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hort-term exposure to a treated sewage effluent alters reproductive behaviourn the three-spined stickleback (Gasterosteus aculeatus)
arion Sebirea,b,∗, Ioanna Katsiadakia, Nick G.H. Taylora, Gerd Maackb,c, Charles R. Tylerb
Cefas Weymouth Laboratory, Barrack Road, The Nothe, Weymouth, Dorset DT4 8UB, UKBiosciences, College of Life and Environmental Sciences, The Hatherly Laboratories, University of Exeter, Exeter EX4 4PS, UKFederal Environment Agency, Environmental Risk Assessment of Pharmaceuticals, Wörlitzer Platz 1, 06844 Dessau-Roßlau, Germany
r t i c l e i n f o
rticle history:eceived 24 January 2011eceived in revised form 13 May 2011ccepted 17 May 2011
Some UK sewage treatment work (STW) effluents have been found to contain high levels of anti-androgenic activity, but the biological significance of this activity to fish has not been determined. Theaim of this study was to investigate the effects of exposure to a STW effluent with anti-androgenicactivity on the reproductive physiology and behaviour of three-spined sticklebacks (Gasterosteus aculea-tus). Fish were exposed to a STW effluent (50 and 100%, v/v) with a strong anti-androgenic activity(328.56 ± 36.83 �g l−1 flutamide equivalent, as quantified in a recombinant yeast assay containing thehuman androgen receptor) and a low level of oestrogenic activity (3.32 ± 0.66 ng l−1 oestradiol equiva-lent, quantified in a recombinant yeast assay containing the human oestrogen receptor) for a period of 21days in a flow-through system in the laboratory. Levels of spiggin, an androgen-regulated protein, werenot affected by the STW effluent exposure, nor were levels of vitellogenin (a biomarker of oestrogenexposure), but the reproductive behaviour of the males was impacted. Males exposed to full strengthSTW effluent built fewer nests and there was a significant reduction in male courtship behaviour forexposures to both the 50 and 100% STW effluent treatments compared with controls. The effect seen onthe reproduction of male sticklebacks may not necessarily have been as a consequence of the endocrine
active chemicals present in the STW effluent alone, but could relate to other features of the effluent, suchas turbidity that can impair visual signalling important for courtship interactions. Regardless the specificcausation, the data presented show that effluents from STW have an impact on reproductive behaviour inmale sticklebacks which in turn affects reproductive performance/outcome. The study further highlightsthe use of fish behaviour as a sensitive endpoint for assessing potential effects of contaminated water
on.
bodies on fish reproducti
. Introduction
To date, the majority of studies assessing the biological effectsf sewage treatment work (STW) effluents on fish have been con-erned with their oestrogenic activity. Induction of vitellogenina female specific yolk protein precursor) has been shown in aide range of fish exposed to UK STW effluents, e.g. in rainbow
rout (Oncorhynchus mykiss, Routledge et al., 1998a), roach (Rutilusutilus, Jobling et al., 1998; Liney et al., 2006; Rodgers-Gray et al.,000; Routledge et al., 1998a), sand goby (Pomatoschistus minu-us, Robinson et al., 2003) and gudgeon (Gobio gobio, van Aerle
t al., 2001). Feminised gonads have also been documented inild fish (roach and gudgeon) populations living in effluent con-
aminated UK Rivers (Jobling et al., 1998; van Aerle et al., 2001).
Importantly, wild male roach that have been found to be femi-nised as a consequence of exposure to oestrogenic effluents have areduced reproductive capability, providing potential for populationlevel effects (Harris et al., 2011; Jobling et al., 2002). STW effluentshave also been shown to be oestrogenic in other parts of Europe,in North America, Australia, China and Japan (Batty and Lim, 1999;Flammarion et al., 2000; Folmar et al., 1996; A. Johnson et al., 2007;Knudsen et al., 1997; Larsson et al., 1999; Ma et al., 2007; Nakadaet al., 2004; Pettersson et al., 2006; Viganò et al., 2008). The oestro-genic chemicals in STW effluents have been identified and includenatural and synthetic steroidal oestrogens, alkylphenol ethoxylatesand in some instances natural products, preservatives (parabens)and pesticides (Chen et al., 2002; Desbrow et al., 1998; Holland,2003; Larsson et al., 1999; McCarthy et al., 2006; Routledge et al.,1998b; Routledge and Sumpter, 1996; Shore and Shemesh, 2003;
Soto et al., 1991).
Few studies have investigated hormonal activities in the envi-ronment other than for oestrogens, but both androgenic (Chatterjeeet al., 2007; Ellis et al., 2003; Katsiadaki et al., 2002a; Parks et al.,
001; Svenson and Allard, 2004a; Thomas et al., 2002) and anti-ndrogenic (I. Johnson et al., 2007; Tollefsen et al., 2007; Urbatzkat al., 2007) activities have been identified. Androgenic activityccurs in both pulp and paper mill effluents and cattle feedlot efflu-nts (in the US) and has been shown to result in masculinisation ofild fish in the receiving waters (Adams et al., 1992; Ankley et
l., 2003; Ellis et al., 2003; Jensen et al., 2006; Karels et al., 2001;rlando et al., 2007; Parks et al., 2001; Soto et al., 2004; Wilson et al.,002). The androgenic chemical in cattle feedlots was identified asrenbolone (a growth promoter used in beef production, Durhant al., 2006; Schiffer et al., 2001) whilst the androgenic activity ofulp and paper mills appears to result from wood derived-productsSvenson and Allard, 2004b). Androstenedione has been reportedo contribute to the androgenic activity of pulp and paper millsAllinson et al., 2008; Jenkins et al., 2001) – although this find-ng remains controversial (Bandelj et al., 2006; Ellis et al., 2003;ewitt et al., 2006). A report by I. Johnson et al. (2007) showed
ignificant anti-androgenic activity in 41 out of 43 UK STW efflu-nts tested, with potency ranging between 21.3 and 1231 �g l−1
utamide equivalent (FLeq, as determined in the recombinant yeastndrogen screen – rYAS), and in a modelling study, this activity washown to be statistically associated with some of the feminisedesponses seen in wild roach living in those rivers (Jobling et al.,010). Some of these anti-androgens have recently been identifiednd they include dichlorophene and di(chloromethyl)-anthraceneHill et al., 2010), although these two chemicals accounted for only
small fraction of the total anti-androgenic activity as determinedn the rYAS. In vitro anti-androgenic activity (as determined inhe rYAS) has also been shown to occur in the produced waterf oil production platforms in Norway, at concentrations rang-ng between 20 and 8000 �g FLeq l−1 (Tollefsen et al., 2007),nd in the River Lambro, North Italy, at levels between 370 and223 �g FLeq l−1 (Urbatzka et al., 2007). A study by Thomas et al.2009) reported that the anti-androgenic fraction in the producedater from offshore oil production platforms in the North Seaas comprised mainly of naphthenic acids and polycyclic aro-atic hydrocarbons (PAHs). Studies on effluents discharged in the
iver Lambro have identified bisphenol A, iprodione, nonylphe-ol, p,p′-DDE and 4-tert-octylphenol, chemicals all known to haveeak anti-androgenic activity, but the vast majority of the anti-
ndrogenic activity in those effluents has not been chemicallyscribed (Urbatzka et al., 2007). Further studies have shown thatater extracts from the River Lambro induced feminisation andemasculination in the South African clawed toads Xenopus lae-is (gonad and aromatase P450 endpoints; Cevasco et al., 2008;assari et al., 2010) while an early study reported intersex barbel
arbus plebejus derived from this river (Viganò et al., 2001). Suchffects likely reflect responses to the mixture of oestrogens andnti-androgens present in this river (Urbatzka et al., 2007; Viganòt al., 2008). In studies on mosquitofish (Gambusia affinis holbrooki)opulations in Lake Apopka (USA), males have been found withhorter gonopodia and reduced sperm counts when compared withsh from reference lakes and contaminants with anti-androgenicotential in this lake have been implicated in causing these effects,otably p,p′-DDE (Toft et al., 2003).
Assessing for anti-androgenic effects in fish requires specificiomarkers. Many fish species have androgen dependent sex traitssuch as secondary sexual characters), although most of these canlso be affected (inhibited) by exposure to oestrogens (e.g. Ankleyt al., 2010). Sex hormone dependent behaviours have been usedor assessing the impacts of individual EDCs (Jones and Reynolds,997; Kane et al., 2005; Scott and Sloman, 2004). Examples of this
nclude a study on mosquitofish where exposure of male fish to aaper mill effluent in Florida resulted in greater aggressiveness (anndrogen dependent behaviour) in their courtship display (Howellt al., 1980). Contrasting with that study, Bortone et al. (1989) found
ology 105 (2011) 78– 88 79
no difference in behaviour of mosquitofish exposed to a kraft milleffluent, even though females showed morphological signs of mas-culinisation. The same observation (morphological changes but noeffects on behaviour changes) were reported for mosquitofish liv-ing in pollutant contaminated lakes in Florida (Toft et al., 2003). In astudy with fathead minnows (Pimephales promelas), males exposedto an oestrogenic STW effluent for three weeks failed to compete fornesting sites and females against control males, but they spawnedsuccessfully when no competition was involved (Martinovic et al.,2007).
The three-spined stickleback (Gasterosteus aculeatus) offersa number of useful features for studying the effects of (anti-)androgenic EDCs. The first of these is the production of a glueprotein, spiggin, that is specifically dependent on androgens(Björkblom et al., 2009; Hahlbeck et al., 2004; Jolly et al., 2006;Katsiadaki et al., 2002a,b; Sanchez et al., 2008) and inhibited byanti-androgens (Jolly et al., 2006; Katsiadaki et al., 2006; Sebireet al., 2008, 2009). In addition to spiggin production, the repro-ductive behaviour in male sticklebacks is well described and underandrogen control (Borg and Mayer, 1995; Mayer et al., 2004; Sebireet al., 2007; Wai and Hoar, 1963). A series of previous studieshave successfully employed reproductive behaviour of male stick-lebacks to assess effects of thyroid disrupting chemicals (Bernhardtand von Hippel, 2008), oestrogens (Bell, 2001; Brian et al., 2006;Wibe et al., 2002) and anti-androgens (Sebire et al., 2008, 2009). Inthe two latter studies, exposure to the model anti-androgen, flu-tamide, and to the anti-androgenic pesticide, fenitrothion, resultedin reduced nest building activity and reduced male courtshipbehaviour towards the female. In this study we investigated theeffect of exposure to an anti-androgenic (and weakly oestro-genic) treated sewage effluent on the reproductive physiology andbehaviour of sticklebacks.
2. Materials and methods
2.1. Animals
The parents of the experimental broods were originally obtainedfrom a wild population from a site in Dorset (WinterbourneHoughton). We regularly use parents from this site as the watersupply is of high environmental quality with very low levels of met-als, PAHs and no endocrine activity. Two broods were used for thisstudy (brood 1 and brood 2) and were raised to adulthood in thelaboratory under a photoperiod of 12L:12D and a temperature of16 ± 2 ◦C. For the breeding test employed in this study, the sepa-ration of adult fish into males and females is vital as the holdingconditions prior to the test are different for the sexes. At the startof the test the females need to be gravid and ready to spawn, whilethe males need to be in a non-breeding condition. Since stickle-backs lack sexual dimorphism outside the breeding season, the sexwas determined using a molecular sex probe.
Genetically identified males from brood 1 were moved to a hold-ing tank, under winter conditions to avoid the onset of breeding(8L:16D and 10 ± 2 ◦C) and kept for 4 months prior to the exposure.The females, originating from brood 2 – in order to avoid siblingeffects on reproductive behaviour (Frommen and Bakker, 2006),were kept at 16 ± 2 ◦C under a summer photoperiod of 16L:8D hoursduring the same period. The fish were fed daily with frozen blood-worm and any remaining debris was removed from the tanks. Onlyfish weighing ≥0.8 g were used in the experiment.
2.2. Sex determination of the fish
When the fish were six months old, and before the develop-ment of any external secondary sexual characters, the sex of thefish was identified by the use of sex-linked DNA-markers (Peichel
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t al., 2004). Two dorsal spines were clipped from each fish and DNAas extracted from them using DNAzol reagent. The PCR proce-ure used to detect the presence of sex-specific genetic sequencesas adapted from Peichel et al. (2004) and the specific primers
mplified a 302 bp fragment in females and two (302 bp and 271 bp)ragments in males.
.3. Site selection
The treated sewage effluent was derived from STWs in Devon,K. The STW has a population equivalent (PE) of approximately4,585 and 43.1% of the influent was derived from industrialources (in practice, it is assumed that 1 PE unit equals to 54 g ofiochemical oxygen demand per day). The treatment process at theTWs comprised of a primary settlement followed by a secondaryiological filter (there was no tertiary treatment). The effluent waselected for use as it had previously been shown to contain a highnti-androgenic activity and a low level of oestrogenic activityusing the rYAS and recombinant yeast oestrogen screen – rYESassays, respectively; I. Johnson et al., 2007). The treated sewage
ffluent for the exposure was delivered to the laboratory twice aeek and maintained at a temperature of 7.9 ± 0.1 ◦C in a chilled
ontainer.
.4. Exposure
Randomly chosen groups of 5 males and 8 females per tankere exposed to each of the treatments for a period of 21 days.
ach experimental tank comprised of six interconnected compart-ents; five of these were small, housing one male fish each and
he sixth compartment was larger housing the eight communallyeld females. The males were physically and visually separated
rom each other but they all had visual access to the female com-artment (for a schematic representation of the tank design seeebire et al., 2008, 2009). The treatment tanks were supplied with0 and 100% (v/v) STW effluent (2 replicate tanks for each treat-ent), the former of which was produced by mixing the effluentith dechlorinated water. Two additional tanks were supplied withechlorinated water only as controls. The flow rate through eachf the tanks was 75 ml min−1 (2.5 changes/24 h), and flow rate andater temperature were monitored daily. One litre of water was
ampled from each tank once a week over the three-week expo-ure period and from the collected treated STW effluent samples oneceipt of the effluent to the laboratory, to assay for anti-androgenicnd oestrogenic activity (using rYAS and rYES). There were 7 efflu-nt deliveries over the 21-day exposure period.
Fish were transferred to the experimental tanks two days prioro the start of the experiment (day −2, 8D:16L, 17 ± 1 ◦C). Theffluent was connected to the flow through exposure system theollowing day (day −1). The photoperiod was then adjusted to6L:8D 24 h post effluent delivery, marking the onset of the experi-ental period (day 0). Ten males from the stock population (brood
) were sacrificed on day 0 in order to assess levels of spiggin andence breeding status in the study population.
.5. Behavioural trials
The behavioural trial methodology has been described else-here (Katsiadaki et al., 2007; Sebire et al., 2008, 2009) and is based
n assessing the two first phases (nest building and courtship)f the reproductive cycle in the male stickleback. Briefly, theehavioural attributes studied were nest-building behaviour (dig-
ing activity and nest presence) from day 10 (after the introductionf nest material and gravel) to day 21 (study termination); andourtship behaviour of the male (bites, zigzag movements, prox-mity to the female, dorsal pricking) over a 15-min encounter with
ology 105 (2011) 78– 88
the female on day 19 (video recorded) and 20 (direct observationonly). During the mating, leading behaviour by the male towardsthe female was also recorded (control males show this activitywithout the presence of a nest; our own personal observations).Although the behavioural assay focused on the exposure effects onmale reproductive behaviour, some behavioural parameters werealso recorded for the females, including the ‘head-up’ posture (char-acteristic signalling of the readiness of the female to spawn), thetime spent in a motionless position, whether the female followedthe male to the nest, and if she spawned. The number of eggs laidwas also counted.
2.6. Biomarker measurement
All fish were sacrificed humanely after 21-day exposure inaccordance with UK Home Office regulations, and measured forlength and weighed to the nearest mg. Blood was collected via cau-dal severance using a pre-heparinised micro-haematocrit capillarytube, transferred to an Eppendorf® tube, and placed on ice. Wholeblood samples were centrifuged at 13,000 rpm for 5 min at 4 ◦C;the supernatant was drawn off and stored at −80 ◦C until deter-mination of vitellogenin (VTG) by enzyme-linked immunosorbentassay (ELISA) (Hahlbeck et al., 2004; Katsiadaki et al., 2002b). Thekidneys were dissected out to allow measurement of spiggin viaELISA (Katsiadaki et al., 2002a). On day 0, 10 males from the pop-ulation used to supply the exposure study were also sacrificed andtheir spiggin levels measured to estimate the breeding status ofthe fish population prior to the test. Ovaries were also dissected andweighed to assess for reproductive development status by calculat-ing the gonadosomatic index (GSI, gonad mass/body mass × 100).Testes were not weighed, since testes’ mass is low and is not a goodindicator of sexual maturation in the male stickleback (Borg, 1982).
2.7. Solid phase extraction, chemical analysis and recombinantyeast screens
Water and effluent samples (1 L) collected for analyses werefirst filtered (4–12 �m pore size, Millipor filter disc) via a vac-uum pump to remove any large particles. The filters were rinsedwith methanol (10 ml) into the pre-filtered samples. The sampleswere then concentrated onto methanol primed C18 solid phaseextraction cartridges as detailed previously in Sebire et al. (2008,2009) and stored at −20 ◦C. The cartridges were eluted with 5 mlmethanol, which was then evaporated to dryness under a streamof nitrogen at room temperature before the extracts were re-suspended in 1 ml of absolute ethanol and stored at −20 ◦C priorto analysis.
The effluent samples were analysed for anti-androgenic andoestrogenic activities using the rYAS and rYES, respectively(Routledge and Sumpter, 1996; Sohoni and Sumpter, 1998). Yeastassays were performed in a type II laminar air flow cabinet. Briefly,125 �l of yeast (from the yeast stock stored at −20 ◦C) was added tothe growth medium and grown on an orbital shaker for about 24 hat 28 ◦C. The following day, samples were diluted in a 1:2 seriesin ethanol for a total of 12 concentrations, and 10 �l aliquots ofeach concentration transferred in duplicate to an optically flat 96-well microtiter plate (Linbro/Titertek; ICN FLOW, Bucks, UK). Theplates were then left at room temperature to allow the ethanolto evaporate. Aliquots (200 �l each) of assay medium (containingthe recombinant yeast and the chromogenic substrate chlorophe-nol red-�-D-galactopyranoside – CPRG, yellow in colour) weredispensed into each sample well. The plates were sealed with auto-
clave tape, shaken for 2 min, and then incubated at 32 ◦C (rYES)or 28 ◦C (rYAS). After an incubation period of 2 (rYAS) or 3 days(rYES), colour development (yellow to red) in the medium was mea-sured at an absorbance of 540 nm, and turbidity of the yeast (as a
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Table 1Flutamide equivalent (�g l−1) and oestradiol equivalent (ng l−1) in the STW effluent-treated tanks.
easure of cell viability) was measured at 620 nm using a Spec-ramax Plus, microtiter plate reader (Molecular Devices, Berkshire,K). Untreated cells (blank) and a solvent control were included
n the experimental design. Data were corrected for turbidity tonsure that the change of colouration of �-galactosidase (in CPRG)as caused by true agonistic/antagonistic responses and not due
o cytotoxicity; the correction was done as follows: correctedalue = chemical absorbance at 540 nm − (chemical absorbance at20 nm − blank and solvent controls mean absorbance at 620 nm).
Dihydrotestosterone (DHT), flutamide (FL), 17�-oestradiol (E2)nd 4-hydroxytamoxifen (HT) used as standards in the recombi-ant yeast assays were purchased from Sigma Chemical Co. Ltd.Dorset, UK). DHT is a model androgen and FL a model anti-ndrogen, both binding to the human androgen receptor integratednto the yeast, while E2 is a model oestrogen and HT a modelnti-oestrogen, binding to the human oestrogen receptor. For thecreening of agonist activity, a dilution series of E2 (stock con-entration 2 × 10−7 M; concentration range from 1 × 10−8 M to.88 10−12 M) was used as positive control for oestrogenic activ-
ty and DHT (stock concentration 2 × 10−6 M; concentration rangerom 1 × 10−7 M to 4.88 × 10−11 M) for androgenic activity. Doseesponse curves were determined for each plate and the con-entrations of the samples were calculated by comparing theestrogenic/androgenic activity of the sample extracts with theeries concentrations of E2/DHT standards, respectively.
For the screening of antagonistic activity, a standard curvef a constant concentration of DHT at 1 × 10−7 M (75% effectiveoncentration) or E2 at 1 × 10−8 M coupled with a serial dilu-ion of FL (stock concentration 2 × 10−3 M; concentration rangerom 1 × 10−4 M to 4.88 × 10−8 M) or HT (stock concentration
× 10−4 M; concentration range from 1 × 10−5 M to 4.88 × 10−9 M)n rYAS or rYES, respectively, were transferred to microtiter plates.ose response curves were determined for each plate and theoncentrations of the samples were calculated by comparing thenti-androgenic/anti-oestrogenic activity of the sample extractsith the series concentrations of FL/HT standards, respectively.
.8. Statistical analyses
All models were fitted in R version 2.12.2 (R Development Coreeam, 2011). Spiggin levels on day 0 and control fish were com-ared using a T-test to validate the assay by ensuring spiggin levelsere significantly lower in the stock population males prior to com-encing the exposure. The effect of exposure to different levels of
TW effluent on the behavioural and physiological outcomes stud-ed was investigated through the use of linear or generalised linear
ixed models (LMMs or GLMMs), where STW effluent exposureas modelled as a fixed effect, and Tank was modelled as a ran-om effect in order to adjust for pseudo-replication occurring dueo multiple measurements being taken within a tank. Continuousutcome variables (spiggin, VTG, condition factor, GSI, proximity tohe female, zigzags duration, dorsal pricking duration and female
otionless position) were modelled assuming that data were nor-ally distributed (using the lme function in the nlme library).here necessary, data were first transformed in order to normalise
he data and homogenise variances. Count data (nest-building starteriod, zigzags frequency, zigzags bouts, bites frequency and dor-al pricking frequency) were analysed assuming the data wereither poisson (using the lmer function in the lme4 library) oregative binomially distributed (using the glmm.admb function inhe glmmADMB library) depending on the level of over-dispersionbserved in the dataset. Binary outcomes and proportions (digging,
est presence, leading and female head-up posture) were againodelled using a GLMM that this time assumed a binomial error
istribution (using the lmer function in the lme4 library). The effectf each level of exposure to sewage treated water was compared
ology 105 (2011) 78– 88 81
to the control group in each model. Differences where the p-valuewas ≤0.05 were considered statistically significant. In addition toexamining behaviours individually, a principle components analy-sis (PCA) was conducted in MVSP version 3.1 (Kovach ComputingServices, Pentraeth, UK) to summarise the data into the dominantbehavioural combinations observed. Data were centred and stan-dardised (thus analysis used a correlation matrix), and case scoresfor individuals in each treatment group were plotted and comparedusing a LMM for each axis with an eigenvalue greater than one.Eight courtship behaviours observed for 39 male sticklebacks weresummarised using this method.
3. Results
3.1. Survival, condition factor and gonadosomatic index
In replicate 1 of the 50% treated STW effluent exposure, therewas one male mortality on day 14 and this fish was excluded fromall data analysis (N = 9 males for this treatment). The condition fac-tor of males and females in the 50% and 100% STW effluent groupsdid not differ from the control groups (control vs. 50%: p = 0.13,control vs. 100%: p = 0.06). Similarly, in females, the GSIs were14.49 ± 1.47, 12.42 ± 0.8 and 14.99 ± 1.86% for controls, 50% and100% STW effluent groups, respectively (control vs. 50%: p = 0.46,control vs. 100%: p = 0.85).
3.2. Hormone activity in the exposure effluent – rYAS and rYES
The freshly delivered treated effluent exhibited anti-androgenicactivity (mean measured, 328.56 ± 36.83 �g FLeq l−1) and a lowlevel oestrogenic activity (mean measured, 3.32 ± 0.66 ng E2eq l−1).The limit of detection (LOD) determined for the anti-androgenicactivity was 29.6 �g FLeq l−1 and the LOD for the oestrogenicactivity was 0.36 ng E2eq l−1. The hormonal activities in theexperimental tanks are presented in Table 1. No androgenic or anti-oestrogenic activities were found in any of the effluents samples(data not shown).
3.3. Biochemical marker responses
The plasma concentrations of VTG in male sticklebacks were lowfor the control and 50% STW effluent groups with 1.59 ± 0.79 and1.75 ± 0.86 �g ml−1, respectively (p = 0.85). In the 100% STW efflu-ent group, the mean VTG concentrations was 40.8 ± 31.8 �g ml−1,
Rep 1, replicate 1; Rep 2, replicate 2; LOD, limit of detection (29.6 �g FLeq l−1 and0.36 ng E2eq l−1).
a The low level anti-androgenic activity may be due to a contamination from a100% STW effluent tank that was situated immediately above that control tank.
82 M. Sebire et al. / Aquatic Toxicology 105 (2011) 78– 88
Fig. 1. Spiggin units/g body weight in male stickleback exposed to a treated sewageeffluent. N = 10 males in all treatments except for the 50% treated sewage efflu-ent (N = 9). 50% and 100% represent the sewage effluent at 50% dilution and at fullstrength effluent, respectively. Box-whisker plots represent percentiles; the medianline represents the 50th percentile of the y-value range with the box borders beingthe 25th and 75th percentiles. Whiskers extend from the box to the upper and loweradjacent values that are calculated utilising the interquartile range (1.5 × [75th val-ua*
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Fig. 2. Nest-building activity exhibited by male sticklebacks after 21 days of expo-sure to a treated sewage effluent. Number of males showing digging activity (light
Axis 2 explained 19% of the behavioural differences between indi-viduals. In this axis there was a negative correlation between thevariables ‘zigzags’ and ‘bites’, ‘leading’, ‘dorsal pricking’ and ‘prox-
Table 2Principal component analysis variable loadings, eigenvalues and percentage vari-ability explained by each axis for male stickleback reproductive behaviours.
Axis 1 Axis 2
Proximity to the female (duration) 0.26 −0.31Bites −0.04 −0.51Dorsal pricking (frequency) 0.35 −0.44Dorsal pricking (duration) 0.37 −0.34Number of zigzag bouts 0.43 0.30Total number of zigzags 0.43 0.30Zigzag bouts (duration) 0.41 0.35
es − 25th values]). Black dots are significant outliers – values exceeding the uppernd lower adjacent values – within the data. Significant differences from the control,
*p < 0.01.
ontrol, 50% and 100% STW effluent groups, respectively, with noignificant differences between them (control vs. 50%: p = 0.07, con-rol vs. 100%: p = 0.24).
Spiggin levels in the male fish sacrificed on day 0 were signif-cantly lower than those in control males sacrificed at the end ofhe experiment (t = 5.62, df = 10.96, p = 0.0002), showing that thehotoperiodic manipulations were successful in both keeping stockales in a non-breeding stage and inducing breeding condition in
xperimental males during the study (Fig. 1). There was no signifi-ant difference in spiggin levels between control and exposed fishn both the 50% and 100% STW effluent exposed groups (control vs.0%: p = 0.81, control vs. 100%: p = 0.22; see Fig. 1). Spiggin levels
n females did not differ between the treatments and controls andere low – up to 145 units – across all treatments (control vs. 50%:
= 0.15, control vs. 100%: p = 0.29); this confirmed that there waso significant androgenic activity in the sewage effluent.
.4. Nest-building behaviour
Most of the control males started building a nest on averageithin 4.7 days following the introduction of nest material and
ravel, while the exposed males started on average within 5.8 daysnd 8 days for 50% and 100% STW effluent groups, respectively.his delay in the onset of nest building was significant for the fulltrength STW effluent treatment (control vs. 50%: p = 0.33, controls. 100%: p = 0.01).
Although exposure to the STW effluent did not significantlyffect nest building behaviour overall (control vs. 50%: p = 0.51, con-rol vs. 100%: p = 0.17), it should be noted that there were fewer
ales showing digging activity and completing a nest in the fulltrength effluent exposure group (Fig. 2).
.5. Courtship behaviour
The females presented the head-up posture initially in all thereatments (control vs. 50%: p = 0.51, control vs. 100%: p = 0.34);his was in accordance with their breeding status (well developedvaries indicated by the high GSI), showing that the females were
grey) and/or had a nest (dark grey). N = 10 males in all treatments except for the 50%treated effluent (N = 9). 50% and 100% represent the sewage effluent at 50% dilutionand at full strength effluent, respectively.
in breeding condition and ready to spawn. Although there weresignificant differences in the female activity (percentage of timemotionless) between replicate tanks for the control and 50% efflu-ent treated groups, overall the sewage exposed females were lessactive. In the 50% and 100% STW effluent exposures, the femalesspent 29.1 ± 6.2% and 30.9 ± 2.5% in a motionless position (com-pared with 5.8 ± 1.1% in controls) during their 15 min encounterwith the males. This reduced activity was statistically significant forthe 50% effluent exposure group (p = 0.0006), and the 100% effluentexposure group (p = 0.0001). Across all treatments, when a male ledthe female to the nest (or an empty pit), the female almost invari-ably followed (in total 33 follow displays over 40 leading displays).However, in the 100% STW effluent treatment, no males displayedleading behaviour and thus for this group inevitably the femalesnever followed the males (p = 0.04). Only one female spawned inthe whole experiment (control group, replicate 2), which occurredon day 20 with a batch size of 58 eggs.
PCA summed the variability in behavioural activities displayedbetween individual sticklebacks into two axes with eigenvaluesgreater than one. These two axes explained 74% of variability inbehavioural activities observed between individuals with axis 1explaining 55% of this variability. Axis 1 related to the general levelof courtship behaviour displayed by individuals, with all courtshipbehavioural activities being positively correlated (i.e. individualsthat performed a lot of one behaviour would perform a lot of all theother behaviours assessed) with the exception of biting (aggressivecomposite), which made little contribution to this axis (Table 2).
Fig. 3. Principal component analysis case scores for the dominant two axes sum-marising reproductive behaviours displayed by male sticklebacks in different STWtreatment groups during the courtship phase on day 19. Controls (C) are indicatedby triangle symbols, and the exposed groups at 50% and 100% STW effluent areindicated by circle and square symbols, respectively. Repartition ellipses of data areindicated for each treatment. Behavioural combinations associated with each axisscore are presented in Table 2.
Fig. 4. Courtship behaviour towards a gravid female by males exposed to treated sewage(where N = 9). 50% and 100% represent the sewage effluent at 50% dilution and at full strline represents the 50th percentile of the y-value range with the box borders being the
adjacent values that are calculated utilising the interquartile range (1.5 × [75th values −
lower adjacent values – within the data. Significant differences from the control, *p < 0.05
ology 105 (2011) 78– 88 83
imity to the female’, meaning that individuals with scores varyingalong this axis did more of one set of behavioural activities thanthe others. Plotting the individual case scores by treatment groupagainst each axis showed that control fish exhibited a wider rangeof behavioural activities in relation to axis 1 than either of the STWtreatment groups (Fig. 3). No obvious differences between groupswere present in relation to axis 2.
The LMM confirmed a significant difference in score betweengroups with regard to axis 1 (F = 10.43, df = 2, p = 0.0005), and that adose response was present with the 50% STW effluent group havinga lower average score than controls (slope = −0.31951, se = 0.12999,t = −2.458, p = 0.0210) and the 100% group having a lower score still(slope = −0.57673, se = 0.1265, t = −4.558, p = 0.0001). No significantdifferences were detected between groups with regard to axis 2(F = 1.701, df = 2, p = 0.2022).
Fig. 4 portrays the results for the main behavioural variablesindividually. No significant differences were found between treat-ments regarding the time a male spent near to a female (controlvs. 50%: p = 0.85, control vs. 100%: p = 0.63) and the frequency oftheir bites towards her (control vs. 50%: p = 0.10, control vs. 100%:p = 0.076), although a slight increase in the number of bites was
observed in the effluent treated groups (Fig. 4). Males exposed toboth STW effluent concentrations showed reduced zigzag move-ments (zigzag duration: p = 0.015 at 50% and p = 0.0005 at 100%;zigzag bouts: p = 0.075 at 50% and p = 0.0009 at 100%; zigzag fre-
effluent (day 19). N = 10 males in all treatments except for the 50% treated effluentength effluent, respectively. Box-whisker plots represent percentiles; the median25th and 75th percentiles. Whiskers extend from the box to the upper and lower25th values]). Black dots are significant outliers – values exceeding the upper and
and **p < 0.01.
8 Toxic
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4 M. Sebire et al. / Aquatic
uency: p = 0.0548 at 50% and p = 0.0028 at 100%; see Fig. 4).lthough the differences in the number of zigzag bouts and fre-uency were not statistically significant for the 50% STW effluentt the p = 0.05 cut-off, the observed p-values were very smallespecially given the relatively small sample size), suggesting thebserved differences were not chance observations. Males in theull strength effluent did not display the dorsal pricking activityhat was observed in control tanks (p = 0.031 for dorsal prickinguration and p = 0.002 for dorsal pricking frequency). No statisti-ally significant differences were observed in males exposed at 50%TW in terms of the frequency and duration of dorsal pricking activ-ty (p = 0.41 and p = 0.48 respectively). There was also no significantifference in the number of males leading females to the nestingrea between control and 50% exposed tanks (p = 0.25). However,o males showed this leading behaviour in the full strength STWffluent exposure replicates (p = 0.079). Although this result wasot significant at the p = 0.05 cut-off, an observation of this natureould only be expected by chance in less than 8% of studies.
. Discussion
The most important outcome of this study is that exposureo a STW effluent with anti-androgenic activity (determined viaYAS) and a low oestrogenic activity (determined via rYES) affectedhe reproductive behaviour of the three-spined stickleback in thebsence of any significant effects on VTG or inhibition of spigginroduction (biomarkers for oestrogen and anti-androgen exposure,espectively). The higher sensitivity of reproductive behaviour inomparison to spiggin responses has been observed previously inale sticklebacks exposed to the model anti-androgen flutamide
Sebire et al., 2008) and to the anti-androgenic organophosphateerbicide, fenitrothion (Sebire et al., 2009). In previous studies, the
owest observed effect concentration (LOEC) for spiggin reductiony flutamide (FL) exposure was 50 �g l−1 in non-photoperiodicallytimulated males (Katsiadaki et al., 2006) and 500 �g l−1 in breed-ng males (Sebire et al., 2008), as studied in this experiment. Thisifference is most likely due to the endogenous androgens thateak a few days after photoperiodic stimulation of the sticklebackreeding cycle (Borg and Mayer, 1995). Sebire et al. (2008) exposedticklebacks to 100 �g l−1 FL, a level of anti-androgenic activityimilar to that measured in the full strength STW effluent in thexposure tanks in this study (average 125.18 and 146.59 �g FLeq l−1
or replicates 1 and 2, respectively), and found no effects onpiggin but both nest-building and courtship behaviours wereffected. For VTG induction, the reported effective concentrationsor 17�-oestradiol in sticklebacks for a 3 week exposure rangeetween 20 ng l−1 and 50 ng l−1 (Allen et al., 2008). These lev-ls were much higher than the E2eq measured in the exposureanks in this study (3.32 ± 0.66 ng E2eq l−1) and thus this biomarkeresponse is again consistent with previous findings for relevantxposures.
The rYAS is regularly being employed to quantify anti-ndrogenic activity in environmental samples and this is resultingn an increasing number of reports indicating evidence for the
idespread presence of anti-androgenic activity in dischargedastewaters (Hill et al., 2010; I. Johnson et al., 2007; Thomas et al.,
009; Tollefsen et al., 2007; Urbatzka et al., 2007). Spiggin reductionn breeding male sticklebacks however, has been shown to occurt FLeq exposures exceeding 100 �g l−1, emphasising the need for
better understanding on the relationship between FLeq (as deter-
ined in the rYAS, employing the human androgen receptor) and
nti-androgen specific responses in fish. Data of this nature willn turn enable more meaningful interpretations on the biologicalignificance of FLeq levels derived from the rYAS in fish.
ology 105 (2011) 78– 88
The variable nature of the STW effluent delivered to the labora-tory with respect to the anti-androgenic and oestrogenic activitiesmeasured (see Table 1) may be due to differences in the influentto the STW from the industrial discharges over the study period,and/or differences in degradation rates of the components due tomicro-organism activity, or other factors (Panter et al., 1999; Tylerand Jobling, 2008). In the experimental tanks, natural steroids fromthe fish may have contributed to the measured oestrogenic activ-ity, as has been reported previously in other fish species (fatheadminnow, Thorpe et al., 2009).
It was not the purpose of this study to identify the causativechemicals, responsible for, or at least contributing to the anti-androgenic activity in the test STW effluent. However, an analysisundertaken for a spectrum of specific chemicals on one of thebatches of effluent used identified two phthalates: di-n-butylphthalate (DBP), at a relatively high concentration of 40 �g l−1, anddiethyl phthalate (DEP) at 1.23 �g l−1), both of which are knownto have anti-androgenic effects in mammals (Ema et al., 2000;Gray et al., 1999; Howdeshell et al., 2007; Mylchreest et al., 1999;Sharpe, 2001) as well as weak oestrogenic effects in vitro (Joblinget al., 1995; Zacharewski et al., 1998). In fish the effect of phtha-late exposure has not been investigated thoroughly but a recentstudy showed that DBP at 35 �g l−1 reduced spiggin levels in malethree-spined sticklebacks (Aoki et al., 2011). It would therefore beexpected that the STW effluent in this study (that at least at a sin-gle point of analysis contained 40 �g l−1 DBP) could affect spigginlevels, but this was not the case. However, it is widely acceptedthat the anti-androgenic effect of phthalates in mammals is notmediated via the androgen receptor but from their ability to dis-rupt steroidogenesis in early development (Gray et al., 2000; Foster,2006). Hence, the differences in exposure system (life stage andduration) between our study and this of Aoki et al. (2011) can fullyaccount for the differences in spiggin responses. Comparisons oneffect concentrations for specific chemicals when alone or as partof complex mixtures, are in general complicated by the fact that themeasured concentrations derived from a complex mixture such asan effluent at one particular point in time do not necessarily reflectwhat is present or bioavailable.
Overall the STW effluent tested affected the reproductivebehaviour of male sticklebacks (reduced nest building activities,reduced zigzag movements, affected dorsal pricking display andleading activity). High levels of 11-ketotestosterone are stronglyassociated with the nest-building and courtship phases (Borg andMayer, 1995; Borg et al., 1987; Páll et al., 2002a,b; Sebire et al.,2007). Both aggressiveness and territoriality in the stickleback havebeen shown to be under control of androgens (Baggerman, 1990;Hoar, 1962; Rouse et al., 1977) as observed in other seasonal repro-ducing species (review by Oliveira et al., 1999). As androgens playa key role in the control of the nest-building, aggressiveness andcourtship behaviours, anti-androgens were more likely to directlyaffect these behaviours than other endocrine disrupter chemicalspresent in the STW effluent.
In our analyses on male behaviours, the dorsal prickingbehaviour was based on assessing the up and down movementsof the basal musculature of the spines, as the spines themselveshad been removed to extract DNA for sex genotyping. Removalof the spines may have affected the normal response for thisbehaviour feature; dorsal pricking is used to keep the female dis-tant while the male is performing nest activities – maintenanceof the nest (Wilz, 1970). When we compared our data for dorsalpricking in this study with a previous study of essentially the samedesign (equal study duration, identical breeding matrix) where the
spines were left intact, the results for controls are highly compa-rable; dorsal pricking incidence occurred 4.10 ± 2.18 times in thestudy by Sebire et al. (2008) and 6.60 ± 2.07 times in this study(p = 0.28).
It is entirely possible that the impairment of male reproduc-ive behaviour observed in this study may have included effectsf other physicochemical features of the effluent, rather than aonsequence of the anti-androgens/oestrogens alone. Turbidity ofhe effluent is perhaps one of the most likely factors that mayave contributed to the effects seen on the reproductive behavioury altering the ability of the fish to send/receive visual cues thatre fundamental for mating, stimulating secondary sexual char-cters (Reisman, 1968) or spawning (Chien, 1973) (see reviewy Rowland, 1999). This is particularly important considering the
mportance of visual signalling in the stickleback; it has been showno be a key factor in mate choice for both females and malesBarber et al., 2000; Rowland, 1982, 1995, 1999). Although no direct
easures of light/spectral penetration through the effluent wereonducted, the ability of the fish to see in the aquaria containingffluent would have been reduced notably. Total suspended solidsTSS) measurements (using Secchi disk method) on the effluenttudied are conducted regularly by the company that supplied theTW effluent and over the exposure period, the overall TSS was2.47 ± 0.28 mg l−1 (50–107 measurements/day, averaged over 12ays). In clear water the TSS are assuming to be <20 mg l−1. Foromparison with other studies, one TSS unit is estimated to bequivalent to 1.0–1.5 NTU (turbidity measured with a nephelome-er – Nephelometric Turbidity Units) in static conditions such asor a lake; thus here for our study the NTU values for the efflu-nt would be somewhere between 22.47 ± 0.28 and 33.71 ± 0.42.n the present system, the conditions were not static and the NTUherefore might be higher. During a UK catchment study in a rivertretch receiving effluent discharges, studied over a period of 2ears (2007/2008) the average turbidity near the STW effluent dis-harges was more than double that of the reference river, exceeding00 NTU in both successive years (Pottinger et al., 2011). Such tur-idity conditions would affect the reproductive behaviour of maleticklebacks in the wild. In Newcombe and Jensen (1996), empiricalquations of a meta-analysis on fish response to suspended sedi-ent in estuaries showed that an exposure at 20 mg l−1 TSS during
days was linked with harmful behavioural effects such as reduc-ion of alarm reaction and abandonment of cover. Furthermore,ngstrom-Öst and Candolin (2007) noticed that male sticklebackshat were reared in clear water (1.1 ± 0.2 NTU), and so wereot adapted to a turbid-type environment, showed no courtshipzigzag dance) behaviour under turbid conditions (51.4 ± 9.4 NTU).n dense vegetation, male sticklebacks have been shown to have aonger latency before starting nest building (Candolin and Salesto,006). Studies on sticklebacks in the Baltic Sea have observed thateduced visibility (as a consequence of algal turbidity) weakenedale competition (males were not able to see each other) and pro-oted dishonest sexual signalling and reduced the evolutionary
otential of sexual selection (Candolin et al., 2007; Wong et al.,007). In other species where visual cues are also important, tur-idity has been shown to have an adverse effect on reproduction by
nterfering with mate choice (dull body colouration and few colourorphs in populations of cichlids in Lake Victoria; Seehausen et al.,
997) and by reducing selectivity towards females of male sail-n mollies Poecilia latipinna (Heubel and Schlupp, 2006) and maleipefish Syngnathus typhle (Sundin et al., 2010).
Regardless of the mechanism(s) involved in the alteration of sig-ificant features of the male stickleback reproductive behaviourchemicals or/and turbidity), the important fact is that exposureo the STW effluent did affect the expression of vital reproduc-ive parameters. Males with impaired courtship behaviour wouldlmost certainly fail to mate. It is unknown whether this effect
s permanent (i.e. the exposed fish would never recover whilexposed) or transient (i.e. the exposed males were simply laggingehind the control males and might have resumed a successfulreeding cycle given a longer period). The latter hypothesis seems
ology 105 (2011) 78– 88 85
to be supported by both the delay in the onset of nest building andthe apparent (although not significant at 5%) aggressiveness (num-ber of bites) seen towards the females during mating observed inthe effluent exposed male fish.
The implications of an impaired or delayed reproductive perfor-mance by the male stickleback due to effluent exposure may havesignificant effects in wild stickleback populations. In some UK riversthe flow is often comprised of at least 10% treated STW effluent.However, during the late spring and summer months (when nat-ural breeding of sticklebacks occurs) the flow of rivers can consistof more than 50% STW effluent (Jobling et al., 1998; Johnson, 2010;Tyler and Jobling, 2008), creating an unfavourable environment forstickleback breeding.
5. Conclusion
In summary, we cannot draw firm conclusions regarding themechanisms of effect for the STW effluent on male sticklebackreproductive behaviour, but it seems that both endocrine disrup-tion mechanisms and the physical features of the effluent wereinvolved together highlighting again the complexity of mixtureeffects of STW effluent. Overall, exposure to STW effluent clearlyhad an adverse affect on the reproductive behaviour of the malethree-spined stickleback and impaired reproduction. These datafurther support the use of stickleback’s reproductive behaviour as asensitive endpoint for assessing environmental chemical exposure.
Acknowledgments
The study was funded by the Department for Environment, Foodand Rural Affairs (Defra, Contract number CT20051). The Universityof Exeter and NERC (NE/E017363/1) supported Exeter personnel inthis work. We greatly acknowledge Jan Shears (University of Exeter,UK) for providing technical help. We also thank Dr. Catherine L.Peichel (Fred Hutchinson Cancer Research Center, USA) for provid-ing the DNA markers and Dr. David M. Stone (Cefas WeymouthLaboratory, UK) for his help with the sex determination analyses.
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