2003-生命早期接触芳香族抑制剂导致雌鱼雄性化和不可逆性逆转

Aquatic Toxicology 65(2003)309–316

Early life exposure to

levels of aromatase inhibitor tributyltin causes masculinisation and irreversible

sperm damage in zebra?sh (

Danio rerio )

Brian G.McAllister,David E.Kime ?

Department of Animal and Plant Sciences,University of Shef?eld,Western Bank,Shef?eld,S102TN,UK

Received 14November 2002;received in revised form 8June 2003;accepted 9July 2003

Abstract

To determine whether early life exposure to tributyltin (TBT),an aromatase inhibitor,impaired reproductive function in ?sh,Danio rerio were exposed to environmentally realistic levels (0.01–100ng l ?1)of TBT from 0to 30,30to 60,and 0to 70days post-hatch,and the sex ratio and sperm motility of the adults examined 3–5months after cessation of exposure.Fish exposed for 70days to 0.1ng l ?1of TBT,a concentration presently below the detection limit in water,showed a male biased population which produced a high incidence of sperm lacking ?agella.At 1ng l ?1,the motility of sperm was signi?cantly lower than that of control ?sh,while at 10ng l ?1,all sperm lacked ?agella and,at 100ng l ?1,milt volume had increased.The effect of exposure on sex ratio was similar after exposure from 0to 70and 0to 30days,but even 100ng l ?1gave only 65%males after exposure from 30to 60days.Effects on sperm motility and morphology and on milt volume were less pronounced after 30day than 70day exposure.Our data suggest that screening for aromatase inhibiting activity and assessment of its risks in early life to human and wildlife fertility needs to be urgently addressed,and that the reproductive toxicity of TBT may presently be underestimated.?2003Elsevier B.V .All rights reserved.

Keywords:Tributyltin;TBT;Aromatase inhibitor;Sperm;Zebra?sh;Endocrine disruptor;Danio rerio

1.It is that a range of pollutants found in the environment,known as endocrine dis-ruptors,compromise the reproductive capacity and sexual development in numerous species (Crisp et al.,1998).Organisms are particularly sensitive to these

?

Corresponding author.Present address:Druimnich,Glenborrodale,Acharacle,PH364JP,Scotland,UK.Tel.:+44-1972-500-272;fax:+44-1972-500-272.

E-mail address:kime@https://www.360docs.net/doc/999859383.html, (D.E.Kime).

compounds during early life,and exposure can result in genital defects,abnormal gamete development,infertility or complete sex reversal (Kime,1999;Jobling et al.,2002).Moreover,an increasing inci-dence of poor sperm quality resulting from abnormal development of Sertoli cells,and genital defects in human males has been linked to possible prenatal and/or early life exposure to estrogenic compounds (Skakkebaak et al.,2001).While numerous screening tests are available for these estrogen mimics and their effects are well documented,far less research effort has focused on other endocrine disruptors,for which

0166-445X/$–see front matter ?2003Elsevier B.V .All rights reserved.doi:10.1016/S0166-445X(03)00154-1

310 B.G.

Aquatic Toxicology 65(2003)309–316

no

routine tests are available.an

aromatase inhibitor that

masculinises marine gas-tropods at levels (1–100ng l ?1)found widely in the aquatic environment (Matthiessen

and Gibbs,1998).These concentrations arose largely from its use in anti-fouling paints

for shipping and marine structures,

and although such use is now under strict legislation,a complete phase out is not due until 2008(Evans et al.,2000).TBT is not readily biodegradable.The highly toxic effect of TBT on animals,including ?sh is well documented (Fent,1996)but there are limited data on its sub-lethal effects on vertebrates in vivo,particularly during sensitive juvenile stages of development,and at environmentally relevant concentrations.Aromatase inhibiting compounds may be particularly important endocrine disruptors since estrogens are essential for normal male and female reproductive development (Sharpe,2001).However,studies of the effects on sperm of either aromatase inhibitors or estradiol de-privation have given mixed results depending on the animal model used,the compound employed and level of exposure (O’Donnell et al.,2001).Hess et al.con?rmed the importance of estradiol on male

development in rodents by showing that trans-genic mice lacking estrogen or aromatase receptors (ER ??KO/ArKO)were infertile due to disruptions to spermatogenesis,reduced epididymal sperm reduced sperm motility and elevated ?uid production

in the efferent ducts of the testes.The administration of aromatase inhibitors to several species of male mammalian organisms,including humans,caused an increase in circulating testosterone,luteinizing hor-mone (LH)and follicle stimulating hormone (FSH)suggesting that estradiol has an important role in the regulation of the hypothalamo–pituitary–gonadal axis in both sexes (O’Donnell et al.,2001).Chang et al.(1999)have demonstrated high gonadal aro-matase activity during sexual differentiation in the grey mullet (Mugil cephalus )and showed that this activity was signi?cantly reduced by exposure to the non-steroidal aromatase inhibitor Fadrozole.The ef-fects of aromatase inhibition on spermatogenesis or

sperm quality,which may provide a useful biomarker

of effects on fertility,has not,however,been studied in any depth.Kwon et al.(2000)clearly showed the effects of aromatase inhibition on sexual develop-ment in juvenile ?sh by exposing genetically female Nile tilapia (Oreochromis niloticus )during the pe-

riod of sexual differentiation to Fadrozole via their daily diet.Exposure to 0–200mg kg ?1from 7to 37days post-hatch gave a dose dependent increase in the numbers of males.At 500mg kg ?1over 80%of exposed ?sh were masculinised.These experiments not affect sperm motility or morphology.Spermato-genesis relies on the normal tion of the hypothalamo–pituitary cells (including Leydig cells,Sertoli cells).During spermatogenesis,the a crucial role in the development from to the mature spermatid (Jegou,tion may lead to a reduction in sperm possible infertility.Sertoli cells are equipped with high numbers of estrogen receptors and immunohis-tochemical research studies with mice have shown that these are active during the early stages of sexual differentiation (O’Donnell et al.,2001).Relatively low levels (11.2ng l ?1)of TBT exposure for 21days decreased sperm counts in guppies (Poecilia reticu-lata )by 40–75%but ?agellar length was unaffected (Haubruge et al.,2000).However,this exposure in-volved adult ?sh and effects on sperm were short term.Since limited data exist on possible effects of TBT induced aromatase inhibition on sexual devel-opment in ?sh,particularly in vivo and at environ-mentally relevant levels we have exposed zebra?sh to environmental levels of TBT in vivo during early developmental stages (0–70,0–30and 30–60days post-hatch)and examined effects on sexual devel-opment and sperm quality 3–5months following termination of exposure.

2.Materials and methods 2.1.Animals and ?sh maintenance

Wild type zebra?sh eggs were obtained (by kind permission of Professor Philip Ingham (Develop-mental Genetics Institute,University of Shef?eld),

B.G.McAllister,D.E.Kime/Aquatic Toxicology65(2003)309–316311

using the method described by Wester?eld(1989). The following morning unfertilised eggs were dis-carded and random batches of50embryos placed in ?ow-through fry rearing chambers.Each treatment was duplicated.Water temperature was maintained between27and28?C and pH8.0.Dissolved oxygen was≥95%.Mains water was dechlorinated by car-bon?ltration and ammonia,nitrite and nitrate kept below detection limits.Photoperiod was14h light and10h dark.Exposure to TBT commenced1day before hatch.During the?rst week of exposure larvae were fed(3mg per fry chamber twice a day)Tetra AZ-300fry feed,(Tetrawerke;Melle,Germany).Fry were fed live artemia nauplii and Tetra AZ300on days16,17post-hatch and artemia nauplii thereafter. From2months post-hatch,?sh were given a bal-ance of artemia nauplii and crushed commercial?ake feed(Tetra?ake)twice daily.Fifteen?sh per repli-cate(30per treatment)were transferred to50l tanks. Waste feed and faeces were siphoned from the tanks and rearing chambers daily.Care was taken to avoid stressing?sh throughout the exposure period.

2.2.TBT exposure

TBT was dissolved in acetone(100mg ml?1)and at the top stock dilution the concentration of sol-vent did not exceed10ng l?1,well below that shown to affect the development of zebra?sh juveniles (100ng l?1)(Gorge and Nagel,1990).Control tanks received an identical acetone concentration into the water(10ng l?1).Fish were exposed to speci?c TBT concentrations using a?ow through system,with a continuous input of TBT from a stock solution held in2.5l brown glass bottles suspended1m above the tanks.The pollutant was fed through2mm silicon capillary tubing and?ow rate adjusted using perspex drop meters(Scienti?c Laboratory Supplies,Notting-ham)following the method described by Sprague (1973).Stocks were prepared at ten times the required ?nal concentration and stored in25l plastic carboys in a dark room at4?C.Stocks were diluted at1:9 with clean aquarium water in the in?ow tube50cm upstream of each tank to ensure a constant dose to each tank.Aqueous concentrations of TBT in aquaria at the top dose(100±5ng l?1)were veri?ed by gas ?ame atomic absorption spectophotometry using a Waters HPLC system(Waters,USA)and kindly car-ried out by Rick Dunn at the University of Shef?eld. Other dose levels were below detection limits.Each tank was primed for3days at the required dosage prior to introduction of embryos.Following the ex-posure period,?sh were transferred to clean recovery tanks and grown on to full maturity.Mortality was not signi?cantly different from controls at all doses employed(<5%).Fish showed no apparent stress or harm throughout the exposure period.In the?rst experiment,?sh were exposed to0,0.01,0.1,1,10 or100ng l?1for70days from hatching.To further de?ne the critical period of exposure,a second exper-iment was performed in which?sh were exposed to0, 0.01,1or100ng l?1TBT from either0to30or from 30to60days post-hatch.At the end of the exposure period?sh were transferred to clean water.Sex ratio and sperm motility were examined3–5months after the termination of exposure.

2.3.Sperm motility

Zebra?sh sperm movement was recorded from5s after activation(to allow for focusing and stabilisa-tion of?uid movement)for successive15s periods using a CCD camera(Sony SPT-M108CE)attached to a Nikon Alphaphot-2microscope with a phase contrast objective lens(20×0.40).Videotapes were analysed using Computer Assisted Sperm Analysis (CASA)(Hobson Vision Ltd.,Baslow,UK)with in-strument parameters optimised for zebra?sh(search radius,7.5?m;trail,50;prediction,off;pause win-dow,0.5s;video,pal(50);aspect,1.5;refresh time, 1s;thresholds,+15/?100;?lter weightings,1:1,2:1, 2:1,3:1;immotile process,shape;maximum size, 7.5?m;minimum size,1.8?m;ratio,1;minimum track time,1s).The Hobson Sperm Tracker produces a large number of different parameters of motility, but for brevity only the curvilinear velocity(VCL),is presented since this parameter is most useful for the analysis of?sh sperm motility(Kime et al.,2001).

2.4.Sperm counts and image recording

Sperm was stripped from mature males following anaesthesia in2-phenoxy ethanol(Sigma,Poole)us-ing a5?l graduated capillary pipette(±0.1?l;Sigma, UK)and kept on ice.One microlitre of sperm was di-luted in97?l extender solution(NaCl5.8g l?1,KCl

312 B.G.McAllister,D.E.Kime /Aquatic Toxicology 65(2003)309–316

0.2g l ?1,CaCl 20.22g l ?1,MgCl 2·6H 2O 0.04g l ?1,NaHCO 3 2.10g l ?1,NaH 2PO 4·2H 2O 0.04g l ?1,glycine 3.75g l ?1in distilled water,pH 8.64)at a 1:100dilution for even dispersal and ease of counting.Two microlitres of 0.5%Rose Bengal dye (Sigma,UK)in extender was added to the solutions to make sperm more visible for analysis.The total numbers of sperm with or without ?agella were calculated by multiplying the mean sperm count of ?ve 10?l sub-samples by the dilution factor.2.5.Sperm image analysis

Two microlitres were taken from each sperm sample and placed in a 200?l Eppendorf vial.A fur-ther 2?l of ?ourescent live/dead sperm dye (Molecu-lar Probes,Eugene,Oregon USA)was added and the samples incubated in darkness for 20min.They were then observed under epi?ourescent microscopy (Leica DLMB)using a 40×objective.Images were captured using an Aequitas image database and image archive management software (Dynamic Data Links,Avebury

TBT (ng l -1)

0.01

0.1

1

10

100

m i l t v o l u m e (m l )

12

3

400.010.1110100

V C L (μm s -1)

20406080

100

TBT (ng l -1

)

0.01

0.1

1

10

% a b n o r m a l s p e r m

20406080

1000

0.010.1110100

% m a l e s

40

6080

100

(a)

(b)

(c)

(d)

Fig.1.The effect of TBT exposure on zebra?sh exposed from hatch for 70days:(a)percentage males at sexual maturity;(b)sperm motility (VCL =curvilinear velocity ±S .E .M .);(c)percentage abnormal sperm (lacking ?agella)±S.E.M.assessed 3–5months after termination of exposure;and (d)volume of milt expressible 3–5months after termination of exposure.For graph (a)(?)indicates that the sex ratio was signi?cantly different (P <0.05)from 50%after exposure using the χ2-test.For graphs (b)–(d)bars with different letters are signi?cantly different (P <0.05)by one-way ANOV A and Tukey pairwise test (for each group n =30).

House,Cambridge,UK)and RT colour camera model 2.2.0(Diagnostic Instruments,UK).3.Results

3.1.Exposure from 0to 70days post-hatch The sex ratio was signi?cantly altered (P <0.05;Chi-squared test)with 60%males at 0.1ng l ?1(18males,12females;Fig.1(a)).At the highest doses em-ployed (100ng l ?1)93%of ?sh were male (28males and 2females).Sperm motility was signi?cantly re-duced at 1ng l ?1and all sperm were immotile at 100ng l ?1(Fig.1(b)).The incidence of sperm lack-ing ?agella,was signi?cantly higher than controls at 0.1ng l ?1,and at 10ng l ?1all sperm were abnormal (Fig.1(c)).Milt volume was signi?cantly higher than control after exposure to 100ng l ?1(Fig.1(d)).The images of sperm (Fig.2)clearly demonstrate that TBT induced a progressive increase in the number of spermatazoa without ?agella with increasing dose exposure 3–5months after exposure was terminated.

B.G.McAllister,D.E.Kime/Aquatic Toxicology65(2003)309–316

313

Fig.2.Photographic images of sperm taken3–5months after0–70days post-hatch TBT exposure:(a)control;(b)1ng l?1;(c)100ng l?1.

3.2.Exposure from0to30and30to60days

post-hatch

The sex ratio was male biased after exposure to

1ng l?1from0to30days post-hatch exposure(68%;

20males and10females),and to100ng l?1from30

to60day exposure post-hatch(68%;20males and10

females;Fig.3(a)).Sperm motility

was signi?cantly

lower than controls after exposure to1ng l?1during both developmental periods(Fig.3(b)).Incidence of abnormal sperm was signi?cantly higher than con-trols at1ng l?1for0–30day exposure and100ng l?1 for30–60day exposure(Fig.3(c)).Milt volume was highly variable in controls and was not signi?cantly different at any of the exposures used(Fig.3(d)). 4.Discussion

The masculinising effect of TBT on sexual de-velopment in?sh demonstrated in the present study correlates well with the induction of imposex in aquatic invertebrates exposed to environmental TBT (Matthiessen and Gibbs,1998).The effect of TBT(an aromatase inhibitor)in zebra?sh is also in accord with the?ndings of Piferrer et al.(1994)that genetically female chinook salmon(Onchorhynchus tshawytscha) developed into phenotypically normal males if brie?y exposed(2h)to Fadrozole,a non-steroidal aromatase inhibitor,during sexual differentiation.While TBT is extremely toxic,the absence of any apparent ill health or excessive mortality even at the highest ex-posure doses and the similarities of effects with those of Matthiessen and Gibbs(1998)and Piferrer et al. (1994)suggests that the effects observed in the present study are induced by its effect as an aromatase in-hibitor.Recent studies(McAllister et al.,unpublished data)have indeed shown that TBT inhibits ovarian aromatase in?sh in vitro.Our data show comparable effects of masculinisation after exposure from0to70 and0to30days post-hatch,while?sh exposed only from30to60days post-hatch had considerably lower masculinisation(Figs.1(a)and3(a)).This is consis-tent with the?ndings of Takahashi(1977)that sexual differentiation in zebra?sh occurred between10and 25days post-hatch.It is,however,unclear whether there exists in zebra?sh,as in salmon,a very short window during which even a2h exposure to an aro-matase inhibitor can induce complete masculinisation. The genotypically female salmon masculinised by2h exposure to Fadrozole in early life produced viable sperm capable of fertilisation(Piferrer et al.,1994) and production of viable all female progeny.This con-trasts with our data on TBT exposed zebra?sh which showed poor quality sperm which were evidently in-capable of fertilisation.In our studies,however,the ?sh were exposed to TBT for a much longer period than in those of Piferrer et al.(1994)which suggests that the period required for induction of abnormal sex differentiation is considerably shorter than that which

314

B.G.McAllister,D.E.Kime /Aquatic Toxicology 65(2003)309–316

0.01

0.1

1

10100

% a b n o r m a l s p e r m

20406080

10000.010.1110100

m i l t v o l u m e (m l )

123

40

0.010.1110100

V C L (μm s -1)

20406080

1000

0.010.1110100

% m a l e s

40

6080

100

(a)(b)

(c)

(d)

TBT (ng l -1)

TBT (ng l -1)

Fig.3.The effect of TBT exposure on zebra?sh exposed to 0,0.01,1or 100ng l ?1from 0to 30(black bars)and 30to 60(open bars)days post-hatch:(a)percentage males at sexual maturity;(b)sperm motility (VCL)±S.E.M.assessed 3–5months after termination of exposure,(c)%abnormal sperm (lacking ?agella);and (d)volume of milt expressible (±S.E.M.)3–5months after termination of exposure.For graph (a)(?)indicates that the sex ratio was signi?cantly different (P <0.05)from 50%after exposure using the χ2-test.For graphs (b)–(d)bars with different letters are signi?cantly different (P <0.05)by one-way ANOV A and Tukey pairwise test (for each group n =30).

is necessary for production of abnormal sperm.This suggestion is supported by our exposure for 0–30and 30–60days post-hatch in which the maximal percent-age abnormal sperm were 50and 30%,respectively,compared with 95%in ?sh exposed for a full 70days.The apparent decline in human sperm counts and the increased incidence of male genital abnormalities have been attributed to environmental estrogen-mimics acting on the developing Sertoli cells during pre-natal or early life (Skakkebaak et al.,2001).An important ?nding in study was that the effects of TBT on ze-bra?sh in early life only became apparent at adulthood and were irreversible even after 5months in clean water.Our study

suggests that aromatase inhibitors may

have had a similar long-term effect to that sug-gested for estrogens in mammals by interfering with Sertoli cell development in early life and induce pro-duction of abnormal sperm lacking ?agella when the organism becomes an adult.Estrogens are essential for normal functioning of the Sertoli cells (O’Donnell et al.,2001)which are involved in attachment of the ?agellum to the mature spermatid.Our ?ndings,therefore,have important implications for human health as well as for that of aquatic organisms and suggest that further study of the effects of early life exposure to aromatase inhibitors is urgently needed.In addition to the in vivo effects of TBT described in the present communication,it can also affect sperm directly (Rurangwa et al.,2002;Huyskens et al.,un-published data).In such in vitro exposures,however,there was no apparent morphological abnormality in the sperm even at the highest doses of TBT.Rurangwa Haubruge et al.(2000)found that adult guppies ex-posed for 21days to 11ng ?1TBT had a short-term decrease in sperm counts but the ?agella length was unchanged.Such effects clearly differ from those de-scribed in the present study in that they are immediate or short-term effects and do not induce loss of ?agella.The present data,therefore,indicate that the effects observed are developmental and irreversible,and can-not be attributed to a direct effect of TBT on sperm such as those found in vitro.

B.G.McAllister,D.E.Kime/Aquatic Toxicology65(2003)309–316315

We considered the possibility that the effects ob-served were a direct result of bioaccumulation,since TBT is a potent bioaccumulator in?sh tissue(Yamada et al.,1993).Since the exposures were of very short duration and before the major growth phase of the?sh in clean water biomagni?cation would have had to be implausibly high to cause an effect6months after cessation of exposure.Such effects would also most likely have had more similarity to the in vitro studies of Rurangwa et al.(2002).The lowest observable effect concentration(LOEC;0.1ng l?1)found in our study is a major cause for concern.The LOEC found in this ex-posure is below standard detection limits(>1ng l?1), and well below the EC permitted levels of receiving waters(2ng l?1;Environment Agency,1998).Mea-sured environmental levels range between the detec-tion limit and100ng l?1,although these may often be greatly exceeded in busy industrial ports and docking areas(Bosselmann,1996).Gibbs and Bryan(1986) found that TBT exposure levels above1ng l?1were enough to inhibit steroids and masculinise female gas-tropods in the wild.It is apparent from our study that juvenile?sh are at least as sensitive to the potent aro-matase inhibitive properties of TBT,and although the effects of the pollutant on commercial species of ma-rine invertebrates was well publicised during the1980s and1990s the possibility that?sh populations might also be affected has not been considered.Our?ndings that TBT,at concentrations below the limits of present detection methods,can cause a high incidence of ab-normal sperm and a skewing of the sex ratio in?sh may have major implications for the ecology and com-mercial?sheries in areas,such as the Mediterranean and North Seas,subjected to heavy marine traf?c. Acknowledgements

The authors wish to acknowledge the assistance of Andy Skinner and Dr.Penny Watt of the University of Shef?eld.

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