Chemosphere 93 (2013) 1064–1069

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Differential survival and reproductive performance across threemitochondrial lineages in Melita plumulosa following naphthaleneexposure

0045-6535/$ - see front matter Crown Copyright � 2013 Published by Elsevier Ltd. All rights reserved.http://dx.doi.org/10.1016/j.chemosphere.2013.05.079

⇑ Corresponding author. Tel.: +61 2 9995 5081; fax: +61 2 9995 5183.E-mail addresses: pann.chung@unsw.edu.au (P.P. Chung), w.ballard@unsw.

edu.au (J. William O. Ballard), ross.hyne@environment.nsw.gov.au (R.V. Hyne).

Pann Pann Chung a, J. William O. Ballard a, Ross V. Hyne b,⇑a School of Biotechnology and Biomolecular Sciences, University of New South Wales, Sydney 2052, NSW, Australiab Centre for Ecotoxicology, NSW Office of Environment and Heritage, P.O. Box 29, Lidcombe 1825, NSW, Australia

h i g h l i g h t s

�We test the response of an invertebrate bioindicator to naphthalene exposure.� Animals of different mitochondrial lineages demonstrated differential survivorship.� Different mitochondrial lineages also showed differential reproductive performance.

a r t i c l e i n f o

Article history:Received 5 April 2013Received in revised form 24 May 2013Accepted 25 May 2013Available online 22 June 2013

Keywords:CrustaceanAmphipodSedimentBiomonitoringPAH

a b s t r a c t

Populations subject to anthropogenic contaminants often display altered patterns of genetic variation,including decreased genetic variability. Selective pressures of contaminant exposure are also reflectedin differential tolerance between genotypes. An industrial chemical spill in a major eastern Australianwaterway in July 2006 resulted in altered patterns of genetic variability in a nearby population of theamphipod, Melita plumulosa for up to one year post-spill, despite the site being declared clean after48 h. Here, we investigate the toxicant response of three mitochondrial lines naturally occurring at theimpacted site by comparing survivorship and life-history trait variables following naphthalene exposure.Overall, M. plumulosa demonstrated differential survivorship between mitochondrial lines under expo-sure to high concentrations of naphthalene. In addition, we identified differential fecundity and frequen-cies of gravidity in female amphipods between the mitochondrial haplotypes examined. These findingssuggest that the patterns of genetic variability previously identified may be linked with differential tol-erance and/or reproductive performance between mitochondrial lineages.

Crown Copyright � 2013 Published by Elsevier Ltd. All rights reserved.

1. Introduction

Populations are subject to a variety of environmental stressorsthat exert selective pressures on their genetic structure, both nat-ural and anthropogenic in origin. The impact of anthropogenicstressors such as toxicants and contaminants on the genetic struc-ture of natural populations has come under increasing scrutiny asmore light is shed on the effects and mechanisms of contaminants.In some cases, exposure to anthropogenic contaminants has beenshown to increase genetic variability through increased mutationrates (e.g. Theodorakis and Shugart, 1997; Rinner et al., 2011).Conversely, other studies have shown contaminants to act as aselective pressure resulting in genetic depauperation of popula-tions (e.g. Keklak et al., 1994; Cohen, 2002). Where contaminants

were shown to exert selective pressure on a population, differen-tial survival between genotypes can also be demonstrated (Duanet al., 2001; Schizas et al., 2001). For example, three naturallyoccurring mitochondrial lineages in the copepod Microarthridionlittorale demonstrated differential survivorship following pesticideexposure (Schizas et al., 2001). Contaminant exposure can also re-sult in population changes at the organismal level by impactinglife-history traits including reproductive performance (Clarkeet al., 2009; Ringwood et al., 2009). Clarke et al. (2009) found thatramshorn snails exposed to treated sewage effluent were heavierand larger than unexposed animals, as well as being more fecund.The aim of this study was to determine whether different geneticlineages display differential tolerance and/or reproductive perfor-mance following toxicant exposure in the amphipod Melitaplumulosa.

One group of chemicals commonly present in many anthropo-genic contaminants is polycylic aromatic hydrocarbons (PAHs).PAHs encompass a large class of toxic compounds comprised of

P.P. Chung et al. / Chemosphere 93 (2013) 1064–1069 1065

two or more adjoined benzene rings, and are prevalent in manypetrochemical and crude oil products. As such, PAHs are commonthroughout many aquatic environments as a result of industrialeffluent discharge, petrochemical spills and petroleum refineryand combustion (Walker et al., 2001). Low molecular weight frac-tions such as naphthalene and naphthalene compounds are knownto be acutely toxic to many aquatic species, although chronic tox-icity is unlikely due to the volatile nature of these compounds (Wuet al., 2012). Naphthalene and naphthalene compounds readily dif-fuse through tissues such as gills and have been shown to bioaccu-mulate in various tissues in numerous aquatic invertebrate species(Laurén and Rice, 1985; Mhadhbi et al., 2010). Juvenile animals areparticularly sensitive to naphthalene due to their smaller body sizeand higher diffusion rates (Aas et al., 2000). The reproductive per-formance of a population as well as other population and life-his-tory traits have also been shown to be impacted by naphthaleneexposure (Marquis et al., 2006; Krång, 2007; Pollino et al., 2009).In one study, sublethal exposure of crabs to naphthalene wasshown to disrupt male pheromone reception resulting in decreasedmate search behaviors (Krång, 2007). Another study on rainbowtrout found that naphthalene exposure resulted in the disruptionof sex steroid production leading to increased testosterone levelsin both male and female fish, and thus an increase in male individ-uals throughout the population (Pollino et al., 2009).

The amphipod M. plumulosa is endemic to many estuarine envi-ronments along the eastern coast of Australia (Zeidler, 1989) and iscurrently being utilized as the major invertebrate indicator formonitoring the health of estuarine sediments. This invertebratespecies is relatively abundant and simple to sample, as well asbeing amenable to laboratory culture (Hyne et al., 2005). Itsemployment as an indicator of estuarine health is due to its sensi-tivity to a range of contaminants including metals and PAHs inboth aqueous and sediment-bound phases (King et al., 2005; Mannet al., 2009; Simpson and Spadaro, 2011). Juvenile animals in par-ticular show greater sensitivity to contaminants than adult animalsin both aqueous and sediment exposures (King et al., 2006).

In July 2006, an industrial toxicant spill occurred in the upperreaches of the Parramatta River in Sydney, Australia. This resultedin approximately 3000 L of a highly flammable tile spray contain-ing an acrylate/methacrylate co-polymer and a variety of aromatichydrocarbons in solvent naphtha to leech into surrounding watersand sediments (Land and Environment Court of NSW, 2008). A pre-vious study found that the genetic structure of M. plumulosa at asite 1 km downstream of the spill were significantly impacted fol-lowing the chemical spill (Chung et al., 2011). A significant increasein individuals harboring mitochondrial the cytochrome c oxidasesubunit I (COI) defined haplotype 1 and a significant decrease inthe level of genetic variability as determined by Tajima’s D werefound for up to one year post-spill, a pattern not found at othermore distant localities along the Parramatta River (Chung et al.,2011).

We hypothesize that amphipods harboring different mitochon-drial types will be differentially affected in their survival and/orreproductive performance in response to naphthalene exposure.Specifically, we hypothesize that amphipods harboring haplotype1 which was found in excess following the spill are likely to havehigher rates of survival or higher reproductive performance follow-ing exposure. To test this hypothesis, cultured amphipods harbor-ing haplotypes 1, 3 and 8 naturally occurring at the site of the spillwere subject to a pulse exposure of naphthalene. These exposureand recovery conditions are designed to mimic the conditions ofthe chemical spill of July 2006 for the purpose of examining possi-ble molecular causes for the altered pattern of genetic variabilityidentified previously (Chung et al., 2011). Although the specificchemical composition of the tile spray from the chemical spill isunknown, we chose to test the effect of naphthalene, as it makes

up much of the water soluble component of the tile spray and isknown to be acutely toxic to aquatic organisms. Here, we identifydifferential survival between mitochondrial haplotypes followingexposure to a relatively high concentration of naphthalene, anddifferences in fecundity following recovery from low concentrationexposure. We also identify differences in the frequencies of femalegravidity between mitochondrial lineages. These results may par-tially account for the patterns of genetic variability previouslyidentified (Chung et al., 2011).

2. Materials and methods

2.1. Test organisms

To investigate possible molecular links between contaminantexposure and the altered patterns of genetic variability in amphi-pods impacted by the chemical spill of July 2006, we examinedthree mitochondrial haplotypes found at the impacted localitypost-spill. Individuals harboring haplotype 1 were found to be inexcess within the 12 months immediately post-spill, whereas thesecond most common haplotype at Duck River (haplotype 3) waslargely absent (Chung et al., 2011). We also chose to test amphi-pods harboring haplotype 8 as a control. It is a rare haplotype iden-tified in samples collected one year after the chemical spill (Chunget al., 2011). Haplotypes 1 and 3 differ by five synonymous basechanges; haplotype 8 is an intermediate haplotype between haplo-types 1 and 3, and differs from haplotype 1 by three synonymousbase changes and from haplotype 3 by two synonymous basechanges.

Adult M. plumulosa were sampled at the Duck River site(33�49028.7400S, 151�03005.1800E) in Sydney, Australia in December2008. This locality corresponds to the site previously found to beimpacted by the chemical spill of July 2006, and is where amphi-pods harboring mitochondrial haplotypes 1 and 3 are most com-mon (Land and Environment Court of NSW, 2008; Chung et al.,2011). The Duck River site is subject to chronic contamination asit is located in a heavily industrialized region and lies immediatelydownstream of a petrochemical refinery. Melita plumulosa werealso sampled from a single locality from the George Rivers at DavyRobinson Point (33�55050.1000S, 150�58007.7000E) in Sydney, Austra-lia, where amphipods harboring haplotype 8 are common (unpub-lished data). The Davy Robinson Point locality adjoins a park alongthe middle reaches of the Georges River and is considered rela-tively uncontaminated.

Cultures from single female and male pairs (isofemale cultures)were established from sampled animals, with founding animals re-moved after 4 weeks to minimize environmental carry-over effects(Hercus and Hoffmann, 2000; Hyne et al., 2005). A 630 bp region ofthe mitochondrial COI was then sequenced according to Chunget al. (2008) to identify the three haplotypes of interest. A singleculture of each haplotype was then retained and maintained understandard clean laboratory conditions (Hyne et al., 2005) for oneyear prior to the establishment of cyto-nuclear introgression lines.

To investigate the relationship between mitochondrial haplo-type and toxicant response, cyto-nuclear introgression lines (i.e.known mitochondrial lines against various nuclear backgrounds)were established using the three mitochondrial lines of interest.This strategy is expected to mimic the naturally occurring popula-tion if there is random mating between females harboring distinctmitochondrial types (Ballard et al., 2002) and enables a more rigor-ous examination of the effect of mitochondrial type on toxicant re-sponse. In January 2010, 5–15 females from each mitochondrialline were placed into fresh culture media with 5–15 males fromeach line in each of the nine possible combinations. Four replicatesper line were established for each cross and subsequently main-

1066 P.P. Chung et al. / Chemosphere 93 (2013) 1064–1069

tained under standard laboratory conditions for one year prior totoxicity assays. As the maximal lifespan of M. plumulosa under lab-oratory conditions is 11 ± 1 SEM months, maintaining cultures fora minimum of two years prior to the study removes the likelihoodof carry-over environmental effects from founding animals (Hercusand Hoffmann, 2000).

2.2. Test solutions and general chemistry

Seawater from Port Hacking, Australia (0.45 lm filtered, salinityadjusted to 25‰ with filtered and dechlorinated Sydney tap waterthen temperature acclimated to 24 �C for a minimum 24 h) wasused for culturing amphipods as well as testing. A 60 g L�1 naph-thalene stock solution was prepared using P99% scintillation-grade naphthalene (Sigma–Aldrich, St. Louis, USA) in 99% ethanolcarrier. Preliminary tests indicated no effect of the ethanol carrieron toxicity response at the concentrations utilized here (maximumethanol concentration of 0.025% v/v in the test solutions). Thenaphthalene stock solution was then diluted with seawater to re-quired concentrations in volumetric flasks. All experimental glass-ware used were washed then soaked overnight in 10% (v/v) HNO3

before thorough rinsing with deionized water.Prior to testing, equal quantities of an artificial fine-milled silica

substrate (Silica-300G, Unimin, Lang Lang, Australia) containing 1%(w/w) a-cellulose (Sigma–Aldrich, St. Louis, USA) were measuredinto test beakers and saturated with respective test solutions (sea-water or naphthalene-seawater solutions) (Mann et al., 2011). Thisartificial substrate mixture has been shown to mimic the proper-ties and total organic carbon content of natural sediments andovercomes potentially confounding effects as a result of differencesin composition in natural sediment (Mann et al., 2011). Beakerswere covered with plastic film and allowed to settle and equili-brate overnight. All solutions were replaced and allowed to settlebriefly immediately prior to the addition of test animals.

To ensure the consistency of test conditions and water chemis-try, measurements of temperature, pH, dissolved oxygen and salin-ity were conducted using calibrated WTW Multi 340i probes(Wissenschaftlich-Technische Werkstätten, Weilheim, Germany).Naphthalene concentrations in both control and test solutionswere monitored by extracting 1 mL seawater samples with 4 mLcyclohexane (RCI Labscan, Bangkok, Thailand) by vigorous mixingin glass scintillation vials. Concentrations were measured using aShimadzu UV-2550 UV–visible spectrophotometer (Kyoto, Japan)at an absorption wavelength of 275 nm employing a 1 cm path-length cell (Maeda et al., 2012). An extraction efficiency of 81.2–85.1% was obtained. Naphthalene concentrations in the overlyingwater of the toxicity tests were compared against a predeterminedstandard calibration curve, which was constructed using a series ofnaphthalene stock solution dilutions in cyclohexane over the range0–20 mg L�1.

2.3. Amphipod mortality tests

To determine appropriate naphthalene test concentrations andsampling time points, preliminary range determination tests wereconducted using juvenile amphipods (King et al., 2006). Firstly,preliminary data showed that naphthalene concentrations in thewater column decreased to half the starting concentration after6 h in an aerated system and were negligible after 24 h due tothe volatility of naphthalene. In a closed system (i.e. test beakerscovered tightly with plastic film, no aeration to minimize naphtha-lene aerosolization) concentrations remained at P75% after 24 h.Secondly, time-to-death tests indicated that 2.5 mg L�1 naphtha-lene was the concentration at which 50% mortality (LC50) wasreached after 24 h exposure within a closed system; at doublethe concentration (5 mg L�1) 50% mortality was reached after 2 h

exposure. We therefore chose to assay mortality rates in a closedsystem at hourly time points for the first 6 h, with a final timepoint at 24 h using the two naphthalene concentrations above.Water chemistry was monitored every 3 h as described above.

Juvenile animals between 1 and 3 weeks old were isolated intoclean cultures using a 212 lm sieve 24 h prior to the commence-ment of tests and fed with Sera Micron� Fry Food (Sera, Heinsberg,Germany) following standard culture conditions (Hyne et al.,2005). A total of 72 assays were conducted. There were two treat-ments (treated with naphthalene and control), three mitochondrialtypes (1, 3 and 8) with three replicates (wild-type and the twointrogression lines) and four blocks (each set of 18 lines repeatedover 4 d). Immediately prior to the commencement of tests, 25juvenile animals were distributed into each test beaker. In caseswhere there were insufficient animals, as many juveniles as couldbe collected were equally distributed across each replicate (mini-mum six amphipods). Amphipods were not fed and test solutionsnot replaced throughout the duration of the tests.

2.4. Life-history trait variables

Naphthalene and other PAHs are known to be acutely toxic toaquatic organisms and impact reproductive performance (Marquiset al., 2006; Pollino et al., 2009). We assayed surviving amphipodsfrom the mortality tests following recovery and maturation underclean conditions in consideration with the post-spill assessment(Land and Environment Court of NSW, 2008) and previous findings(Chung et al., 2011).

To investigate the possible links between naphthalene expo-sure, mitochondrial haplotype and life-history trait variables, weexamined female fecundity (measured as the number of embryosper female), and the proportion of gravid females present. (Chunget al., 2008; Mazurová et al., 2010). All surviving juvenile amphi-pods at the conclusion (24 h exposure) of the 2.5 mg L�1 naphtha-lene tests were transferred to clean culture media and maintainedunder standard laboratory conditions for 8 weeks (i.e. until sexualmaturity). Upon maturity, these amphipods were then harvestedfor life-history trait variable analyses.

2.5. Statistical analysis

Firstly, we analyzed differences in survivorship using a two-way analysis of variation (ANOVA, a = 0.05), examining the interac-tion between naphthalene exposure and mitochondrial haplotype.Preliminary analysis showed no significant differences betweenreplicate day blocks, and therefore day blocks were also pooledfor subsequent analyses. The time point at which 50% survivorshipwas reached was analyzed for both concentrations (2 h time pointfor the 5 mg L�1 exposure and 24 h for the 2.5 mg L�1 exposure).Survivorship frequencies were arcsine square-root transformed tonormalize data (Sokal and Rohlf, 1995). As an additional compari-son, the time to 50% lethality (LT50) was also analyzed using one-way ANOVA for the 5 mg L�1 exposure test. Where statistically sig-nificant differences were detected in the ANOVA models, Tukey’sStudentized post-hoc tests were conducted to determine pairwisedifferences between and within treatments (a = 0.05). All statisti-cal analyses were conducted using JMP v5.0 (SAS Institute, Cary,USA).

For the life-history trait variables analyses, fecundity embryocounts were square root transformed, and proportions of gravid fe-males were arcsine square-root transformed to meet ANOVAassumptions (Sokal and Rohlf, 1995). Tukey’s Studentized post-hoc tests were conducted to determine pairwise differences be-tween and within treatments (a = 0.05) where statistically signifi-cant differences were detected in the ANOVA models. Statisticalanalyses were conducted with JMP v5.0 (SAS Institute, Cary, USA).

ControlsExposed

P.P. Chung et al. / Chemosphere 93 (2013) 1064–1069 1067

3. Results

3.1. Water chemistry

Water quality parameters (temperature, dissolved oxygen,salinity and pH) in the overlying water of all the naphthalene expo-sure tests were similar for the control and treated samples. For theduration of the tests, temperature remained between 24.3 and24.6 �C, dissolved oxygen persisted at P79% saturation, salinityvaried from 24.6‰ to 25.3‰, and pH values ranged from 7.71 to8.01. Naphthalene concentrations did not fall below 75% of thestarting concentration at the termination of the exposure. Nonaphthalene was detected in any of the control solutions.

3.2. Juvenile survival

Survivorship at 5 mg L�1 naphthalene exposure decreased rap-idly across all three mitochondrial lines within the first 6 h ofexposure, and was near or at 0% after 5 h exposure (Fig. 1). For con-trol animals, survivorship was greater than 97% for all lines andtime points over the duration of the test (data not shown). After2 h (i.e. approximately 50% mortality) ANOVA identified a signifi-cant effect of naphthalene exposure (F1,67 = 997.13, p < 0.001) andmitochondrial haplotype (F2,67 = 5.94, p < 0.01) but no significantnaphthalene exposure � haplotype interaction effect (F2,67 = 2.86,p = 0.07). Examining only data for the exposed animals, the signif-icant difference in survivorship between mitochondrial lines wascorroborated (F2,33 = 4.51, p = 0.02). Tukey’s post-hoc test identifiedamphipods harboring haplotype 3 to be significantly greater innumber than those harboring haplotype 8 after 2 h naphthaleneexposure (Fig. 1). The LT50 of haplotype 1 was 1.26 h ± 0.17 SEM,for haplotype 3 it was 1.63 h ± 0.16 SEM, and for haplotype 8 itwas 0.95 ± 0.09 SEM. ANOVA conducted on the LT50 of exposedanimals corroborated the findings (p = 0.048), with Tukey’s testidentifying amphipods harboring haplotype 3 showing a higherLT50 than those harboring haplotype 8.

At 2.5 mg L�1 naphthalene exposure, mean survivorship wasgreater than 90% across all three mitochondrial lines at all timepoints assayed within the first 6 h (Supp. 1). Mean survivorshipacross the three mitochondrial lines after 24 h exposure ranged be-tween 55% and 73%. Survivorship among control animals wasgreater than 98% for all lines and time points over the durationof the test (data not shown). A significant effect of naphthaleneexposure was identified (F1,69 = 149.59, p < 0.001). There was nosignificant difference between mitochondrial lines after 24 h expo-

0

0.2

0.4

0.6

0.8

1

0 1 2 3 4 5 6 24

Haplotype 1Haplotype 3Haplotype 8

Prop

ortio

n su

rviv

ors

Exposure (h)

AAB

AB

Fig. 1. Proportion of surviving juvenile Melita plumulosa by mitochondrial haplo-type over a 24 h exposure to a 5 mg L�1 concentration naphthalene. Error barsrepresent standard error of the mean. Tukey’s post-hoc test was employed to detectdifferences between mitochondrial haplotypes after 2 h exposure. Letters aboveeach haplotype denote statistical significance, where haplotypes with the sameletter do not differ significantly.

sure (F2,69 = 2.18, p = 0.12), and no significant interaction effect(naphthalene exposure � haplotype, F2,69 = 1.74, p = 0.18). For dataincluding only the exposed animals, no significant differences be-tween mitochondrial lines were identified (F2,34 = 2.12, p = 0.14).

3.3. Life-history variables

To draw links between naphthalene exposure, mitochondrialbackground and organismal level responses to exposure, we as-sayed two life-history traits after an eight-week period of recovery.Female fecundity as measured by the number of embryos per fe-male was found to be significantly impacted by naphthalene expo-sure (F1,417 = 10.99, p < 0.01), but not haplotype (F2,417 = 0.69,p = 0.50) or a naphthalene exposure � haplotype interaction(F2,417 = 1.70, p = 0.18). Tukey’s test on exposed and unexposed ani-mals identified female fecundity to be significantly higher post-exposure among amphipods harboring haplotype 1 (Fig. 2); no sig-nificant differences were detected within haplotypes 3 and 8.

The gravidity of amphipods following naphthalene exposurewas found to decrease across all mitochondrial lines (Fig. 3).Amphipod gravidity was found to be significantly affected bynaphthalene exposure (F1,66 = 5.94, p = 0.02) and mitochondrialhaplotype (F2,66 = 3.62, p = 0.03). No significant interaction effectbetween exposure and mitochondrial type was identified(F2,66 = 0.70, p = 0.50). Tukey’s post-hoc test suggested amphipodsharboring haplotype 3 had significantly more gravid females thanthose harboring haplotype 1 (Fig. 3).

4. Discussion

Where contaminants exert selective pressures on populationsdifferent genotypes may confer differential susceptibility to thecontaminants (Keklak et al., 1994; Schizas et al., 2001), or resultin differential life-history trait trade-offs, such as differentialreproductive success (Clarke et al., 2009; Ringwood et al., 2009).These may in turn result in altered patterns of genetic variabilityin natural populations following contaminant exposure. Wehypothesized that amphipods harboring the mitochondrial haplo-type found to be in excess following a solvent naphtha chemicalspill are likely to demonstrate greater survivorship, or greaterreproductive performance when subject to naphthalene exposure.

4

6

8

10

Haplotype 1 Haplotype 3 Haplotype 8

Embr

yos

/ fem

ale

Mitochondrial background

B ABA AB AB A

104 46 91 71 71 35

Fig. 2. Fecundity of Melita plumulosa by mitochondrial haplotype following a 24 hpulse exposure to a 2.5 mg L�1 concentration naphthalene. Juvenile test animalswere allowed to recover and reach sexual maturity (8 weeks post-exposure) underclean conditions then fecundity as measured by the number of embryos per femalewas assayed. Error bars represent standard error of the mean. Tukey’s post-hoc testwas employed to detect differences in the interaction between naphthaleneexposure and mitochondrial haplotypes. Letters above each bar denote statisticalsignificance, where bars with the same letter do not differ significantly. Numberswithin each bar represent the number of gravid females counted.

0

0.2

0.4

0.6

0.8

Haplotype 1 Haplotype 3 Haplotype 8

ControlsExposed

Prop

ortio

n gr

avid

Mitochondrial background

B A AB

104 46 91 71 71 35

Fig. 3. Gravidity of female Melita plumulosa by mitochondrial haplotype following a24 h pulse exposure to a 2.5 mg L�1 concentration naphthalene. Juvenile testanimals were allowed to recover and reach sexual maturity (8 weeks post-exposure) under clean conditions then the proportion of gravid females wasassayed. Error bars represent standard error of the mean. Tukey’s post-hoc test wasemployed to detect differences between mitochondrial haplotypes followingexposure. Letters above each mitochondrial haplotype denote statistical signifi-cance, where haplotypes with the same letter do not differ significantly. Numberswithin each bar represent the number of gravid females counted.

1068 P.P. Chung et al. / Chemosphere 93 (2013) 1064–1069

The results from this study suggest a link between mitochon-drial haplotype and toxicant exposure survivorship in M. plumul-osa. Most generally, amphipods harboring haplotype 3 showedhighest survivorship over those harboring haplotypes 1 and 8. Fur-ther, haplotype 3 survivorship was significantly higher than haplo-type 8 over the course of the naphthalene exposure at the higherconcentration. No significant differences in survivorship were de-tected across the mitochondrial lines at the lower naphthaleneconcentration though this same trend was observed. Amphipodsharboring both haplotypes 1 and 3 were sampled from the chron-ically contaminated Duck River locality where these haplotypes aremost common (Chung et al., 2011). Amphipods harboring haplo-type 8, however, were sampled from the clean Georges River,and are rare at the Duck River locality (Chung et al., 2011). Chronicexposure may result in the elimination of individuals more suscep-tible to toxicants and may explain why amphipods harboring hap-lotypes 1 and 3 are more prevalent at the Duck River locality. Forinstance, a study by Keklak et al. (1994) found that mosquitofishsubject to chronic uranium exposure demonstrated significantlyhigher tolerance to uranium than mosquitofish from a cleanenvironment.

Naphthalene exposure was also found to significantly impactthe fecundity of M. plumulosa. Following recovery from naphtha-lene exposure all three mitochondrial lineages demonstrated in-creased female fecundity as compared to unexposed controlanimals. Furthermore, amphipods harboring haplotype 1 werefound to be significantly more fecund post-exposure than controlanimals. Increased fecundity following toxicant exposure has beenobserved in other aquatic species and has been suggested to be acompensatory mechanism in response to increased mortality ordecreased egg quality (Cooley, 1973; Brausch et al., 2009). We alsoidentified differences in gravidity across the three mitochondriallines tested. Amphipods harboring haplotype 3 had a greater pro-portion of gravid females than those harboring haplotype 1. A de-crease in the proportion of gravid females was also found across allthree mitochondrial lineages following naphthalene exposure.Combined, these results suggest increased fecundity post-exposuremay serve as a compensatory effect for the decrease in overall rateof gravidity of the population. To critically examine the potentialfor a compensatory response future studies should test the mito-chondrial functions of amphipods following toxicant exposure.

The results of this laboratory study go some way to explain thealtered patterns of genetic variation among impacted amphipodsfollowing the chemical spill of July 2006 in the Parramatta River(Chung et al., 2011). From the findings of the previous study, wehypothesized that amphipods harboring mitochondrial haplotype1 found in excess for up to one year post-spill (Chung et al.,2011) would show greater survivorship or reproductive perfor-mance following naphthalene exposure. The data here indicatethat although amphipods harboring haplotype 1 are not the mostresistant to naphthalene exposure at either concentration tested,these amphipods show increased fecundity post-exposure andmay repopulate faster being more numerous even prior to theindustrial spill (Chung et al., 2008, 2011).

A limitation of this study is that the mitochondrial lines em-ployed here are based on a �630 bp region of the COI gene (Chunget al., 2008, 2011). To directly assess the functional properties ofthe mitochondrial lineages examined, bioenergetics assays willserve to elucidate the metabolic differences between each lineagein both the short-term toxicant exposure and long-term recoveryresponse (e.g. Coghlan and Ringler, 2005; Pichaud et al., 2012).Alternatively, assaying nuclear loci will likely identify other infor-mative markers of exposure to contaminants such as the cyto-chrome P450 genes (see Guengerich, 2008 for a review), whichhave been linked with response to PAH and naphthalene com-pound exposure in copepods and fish (Aas et al., 2000; Hansenet al., 2008). Another limitation is that we only tested the toxicityresponse of the naphthalene fraction of the tile spray from the July2006 spill. As neither the tile spray nor its exact chemical compo-sition are available, we chose to test the naphthalene fraction be-cause it was a major constituent and has a known toxicity toaquatic life. Investigation of the toxicity response to other knownfractions of the tile spray may elucidate the strongest selectiveagents in this and other industrial chemicals. These measuresmay give additional insight into the pattern of genetic variationidentified after the chemical spill (Chung et al., 2011).

5. Conclusion

The results of this study show differential response to toxicantexposure as well as intrinsic organismal-level differences betweenmitochondrial lineages in the amphipod M. plumulosa. These dataalso show correlation with the patterns of genetic variation previ-ously identified following a chemical spill, and suggest that thechemical spill did significantly impact the genetic structure ofamphipods in the vicinity. However, we have not demonstrated acausal link between the genetic locus upon which our mitochon-drial lineages have been classified and the phenotypic variationsobserved. Testing the mitochondrial functions of amphipods har-boring different mtDNA types before and after exposure to naph-thalene has the potential to give a mechanistic link betweencause and effect.

Acknowledgements

The authors wish to express their gratitude to R.M. Mann (Uni-versity of Technology Sydney) for assistance with sample collec-tions. C.C. Correa, K.M. Cairns and W.C. Aw (University of NewSouth Wales) are also thanked for their comments on earlier ver-sions of the manuscript. The authors also wish to thank the twoanonymous reviewers who provided constructive comments thatimproved the manuscript. Sequencing was performed at TheRamaciotti Centre for Gene Function Analysis. This study was sup-ported in part by funds provided by the OEH/UTS Centre for Eco-toxicology, and P.P. Chung is a recipient of the AustralianPostgraduate Award.

P.P. Chung et al. / Chemosphere 93 (2013) 1064–1069 1069

Appendix A. Supplementary material

Supplementary data associated with this article can be found, inthe online version, at http://dx.doi.org/10.1016/j.chemosphere.2013.05.079.

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