Experimental Designs to Assess Endocrine Disrupting Effects in Invertebrates A Review

Ecotoxicology, 13, 511–517, 2004 Ó 2004 Kluwer Academic Publishers. Manufactured in The Netherlands.

Experimental Designs to Assess Endocrine Disrupting Effects in Invertebrates A Review CARLOS BARATA,1,2,* CINTA PORTE2 AND DONALD J BAIRD3 1

Laboratory of Environmental Toxicology, INTEXTER-UPC. Carretera Nacional 140 Km. 14.5. Terrassa 08220, Barcelona, Spain 2 Department of Environmental Chemistry, IIQAB-CSIC, Jordi Girona 18, Barcelona 08034, Spain 3 Institute of Aquaculture, University of Stirling, Stirling FK9 4 LA, UK Accepted 30 June 2003

Abstract. In order to gain basic understanding of the ecological effects of vertebrate Endocrine Disrupting Chemicals (EDCs), many research groups are currently testing these chemicals using aquatic invertebrates. Small crustaceans, such as cladocerans and copepods, are of particular interest since they are ecologically important and their short life cycles allow obtaining information on demographic parameters. Despite the existence of diverse literature on the development, growth and reproductive effects of EDCs on these crustaceans, only a few studies have unambiguously assessed a truly endocrine disrupting effect. This review discusses new experimental designs to differentiate between endocrine disruption and other causes of reproductive and developmental impairment. Our findings clearly illustrate that many studies may have falsely concluded that chemicals have endocrine disrupting modes of action when in fact a much simpler explanation was not previously ruled out (e.g., egg mortality, feeding inhibition). This means that there is an urgent need for integration of toxic effects on energy intake to toxicity assessments. Such an approach would permit different ectotoxicological models of action, including endocrine disrupting effects, to be distinguished and their relative roles in the overall toxic response to be clarified. Keywords: endocrine disruption; reproduction; development; Daphnia, copepods

Introduction In order to gain basic understanding of the ecological effects of vertebrate endocrine disrupting chemicals (EDCs), many research groups are currently testing these chemicals using *To whom correspondence should be addressed: Laboratory of Environmental Toxicology (INTEXTER-UPC), Carretera Nacional 150, Km 14.5 Tel.: +34-93-7398396; Fax: +34-93-7398392; E-mail: [email protected]

aquatic invertebrates (Hutchinson et al., 2000; Hutchinson, 2002). Small crustaceans such as cladocerans and copepods are of particular interest since they are ecologically important and their short life cycles allow to obtain information on survival and reproductive responses and hence to predict population responses of toxic exposure (Hutchinson, 2002). Nevertheless, despite the existence of diverse literature on the development, growth and reproductive effects of EDCs on crustaceans, only a few studies have unambiguously assessed a truly endocrine

512 Barata et al. disrupting effect. According to Hutchinson (2002), toxicity of EDCs to crustaceans such as copepods are in most cases more likely due to other undefined modes of actions. Thus, it is important to develop bioassay designs which can be used to differentiate between endocrine disruption and other causes of reproductive and developmental impairment. Studies on EDCs conducted in small crustaceans such as copepods and Daphnia often assume that these chemicals only produce effects on developmental rate, growth and reproduction by disrupting hormones controlling maturation and reproduction. Nevertheless, other toxic effects such as stage- or sex-related differences in sensitivity (i.e. survival) to EDCs and/or reduced resource acquisition may also affect the above mentioned lifehistory traits. For this reason, it is important to develop bioassays able to discriminate between hormone-mediated and other toxicological modes of action. Here we propose experimental designs which permit endocrine disrupting and other modes of action to be separately diagnosed.

Ecotoxicological modes of action of toxic chemicals In chemical risk assessment, population responses have been proposed as more ecologically relevant measures of toxicant effect than mortality or reproduction. This is argued from the viewpoint that population responses integrate life-history responses of individuals, providing insight into potential effects at higher levels of ecological organisation (Barata et al., 2002c). Population responses are often inferred from observations made at the level of the individual (De Angelis and Gross, 1990). To make predictions concerning an individual’s contribution to population responses, age-specific survival, fecundity and timing of reproduction must be measured and combined to estimate demographic parameters (e.g. Barata et al., 2002c). Thus, toxic substances can affect population responses in three different ways, hereafter referred as ecotoxicological modes of action. Figure 1 shows briefly the relationship between ecotoxicological modes of action, individual energy processes and life-history responses. Toxic chemicals can affect population responses by impairing survivorship (lethal effects) of eggs, juveniles, males or female adults by disrupting

specific key physiological functions (i.e. acetylcholinesterase by organophosphorous pesticides; Barata et al., 2001; Fig. 1A). Alternatively, toxic substances may also exert sublethal effects on demographic responses impairing reproduction and growth related traits including offspring production, growth and maturation age by either altering individual energy budgets or disrupting hormonal levels. Energy budgets can be altered by either impairing energy acquisition (i.e. feeding rates by metals; Barata and Baird, 2000; Fig.1B1) or/and increasing energy demands (i.e. increasing maintenance costs by 2,4-dichloroaniline; Barata and Baird 2000; Fig. 1B2). Since the amount of energy that an individual can invest in maintenance, growth and reproduction is limited (Sibly and Calow, 1986), reductions in energy acquisition and/or increasing energy demand will resolve in decreasing growth and offspring production (Barata and Baird, 2000). Finally, toxic exposure may affect reproduction and growth responses disrupting hormonal levels without affecting individual energy budgets or other physiological functions (Fig. 1C1 and C2). For example, estrogenic and androgenic substances are likely to enhance and reduce reproduction of female individuals at expenses of survival or growth, respectively (Tillman et al., 2001). Additionally, ecdysteroid and juvenile hormone agonist and antagonists substances may alter growth and the onset of reproduction by disrupting moulting and metamorphosis processes (Chang and O’Connor, 1978; Hertz and Chang, 1986). It is worth noting that chemicals that affect energy budgets or disrupt endocrine levels may also impair survival at high exposure levels. For example, high levels of feeding inhibition and ecdysone agonists may cause mortality due to toxic anorexia or incomplete ecdysis, respectively (Barata and Baird, 2000; Baldwin et al., 2001). Nevertheless, the effects that these chemicals exert on reproduction and growth-related traits at sublethal exposure levels are of special concern to environmental risk assessors (Hutchinson, 2002).

Toxicity bioassays to measure ecotoxicological models of action Toxic substances can have multiple toxicological modes of action. It is therefore crucial to focus our toxicological investigations to study the

Bioassays to Assess Endocrine Disrupting Effects 513 EA

M (s)

R (n)

ED

G (t) A

B1

s

s

C1 n

n

s

t

t C2

B2 s

n

t

s

n

t

Figure 1. Pseudobond diagrams describing the relationship between ecotoxicological modes of action, energy budgets and demographic traits. Top graph describes how the energy acquired from food (EA) is allocated into the three main energy allocation processes that constitute the energy demand (ED) compartment (M, maintenance; G, growth; R, reproduction), and how energy flow is related with demographic traits (s, survival; t, developmental rate or maturation age; n, offspring production). Arrows indicated the direction of energy flow relative to the organism (depicted as a diamond symbol). The wider the arrow the greater the energy flow. Bottom graphs describe five different scenarios including: Toxic effects on egg, juvenile, adult female or male survival due to specific modes of action without effecting energy budgets, neither reproduction nor growth (A). Toxic effects on developmental rates, offspring production and survival due to altered energy budgets by either reducing energy acquisition (B1) or increasing maintenance costs and hence energy demand (B2). Endocrine disrupting effects including enhanced reproduction (C1) and growth (C2) at expenses of the other energy allocation processes. Alternative endocrine disruption effects including reductions of growth and reproduction are also possible. In graphs A–C life-history traits that are likely to be affected are also depicted. For further explanation see text.

dominant ecotoxicological (sensu Barata and Baird, 2000) mode of action, based on the concentration at which various toxicological effects become potentially operative at population or ecosystem level. Although it is possible to obtain information on lethal and sublethal toxic effects from standard reproduction and life-cycle tests, no existing test yields a complete picture of the overall response. Tests based on putative sublethal responses such as the Daphnia magna and copepod chronic tests (ASTM, 1995; OECD, 1997; Bechmann, 1999; Hutchinson et al., 1999a, b; Andersen et al., 2001), may allow collection of information on lethal and sublethal effects, but this information is incomplete and may be

confounded, for example by failing to distinguish between effects on embryonic survivorship during development or energy allocation to eggs (Barata and Baird, 2000). Furthermore, existing reproduction and demographic tests are also unable to discriminate if sublethal effects on reproduction or growth are mediated by reductions in energy acquisition. By developing test designs which allow us to quantify and separate lethal and sublethal effects it should be possible to group chemicals by the effects that they exert on individuals in terms of their population or ecosystem consequences. This could be simply achieved from existing reproduction and demographic tests, for example by measuring a few

514 Barata et al. additional variables such as feeding and egg mortality rates. Furthermore, by exposing eggs, nauplii and adults in separate tests it should be also possible to shorten the duration of these tests to as little time as is the production of few broods. Indeed, recent studies have shown that it is possible to successfully measure both lethal effects on egg, juvenile and adult stages and nonlethal effects on food acquisition and production rates of Daphnia and copepod species by using a set of ecological relevant short term toxicity bioassays (Barata and Baird, 2000; Barata et al., 2002 a, b).

Proposed experimental design To truly assess endocrine disruption effects of potential EDCs on populations we should be confident that toxicity is manifested in life-history responses which include maturation, growth and reproduction and that the following required criteria are satisfied: 1. Toxic effects should occur at relevant environmental–physiological concentrations and effects on maturation, growth and/or reproduction should occur at lower concentrations that those causing egg, juvenile and/or adult mortality. Although many substances can disrupt hormonal levels at high exposure levels and impair survival, those that affect growth and reproduction-related traits at low exposure levels are of special interest since they may affect populations over the long term. 2. Energy intake should not be affected (feeding or ingestion rates). Many substances including PAHs, metals and pesticides affect energy intake and hence exert sublethal effects on growth and reproduction in copepod and Daphnia species (Barata and Baird 2000; Barata et al., 2002 a, b). Thus it is important to discriminate between sublethal effects on energy intake from those related to endocrine disruption per se. This could be achieved experimentally by using toxicity bioassays which allow us to quantify and separate lethal effects on egg, juvenile, male and female survival and sublethal effects on food intake, growth and reproduction. Although toxic effects on energy demand (i.e. metabolic rates) could also be measured, previous studies have suggested

that energy supply is usually affected at lower exposure levels and to a greater extent than energy demand. As a result toxicity effects on energy demand tend to contribute little to the total energy budget (Baird et al., 1990; Barata and Baird, 2000). 3. Life-history responses should be coupled with biochemical effects including evidence for mimicking and antagonising hormonal effects, altered patterns of synthesis and metabolism of hormones, or interaction with hormonal receptor levels (Baldwin and LeBlanc, 1994; Depeldge and Billinghurst, 1999; Dinan et al., 2001) Although biochemical effects of EDCs have been exhaustively studied in vertebrate species, information on invertebrate species are mostly limited to larger invertebrate species (Fingerman, 1997; Depeldge and Billinghurst, 1999). Thus more research should be done in studying endocrine control of key processes such as moulting or reproduction in small crustaceans such as copepods and Daphnia. A promising approach, however, is the use of receptor-based in vitro and in vivo hormone metabolism bioassays to screen potential EDCs (Baldwin and LeBlanc, 1994; Baldwin et al., 1995, 1998; Dinan et al., 2001).

Testing the proposed experimental design against published demographic bioassays In order to assess whether recent published ecotoxicological studies on EDCs were able to truly assess endocrine disrupting effects on demographic responses, the proposed experimental design was tested against ecotoxicological studies conducted in Daphnia, Tisbe battagliai and Acatia tonsa species. Due to the large number of studies performed on invertebrate and small crustacean species, we restricted the study to three species that are widely used in Ecological Risk Assessment of toxic chemicals. From the 16 demographic studies analysed (Table 1), one failed to truly differentiate between lethal and sublethal effects, and none measured toxic effects on food intake. This means that although 11 out of 31 substances examined exerted significant sublethal effects in development, growth or/and reproduction related traits (i.e. diethyl phthalate, Aroclor 1242, PCB29, 4-nonylphenol, pentachlorophenol, cyproterone

Bioassays to Assess Endocrine Disrupting Effects 515 Table 1. Concordance of reported studies on EDC in Daphnia, Acartia tonsa and Tisbe battaglai species against the three premises of the proposed experimental design 1 LeBlanc and McLachlan (2000); 2 Oberdorster et al. (1998); 3 Baldwin et al. (1998); 4 Zou and Fingerman (1997); 5 Baldwin et al. (1997); 6 Parks and LeBlanc (1996); 7 Baldwin et al. (1995); 8 LeBlanc and MaLachlan (1999); 9 Baldwin et al. (2001); 10 Olmstead and LeBlanc, (2000); 11 Peterson et al. (2001); 12 Mu and LeBlanc (2002); 13 Andersen et al. (2001); 14 Hutchinson et al. (1999a); 15 Hutchinson et al. (1999b); 16 Bechmann (1999). 1 – reporting sublethal effects on growth (G) and reproduction (R) related life-history traits including effects on body size (l), moulting frequency (m), development (t), offspring number (n) and offspring sex (x) at relevant environmental concentrations, 2 – showing that sublethal effects are not due to reduced energy acquisition, 3 – relating life-history effects with pharmacological effects (i.e. hormonal metabolism, HM; endogenous hormonal levels, EH). – Untested premise due to an incomplete experimental design. Reported studies are depicted between brackets Species and chemicals tested Daphnia Tributylin [1] Tributylin [2] 4-Nonylphenol [3] 4-Octylphenol, lindane [4] Diethyl phthalate, Aroclor 1242, PCB29 [4] 4-Nonylphenol [5] Pentachlorophenol [6] Diethylstilbestrol [7] Cyproterone acetate [8] 20-Hydroxyecdysone [9] Ponasterone A [9] Diethylstilbestrol, methroprene, androstenedione [10] Methroprene, 20-hydroxyecdysone [11] 9-Cis and trans retinoic acids [11] Testosterone [12] Acartia tonsa 17 b-Estradiol, estrone, testosterone, progesterone, 10-hydroxyecdysone, juvenile hormone-III, flutamide, tamoxifen, hydroxyflutamide, 17a-ethinylestradiol, 4-octylphenol, bisphenol A, 4-nonylphenol, diethyl phthalate, potassium dichromate, 3,5-dichlorophenol, 3,4 dichloroaniline [13] Tisbe battagliai Diethylstilbestrol, 17b-estradiol, estrone, 17a-ethinylestradiol [14,15] 20-Hydroxyecdysone [14,15] Nonylphenol [16]

1

2

3

– No (G-m, R-n) No (R-n) No (G-m) Yes (G-m)

– – – – –

Yes (HM) Yes (HM) Yes (HM) – –

Yes (R-n) Yes (R-n) Yes (G-l-m,R-n) Yes (G-l-m-t,R-n) No (G-m, R-n-x) Yes (R-n) Yes (G-l)

– – – – – – –

Yes Yes Yes – Yes Yes –

Yes (R-x)

–

–

– –

– No (EH)

–

–

No (G-t, R-n)

–

–

Yes (R-n) No (G-t, R-n)

– –

– –

No (R-x) No (R-n) –

a

b

(HM) (HM) (HM) (HM) (HM)

a

Reduced offspring number were due to lethal effects on embryos rather than on egg number. Lethal effects were measured in adults whereas sublethal effects on developmental rates were assessed in nauplii.

b

acetate, 20-hydroxyecdysone, methroprene, diethylstilbestrol, ponasterone A, androstenedione), incomplete experimental designs prevented the discrimination between endocrine disruption and energy-mediated effects. Moreover, in only eight studies were demographic responses complemented by evidence of biochemical effects (i.e. altered hormonal metabolism, endogenous

hormonal levels). More specifically, for only 4 substances, 4-nonylphenol, pentachlorophenol, diethylstilbestrol, ponasterone A were sublethal demographic responses coupled with biochemical effects. It is worth emphasising that in all the studies considered in this review, sublethal effects were always associated with reduced reproduction

516 Barata et al. and/or growth related traits, a response which is also invariably associated with reduced energy intake. Alternative effects such as enhanced reproductive output following exposure to xenoestrogens, a typical response found in prosobranch snails (Oehlmann et al., 2000), was not found in any of the reported studies. Furthermore, several studies conducted with Daphnia misinterpreted endocrine disrupting effects on qualitative traits such as moulting and sex (Zou and Fingerman, 1997; LeBlanc and McLachlan, 1999; Olsmtead and LeBlanc, 2000; Baldwin et al., 2001; Peterson et al., 2001). Although traits like incomplete ecdysis, altered moulting frequencies, changing sex ratios and altered sex characters are good indicators of endocrine disrupting effects, alternative modes of action may also affect the previous endpoints. Daphnia needs to moult to growth and ultimately to survive (Barata and Baird, 1998), hence energy-mediated impairments in growth and survival in addition to endocrine disruption are also likely to affect the moulting process. Furthermore, sex in Daphnia including male production and sexual characters are environmentally determined by many factors and are also related with growth (Mitchell, 2000; Zhang and Baer, 2000). Therefore, chemicals that alter individual energy budgets and hence impair growth may also affect sex characters in Daphnia.

Conclusions and recommendations It is clear from the above that many studies may have falsely concluded that chemicals have endocrine-disrupting modes of action when in fact a much simpler explanation was not previously ruled out (e.g. egg mortality, feeding inhibition). From the principle of parsimony, it is generally accepted that simpler hypotheses, based on well-understood mechanisms should have primacy over more complex, mechanistically opaque assertions. Our findings clearly illustrate the urgent need for integration of toxic effects on energy intake to toxicity assessments. Such tests should permit different ecotoxicological modes of action including endocrine disrupting effects to be distinguished, and their relative roles in the overall toxic response to be clarified. Also, it is clear that more effort

should be devoted to studying the biochemical effects of EDCs on target species. Although our conclusions are restricted to bioassays performed on only three species, our arguments can be applied more generally to studies conducted in other invertebrate species.

Acknowledgement The senior author was partially supported through the European Large Scale Facility Programme and European Union Improving Human Potential Programme (contract HPRI-CT-1999-00106).

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