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Aquatic Toxicology 86 (2008) 426–436
Evidence of population genetic effects of long-term exposure tocontaminated sediments—A multi-endpoint study with copepods
Johanna Gardestrom a,∗, Ulrika Dahl b, Ola Kotsalainen a, Anders Maxson b,Tina Elfwing a, Mats Grahn c, Bengt-Erik Bengtsson b, Magnus Breitholtz b
a Department of Systems Ecology, Stockholm University, SE-10691 Stockholm, Swedenb Department of Applied Environmental Science (ITM), Stockholm University, SE-10691 Stockholm, Sweden
c Department of Natural Sciences, Sodertorn University College, SE-14189 Huddinge, Sweden
Received 24 October 2007; received in revised form 9 December 2007; accepted 11 December 2007
bstract
In the environment, pollution generally acts over long time scales and exerts exposure of multiple toxicants on the organisms living there.ecent findings show that pollution can alter the genetics of populations. However, few of these studies have focused on long-term exposure ofixtures of substances. The relatively short generation time (ca. 4–5 weeks in sediments) of the harpacticoid copepod Attheyella crassa makes it
uitable for multigenerational exposure studies. Here, A. crassa copepods were exposed for 60 and 120 days to naturally contaminated sedimentsi.e., Svindersviken and Trosa; each in a concentration series including 50% contaminated sediment mixed with 50% control sediment and 100%ontaminated sediment), and for 120 days to control sediment spiked with copper. We assayed changes in FST (fixation index), which indicates ifhere is any population subdivision (i.e., structure) between the samples, expected heterozygosity, percent polymorphic loci, as well as abundance.here was a significant decrease in total abundance after 60 days in both of the 100% naturally contaminated sediments. This abundance bottleneck
ecovered in the Trosa treatment after 120 days but not in the Svindersviken treatment. After 120 days, there were fewer males in the 100%aturally contaminated sediments compared to the control, possibly caused by smaller size of males resulting in higher surface: body volume ration contact with toxic chemicals. In the copper treatment there was a significant decrease in genetic diversity after 120 days, although abundanceemained unchanged. Neither of the naturally contaminated sediments (50 and 100%) affected genetic diversity after 120 days but they all hadigh within treatment FST values, with highest FST in both 100% treatments. This indicates differentiation between the replicates and seems toe a consequence of multi-toxicant exposure, which likely caused selective mortality against highly sensitive genotypes. We further assayed two
rowth-related measures, i.e., RNA content and cephalothorax length, but none of these endpoints differed between any of the treatments andhe control. In conclusion, the results of the present study support the hypothesis that toxicant exposure can reduce genetic diversity and causeopulation differentiation. Loss of genetic diversity is of great concern since it implies reduced adaptive potential of populations in the face ofuture environmental change.2007 Elsevier B.V. All rights reserved.
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eywords: Long-term exposure; Contaminant mixtures; Biodiversity; Genetic
. Introduction
The conservation of biodiversity has been on both the sci-ntific and the political agenda ever since the Rio de Janeiroarth Summit in 1992. The emphasis is often on protecting
articular species and habitats but the declaration also com-rises the maintenance of functional diversity of ecosystemsnd genetic diversity of species. Relatively little attention has,∗ Corresponding author. Tel.: +46 8 16 37 04; fax: +46 8 15 84 17.E-mail address: [email protected] (J. Gardestrom).
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166-445X/$ – see front matter © 2007 Elsevier B.V. All rights reserved.oi:10.1016/j.aquatox.2007.12.003
ity; Genetic differentiation; RNA; Environmental risk assessment
owever, been drawn to changes in genetic diversity, especiallyaused by indirect or direct exposure to toxicants (Bickham etl., 2000). Recent findings show that pollution in fact can causeapid genetic changes in exposed populations; changes that maye complex and that may take place within very short time scalesi.e., over a few generations) (Gardestrom et al., 2006; Medinat al., 2007 and references therein). Direct effects may, e.g.,ccur when a substance cause damages on the molecular struc-
ure of the DNA, i.e., mutagenic effects, while indirect effects ofxposure are population-mediated processes that include alter-tions of the genetic variability in the population (De Wolf etl., 2005). Field studies have shown that mutations accumulate
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J. Gardestrom et al. / Aquat
ore rapidly in more polluted environments (Yauk and Quinn,996; Rogstad et al., 2003) and a result of this could be increasedenetic diversity. However, most mutations that accumulate inhe genome are deleterious, which result in lower viability ofhe population (Berckmoes et al., 2005). Another, and morerequently reported indirect outcome of toxicant exposure iseduced genetic diversity (Ma et al., 2000; Van Straalen andimmermans, 2002; Ross et al., 2002; Gardestrom et al., 2006).trong selective pressures may act to lower genetic diversity byliminating less fit phenotypes and eventually their associatedenotypes from a population (Van Straalen, 2003). Popula-ion bottlenecks (severe reduction in numbers of individuals)esulting from the toxic effects of pollutants may also lead to
decrease in genetic variation within populations inhabitinghronically polluted environments (Bickham et al., 2000).
At the individual level, a change in RNA content has proven toe a sensitive tool in understanding growth-related responses innvertebrates (e.g., Ibiam and Grant, 2005; Dahl et al., 2006).tress proteins are, for instance, produced upon exposure toeveral pollutants, a process which is preceded by increasedNA production (e.g., Feder and Hoffman, 1999; Becker et al.,000). Reduced somatic growth as a response to toxicant expo-ure may on the other hand result in decreased RNA productionnd protein synthesis (Dahl et al., 2006). Hence, by measuringndividual RNA content, normalizing the content against, e.g.,ody length, and comparing the normalized content betweenxposed and unexposed individuals, it is possible to determinef the toxic response, at a certain concentration, is a result ofeduced growth or stress-induced protein synthesis (Dahlhoff,004; Dahl et al., 2006).
In the environment, organisms are often exposed to sub-tances over a time scale that is in most cases unrealistic to mimicn the laboratory. In order to catch (realistic) population geneticffects, the exposure time has to extend over multiple genera-ions, which is a great challenge in laboratory ecotoxicologicaltudies. Further, in the environment toxicants act simultaneously,nd not as single substances. The single substance tests isolatehe effects of the compound of interest but in the natural environ-
ent it is not often that chemicals act in isolation. Another greathallenge in ecotoxicology is therefore to determine toxicity ofontaminant mixtures (Eggen et al., 2004).
Sediments accumulate large numbers of anthropogenic sub-tances in our environment and consequently represent a sourcef a complex mixture of potentially toxic compounds. Fornstance, metals and persistent organic pollutants accumulate inediments (Van Leeuwen, 1995; Carriger et al., 2006), therebyosing a risk to benthic and epibenthic organisms (Carriger etl., 2006; Dekker et al., 2006). The harpacticoid copepods arehe second most abundant group of species after the nematodesn the benthic environment (Huys et al., 1996; Coull, 1999) andepresent an important link in the food chain between primaryroducers and secondary consumers (Hicks and Coull, 1983).hey generally have short generation times, which mean that
tress responses can be easily studied over the full life cycle.wing to this, harpacticoid copepods have been used in chronicesting of single substances as well as complex mixtures forany years (e.g., Hutchinson et al., 1999; Breitholtz et al.,
blt(
icology 86 (2008) 426–436 427
003; Chandler et al., 2004; Gardestrom et al., 2006; Ulfsdotteruresson et al., 2007). Hence, these organisms are optimal bothor multigenerational (including population genetic) effects andediment exposure studies.
We have earlier shown that exposure to a single substancei.e., the polybrominated diphenyl ether BDE-47) can induceubtle genetic alterations in populations of the harpacticoidopepod Nitocra psammophila, although abundance remainednchanged (Gardestrom et al., 2006). In this study, with rel-tively controlled experimental conditions and short exposureime (i.e., corresponding to one generation), the results showedhat it is crucial to include endpoints at different levels ofiological organization in order to adequately interpret toxi-ant exposure. The major objective of the present study waso increase the ecological realism and test the hypothesis thatong-term laboratory exposure to contaminated sediments caneduce expected heterozygosity and genetically differentiateopulations. We also wanted to compare these two populationenetic endpoints with both another population-level endpoint,.e., population abundance, and responses at the individualevel, i.e., RNA content and cephalothorax length. To inves-igate this we exposed the fresh water harpacticoid copepodttheyella crassa in a laboratory experiment for about 3 gen-rations (i.e., 120 days) to different types of sediments: (i) aediment from a locality often used as a reference in biomonitor-ng (i.e., control); (ii) the control sediment spiked with copper;nd (iii) two sediments with well documented chemical bur-en.
. Materials and methods
.1. Test organism
A. crassa (Sars) has been cultured in our laboratory sinceugust 2004. The entire culture originates from one oviger-us female sampled August 10, 2004, from Lake Hallbosjon,rena, county of Sodermanland, Sweden (58◦50′N, 16◦40′E).
t has a generation time ranging from 6 to 8 weeks when cul-ured in the laboratory at 20–20.6 ◦C (Sarvala, 1977). However,f provided with sediments, the experiences from our culturesunpublished data) indicate a shorter generation time, closer to–5 weeks. When collecting harpacicoid copepods in the wild,t is more or less unavoidable to get a mix of several species. Inrder to identify one species it is necessary to kill and carefullyissect the individuals and look at a number of morphologicalttributes. In this case we had isolated a successful laboratorytrain (i.e., that reproduced well under laboratory conditions)nd that we considered useful. For that purpose, several individ-als belonging to this strain (originating from the same fertilizedemale caught in the wild) were sent for taxonomic determi-ation (Prof. J. Sarvala, Univ. Turku, Finland). Since it takesime to isolate and identify a single species that can survivender laboratory conditions, the described procedure is proba-
ly as close to natural levels of heterozygosity possible underaboratory conditions. Further information about the species andhe culture conditions is described in Ulfsdotter Turesson et al.2007).
428 J. Gardestrom et al. / Aquatic Toxicology 86 (2008) 426–436
Table 1Levels of analyzed metals in surface sediments (top 0.5–1 cm) from Asko, Svindersviken and Trosa
Site Metal
Cu (�g/g d.w.) Zn (�g/g d.w.) Pb (�g/g d.w.) Hg (�g/g d.w.)
Asko (reference)a 30.5–49.9 123–186 33.1–41.3 0.10–0.13Svindersvikenb 614–864 1110–1140 995–1080 7.8–28.6Trosac 1039 2966 55 n.a.
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SGAB (1998), Eriksson Wiklund and Sundelin (2002).b SGAB (1998), Sundelin and Eriksson (2001).c Eklund et al. (2006).
.2. The sediments
We used sediments collected from three different locations,ith different contaminant burdens (for documentation of somef the chemicals see Table 1). Our reference sediment was col-ected at Asko (58◦46′N, 17◦46′E) in February 2006. This siteas been used as a reference site in the national environmentalonitoring programme (site reference no 6022) for many years.he second sediment was collected in Svindersviken (59◦18′N,8◦7′E) in February 2006. Svindersviken is situated at the har-or inlet of Stockholm and the water has been polluted for over00 years by the scrapping of marine ships and the import ofetroleum products (Bengtson, 2004). Analyses of this sedi-ent have previously shown high levels of hydrocarbons from
etroleum residues (Bengtson, 2004) and heavy metals such asd, Pb, and Hg (Sundelin and Eriksson, 2001; Eriksson Wiklundnd Sundelin, 2002). Both the sediments from Asko and Svinder-viken were collected from the surface (0–2 cm) with a sedimentled (Blomvist and Lundgren, 1996). The third sediment wasollected in Trosa harbor (58◦53′N, 17◦432′E) in March 2006sing a Kajak-type tube corer. This site is located adjacent toboat wash operating during the summer, so this sediment is
resumably polluted by antifouling paints. This sediment hasreviously also been shown to contain Cu, Pb, and Zn (Eklundt al., 2006). The organic matter and water contents were deter-ined for all three sediments. This was done to assure that the
rganic matter content was within acceptable range (Asko, 7%;rosa 8%, and Svindersviken 12%) and to determine the dryeight to wet weight ratio of the respective sediments. All sed-
ments were stored at 4 ◦C until the start of the experiment inpril 2006. Before the initiation of the experiment, the sedi-ents were sieved through a 63 �m sieve and heated for 45 min
o 60 ◦C in order to eliminate any meiofauna present.
.3. Sampling tests
A sampling technique was devised where the content of eachxperimental beaker was sieved through a 63 �m sieve andinsed on a Petri dish, where the animals were counted usingight microscopy. The robustness of the sampling technique wasested by preparing 10 jars with 100 ml water, 1 g sediments,
0 copepods and 10 nauplii. The accuracy of the sampling tech-ique was satisfactory, with a mean recovery of nauplii of 75.5%;V% = 12.7, and a high mean recovery and low mean variabilityor copepodites/adults, i.e., 98% recovery and CV% = 4.9.
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.4. Experimental set up
24 h before the initiation of the experiment, the treatmentsere prepared by adding sediments corresponding to a dryeight of 0.5 g and 100 ml fresh water to 250 ml glass beakers.he water in the fresh water cultures was collected fromake Stensjon (Tyresta National Park; 59◦11′N, 18◦19′E). Itas filtered (0.03-mm paper filter), heated to 80 ◦C and thenF/C-filtered (1.2-�m glass micro-fiber filter) to eliminate
arger particles and organisms but to keep the water chemistrynaffected. Sixty 250 ml glass beakers were prepared, twelveeplicates in each treatment, as follows: 100% control (Asko sed-ment); 100% Svindersviken; 100% Trosa; 50% Svindersviken50% control; 50% Trosa + 50% control. These treatments areereafter referred to as control; Svindersviken100%; Trosa100%;vindersviken50%; and Trosa50%, respectively. Another sixeakers were prepared with 100% control sediment to which00 ml water with copper sulfate was added, correspondingo a concentration of 130 �g Cu/l (i.e., 96 h LC5 for adult A.rassa, unpublished data). Six of the twelve replicates in eachreatment were terminated after 60 days. The remaining six repli-ates were terminated after 120 days. The copper treatment wasampled at termination (i.e., after 120 days). For each treat-ent, an additional test beaker was prepared to serve as internal
ontrol, enabling monitoring of pH and oxygen levels withoutisturbing the copepods in the experimental beakers. These mea-urements were done weekly in the water phase, just above theediment surface and they showed that these levels were constanthroughout the experimental period. Since we used brackishater sediments in a freshwater system we also measured the
alinity in these extra beakers to assure that the water remainedresh throughout the experiment. The measurements showed thathe salinity at all times and in all test beakers was below 0.1‰.
As the design of this study demanded a large number of ani-als, the size of the starting population of each replicate was
estricted by the size of the A. crassa population in the continu-us culture. About 2000 adult copepods were isolated from theulture. 30 randomly chosen adult A. crassa from this isolatedool were then added to each replicate glass beaker. Due to thearge number needed, both males and females were used. Fur-her, the mating behavior of copepods is characterized by the
ale attaching to the female (i.e., precopula). When this hasccurred it is very difficult to separate the two, hence couplesere collected in order not to injure the animals. In all, the 30
nimals included: one ovigerous female, three females without
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ggs, five couples in which the females were ovigerous, andight couples in which the females were not. The experimentan at a constant temperature of 20 ◦C (±1 ◦C) in a daily regimef 14 h light and 10 h darkness. Because of evaporation, doubleistilled water was added once a week to keep a constant volumef 100 ml.
The animals were fed at the initiation of the experiment andhereafter once a week with the micro algae Pseudokirchner-ella subcapitata and Monoraphidium contortum (final densityn experimental beakers 1 × 106 cells/ml for each algal species)see Ulfsdotter Turesson et al., 2007, for further details onandling of cultures). This amount of food was consideredo be in excess. For instance, in a proposed OECD harpacti-oid copepod development and reproduction test guideline, themount of food (i.e., microalgae) given weekly is at maximum× 104 cells/copepod, regardless of stage (OECD, 2007). Given
hat the maximum number of animals in a test beaker in theresent study was about 400 for any treatment, the number offells given weekly was at least 7.5 × 105 cells/copepod (i.e.,100 ml × 3 algal species × 106 cells/ml)/400 animals). After 60ays and 120 days, respectively, the experiments were termi-ated. In each replicate, the animals were counted and classifieds nauplius, copepodite or adult. The adults were further sexetermined. The males were incubated overnight without food inresh water to allow emptying of gut content, in order to removeny non-Attheyella DNA present, and then stored in 99% alcoholt −20 ◦C until DNA extraction.
.5. Genetic analysis
DNA was extracted and purified from individual maleopepods according to Laird et al. (1991). 43 males wereampled from the continuous culture at the start of thetudy and 10 individuals were sampled from each repli-ate after 120 days. In some treatments (Trosa100% andvindersviken100%), however, there were some replicates thatontained fewer than ten males. In these replicates all malesound were sampled. The AFLP (amplified fragment lengtholymorphism) analysis was done according to Vos et al.1995) with minor modifications described by Bensch et al.2002). Genomic DNA was digested with EcoRI and MseI,ollowed by a pre-amplification step before the selective ampli-cation. Four sets of selective amplification primers weresed: (EcoRI + TAG/MseI + CCA, EcoRI + TAG/MseI + CTA,coRI + TCT/MseI + CCA, EcoRI + TCT/MseI + CTA) with thecoRI primers labeled with fluorescent dyes (FAM and NED,pplied Biosystem). In the AFLP-analysis the processing of
ndividual samples was randomized with regard to treatmentsnd replicates to avoid bias introduced by inconsistent restric-ion and amplification across batches and fragment analysisBensch and Akesson, 2005). The fragments were analyzedt Uppsala Genome Center (ABI PRISM® 3700 DNA Ana-yzer). The data were imported into Genographer version 1.6.0
Benham et al., 1999) for band calling. Each AFLP locus wasssessed and scored using the ‘thumbnail’ option of Genogra-her, which enables fluorescence signal strength distributionser locus to be compared across 200 individuals together. Thevgtr
icology 86 (2008) 426–436 429
ands were scored as being present (1) or absent (0), creatingpresent/absent matrix. Presence was assigned if an individ-
al had a band ≥100 fluorescence unit. Only bands above theize of 70 base pairs were scored since smaller bands should bereated with caution (e.g., Vekemans, 2002; Papa et al., 2005).nly loci where the frequency of band absence is higher than/N (where N = total sample size) was included in the analysisn order to avoid uncertain estimates of heterozygosity (Lynchnd Milligan, 1994).
.6. Individual growth analysis
Individual growth-related endpoints (i.e., cephalothoraxeasurements (�m) and RNA content (ng/ind)) were studied
ased on the rationale that rRNA levels (70–80% of total RNAontent) at any given time are directly related to the protein syn-hesis of a cell (Elser et al., 2000), which in turn may be inducedy, e.g., somatic growth or stress protein synthesis. RNA con-ent has been used as an estimate of growth in small copepodsor a long time (Sutcliffe, 1965) and changes in RNA contentDahlhoff, 2004) can be used to determine if an organism haseen exposed to a stressor, including environmental pollutantsYang et al., 2002).
For RNA content analysis, microplate fluorometric high-ange assay measurements of quantitative RNA (ng/ind) inndividual copepods were performed according to Gorokhovand Kyle (2002) and Dahl et al. (2006). Briefly, RNAlater pre-erved copepodites stage three (CIII) were stored for maximummonths at 4 ◦C (Gorokhova, 2005). CIII was used since it haseen shown to be the copepodite stage with the lowest indi-idual variability (Gorokhova, unpublished). Extraction using-laurylsarcosine and ice-cold ultra sonic bath was followedy RNase digestion. All standards (0.01, 0.08, and 0.12 �g/ml)nd samples contained a final sacrosyl concentration of 0.2%.luorescence measurements were performed using fluorometerLUOstar Optima (filters: 485 nm for excitation and 520 nm formission) and black solid flat-bottom microplates. Plates werecanned with 0.2 s well measurement time, 10 measurements perell. Copper-treated animals were not analyzed due to impor-
ance of prioritizing the animals from the naturally contaminatedediments. At time of analysis of copper-treated copepods, thenimals had exceeded the time of sustainable RNA preservationGorokhova, 2005). Cephalothorax length was measured in aicroscope slide with the imaging software Leica IM50, Imageanager (Leica DMIL).
.7. Statistical analysis
Statistical calculations on population abundance, cephalotho-ax lengths and RNA contents were performed using SPSSersion 15 (Lead Technologies Inc.). The Svindersviken, Trosa,nd copper treatments were analyzed separately, although theame control was used for all exposures. Homogeneity of
ariances was tested with the Levene test. Variances were homo-eneous in abundance in the Svindersviken100%, Trosa100%reatments and in all treatments with regards to cephalotho-ax length and RNA content and were hence analyzed with
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30 J. Gardestrom et al. / Aquat
ne-way ANOVA. Significant differences between specific treat-ents and the control were analyzed with Dunnet’s Post
oc test. When variances were heterogeneous (i.e., in thevindersviken50%, Trosa100%, and copper treatments), the non-arametric Kruskal–Wallis test was used. The Mann–Whitneyest was used to test significant differences between specificreatments and the control; Bonferroni corrections were madeo correct for compounded alpha error from multiple compar-sons with the same control. Abundance of various life stages inhe copper treatment was analyzed against the same life stagesn the control using either the t-test (homogenous variances)r the Mann–Whitney test (heterogeneous variances). Signifi-ant differences from the controls were generally accepted when< 0.05; when Bonferroni corrections were used the signifi-ance level was however corrected to 0.025 (i.e., equals twoomparisons against the same control).
The genetic data (the present/absent matrix) were importedo the program Hickory version 1.0.4 (Holsinger and Lewis,005), which use Bayesian statistics to obtain values of expectedeterozygosity across loci (Hs, i.e., genetic diversity) and theartitioning of the genetic diversity between samples (FST). TheST (fixation index) is an index that compares the expected het-rozygosity over all samples with the expected heterozygosityithin respective sample and thus indicates if there is any popu-
ation subdivision (i.e., structure) between the samples. Analysesere carried out on the individual level by using the individu-
ls within each treatment, i.e., the starting population (n = 43),ontrol (n = 60), Svindersviken50% (n = 60), Svindersviken100%n = 50), Trosa50% (n = 58), Trosa100% (n = 44), and coppern = 60), respectively, as replicate samples. The analysis byickory was also carried out for each treatment one by one,
till at the individual level, but in that analysis each replicateas treated as a population to test for subdivision within respec-
ive treatment, i.e., among the replicates. The Bayesian methodses standard Monte Carlo Markov Chain (MCMC) methodso approximate the posterior distribution of θ (refers to FST),nd F (an estimate of the inbreeding parameter FIS) (Holsingernd Wallace, 2004). It does it without treating dominant mark-rs as haplotypes and it does not assume that genotypes are inardy–Weinberg proportions within populations (Holsinger et
l., 2002). The MCMC method depends on convergence of thearkov chain to its stable distribution. The program takes sam-
les from the simulation chain after discarding a fixed number ofnitial iterations (the burn-in period) to ensure that the chain haseached a stable distribution. After the burn-in period, samplesre retained at fixed intervals (the thin) to minimize autocorrela-ion among samples. The sampler parameters used were; burn-in70,000 iterations), sampling (100,000), and thinning (5). Wesed the default prior, beta (1.1) for F, which is equivalent to aniform (0, 1) non-informative prior with a uniform distribution.he output of the run gave us four different models.
When using dominant markers, the F estimates can some-imes give unreasonable values in the “full model” which seems
ntrinsically related to the weak potential identifying FIS withominant markers (Holsinger and Lewis, 2005). The authorsuggest the use of “f free model” when that is suspectedHolsinger and Lewis, 2005). In this model, the program doesbSpp
icology 86 (2008) 426–436
ot attempt to estimate FIS so the estimate of FST is unaffected bynreasonable estimates of FIS. Since the FIS value we obtainedas not biologically realistic (0.98), we chose the “f free model”
nd the data were run three times to assure that the resultsbtained were consistent. Bayesian statistics do not produce theonfidence interval of classical statistics; instead the data areresented as 95% CI (credible intervals, Holsinger and Lewis,005). In order to test if the treatments differed from the controls,he mean and the 95% credible interval for the difference (con-rols − exposed) was calculated according to Berry (1996). The5% credible interval for the difference (controls − exposed)as calculated as: ±1.95 × the square root of the sum of the
quared standard error, i.e.,
((S.E. control)2 + (S.E. exposure)2).
To ensure that the estimate of FIS given by Hickory’s “fullodel” was unreasonable, the data were tested with AFLP-URV 1.0 (Vekemans, 2002). This program assumes thatardy–Weinberg genotypic proportions (FIS = 0) are true for
he data set so if the levels of heterozygosity from Hickory cane reproduced, FIS could not be 0.98. The sampler parametersere: 2000 permutations for test of FST and 1000 bootstraps
or genetic distance. The same data set was also run in a hier-rchical AMOVA (Analysis of Molecular Variance) with thexcel add-in package GenAlEx (Peakall and Smouse, 2006)sing each replicate beaker as a replicate unit to catch the par-itioning of the molecular variance between individuals withineplicates, between replicates within treatments, and betweenreatment components. Genetic diversity was also described asercent variable bands, calculated as the proportion of all bandshat were variable.
. Results
.1. Population abundance
The mean total abundance was significantly reduced invindersviken100% (p < 0.01) and Trosa100% (p < 0.001) after 60ays as compared to the control. However, there were no dif-erences in total abundance between the control and the mixedediments (i.e., Svindersviken50%, Trosa50%; Fig. 1A and B).fter 120 days, the mean total abundance did only differ betweenvindersviken100% and the control (p < 0.01, Fig. 1A).
The mean numbers of males after 60 days was signifi-antly lower in the Trosa100% treatment compared to the controlp < 0.001, Fig. 1B). After 120 days, the numbers of malesere significantly lower in both Trosa treatments and in thevindersviken100% treatment (Trosa50% p < 0.001; Trosa100%< 0.001; Svindersviken100% p < 0.001) compared to the con-
rol, while the number of females did not differ between theontrol and any of the treatments (Fig. 1A and B). Still, therosa100% treatment had quite few females (p = 0.071).
It was only at day 60 that a reduced mean num-
er of copepodites were detected. Both the Trosa100% andvindersviken100% treatments had significantly fewer cope-odites compared to the control (Trosa p < 0.05; Svindersviken< 0.05). There were no significant effects on the mean number
J. Gardestrom et al. / Aquatic Toxicology 86 (2008) 426–436 431
F llutedb ) deno
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ig. 1. Abundance of various life stages of A. crassa populations exposed to poars indicate the 95% confidence intervals of means (α = 0.05, n = 6). Asterisks (*
f nauplii. However, at day 60 the Trosa100% treatment had closeo significantly lower number nauplii compared to the controlp = 0.058).
No significant differences could be observed between theopper treatment and the control in mean total abundance. How-ver, the copper treatment had more females (p < 0.05, Fig. 2).
.2. Genetic variability
The genetic diversity could only be assessed after 120 days,imply because there were not enough males to sample inost of the replicates after 60 days. The Hickory analysis
f all treatments and the control together gave an FST (θ)
alue of 0.050 (95% CI 0.0383–0.0620). The 95% crediblenterval did not include 0, which means that a genetic struc-ure between treatments had developed during the coursef the experiment, a time span covering at the most 3.5ig. 2. Abundance of various life stages of A. crassa populations exposed toopper (corresponding to 130 �g/l in water) after 120 days. Error bars indicatehe 95% confidence intervals of means (α = 0.05, n = 6). An asterisk (*) denotessignificant difference from the control (*p < 0.05).
v(osio(oT(0wwthte
et
sediments from (A) Svindersviken and (B) Trosa after 60 and 120 days. Errorte significant differences from the controls (*p < 0.05; **p < 0.01; ***p < 0.001).
enerations of the test organism. The amount of emergentenetic structure within treatments was less pronouncedn the controls (FST = 0.0394) and increased in strength inhe treatments, with the strongest genetic structure foundn the Svindersviken100% treatment (FST = 0.146, Table 2).nterestingly, the genetic differentiation in the naturally con-aminated sediments showed a concentration-related responseattern, with highest FST values in the 100% treatments:ST[control] < FST[Svindersviken50%] < FST[Svindersviken100%nd FST[control], <FST[Trosa50%], <FST[Trosa100%]. The hier-rchical molecular variance analysis (AMOVA) revealedhat the component of genetic variation due to differencesetween treatments was 4% (p < 0.001) of the total geneticariation while replicates within treatments explained 12%p < 0.001) and the remaining 84% (p < 0.001) was at the levelf individuals within replicates. The Hickory analysis furtherhowed that the value of expected heterozygosity (Table 2)n the controls (0.253, 95% CI 0.232–0.278) did not decreasever time (i.e., 120 days) compared to the starting population0.273, 95% CI 0.259–0.289) since the 95% credible intervalf the difference did include zero (95% CI −0.004 to 0.024).he expected heterozygosity of the copper treatment was 0.215
95% CI 0.195–0.239), while the control had 0.253 (95% CI.232–0.278). The 95% credible interval of the differenceas 0.0331–0.0421 and since it does not include 0, thereas a significant difference in heterozygosity between these
wo treatments. Thus copper exposure decreased expectedeterozygosity. There was no significant difference betweenhe control and any of the naturally contaminated sediments in
xpected heterozygosity (Table 2).The levels of heterozygosity were roughly the same whenstimated with AFLP-SURV (Table 3) as the Hickory estima-ions, with the copper treatment showing the lowest value. The
432 J. Gardestrom et al. / Aquatic Toxicology 86 (2008) 426–436
Table 2Expected heterozygosity (Hs), its standard deviation (S.D.) and its 95% credible interval (CI) of the starting population, control and the treatments when analyzedtogether
Starting population Control Svindersviken50% Svindersviken100% Trosa50% Trosa100% Copper
Hs 0.27 0.25 0.26 0.28 0.25 0.27 0.22S.D. 0.008 0.013 0.012 0.014 0.016 0.011 0.01295% CI 0.259–0.289 0.232–0.277 0.241–0.284 0.258–0.307 0.222–0.278 0.259–0.289 0.195–0.238
FST – 0.0394 0.108 0.146 0.1093 0.144 0.070
A cred
fli
3
ocd
4
ecsbsrirpfaaomitdao
TGn
T
CSSTTC
ctadsttgteeg2nfl(2D
sb6oa(ttoo
95% CI – 0.026–0.056 0.085–0.134
lso, the fixation index (FST) within the control and each treatment and its 95%
our AFLP primers produced totally 161 polymorphic loci. Inine with the heterozygosity pattern, band variability was lowestn the copper treatment (Table 3).
.3. Individual growth
There were no significant differences in cephalothorax lengthr RNA content of A. crassa between the different treatmentsompared to the control, at any time of sampling. For furtheretails, see Table 4.
. Discussion
Many tests used in environmental risk assessment focus onasily observable endpoints at the individual level for logisti-al and economical reasons. However, alterations at this leveleldom have a linear response to alterations at higher levels ofiological organization (Forbes and Calow, 1999). In the presenttudy we found that exposure to a single toxicant (i.e., copper)esulted in reduced genetic diversity after several generationsn contaminated sediments, which is in accordance with theesults in the shorter (fewer generations) study on another cope-od species by Gardestrom et al. (2006) and also with resultsrom studies on other species (Ma et al., 2000; Van Straalennd Timmermans, 2002; Ross et al., 2002; Medina et al., 2007nd references therein). However, this is not a general resultf toxicant exposure since the present study also shows thatultiple toxicants may affect genetic diversity differently, i.e.,
ncreasing differentiation with increasing contaminant concen-
ration. Total abundance was also affected by exposure, andid significantly differ from the control in Svindersviken100%t days 60 and 120 and in Trosa100% at day 60. In contrast, nonef the individual growth measurements (i.e., RNA content andable 3enetic variability given by AFLP-SURV expressed as heterozygosity, andumber of variable loci and percent variable loci in respective treatment
reatment Heterozygosity(AFLP SURV)
# variableloci
% variableloci
ontrol 0.225 158 98vindersviken50% 0.237 157 97.5vindersviken100% 0.256 157 97.5rosa50% 0.211 149 92.5rosa100% 0.254 152 94opper 0.189 134 83
btgoGcceaTaomttm
0.118–0.174 0.087–0.133 0.114–0.174 0.051–0.091
ible interval (CI).
ephalothorax length) resulted in significant differences fromhe control (Table 4). It is possible that the exposure times (60nd 120 days) were too long for a continuous stress protein pro-uction. The levels of stress proteins are normally low whentressors, such as pollutants, are absent. However, at stress, pro-ein synthesis is induced, and the RNA production supportinghis induction is often rapid (Locke and Noble, 2002). This sug-ests that the growth-related endpoints normally used in shorterests, may not be suitable for long-term (i.e., multigenerational)xposure experiments with copepods. It should be noted thatven though RNA/DNA ratios are commonly used to assessrowth of, e.g., fish (Buckley et al., 1999; Fonseca and Cabral,007) and krill (Cullen et al., 2003; Shin et al., 2003), this ratio isot used in the present study due to ontogeny- and growth-relateductuations in the DNA content of small crustacean speciesGorokhova and Kyle, 2002; Rosa and Nunes, 2003; Gorokhova,003), such as A. crassa, which violates the assumption thatNA levels are constant.The copepods that were exposed to copper for 120 days
howed reduced genetic variation, and moderate differentiationetween replicates. Since sampling of copper replicates at day0 was excluded, we do not know how the abundance devel-ped over time in this treatment. From an initial number of 30dults per replicate, the abundance increased to a mean of 125.5S.D. 19.1) in the control and 138.8 (S.D. 33.2) in the copperreatment (Fig. 2). The sole selective force in this “one-toxicant”reatment compared to the controls was copper. The reductionf heterozygosity together with the unaltered total abundancebserved in the copper treatment therefore suggests that there haseen a directed and successful selection for copper tolerant geno-ypes. This finding is supported by a number of studies showingenotype differences in copper sensitivity/tolerance in aquaticrganisms (e.g., in mussels, Hvilson, 1983; in fish, Chagnon anduttman, 1989; in polychaeta, Virgilio and Abbiati, 2004). The
opepod individuals that could cope with the stress imposed byopper likely contributed to the subsequent generations moreffectively (passing on the “right” genes to the next generation)s compared to the individuals with less successful genotypes.hese results are in line with those by Gardestrom et al. (2006)nd suggest that population effects may occur rapidly in the facef a single toxicant, although not obviously apparent since it is
anifested at the genetic level and not in abundance. In fact, inhe copper treatment only the females differed in number fromhe controls. Possibly, this treatment yielded a higher quota of
ore tolerant individuals in the reproductive adult life stage.
J. Gardestrom et al. / Aquatic Toxicology 86 (2008) 426–436 433
Table 4Individual growth in A. crassa represented by cephalothorax length (�m; CIII individuals) and RNA content (ng/ind; CIII individuals)
Treatment Cephalothorax length (�m) RNA content (ng/ind)a
Sampling 1 Sampling 2 Sampling 1 Sampling 2
Control 121 (±4) 123 (±2) * 26 (±4)Svindersviken50% 119 (±2) 123 (±4) * 28 (±7)Svindersviken100% 121 (±1) 120 (±1) * 35 (±10)Trosa50% 119 (±4) 118 (±2) * 29 (±19)Trosa100% 120 (±7) 117 (±4) * 21 (±15)
95% confidence intervals of means are shown in brackets. None of the treatment differed significantly from the control (p > 0.05).ely. W
s e notv
ttlccscHsStsnnr6
iscupstgdttp5haatdlb
iats
aSactheiiboLlt2ab
etlidpettsflRhis
peri
a The RNA measurements on the two sampling occasions were made separatince these RNA values were in general somewhat low. Therefore, these data aralues within a normal range.
Compared to the copper treatment, which only differed fromhe control regarding the presence of copper, the naturally con-aminated sediments were from different locations and did mostikely differ in other factors apart from the contaminants theyontained (e.g., structure, particle size and composition). Wean consequently not assume that the genetic differentiationeen between the replicates within the treatments with naturallyontaminated sediments was exclusively caused by toxicants.owever, the history of large contaminated burden in Svinder-
viken (Sundelin and Eriksson, 2001; Eriksson Wiklund andundelin, 2002) and Trosa (Eklund et al., 2006) together with
he large mortality recorded in the 100% naturally contaminatedediments, suggests that effect of the contaminants could be sig-ificant. What further supports the impact of the toxicants in theaturally contaminated sediments is the concentration-relatedesponse revealed in these treatments in both abundance (days0 and 120) and FST (day 120) (Fig. 1A and B; Table 2).
The recorded population crash (i.e., bottleneck) at day 60n Svindersviken100% and Trosa100% was probably a result ofelective mortality against highly sensitive genotypes, whichaused lower mean total abundances. Hence, the effective pop-lation sizes, which is the number of individuals carrying aarticular allele to the next generation (Newman, 2001), wereignificantly smaller in these replicates compared to the con-rols. Smaller populations are generally more susceptible toenetic drift (Street et al., 1998), which tends to geneticallyifferentiate populations, an observation that was also made inhe present study. In fact, in both treatments with naturally con-aminated sediments there was a concentration-related responseattern coupled to within treatment differentiation (FST). The0% treatments (Svindersviken50% and Trosa50%) displayedigher FST compared to controls, and even higher differenti-tion was reached in the 100% treatments (Svindersviken100%nd Trosa100%) (Table 2). This was probably a short-term evolu-ionary effect of a genetic bottleneck causing increased geneticrift. This type of concentration-related response between pol-ution and genetic marker has, to the best of our knowledge, noteen reported previously.
The copepods exposed to these naturally contaminated sed-
ments showed no reduction in genetic diversity. In contrast,concentration-related increase in the degree of differentia-ion was observed. Both Svindersviken100% and Trosa100% hadignificantly lower total abundance compared to the control
otBa
e suspect that the standards used in sampling 1 were contaminated by RNAseincluded. Other standards were used in the second sampling, which resulted in
fter 60 days. The total abundance was continuously low invindersviken100% at day 120. Both in the Svindersviken100%nd Trosa100% treatments there were significantly fewer malesompared to in the control. In general, higher sensitivity of maleso toxicity is not uncommon and it has previously been shown inarpacticoid copepods (e.g., Klosterhaus et al., 2003). Possiblexplanations for this include, e.g., smaller size of males resultingn higher surface: body volume ratio in contact with toxic chem-cals (Sprague, 1995), and faster elimination rate of toxicantsy females during egg production, as the toxic substances passut with the lipids present in the eggs (McManus et al., 1983;otufo, 1998). Further, it is well known that larval and juvenile
ife stages of copepods are usually more sensitive to toxicityhan the adult life stage (e.g., Forget et al., 1998; Medina et al.,002), which indicates that the lowered numbers of copepoditest day 60 (i.e., in Svindersviken100% and Trosa100%) could haveeen caused by earlier naupliar and copepodite mortality.
Populations having experienced a recent reduction in theirffective population size normally exhibit a correlated reduc-ion of their allele numbers and heterozygosities at polymorphicoci. Contrastingly, we did not observe lower genetic diversityn the populations exposed to naturally contaminated sediments,espite the higher genetic drift as compared to the control. It isossible that the exposure time was not long enough to detectffects on total genetic diversity since rare alleles are lost fasterhan heterozygosity (Maruyama and Fuerst, 1985). An alterna-ive explanation is that that these copepods were exposed to aelective regime that consisted of too many different stressorsor any single genotype to have a direct selective advantage, inine with the niche-width variation hypothesis (Van Valen, 1965;oughgarden, 1972). The upheld level of genetic diversity in theeterogeneous environments presented in the naturally contam-nated sediments is probably highly dependant on the degree oftress.
The ultimate goal of environmental risk assessment is torevent chemical substances causing irreversible damages tocosystems (e.g., European Commission, 2006). Unfortunately,elatively little interest has been devoted to understanding actualmpacts that pollution may have on higher levels of biological
rganization, e.g., related to the adaptive capacity of popula-ions and the functioning of ecosystems (Admiraal et al., 2000;elfiore and Anderson, 2001; Breitholtz et al., 2006; Forbes etl., 2006). To reach this goal in environmental risk assessment,
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34 J. Gardestrom et al. / Aquat
t is, in our view, crucial to increase the knowledge about howollution may affect biodiversity, and if/to what extent alteredenetic diversity, as a results of toxicant exposure, may actu-lly be deleterious to populations. We have here studied the firstf these two issues and we urge for more research concerningoth of them. Hence, we believe that further investigations areeeded to find out whether altered genetic diversity, as results ofoxicant exposure, may actually ultimately affect the resiliencef populations and ecosystems. Nonetheless, the findings fromhe present study indicate that population genetic data and the-ries, may very well be considered in future environmental riskssessment, for a proper decision-making with the purpose torotect the environment from harmful events.
. Conclusion
The present study shows that long-term (multigenerational)xposure to contaminated sediments can genetically erode andifferentiate copepod populations. Alterations in genetic diver-ity and structure are of importance in wild populations sincet imply changed adaptive potential for future perturbations.o fully understand how populations may be affected in thenvironment, we emphasise that it is crucial to gain morenowledge about the different plausible outcomes of toxicantxposure. Other scientific studies have for instance addressedhe importance of high genetic diversity in rendering resistanceo disturbance (Nevo et al., 1986; Hilborn et al., 2003; Hughesnd Stachowicz, 2004; Reusch et al., 2005). We might thus facedilemma in the sense that toxicant exposure may decrease
enetic diversity at the same time, as genetic diversity is neededor coping with perturbations.
cknowledgements
We would like to thank FORMAS (KREBS, 216-2005-1329)or grant to Brita Sundelin et al., which partly financed theresent project. We also thank Stockholm marine science centerSMF) for financial support.
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