Environmental Science & Technology types of endocrine activity, with trans permethrin being a weak estrogen and antiandrogen versus cis permethrin which acts as both a weak estrogen and a weak antiestrogen.92 Differences in endocrine response between fish and mammals could therefore be due to variations in metabolic capacity. Effects in Vitro. Permethrin may activate or antagonize ERs depending upon the cell line and concentration.25,93−96 Metabolites may also be more potent than the parent.27 However, a recent study using human and rat ERα reporter gene assays showed that both permethrin and its metabolite 3phenoxybenzoic acid inhibited the binding of estradiol to its receptors.96 Additionally, permethrin activates expression of an estrogen-linked transcription factor in some studies,25,79 which contrasts other reports.93−97 Brander et al.28 demonstrated this paradox with a comparison of in vivo responses in M. beryllina (estrogen-mediated protein choriogenin) with an in vitro evaluation using the CALUX ER cell line produced by Rogers and Denison81 (natively expressed human ER α and β). Permethrin caused upregulation of choriogenin in vivo, but blocked the binding of estradiol in vitro. As mentioned earlier, this could be due to different estrogen receptor isoforms being present in different cell lines, due to some cell lines having metabolic capacity while others lack this ability, permethrin acting as a partial agonist, or due to fish having an additional estrogen receptor isoform98 that may respond differently to pyrethroids than traditional mammalian cell lines. Furthermore, it is notable that, unless modified to do so, the most commonly utilized yeast YES cell line does not discriminate well between agonists, partial agonists, and antagonists,3 and the YES assay also does not express mRNA for both ER alpha and beta. These issues aside, however, indications are that permethrin metabolites exert greater estrogenicity than their parental counterparts3,24 (see Biotransformation). A three-way comparison of a fish cell line that lacks metabolism but possesses all three fish ER isoforms with a line that has P450 or carboxylesterase activity, alongside an evaluation of in vivo fish response would be necessary to address current gaps in knowledge for both permethrin and bifenthrin. Effects in Mammals. Like with other pyrethroids, effects in mammals suggest that these chemicals act as EDCs via similar cellular mechanisms across vertebrate species. A study using rat sperm suspensions showed that permethrin (and cypermethrin) have a concentration-dependent ability to reduce motility. However, reduction in motility was not correlated with increasing concentrations of 3-phenoxybenzoic acid (3-PBA), a permethrin metabolite previously shown to have estrogenic, antiestrogenic, and antiandrogenic activity.99 Sperm motility and count were significantly reduced, along with plasma and testicular testosterone levels following oral cis permethrin exposure (70 mg/kg/day). Increased abnormalities in testicular morphology and gene expression (StAR, steroidogenic enzymes) were also observed. However, these effects were not observed following trans permethrin exposure.100 Differential impacts of cis and trans permethrin have also been reported in fish and an earlier study in rats, and likely due to greater susceptibility of trans-permethrin to microsomal esterases,27,100,101 and 3-PBA’s lack of impact on sperm function.99 This is evidenced by trans permethrin-exposed mice having three to 7-fold higher 3-phenoxybenzoic acid levels in the testes and urine.100 Although differences in metabolic capabilities do exist between fish and mammals,100,102 similarities in endocrine response across vertebrates indicate

that conserved mechanisms are responsible. Studies on sperm motility in fish are needed to further elucidate said mechanisms. In addition to direct effects on sperm parameters and gene expression, permethrin has also been implicated in causing cross-generational effects. In mixture with the commonly applied insect repellant, diethyltoluamide (DEET), female rats exposed during fetal gonadal sex determination had offspring with an increased pubertal abnormalities and testicular and ovarian disease. Furthermore, these effects were also observed in the F3 generation, which had a sperm epigenome with over 300 differentially methylated DNA regions in comparison to controls.103 Permethrin’s ability to induce effects in future generations requires additional study, considering the magnitude of its use. Studies in fish that evaluate the toxicity of pyrethroids to the epigenome are highly warranted.

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CYFLUTHRIN Effects in Fishes. Cyfluthrin is a synthetic pyrethroid insecticide that is commonly used in many household insecticides, which also causes paralysis and oxidative stress.104 Cyfluthrin is more toxic than permethrin, and lethal at low ppb concentrations in fish,105,106 with toxicity being greater at cold temperatures107 because it is metabolized more slowly by carboxylesterase and/or cytochrome P450.38 While cyfluthrin is known to bioaccumulate in fishes20 and to induce hepatic enzymes in zebrafish,108 to our knowledge no studies examining cyfluthrin’s potential as an EDC in fishes have been conducted. However, cyfluthrin does retard amphibian larval growth, resulting in smaller larvae undergoing metamorphosis.109 Additionally, Marinowic et al.110 reported that beta-cyfluthrin causes DNA damage (strand breakage measured via Comet assay) in fish found in stagnant waters (Bryconamericus iheringii) at exposure levels of 5.6 μg/L, which are reportedly environmentally relevant in Brazil. Effects in Vitro. Many pesticides exhibit both estrogenic and antiandrogenic activity (including cyfluthrin) increasing the potential for population feminization.94 Cyfluthrin is a strong AR antagonist, inhibiting the transcriptional activity of DHT, but also shows some antiestrogenic activity by competing with 1 × 10−9 E2.95 Cyfluthrin, along with other pyrethroids, binds the pregnane-X-receptor (PXR), a ligand dependent transcription factor that regulates multiple genes including those involved in xenobiotic metabolism.111 Effects in Mammals. Zhang et al.112 reported that low doses (6 mg/kg) of cyfluthrin showed no reduction in sex accessory tissue in rats but higher doses (16−54 mg/kg) of cyfluthrin reduced weights of the seminal vesicle, ventral prostate, dorsolateral prostate, LABC, and Cowper’s glands. Beta-cyfluthrin only showed reduction in weights of these sex accessory tissues at the highest dose (36 mg/kg). The seminal vesicle was shown to be the most sensitive to these antiandrogenic compounds. Studies in fish that examine cyfluthrin’s potential to interfere with the development of reproductive tissues or to reduce fecundity, with particular emphasis on detecting molecular initiating events at the androgen, estrogen, and/or pregnane X receptors, would be highly pertinent for comparisons with mammals.

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ESFENVALERATE/FENVALERATE Effects in Fishes. Like other pyrethroids, at higher concentrations esfenvalerate and fenvalerate are acutely toxic

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Environmental Science & Technology to fish.1,113 Decades ago it was discovered that fenvalerate’s toxicity was mainly associated with the 2S and αS enantiomers, the mixture of which is esfenvalerate, the most widely used of the two.9 Sublethal reproductive effects were reported by Werner et al.114 in Japanese medaka (Oryzias latipes), with observations of a significant downward trend in fecundity and fertilization rate, as well as elevations in heat shock proteins (females were more sensitive). Hatching success and fecundity were also significantly impacted by pulse-exposure to esfenvalerate in rainbowfish (Melanotaenia f luviatilis).115 In bluegill sunfish, individuals exposed to 1 μg/L had significantly delayed spawning in comparison to controls, and YOY growth was significantly reduced at concentrations as low as 80 ng/ L.116 Another study evaluating effects at ecologically relevant concentrations in larval delta smelt (Hypomesus transpacif icus) (62.5 ng/L), showed that genes associated with immune response, apoptosis, redox, osmotic stress, detoxification, growth, and development were differentially expressed.117 The use of esfenvalerate has declined as more toxic pesticides such as bifenthrin have emerged; as such more recent studies in fishes are lacking. Effects in Vitro. The in vitro effects of this pyrethroid pair were first reported when Go et al.97 showed that fenvalerate induces expression of the estrogen-responsive gene ps2 in the MCF-7 human breast carcinoma cell line. Notably, estrogenicity was not inhibited by an ER antagonist, indicating action via an alternative pathway to nuclear ER (demonstrated with other pyrethroids). Fenvalerate also increased cell proliferation in a dose dependent manner. Kojima et al.94 then reported that esfenvalerate increased the transcription of ERα in CHO (Chinese hamster ovary) cells, and Chen et al.25 showed that both fenvalerate and esfenvalerate had specific ER agonist effects. Antiandrogenic activity was reported by Sun et al.,26 in that an AR mediated reporter gene was suppressed by esfenvalerate and cypermethrin and 3-PBA (metabolite) reduced gene expression, thus acting as an antiandrogen in CV-1 cells. In 2010, Du et al.95 reported that cyfluthrin and esfenvalerate antagonized AR by inhibiting the transcriptional activity of Dihydrotestosterone (DHT) but also showed an antiestrogenic response when competing with 1 × 10−9 E2. Evidence suggests that esfenvalerate has the potential to cause negative effects via different endocrine pathways. Interestingly, fenvalerate also induces cell proliferation by progressing cells to the s-stage of the cell cycle and inhibits cell apoptosis, via a non-ER based pathway.118 Effects in Mammals. Esfenvalerate toxicity has been studied in mammals as well. Meng et al.119 reported that exposure to esfenvalerate during puberty in mice resulted in gender specific effects on cognitive development. Exposed female mice had greater impairment in spatial learning and memory compared to male mice, and an increase in anxiety. In the male mice, aggressive behaviors were inhibited. Considering the similarities in response between mammals and fish with regard to other pyrethroids, further investigation into esfenvalerate’s potential impacts on the endocrine systems of fishes is needed.

Cypermethrin has been shown to affect olfactory senses via endocrine-mediated pathways. In brown trout, exposure to 1.0 μg/L decreases the male’s ability to detect female priming pheromones by decreasing blood plasma levels of 17,20β-P and 11-KT.121 Development and fertilization in salmonids and zebrafish was impacted following cypermethrin exposure.120,122 In salmon, the priming effect of prostaglandin, a hormone in the urine of fish that stimulates spawning was significantly reduced after exposure to phenothrin > fluvalinate > permethrin > resmethrin.147 Resmethrin (at 1, 3, 10, and 30 μM) also showed negative activity for the human androgen receptor in a CHO 1K cell assay,148 and was not able to displace [3H]testosterone bound to human sex hormone binding globulin (SHBG) at 10−4 M.146

DELTAMETHRIN Effects in Fish. Deltamethrin a confirmed EDC,133 is commonly used as a treatment for sea lice,10 even though it can be acutely toxic to fishes. 134 Deltamethrin has been documented to affect swimming performance at ng/L concentrations.135 Following exposure, overall reproductive state decreased in zebrafish, including decreased fecundity, hatching rate and egg production136 (1.5 mg/kg). Because deltamethrin causes oxidative stress in fish at low ug/L concentrations,10,137 free radicals form and DNA may be damaged. As mentioned above, other potential molecular targets include dopamerinergic neurons, as behavior modifications were observed with diminished dopamine receptor levels in zebrafish.138 Effects in Vitro. In vitro, exposure to 10−5 M of deltamethrin induced antagonistic effects on hERα and hERβ in CHO-1 cells.94 Deltamethrin had no effect on the estrogen receptor transactivation in MCF-7 cells, but downregulated or inhibited androgen receptor transactivation in CHO K1 cells

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ALTERNATIVE ENDOCRINE DISRUPTING MECHANISMS Although much of the focus on the endocrine disrupting properties of pyrethroid pesticides has centered on estrogenic and androgenic effects, a handful of studies have examined alternative endocrine mechanisms. Several studies point to a potential for thyroid disruption, for example. Studies conducted 8984

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Figure 4. A summary of endocrine signaling pathways and outcomes influenced by pyrethroid pesticides.

in both fish and rats suggest that pyrethroids are associated with altered levels of thyroid hormone,83,147,149−151 and experiments conducted by Du et al.94 demonstrated that many pyrethroids (i.e., cyfluthrin, cypermethrin, deltamethrin, permethrin) antagonize the thyroid receptor in vitro. Cypermethrin has also been shown to stimulate the proliferation of GH3 cells in the T-screen assay.151 Effects on thyroid hormone are variable, with two studies using fenvalerate showing an increase in T3 (triiodothyronine) and T4 (thyroxine),149,150 and another that evaluated commercial formulations of bifenthrin (Talstar) and lambda-cyhalothrin (Karate) resulting in significant suppression of serum T3 and T4.83 Two additional studies in fish and mammals, respectively, also showed a decline in T3 and T4 following lambda-cyhalothrin exposure.147,151 Further studies on fish (Labeo rohita Hamilton) confirm a decreases in T3 and T4, and indicate a significant, dose−response increase in serum thyroid-stimulating-hormone (TSH) levels following exposure.153 TSH is a hormone responsible for the synthesis of T3 and T4, therefore an increase in synthesis may be triggered by the downstream reduction in these hormones. It has been

suggested that alteration of circulating thyroid hormone levels and the concomitant increase in calcium levels149 could be partially responsible for the disruption of motor activity caused by pyrethroid exposure. Recent research also points to the aryl-hydrocarbon receptor (AhR) as being another potential target of pyrethroid pesticides. It is well-established that AhR binds dioxins (i.e., TCDD) and related compounds (i.e., PCBs), and is structurally promiscuous for planar aromatic hydrocarbons. There is also cross-talk that occurs between the AhR and ER pathways resulting in downregulation of estrogen-signaling.154,155 The biphenyl structure of pyrethroid pesticides may make them candidates for AhR binding. Cypermethrin, which has been shown to induce the enzyme CYP1A1 without AhR interaction,156 was found to antagonize TCDD-induced AhR transactivation in mouse hepatoma cells at high μM concentrations.152 Kojima et al.111 observed that a number of pesticides, including all pyrethroids used in their study (i.e., permethrin, cyfluthrin, cypermethrin deltamethrin, fenvalerate) acted as PXR (pregnane X receptor) agonists. Since PXR 8985

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steroidogenesis (e.g., P450scc, 3bHSD, StAR)82,100 in response to several other pyrethroids (bifenthrin, permethrin, cypermethrin, cyfluthrin), since gonadotropins directly stimulate steroidogenesis. A decrease in steroidogenic genes was also seen in the ovaries following exposure to bifenthrin,78 which corresponds with decreases in fecundity, egg production, and follicle size (deltamethrin, bifenthrin, esfenvalerate, cypermethrin).100,120,136,169 Prostaglandins, which stimulate the contractions necessary for ovulation to occur, were also found to decrease in the presence of bifenthrin.78 Reductions in egg production and fertilization rates29,114,120,136 may be attributable to a decrease in prostaglandin, for example. Counterintuitively, egg proteins vitellogenin (Vtg) and choriogenin (Chg), which are stimulated by increases in either hypothalamic or gonadal estradiol production, increase after exposure to bifenthrin and permethrin,89,90,27,28 although Chg has also been observed to decrease as fish are exposed to higher concentrations of bifenthrin.29 This may indicate that complex feedback processes are underway as plasma concentrations of estradiol fluctuate, initially contributing to an increase in egg protein production but perhaps ultimately causing negative feedback, resulting in lowered gonadotropin production (indicated by reduced gonadotroph size). Changes in the direction of response could also be influenced by the metabolic processes discussed in depth above. The links between the master glands in the brain, gonads, and liver should be further explored, not only using single pyrethroids, but with mixtures as well. Another end point of interest specific to the pituitary is thyroid stimulating hormone (TSH), which increases in the presence of bifenthrin,83 while both T3 and T4 synthesis appears to decline in thyroid hormone producing tissues,147 potentially due to antagonistic actions of pyrethroids such as bifenthrin and lambda-cyhalothrin documented in vitro at TRβ.95 An increase in TSH is concomitant with decreases in both thyroid hormone isoforms, as the negative feedback resulting from what appears to be reduced hormone synthesis in the thyroid gland or thyroid hormone producing cells (no gland in many fish species) would stimulate an increase in TSH. Although links between pyrethroid exposure and immune response are tenuous, one study indicates that gonadal inflammation increases due to the presence of mast cells.66 Elevated immune response following pyrethroid exposure were shown by Jeffries et al.157 and Connon et al.,116 and Kumar et al.,158 showed an increase in apoptosis in response to deltamethrin, suggesting that disruption of immune response is a potential mechanism accompanying endocrine disruption that should be further explored.

regulates many genes involved in xenobiotic metabolism, this finding indicates that pyrethroids can modulate the expression and activity of P450 enzymes, possibly including those involved in steroidogenesis, via the PXR pathway. Pyrethroids may also exert an effect on immune responses. Both permethrin and esfenvalerate exposure at μg/L concentrations in the delta smelt (Hypomesus transpacif icus), decreased expression of genes involved in immune function in a microarray analysis.117,157 Similarly, deltamethrin appears to induce apoptosis in cells of the thymus in mice,158 likely causing a decrease in immune response. One potential mechanism for such immune disturbance may be via the glucocorticoid receptor, for which cypermethrin caused activation of in the MDA-kb2 luciferase cell line.159 Activation of the glucocorticolid receptor can cause immunosuppression. It is also known that EDC interference with estrogen and androgen signaling causes changes in immune responses, since ER and AR receptors are present on the T and B cells of the immune system in vertebrates.160−162 As such, the potential for modulation of immune response following pyrethroid exposure is not surprising. While both the observed thyroid-disrupting and AhR activities of pyrethroid pesticides should be further investigated, as well as effects on immune response, possibly more concerning is the ability of pyrethroids to cause changes in DNA methylation levels causing alterations in gene expression in offspring (F1 and F2) of exposed adults.163 Permethrin causes a decrease in global DNA methylation in the unexposed offspring of mothers (rats) that were exposed only during early life.164 Studies in fish with rapid developmental life stages may allow further elucidation this mechanism, as it is already established that fish DNA is methylated similarly to mammals in response to EDC exposure.165−167

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ENDOCRINE EFFECTS IN FISH: A SUMMARY The hypothalamic pituitary gonadal and thyroid axes in fishes are affected by pyrethroid exposure in a number of ways, with responses both upstream and downstream likely being influenced by each effect (Figure 4, SI Figure S1). Alterations to the hypothalamus and “master gland” (pituitary), endocrine signaling centers in vertebrates, are typical of other well-studied endocrine disrupting compounds, such as bisphenol A, phthalates, and diethylstibestrol.33 Pyrethroid induced disruption begins with changes in the hypothalamus, such as fluctuations in GnRH or GnRH2 (gonadotropin releasing hormone) or GnRHR (GnRH receptor),80 which control the release of gonadotropins from the pituitary gland (FSH, LH) that in turn stimulate steroidogenesis in the gonads. Notably, pyrethroids such as bifenthrin appear to interfere with dopamine signaling, which in fish controls the release of GnRH and hence production of steroids such as estradiol.75 Another study showed that gonadotroph size in fish decreases following exposure to cypermethrin,123 which could be related to responses triggered by bifenthrin in dopamine receptor levels and GnRH. Brain aromatase (CYP 19b), also found in the hypothalamus where it contributes to feedback processes, increases with bifenthrin or permethrin exposure,69,90 to which observed fluctuations in plasma estradiol72,75 and testosterone100 could also be attributed. Decreases in gonadotroph size observed by Singh and Singh123 in response to cypermethrin could be linked to noted decreases in various sperm parameters,99,100,129 GSI or testis weight,72,131 and the decrease in expression of genes involved in

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CONCLUSION Pyrethroid contamination of aquatic ecosystems has become a global issue.168 Furthermore, indications are that declines in fish abundance in various sites have coincided with the upward trend in pyrethroid usage,169 and that wild fish are contending simultaneously with additional threats such as other contaminants, hypersaline conditions from sea-level rise, warming temperatures, and decreased habitat availability. Effects occurring in combination with other endocrine-active pollutants cannot necessarily be predicted by concentration-addition models. 74 Furthermore, exposure to multiple stressors decreases the ability to physiologically cope with these highly toxic, endocrine active compounds.170 While some of the studies discussed above use concentrations higher than those 8986

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and the Environmental Protection Agency (EPA STAR #835799, to SMB) made this work possible.

commonly measured in the environment, many are consistent with pyrethroid concentrations detected in aquatic habitats, known to range from ng/g in sediments to ng−μg/L in the water column.1,7,17,14,110 Notably, concentrations in suspended sediments have been measured in the mg/L range following storms.171 It is also important to consider that these concentrations may not be representative of those that are actually bioavailable, since pyrethroids have been demonstrated to bioconcentrate in animal tissues possessing ng−μg/g levels.19,20 Furthermore, much work remains to better determine safe levels in the context of multiple stressors, pyrethroid mixtures, metabolites, and the potential for bioconcentration/bioaccumulation, and to develop analytical approaches sufficient to consistently detect pyrethroid presence in water, sediment, and tissues at ng/L and ng/g levels and below, particularly given that pyrethroid usage is on the rise globally.2 An additional complication is the observed tendency of pyrethroids, like other EDCs, to elicit nonmonotonic responses.33,34,171 Several studies have demonstrated stronger responses to pyrethroids at lower concentrations than at higher, indicating that these compounds are acting via different mechanisms depending on the dose.22,28,157,169,172 As mentioned earlier, this may be due in part to the different effects of pyrethroid metabolites versus parent compounds. Studies in fishes that better isolate the molecular- and organism-level effects of pyrethroid metabolites, stereospecific interactions, species-specific response, epigenetic impacts, and the nature of observed biphasic or nonmonotonic responses are greatly needed. Additionally, the development of modeling approaches that allow for the use of data generated from studies conducted at these smaller biological scales, to be used in making predictions of population persistence, is necessary to make the link between the multitude of biological effects observed in the laboratory and the continued observed decline in wild fish populations.173 Many of the molecular mechanisms responsible for such effects appear to be similar between vertebrates (mammals and fish), and as such the use of fish models for biomonitoring and mechanistic studies will inform both human and ecological health.

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(1) Werner, I.; Moran, K., Effects of pyrethroid insecticides on aquatic organisms. In Synthetic Pyrethroids: Occurrence and Behavior in Aquatic Environments, Gan, J.; Spurlock, F.; Hendley, P.; Weston, D. P., Eds.; American Chemical Society: Washington D.C., 2008; ACS Symposium Series 991, pp 310−335. (2) van den Berg, H.; Zaim, M.; Yadav, R. S.; Soares, A.; Ameneshewa, B.; Mnzava, A.; Hii, J.; Dash, A. P.; Ejov, M., Global trends in the use of insecticides to control vector-borne diseases. Environ. Health Perspect. 2012, 120, (4), 577, DOI: 10.1289/ ehp.1104340. (3) McCarthy, A. R.; Thomson, B. M.; Shaw, I. C.; Abell, A. D. Estrogenicity of pyrethroid metabolites. J. Environ. Monit. 2006, 8, 197−202. (4) Barr, D. B.; Olsson, A. O.; Wong, L.-Y.; Udunka, S.; Baker, S. E.; Whitehead, R. D., Jr.; Magsumbol, M. S.; Williams, B. L.; Needham, L. L. Urinary Concentrations of Metabolites of Pyrethroid Insecticides in the General U.S. Population: National Health and Nutrition Examination Survey 1999−2002. Environ. Health Perspect. 2010, 118 (6), 742−748. (5) Brander, S. M.; Werner, I.; White, J. W.; Deanovic, L. A. Toxicity of a dissolved pyrethroid mixture to Hyalella azteca at environmentally relevant concentrations. Environ. Toxicol. Chem. 2009, 28, 1493−1499. (6) Brander, S. M.; Mosser, C. M.; Geist, J. P.; Hladik, M. L.; Werner, I. Esfenvalerate toxicity to the cladaceran Ceriodaphnia dubia in the presence of the green algae, Pseudokirchneriella subcapitata. Ecotoxicology 2012, 21 (8), 2409−2418. (7) Weston, D. P.; Lydy, M. J. Stormwater input of pyrethroid insecticides to an urban river. Environ. Toxicol. Chem. 2012, 31 (7), 1579−1586. (8) Hasenbein, S.; Connon, R.; Lawler, S.; Geist, J. A comparison of the sublethal and lethal toxicity of four pesticides in Hyalella azteca and Chironomus dilutus. Environ. Sci. Pollut. Res. 2015, 22 (15), 11327−11339. (9) Bradbury, S. P.; Coats, J. R. Toxicokinetics and toxicodynamics of pyrethroid insecticides in fish. Environ. Toxicol. Chem. 1989, 8, 373− 380. (10) Guardiola, F. A.; Gónzalez-Párraga, P.; Meseguer, J.; Cuesta, A.; Esteban, M. A. Modulatory effects of deltamethrin-exposure on the immune status, metabolism and oxidative stress in gilthead seabream (Sparus aurata L.). Fish Shellfish Immunol. 2014, 36 (1), 120−129. (11) Haverinen, J.; Vornanen, M. Effects of deltamethrin on excitability and contractility of the rainbow trout (Oncorhynchus mykiss) heart. Comp. Biochem. Physiol., Part C: Toxicol. Pharmacol. 2014, 159, 1−9. (12) CDC. Fourth National Report on Human Exposure to Environmental Chemicals; Atlanta, GA, 2009. (13) Phillips, B. M.; Anderson, B. S.; Hunt, J. W.; Siegler, K.; Voorhees, J. P.; Tjeerdema, R. S.; McNeill, K. Pyrethroid and organophosphate pesticide-associated toxicity in two coastal watersheds (California, USA). Environ. Toxicol. Chem. 2012, 31 (7), 1595− 1603. (14) Kuivila, K. M.; Hladik, M. L.; Ingersoll, C. G.; Kemble, N. E.; Moran, P. W.; Calhoun, D. L.; Nowell, L. H.; Gilliom, R. J. Occurrence and Potential Sources of Pyrethroid Insecticides in Stream Sediments from Seven U.S. Metropolitan Areas. Environ. Sci. Technol. 2012, 46 (8), 4297−4303. (15) Parry, E.; Young, T. M. Distribution of pyrethroid insecticides in secondary wastewater effluent. Environ. Toxicol. Chem. 2013, 32 (12), 2686−2694. (16) Smalling, K. L.; Kuivila, K. M.; Orlando, J. L.; Phillips, B. M.; Anderson, B. S.; Siegler, K.; Hunt, J. W.; Hamilton, M. Environmental fate of fungicides and other current-use pesticides in a central California estuary. Mar. Pollut. Bull. 2013, 73 (1), 144−153.

ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.6b02253. Figure S1: (a) figure demonstrating that the two chiral centers of the pyrethroid bifenthrin allow for two enantiomeric forms, 1R-cis and 1S-cis.; (b) table summarizing pyrethroid endocrine effects per test type (fish, mammal, in vitro) and concentration (PDF)

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REFERENCES

AUTHOR INFORMATION

Corresponding Author

*Phone: 910-962-3786; fax: 910-962-4006; e-mail: branders@ uncw.edu. Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS Funding from the California Department of Fish and Wildlife (contract #E1183010 to REC, SMB), the State and Federal Contractors Water Agency (contract #06-447-300, to REC), 8987

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