Genotoxic, cytotoxic, developmental and survival effects of tritiated water in the early life stages of the marine mollusc, Mytilus edulis

Aquatic Toxicology 74 (2005) 205–217

Genotoxic, cytotoxic, developmental and survival effects oftritiated water in the early life stages of the marine

mollusc,Mytilus edulis

Josephine A. Hagger, Franck A. Atienzar1, Awadhesh N. Jha∗

School of Biological Sciences, University of Plymouth, Drake Circus, Devon, Plymouth PL4 8AA, UK

Received 12 January 2005; received in revised form 27 May 2005; accepted 28 May 2005

Abstract

Using an integrated approach linking different levels of biological organisation, the genotoxic, cytotoxic, developmental andsurvival impact of tritiated water (HTO) were investigated in the embryo-larvae of marine molluscMytilus edulis. One-hour-oldembryos were exposed to a range of concentrations (0.37–370 kBq ml−1) of HTO, which delivered a dose between 0.02 and21.41 mGy over the exposure period for different end points. Detrimental effects, if any, were monitored at different levels ofbiological organisation (i.e. DNA, chromosomal, cellular and individual). Genotoxic effects were assessed using molecular andcytogenetic approaches which included analysis of random amplified polymorphic DNA (RAPD), induction of sister chromatidexchanges (SCEs) and chromosomal aberrations (Cabs). Cytotoxic effects were evaluated by determining the proliferative rate

sults inortality

alitativeentrationwere notmentallyong termin order

i-omna

index (PRI) of the embryo-larval cells. Developmental and survival effects were also monitored every 24 h up to 72 h. Regeneral indicated that HTO significantly increased cytogenetic damage, cytotoxicity, developmental abnormalities and mof the embryo-larvae as a function of concentration or radiation dose. The analysis of RAPD profiles also revealed queffects in the HTO exposed population compared to controls. However, while the embryo-larvae showed dose or concdependent effects for mortality, developmental abnormalities and induction of SCEs, the dose-dependent effectsapparent for Cabs and PRI at higher doses. The study contributes to our limited understanding of the impact of environrelevant radionuclides on non-human biota and emphasises the need for further investigations to elucidate potentially ldamage induced by persistent, low levels of other radionuclides on commercially and ecologically important species,to protect human and ecosystem health.© 2005 Elsevier B.V. All rights reserved.

Keywords:Tritium; Genotoxicity; RAPD assay; Developmental effects; Marine mussels

∗ Corresponding author. Tel.: +44 1752 232978;fax: +44 1752 232970.

E-mail address:[email protected] (A.N. Jha).1 Present address: UCB S.A. Pharma Sector, Chemin du Foriest,

B-1420 Braine-l’Alleud, Belgium.

1. Introduction

In 1977 the International Commission on Radiologcal Protection stated that “If humans are protected frthe effects of ionising radiation, then flora and fau

0166-445X/$ – see front matter © 2005 Elsevier B.V. All rights reserved.doi:10.1016/j.aquatox.2005.05.013

206 J.A. Hagger et al. / Aquatic Toxicology 74 (2005) 205–217

are also adequately protected” (ICRP, 1984). However,this statement is often questioned by the general pub-lic and regulatory authorities as: (a) different speciesvary in their sensitivity to toxicants including radia-tion; (b) different life stages of the same species canshow varying levels of response to contaminants; (c)there are areas where wildlife may be exposed to higherlevels of radiation than are permissible to humans and(d) endangered species with longer generation timesor isolated populations might need special considera-tion, in order to maintain ecosystem quality. At present,an internationally accepted method for assessing theenvironmental impact of ionising radiation does notexist (Pentreath, 1998). Attempts are therefore beingmade to link together the release scenario of radioactivesubstances, potential exposure pathways, dose to eco-logically relevant, commercially important and criticaltarget organisms and subsequently effects at differentlevels of biological organisation (ERICA, 2005).

As a result of nuclear establishments around coastalareas, the marine environment is often the ultimaterecipient of radioactive contaminants. In this context,tritium (3H), a naturally occurring isotope of the ele-ment hydrogen, is manufactured commercially and isalso released as a by-product of nuclear reactions. Anoverview of historic trends in liquid and aerial dis-charges of radioactive substances in the UK, carried outbetween 1979 and 1998, showed that there has been asustained increase in the discharge of3H from differ-ent nuclear facilities (Department of the Environment,T ef anh ingw ell-fi ants( v-i enta erne dt ials,r ce ofb o-c

iotaf ings fil bio-

logical or toxicological impact of3H on commerciallyand/or ecologically important species. In particular,there has been a lack of research on potential detrimen-tal effects of ionising radiation on invertebrates, whichconstitute over 90% of the existing species and playa major role in the maintenance and function of everyecosystem. Consequently, the aim of the present studywas to assess the genotoxic, cytotoxic and developmen-tal effects of�-radiation emitted by HTO on embryo-larvae of an ecologically and commercially importantinvertebrate, the marine bivalve molluscMytilus edulis.

2. Materials and methods

2.1. Exposure scenario

HTO was obtained from ICN Pharmaceuticals,Inc., UK at a concentration of 3.7 GBq ml−1. Therange of concentrations of HTO used in the presentstudy (0.37–370 kBq ml−1) was adopted from pre-vious work carried out on fish and barnacle larvae(Strand et al., 1977; Abbott and Mix, 1979) and gavea dose between 0.02 and 21.42 mGy over the expo-sure period as described in Section3.2. Fig. 1 showsthe schematic representation of the general proto-col used for assessing the genotoxic [random ampli-fied polymorphic DNA (RAPD)], sister chromatidexchanges (SCEs) and chromosomal aberration (Cabs)assays, cytotoxic proliferative rate index (PRI), devel-oe ofs ta ermf ggswst -a ugha toa e-l n-te icht lassb 0 mlw con-t lied

ransport and the Regions, 2000). As HTO can movreely in the biosphere it could be a threat to humealth as it may be ingested not only through drinkater but also by food, especially from fish and shsh that are known to bioaccumulate some toxicKirchmann et al., 1973). In this context, elevated actity concentrations of3H have been reported in sedimnd at different trophic levels of biota from the Sevstuary, UK (McCubbin et al., 2001). It was suggeste

hat the high activity concentrations in these materelative to that in seawater, was due to the presenioavailable organic3H-labelled compounds in radihemical wastes discharged in the vicinity.

Despite recent moves to protect non-human brom exposures to ionising radiations and the growcientific and public concern over the presence o3Hn the environment (Environment Agency, 2001), veryittle work has been carried out to evaluate the radio

pmental and survival effects of HTO onM. edulismbryo-larvae. In brief, following the inductionpawning, as described in detail elsewhere (Jha el., 2000a), eggs from two female mussels and sp

rom one male were collected. The number of eas adjusted to 50 eggs ml−1 prior to adding spermolution at approximately 107 sperm ml−1. Fertilisa-ion (rate 92%) was carried out at 15± 1◦C, wherefter excess sperm was removed by filtering throsieve (30�m) and the embryo density adjusted

pproximately 30 embryos ml−1 as per standard guidines for acute toxicity testing of environmental coaminants (ASTM, 1989). Following fertilisation, thembryos were allowed to grow for 1 h, after wh

hey were divided into different exposure vessels (geakers) where the final volume was adjusted to 50ith seawater. Temperature, salinity and pH were

rolled throughout the exposure and air was supp

J.A. Hagger et al. / Aquatic Toxicology 74 (2005) 205–217 207

Fig. 1. Flow chart illustrating the experimental protocol adopted toevaluate genotoxic, cytotoxic, developmental and survival effects oftritiated water (HTO) on the embryo-larvae ofM. edulis.

via gentle aeration. Embryos were exposed to fourdifferent activity concentrations of HTO (0.37, 3.7,37 and 370 kBq ml−1), 1 negative control (i.e. 10�mfiltered seawater) and 1 positive control containingmethyl methane sulphonate (1.0× 10−3 M MMS) asa reference genotoxic agent. This single concentra-tion of MMS was selected based on our earlier studies

(Jha et al., 2000a,b). HTO concentrations were madeusing a serial dilution with seawater from the originalstock. One hundred microlitres of each dilution wasadded to 500 ml of seawater containing the embryos.MMS was also made using a serial dilution with sea-water and 100�l of sample was added to achieve thedesired concentration for the volume of seawater. Inorder to obtain sister chromatid differential staining(SCD) a minimal concentration (10−5 M) of 5-bromo-2-deoxyuridine (BrdU) was added to all the vessels at12 h post-fertilisation.

2.2. Random amplified polymorphic DNA (RAPD)assay

As described inFig. 1, along with Cabs and SCEsassays, a subset of embryo-larvae were processed foranalysis using the random amplified polymorphic DNAassay. Genomic DNA was extracted and purified usinga conventional phenol/chloroform method as describedelsewhere (Atienzar et al., 1999). DNA concentrationswere determined by electrophoresis by comparisonwith known concentrations of Lambda phage DNA(Sigma, Poole, UK). RAPD was performed with 5 ngof genomic DNA as described elsewhere in details(Atienzar et al., 2002). The 10-mer primers used wereOPA9, OPB1, OPB5, OPB6, OPB7, OPB8, OPB10,OPB11, OPB12, OPB14, or OPB17 (Operon Technolo-gies, Southampton, UK; for the sequences seeAtienzarand Jha, 2004). Thermal cycling parameters and anal-y eend ;AD er-l

2S

2 ryoswO sea-w toa at-m ryosw cen-t

sis of PCR products by electrophoresis have bescribed elsewhere in details (Atienzar et al., 2002tienzar and Jha, 2004). The GeneRulerTM 100 bpNA ladder plus (Immunogen International, Sund

and, UK) was used.

.3. Slide preparation for cytogenetic (i.e. PRI,CEs and Cabs) analyses

Following addition of BrdU, after 112 (Cabs) andcell cycles (SCEs and PRI) two subsets of embere removed (1 cell cycle = 3.8 h;Jha et al., 2000a,b).ne subset was placed into colchicine dissolved inater (0.025%, w/v) for 30 min and then exposedseries of hypotonic (seawater/0.075 M KCl) treents. After the hypotonic treatments, the embere transferred into centrifuge tubes and were

rifuged for 5 min at 200×g. After centrifugation


the supernatant was discarded and a few drops ofcold (4◦C) Carnoys fixative (1:3 (v/v) glacial aceticacid/methanol) was added. The cells were then pro-cessed to produce metaphase spreads, stained withGiemsa (for the analysis of Cabs), or processed forsister chromatid differential staining (for the analysisof SCEs and PRI) as previously described (Harrisonand Jones, 1982; Jha et al., 1995a,b, 1996, 2000a,b)for microscopic analysis.

Prior to scoring, slides were randomised, coded byan independent disinterested individual and then scoredblind. Complete metaphase spreads (2n= 28) wereexamined using a bright field microscope (final magni-fication×1000). Details of classification of PRI stages,scoring of Cabs and SCEs have been described in ear-lier studies (Jha et al., 1995a, 1996, 2000a, 2000b).In brief, metaphases were classified, as first (M1), sec-ond (M2) and third or subsequent (M3+) division cells.For PRI at least 200 cells per treatment were scored,100 each from two replicates. PRI was calculated usingthe formula PRI = ((1× M1) + (2× M2) + (3× M3))/number of cells analysed (Lamberti et al., 1983). Onlysecond (M2) division metaphases were analysed forthe occurrence of SCEs. Fifty metaphases per treatmentwere analysed for SCEs, 25 from each of two replicates,and the results were expressed as the SCEs frequencyper cell. For Cabs 200 cells from each treatment werescored (100 from each of two replicates) as per therecommendation of the United Kingdom Environmen-tal Mutagen Society (UKEMS) (Kirkland, 1990). Cabsw rcent-a ionsi ions( ber-r

2s

tal-i d at2 -o . Thed icpe es ately2 vae,

with a circular band of short cilia (prototroch) and aflagellum (Fig. 2a). Abnormal larvae are usually mis-shapen (not spherical in shape or symmetrical) andmay lack cilia and/or the flagella, resulting in abnormalswimming behaviour (Fig. 2b). The second larvae stageis the veliger, which inM. edulisis usually formed after48 h (15◦C). A dorsal thickening of the ectodermis,secreted by the shell gland, forms an initial organic cuti-cle that spreads over the entire body. A straight dorsalhinge gives the larvae the characteristic capital D shape,hence the name “D-larva” (Fig. 2c). At this stage thelarvae are approximately 60–70�m in size (His et al.,1999). Dead trochophores appear as a mass of degener-ating cells with no cilia or flagellum movement. DeadD-larvae may also lack ciliary movement, lose clar-ity of internal organs and become transparent (Hiset al., 1999). Pollutants present in the marine envi-ronment may lead to morphologically abnormal tro-chophores and D-larvae (Fig. 2b and d, respectively).

2.5. Water quality parameters and measurement ofradioactivity

Water quality parameters (pH, temperature andsalinity) were monitored at the start and end of expo-sure and were relatively stable throughout (mean andS.E.; pH 8.07± 0.05; salinity 35± 1, 15.5± 0.5◦C).Pre- and post-exposure levels of radioactivity weredetermined using 100�l of test solution in 5 mlof liquid scintillation cocktail, Packard Ultima Gold( rec plesw om0 anL s perm te-g ncyfa hasa sw werd

2

atis-t in-d 97.

ere expressed as the percent aberrant cells (pege of cells that contained one or more aberrat

ncluding heavily damaged cells) and total aberratthe number of chromosome plus chromatid type aations per 100 cells excluding gaps).

.4. Analyses of developmental and mortality/urvival effects

Developmental morphology, behaviour and morty of the growing embryo-larvae were assesse4, 48 and 72 h post-fertilisation.Fig. 2 shows varius stages of normal and abnormal developmentevelopment ofM. edulis consists of an embryonhase followed by a larval phase (Bayne, 1976). At thexperimental temperature (15± 1◦C), the first larvatage known as trochophores appears at approxim4 h. This stage is defined as free swimming lar

Sigma–Aldrich Ltd., USA). Five replicates weounted from each activity concentration. Samere read for 10 min (detection limits range fr.06 to 14.5 keV) in a scintillation counter (BeckmS6500, USA). Readings were displayed as countinute (CPM), then automatically converted to disinrations per minute (DPM) using percentage efficie

or 3H and then converted to kBq ml−1 (37 kBq equiv-lent to 2,220,000 DPM). The scintillation counterbuilt in quench correction and14C and3H standardere used to calibrate the counter regularly. The loetection limit for3H was 0.00962 kBq ml−1.

.6. Statistical analyses

Statistical analysis were carried out using the stical software package STATGRAPHICS Plus for Wows 3.1©from Statistical Graphics Corp. 1994–19


Fig. 2. The developmental stages ofM. edulisembryo-larvae. Photos a and c represent normal embryo-larvae at 24 h and >48 h, respectively.Photos (b) and (d) represent morphologically abnormal embryo-larvae at 24 h and >48 h, respectively.

One-way analysis of variance (ANOVA) was carriedout on normalised data. If required, transformation ofdata for developmental effects was carried out usingArc SIN and square root as recommended byBurgeotet al. (1995)whereas SCEs, Cabs and PRI values weretransformed either via log or square root conversions.Non-normally distributed data were analysed using thenon-parametric Kruskal–Wallis test (Sparks, 2000).

3. Results

3.1. Determination of radioactivity

Table 1presents the levels of radioactivity in dif-ferent test vessels for the exposure scenario. There

was neither significant variation in the percentage ofradioactivity lost during the exposure period (differ-ence between pre- and post-exposure), nor any trend inthe percentage loss for different concentrations of HTOused. Each vessel containing HTO lost similar percent-ages of radioactivity although the samples containinglarger amounts showed a greater loss in terms of theactual concentration over the exposure period.

3.2. Calculation of dose received byembryo-larvae

The estimated dose to growing embryo-larvaeexposed to HTO was calculated by a method used byStrand et al. (1977)on fish larvae. The dose equa-tion is as follows:Dß = 2.13ε�C, where, Dß = dose


Table 1Nominal and measured radioactivity (kBq ml−1) of tritiated water (HTO) in seawater before and after exposure ofM. edulisembryo-larvae

Nominal Concentration(kBq ml−1)

Measured pre-exposureactivity (kBq ml−1)

Measured post-exposureactivity (kBq ml−1)

Percentage loss

Seawater (SW) 0.003± 0.001 0.002± 0.001 33.33SW + MMS 0.003± 0.001 0.002± 0.001 33.330.37 0.496± 0.010 0.361± 0.008 27.203.7 2.460± 0.025 2.320± 0.023 5.7037 30.200± 0.300 22.610± 0.220 25.13370 284.180± 2.67 220.200± 1.960 22.51

Mean± S.E. from five replicate counts from different beakers at each concentration.

in rads per hour,ε� = average beta-ray energy (3H =0.00569 MeV),C= concentration of3H (in �Ci ml−1:where 1�Ci is equal to 37 kBq) and 2.13 is the con-version factor. These calculations were based on theassumptions that: (a)3H was uniformly distributedwithin the embryo-larvae over the exposure period and(b) that no concentration of3H above a water equilib-rium level occurred in the embryos. The predicted dosereceived during the exposure to3H ranged between0.02 and 27.83 mGy as presented inTable 2.

3.3. Analysis of RAPD results

Rather than showing individual changes in RAPDpatterns related to HTO treatments, we decided to showsome variations in profiles that occurred at differentdoses.Fig. 3a and b are an illustration of such changes.For instance with primer OPB6, the band 6-1 was moreintense after 1.5 and 2 cell cycles than the control at alldoses of3H and even in the positive control (MMS)(Fig. 3a). RAPD profiles generated by primer OPB12show that band 12-1 present in the control profile nearlydisappeared after 1.5 and 2 cell cycles in embryosexposed to 3.7, 37 and 370 kBq ml−1 (Fig. 3b). The use

of two DNA concentrations (5 and 20 ng) confirmedthe reproducibility of the RAPD profiles with primerOPB6 and OPB12 (data not presented). Finally, not allprimers showed differences in RAPD profiles gener-ated by the HTO treated population in comparison tocontrol patterns (data not shown).

3.4. Analysis of sister chromatid exchanges

The induction of SCEs in the embryo-larvae ofM. edulis is shown in Fig. 4a. The frequency ofSCEs increased with increasing activity concentra-tion of HTO. Statistical analysis using Kruskal–Wallistest showed that there was a significant differencebetween the controls and the samples treated with HTO(P< 0.005). The controls were not significantly differ-ent to the lowest activity concentration (0.37 kBq ml−1)but there was significant difference among the con-trols and the other three concentrations of HTO. It wasalso suggested that there was a significant differencebetween all activity concentrations of HTO, with 0.37being different to 3.7, 37 and 370 and 3.7 being differentto 37 and 370 kBq ml−1. In addition, there was a sig-nificant difference for the induction of SCEs between

Table 2Dose (mGy) received byM. edulisembryo-larvae for different exposure period and endpoints following exposure to tritiated water (HTO)

Concentration(kBq ml−1)

Cabs (exposuretime = 16.7 h) (mGy)

SCEs andPRI (exposuretime = 18.6 h) (mGy)

Developmental and survival effects(exposure time = 23 h) (mGy)

S 0.00M 0.000 0.032 0.153 1.842 17.31

A ed inable 1.

eawater (0.00) 0.00 (0.00)MS (0.00) 0.00 (0.00).49 (0.37) 0.03 (0.02).46 (3.70) 0.13 (0.21)0.20 (37.00) 1.65 (2.14)84.18 (370.00) 15.54 (21.42)

ctual based on measured mean HTO concentrations presentT

(0.00) 0.00 (0.00)(0.00) 0.00 (0.00)(0.02) 0.04 (0.03)(0.24) 0.19 (0.28)(2.37) 2.27 (2.78)(23.72) 21.41 (27.83)

Values in parentheses represent the predicated values.


Fig. 3. RAPD profiles of embryo-larvaeM. edulisexposed to tritiated water (HTO) and MMS. RAPD profiles were generated using 10-merprimers OPB6: (a) and OPB12 (b). M: DNA molecular size marker (GeneRulerTM 100 bp DNA ladder plus, Immunogen International). Themolecular sizes (in kilobases) of selected bands are indicated on the left of each panel. Bands visualised were from top to bottom 3000, 2000,1500, 1200, 1031, 900, 800, 700, 600, 500, 400, 300, 200 and 100 bp. (−): No DNA control (negative control). The number of cell cyclesare indicated in both panels. Embryos were exposed to 0 (line 1), 0.37 (line 2), 3.7 (line 3), 37 (line 4) and 370 kBq ml−1 (line 5) as well as1.0× 10−3 M of MMS. Selected changes are indicated by arrows.

embryo-larvae exposed to 37 and 370 kBq ml−1

HTO.

3.5. Analysis of chromosomal aberrations

Fig. 5 represents the induction of Cabs in embryo-larvae ofM. edulisas a function of dose delivered by

each concentration of HTO over the exposure period.ANOVA indicated that there was a statistically sig-nificant increase for the induction of both aberrantcells (P< 0.001) and the number of total aberrations(P< 0.005) in HTO exposed samples compared to sea-water controls. However, there was no significant dif-ference among the three highest concentrations of HTO


Fig. 4. (a) Induction of sister chromatid exchanges (SCEs) and (b) Proliferative rate index (PRI) inM. edulisembryo-larvae after exposure totritiated water (HTO). Error bars represent standard error. (*) Indicates means significantly different from control (P< 0.05).

for the induction of aberrations. The baseline level ofaberrations in the controls was 7± 2 (mean and S.E.)aberrant cells containing 11± 6 (mean and S.E.) aber-rations in total. MMS exposed cells contained 31± 2aberrant cells and 32.5± 15 total aberrations (data notincluded inFig. 5). The frequency of aberrant cells fol-lowing MMS exposure was greater compared to HTO,however, the overall frequency of total aberrations wasgreater in the HTO exposed embryo-larvae. Thus, HTOexposed embryo-larvae had less cells containing aber-rations than MMS exposed embryo-larvae but HTOexposed cells that did have aberrations contained more

aberrations than MMS exposed cells. As the majorityof the aberrations in this study were acentrics (chro-mosome/chromatid breaks) the distribution pattern wasover dispersed for acentrics as have been observed formost studies carried out for external low LET radia-tion (Lloyd et al., 1986). In the (untreated) controls,31.8% of aberrations observed were chromatid typeaberrations while in the HTO exposed cells the fre-quency of chromatid type aberrations ranged from 45.2to 67.9%. There was no trend for the ratio of chro-mosome and chromatid type aberrations for differentdoses.

F bryo-l so

ig. 5. Induction of chromosomal aberrations (Cabs) inM. edulisemf actual and predicted doses. Line of best fit = logarithmic.

arvae as a function of predicted radiation dose.Table 2shows detail


Fig. 6. Percentage of: (a) normal and (b) deadM. edulisembryo-larvae following exposure to tritiated water (HTO). Error bars represent standarddeviation. (*) Indicates means significantly different from control (P< 0.05).

3.6. Analysis of cytotoxic effects (i.e. proliferativerate index or PRI)

Fig. 4b shows the values for PRI after exposure toHTO in comparison with controls. Statistical analy-sis of the results using ANOVA showed that all fourconcentrations of HTO have statistically significantlylower values for PRI compared to the controls. How-ever, there was no significant difference among differ-ent concentrations of HTO except between 0.37 and370 kBq ml−1.

3.7. Analysis of developmental andsurvival/mortality effects

Fig. 6a and b represent the percentage of normaland deadM. edulis embryo-larvae after 72 h post-fertilisation following exposure to HTO and MMS. Adose dependent decrease for the percentage of normalembryo-larvae occurred at 24, 48 and 72 h (P< 0.001for all times) with all the concentrations of HTO andMMS being significantly different only from the con-trol as indicated using ANOVA. By 72 h only the low-est concentration of HTO (0.37 kBq ml−1) containeda small frequency of normal embryo-larvae and all theother concentrations were devoid of normality. The per-centage of dead embryo-larvae in the seawater controlsdid not vary greatly over the three days with approxi-mately 10% of the embryos being dead. At 72 h therewas a large increase in mortality in embryo-larvae

exposed to HTO and MMS. With all the concentrationsof HTO and the MMS being significantly different fromthe control (P< 0.001). Furthermore, the lowest con-centration (i.e. 0.37 kBq ml−1) was also significantlydifferent from all the other concentrations and theMMS. There was no difference in mortality between3.7 and 37 kBq ml−1 but they were statistically differ-ent from both the 370 kBq ml−1 and MMS exposedlarvae. The regression coefficient indicated a moderatedose–response relationship (R2 = 42.69%). One hun-dred percent mortality was exhibited at 72 h in thehighest HTO concentration as well as the MMS concen-tration. Consequently, the 72 h LC50 forM.eduliscouldnot be predicted. The predicted 48 h LC50 was, how-ever, 0.27 Bq ml−1 with a received dose of 0.9 mGy.

4. Discussion

The analysis of RAPD profiles in the present studyshowed changes in patterns due to3H induced DNAalterations even at the lowest HTO concentration (i.e.0.37 kBq ml−1). Although it is difficult to decipherthe exact mechanisms responsible for the changes inRAPD profiles following HTO exposure, it has beendemonstrated that DNA damage and mutations caninduce changes in RAPD patterns after exposure togenotoxic agents (Atienzar et al., 2002). For instance,damaged DNA can block or reduce the processivityof theTaqDNA polymerase leading to disappearance


of bands in the RAPD profiles. Alternatively, con-formational changes in the DNA could also improvethe access of the primer(s) to the binding site(s) lead-ing to appearance of new bands in the correspondingRAPD patterns. In fact, such changes in the RAPD pro-files have been correlated with other ecotoxicologicalparameters (i.e. growth, development and reproductivesuccess) in aquatic invertebrates (Atienzar et al., 1999,2001; Atienzar and Jha, 2004). Combined with robustprotocols, such molecular approaches could play anincreasingly important role in determining the impactof genotoxins in the aquatic environment (Jha, 2004).

Despite the fact that low-LET ionising radiationsdelivered by external sources are suggested to be poorinducer of SCEs (Natarajan et al., 1993), in the presentstudy a clear dose-dependent increase for SCEs wasfound between 3.7 and 370 kBq ml−1 of HTO com-pared to controls.Straume and Carsten (1993)reporteda statistically significant increase for the incidence ofSCEs on bone marrow cells of mice maintained on111 kBq ml−1 of ingested HTO for 81–216 days. Sinceinduction of SCEs is S-phase dependent phenomenon,it could be suggested that3H interferes either with thereplication process or inhibits the actions of associatedenzymes, in common with other chemical genotoxins(Jha et al., 1992, 1996, 2000a,b).

Suyama et al. (1981)reported increased frequencyof Cabs (i.e. chromosome bridges) when eggs of thefreshwater teleost,Oryzias latipes, were treated withHTO delivering doses higher than 0.19 Gy. This iss ntlyee rela-t andm erew forC k ofd ld beh ellsa izedo mei onwb rvall aree ofa s ofd ose-

dependent increase for Cabs and PRI in the very earlylife stages of aquatic organisms could be related totheir known capacity to undergo apoptosis, thereby effi-ciently eliminating cells with high frequencies of Cabs,as observed in mammalian cells (Jha et al., 1995b). Inaddition, it is also possible that if highly damaged cellsare eliminated from the analysis, only the ‘bystandereffects’ are analysed (Morgan, 2003). Furthermore,apart from unstable Cabs as observed in the presentstudy, intake of HTO can also induce stable or balancedCabs that could be detected several years after intakeby humans (Lloyd et al., 1998). In parallel with humanhealth risk assessment, it is therefore important to takeinto account such stable genetic damage and potentialdelayed effects (Mothersill and Seymour, 1998) to pro-tect the environment.

As well as alterations in RAPD profile, induc-tion of SCEs and Cabs, HTO exposed embryo-larvaealso experienced lower values of PRI due to a reduc-tion in cell cycle progression probably caused by anincrease in toxicity. However, as mentioned above, nomarked difference for the PRI values could be foundin the embryo-larvae exposed to different concentra-tions of HTO. With respect to this apparent lack ofdose–response relationship for PRI, in addition to pos-sible explanations provided above, it should be pointedout that although originally proposed byLamberti etal. (1983)for application to actively proliferating andhighly differentiated mammalian cells in vitro, the PRIhas been found to provide an useful means to quan-t tico nts(e pre-s lationo cr ,1 rateo ndp ovet d tos ow-e icalg dif-f ionso nate,b ents

lightly higher than the dose that induced significalevated levels of Cabs in the embryo-larvae ofM.dulisin the present study. While a dose–response

ionship for SCEs, developmental abnormalitiesortality, were observed in the present study, thas a lack of clear dose-response relationshipsabs and PRI. The reasons for this apparent lacose–response relationships are not clear. It couypothesized that only critical sites or sensitive cre targets for radiation action, when these small-srganisms, undergoing cellular differentiation co

nto contact or uptake the radionuclides in commith higher organisms (Gossner, 2001). It could alsoe speculated that in the very sensitive embryo-la

ife stages, highly damaged cells leading to deathliminated from analysis, affecting both the yieldberrations and analysis of cells in different stageivision. In other words, the absence of a clear d

ify the cytotoxicity in embryo-larval stages of aquarganisms following exposure to different toxicaJha et al., 1996, 2000a,b; Hagger et al., 2002). How-ver, the PRI for the embryo-larvae probably reents an average figure, based on a mixed popuf highly differentiating cells with differing mitotiates over the exposure period described (Jha et al.996). Therefore, even a slight perturbation in thef cell cycle progression, cellular differentiation arogrammed cell death, for which as mentioned ab

hese cells have a well known capacity, could leaignificant changes in the PRI values. This is, hver, in contrast to our earlier studies using chemenotoxins, where PRI values showed significant

erences from controls only at the higher concentratf test compounds (e.g. methyl methane sulphoenzo (a) pyrene, tributyltin, contaminated sedimamples) in the embryo-larvae ofM. edulis and P.


dumerilii (Jha et al., 2000a,b; Hagger et al., 2002). Alower plateau-value for PRI at higher concentrationsof HTO in the present study therefore appears to berelated to the mechanisms of action of ionising radia-tion, which, in contrast to chemicals, could impart theireffect continuously the very moment cells come in con-tact with the radionuclides.

In contrast to induction of Cabs and PRI values,compared to unexposed controls, the study showeda significant dose-dependent increase for mortality inHTO exposed embryo-larvae by the end of the exper-iment. Abbott and Mix (1979)also observed similareffects on survival in the larval stage of goose barna-cles following exposure to seawater containing only0.2 Bq ml−1 of HTO. In contrast to the sexually matureadults, the embryo-larval stages of aquatic organismsare considered to be several orders of magnitude moresensitive in terms of toxicological injury. The high-est detrimental effects of radiation in both mammalsand aquatic organisms have been observed in exper-iments when irradiation occurred immediately afterfertilisation. It has also been suggested that the irra-diation of cells immediately after fertilisation mayalso encompass the sensitive period of second mito-sis (Anderson and Harrison, 1990). It is thereforenot surprising that 1h old embryo showed extensivedevelopmental abnormalities and mortality followingexposure to3H, an internal emitter. In this context,Knowles and Greenwood (1997)also found that thedevelopment and survival of eggs to larvae of thep -n lorhe( tageo n-t or-t nts

n-t esents ifi-c (cor-r PRI( erc ntra-

tion (0.37 kBq ml−1) of HTO, no significant increasefor SCEs and aberrant cells were found at this concen-tration. These discrepancies in ‘concentration–effect’relationships for the lowest concentration of HTOappear to be odd. In fact, studies in the literature simul-taneously evaluating concentration dependent effects atdifferent levels of biological organisation (i.e. chromo-somal, cellular and individual) are very scarce. Fromthe observations made in the present study, it appearsthat there is not a direct, linear relationship betweenobserved effects at lower (i.e. chromosomal and cellu-lar) and higher (e.g. developmental abnormalities andmortality) levels of biological organisation. It is also tobe remembered here that while cytogenetic effects (i.e.SCEs, Cabs and PRI) were evaluated soon after HTOexposure (one and half to two cell cycles), the develop-mental abnormality and survival effects were analysedat later stages (data presented for 72 h post-fertilisationperiod). It appears therefore that induced damage atlower levels of biological organisation, which couldhave knock-on effects as a function of time, couldprecipitate cumulatively and thus be manifested laterduring the developmental process at the individuallevel.

In conclusion, our integrated study linking geneticand cellular biomarkers with embryo-larval bioassaysuggests that extremely low doses of�-radiation emit-ted by HTO can induce genetic damage in the embryo-larval stages of marine mussels and that this damagep ev-e lifes ponser theb log-i ent( ew

A

byt C),E ina L),D theM RB-4

olychaete worm,Ophryotrocha diadema, was sigificantly reduced when exposed to 2.24 MBq m−1

f HTO but was not affected by similar doses of�-adiation delivered by137Cs. Indeed,�-rays from HTOave been shown to be more effective than�-raysmitted by90Sr, 90Y (in solution), 60Co and X-raysexternal sources) in inducing Cabs in the blastula sf Oryzias latipes(freshwater teleost) eggs, pote

ially leading to an increase in abnormality and mality (Suyama et al., 1981), as observed in the presetudy.

While having a global view of all the conceration or dose-dependent responses in the prtudy, it is interesting to observe that while signantly decreased values of normal embryo-larvaeespondingly increased frequencies of deads) andwhich had a ‘plateau level’ effect at all the highoncentrations) were found at the lowest conce

otentially accumulates at cellular and individual lls. Such an approach could be adopted to earlytages of other invertebrates to establish dose–reselationships since estimation of dose received byiota and resultant effects are important in radio

cal protection criteria for the natural environmPentreath, 1998; ERICA, 2005). Our study goes somay towards achieving this goal.

cknowledgements

This work was supported financially, mainlyhe Natural Environmental Research Council (NERnvironment Agency (EA) (GT22-1998-EA-0003)ssociation with Devonport Royal Dockyard (DMevon, UK. The RAPD work was supported byarie Curie Grant of the European Commission (E001-GT-97-0136).


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