THE IMPACT OF MUNICIPAL WASTEWATER EFFLUENT ON FIELD-DEPLOYEDFRESHWATER MUSSELS IN THE GRAND RIVER (ONTARIO, CANADA)

PATRICIA L. GILLIS,*y FRANÇOIS GAGNÉ,z RODNEY MCINNIS,y TINA M. HOOEY,y EMILY S. CHOY,yCHANTALE ANDRÉ,z MD EHSANUL HOQUE,x and CHRIS D. METCALFEx

yAquatic Contaminants Research Division, Water Science & Technology Directorate, Environment Canada, Burlington, Ontario, CanadazAquatic Contaminants Research Division, Water Science & Technology Directorate, Environment Canada, Montreal, Quebec, Canada

xDepartment of Biology, Trent University, Peterborough, Ontario, Canada

(Submitted 8 April 2013; Returned for Revision 16 June 2013; Accepted 19 September 2013)

Abstract: To examine effects of municipal wastewater effluent (MWWE) on sentinel organisms, the authors deployed caged freshwatermussels (Lasmigona costata) in the Grand River (ON, Canada) upstream and downstream of anMWWEoutfall. Passive sampling deviceswere deployed alongside caged mussels to confirm exposure. Biomarkers of xenobiotic biotransformation, oxidative stress, estrogenicity,and immunomodulation were investigated. Elevated concentrations of selected pharmaceutical and personal care products (PPCPs) and anatural estrogen (estrone) were found at the downstream sites. Mussels caged downstream of the effluent for 2 wk showed minimalevidence of exposure, while those deployed for 4 wk exhibited significantly higher levels of lipid peroxidation (LPO) and glutathione S-transferase (GST) activity, demonstrating that MWWE-exposed mussels exhibit increased activity in xenobiotic conjugation andoxidative stress. With respect to immune responses, a significant increase in plasma lysozyme activity and hemocyte viability wasobserved in MWWE-exposed mussels. Vitellogenin (vtg)-like protein in male mussels showed a trend toward induction after 4 wk ofdeployment at the first downstream site, but mean levels were not significantly different. Discriminant function analysis indicated thatmussels deployed for 4 wk upstream and downstream of the MWWE discharge could be discriminated on the basis of LPO, GST, plasmalysozyme, and vtg responses. The physiological stress observed in caged mussels indicates that wild mussels chronically exposed toMWWE in this ecosystem would also be negatively impacted. Environ Toxicol Chem 2014;33:134–143. # 2013 SETAC

Keywords: Freshwater mussels Municipal wastewater Biomarkers Semipermeable membrane devices polar organiccontaminants integrated samplers

INTRODUCTION

Contaminants of emerging concern detected in the effluentsfrom Canadian municipal wastewater treatment plants(WWTPs) and in receiving waters impacted by these dischargesinclude pharmaceuticals, personal care products, and estrogens[1–7]. However, monitoring of receiving waters using analyticaltechniques does not provide information on the bioavailabilityand the resulting biological impacts of exposure to thesecontaminants [8]. Monitoring of aquatic organisms can be usedto evaluate whether contaminants in wastewater are havingecological impacts. Organisms of particular utility from amonitoring perspective include bivalves [9], which can be caged(i.e., field-deployed) at varying distances from the source ofcontamination and compared with mussels caged at upstreamand/or reference locations [10,11].

The effects of municipal wastewater effluents (MWWEs)discharged into aquatic environments are especially evident inareas with high population density. The Grand River in southernOntario, Canada, is an example of an urban-impacted watershedthat receives inputs from 30municipalWWTPs that treat sewagefrom the surrounding population of nearly 1 million people.Recently, feminization of wild fish [12] and alterations in fishcommunity structure [13] were observed in regions of the GrandRiver watershed that are impacted byMWWEs. In addition, wildfreshwater mussels living downstream of a large urban area onthe Grand River have reduced condition factor and do not live as

long as mussels living upstream of the cities [14]. Because theGrand River also receives inputs from agriculture, road runoff,and industry, it was not clear whether MWWE exposurecontributed to the impacts observed in wild Grand River mussels[14].

The Grand River watershed has historically supported adiverse population of freshwater mussels, including 25 species[15], 9 of which are Species at Risk [16]. In fact the Grand Riverwatershed is among the most important habitats for freshwatermussels in Canada, with more than 60% of Canadian speciesfound there [15]. However, population surveys have shown thatcompared with historical reports, mussel diversity in thewatershed has declined, citing that poor water quality, andspecifically sewage pollution, contributed to the loss of diversity[15]. Although this conclusion was based on populationassessments, the specific impact of MWWE on freshwatermussels in this watershed had not been investigated.

The goal of the present study was to determine the directimpact of MWWE on freshwater mussels in the Grand River.Therefore, mussels were deployed at locations upstream anddownstream of a municipal WWTP effluent outfall. Exposure toMWWE-associated compounds was characterized by deployingpassive sampling devices alongside mussel cages. At retrieval,mussel tissues were removed to assess physiological indicatorsof exposure by using a series of biomarkers and immunologicalmeasurements. Because MWWE contains a complex mixture ofchemicals, detoxification functions can be induced in variousorgans, and the effects of exposure may be evident acrossmultiple biological systems, including the immune, reproduc-tive, and digestive systems. Immune responses were character-ized at both the cellular and humoral levels through hemocyte

* Address correspondence to patty.gillis@ec.gc.ca.Published online 25 September 2013 in Wiley Online Library

(wileyonlinelibrary.com).DOI: 10.1002/etc.2401

Environmental Toxicology and Chemistry, Vol. 33, No. 1, pp. 134–143, 2014# 2013 SETAC

Printed in the USA

134

viability (cellular) and lysozyme activity (humoral) assessments.Estrogenic effects were investigated by quantifying levels ofvitellogenin-like proteins, and oxidative stress was examined byquantifying cellular membrane damage (lipid peroxidation).Finally, phase 1 and 2 biotransformation activities weredetermined by quantifying the induction of cytochrome P450-3A (P4503A) and glutathione S-transferase (GST) activities,respectively.

MATERIALS AND METHODS

Mussel deployments

Adult Lasmigona costata (fluted-shell mussel; Rafinesque,1820) were collected from a local reference site and held in thelaboratory for 2 wk prior to deployment. During this holdingperiod, mussels were fed a commercial shellfish diet (InstantAlgae Shellfish Diet 1800). Because larger mussels are preferredfor hemolymph sampling for immune assessment, only adultmussels were deployed in the cages.

The Grand River watershed is the largest drainage basin onthe northern shore of Lake Erie (ON, Canada; Figure 1). Thirtymunicipal WWTPs are found in the watershed, including 6upstream of the study area [14]. The closest WWTP upstream ofthe study area (serving a population of 128 000) is located 21 kmupstream [17]. The WWTP examined in the present studyemploys secondary (activated sludge) treatment and serves apopulation of approximately 230 000 [17]. Although theupstream caging site is located upstream of the WWTP beinginvestigated, mussels deployed at that site are potentiallyexposed to a mixture of anthropogenic compounds from theupstream MWWEs and the general urban area (road runoff,industrial inputs).

Mussels were deployed on 28 September 2010 upstream anddownstream of the Kitchener (ON, Canada) municipal WWTP(Figure 1). The first site (UP) was immediately upstream of theWWTP discharge. The 2 downstream sites, referred to as D1and D2, were located approximately 0.5 km and 1.5 kmdownstream of the WWTP outfall, respectively. Mussel cagedesign and construction was modeled after the ASTM guide[18], with some modifications. Cages consisted of a 1-m2

polyvinyl chloride frame with mesh netting (i.e., mussel socks)strung across the frame. Cages were anchored to iron rods, andsubmerged to a depth of 40 cm to 70 cm. Passive samplingdevices were deployed alongside mussel cages at each site, asdescribed below. Deployed equipment was monitored at leastonce a week to confirm submersion and to remove anyaccumulated debris.

Mussels were removed from the cages after 2 wk (20mussels) and 4 wk (15 mussels) of deployment (12 October and25 October, respectively) and transported for 1 h in a coolercontaining dechlorinated tap water (condition of collectionpermit) to the lab in Burlington (ON, Canada) for processing.Total depuration time in clean water was approximately 2 h.External parameters of shell length, height, and width weredetermined with digital calipers as well as total (whole animal)wet weight. Hemolymph was collected from the sinus of theposterior abductor muscle with a 22-gauge needle andimmediately distributed into either 96-well black polystyrenemicroplates (six 100 mL replicates per mussel) for viability andprotein assessment (details below) or centrifuged (1 mL; 200 gfor 5 min at 4 8C) to separate plasma (supernatant) from the cellpellet. Plasma was preserved at –80 8C for lysozyme activityassessment (details below). Mussels were dissected to obtaindigestive gland and gonad tissues for biomarker analysis.

200m1000 ft

N

Ontario

Cambridge

Kitchener

Waterloo

Municipal wastewatertreatment plant

Mussel caging sites

River flow

UP

D1

Cagingstudyarea

Grand River

D2

Municipal wastewatertreatment plant outfall

Figure 1. Map of study area illustrating the locations where mussels and passive samplers were deployed. Direction of river flow is indicated.

Effect of municipal wastewater on caged freshwater mussels Environ Toxicol Chem 33, 2014 135

Gravidity (presence of brooding glochidia) of the mussels wasnoted during dissection and was used as a surrogate for genderbecause this species is not externally sexually dimorphic. Thereliability of this assumption (i.e., gravid¼ female) wasconfirmed by microscopic examination of gonad tissue. Tissuesamples were stored at –80 8C until analysis for biomarkers.

Contaminants

Semipermeable membrane devices (SPMDs) for monitoringnonpolar (water-insoluble) compounds and polar organiccontaminants integrated samplers (POCIS) for monitoring polar(water-soluble) compounds were deployed alongside musselcages and retrieved after 2 wk. The SPMDs were prepared ina clean lab at Trent University (Peterborough, ON, Canada)as described previously [19], and POCIS containing OasisHydrophilic–Lipophilic Balance sorbent were purchased fromEnvironmental Sampling Technology. The SPMDs (n¼ 3) andPOCIS (n¼ 2) were mounted in perforated stainless steel cagesas described previously [20] and anchored underwater to the ironrods that held the mussel cages.

Nonpolar contaminants were extracted from the SPMDsfollowing methods described by O’Toole et al. [19], and polarcontaminants were extracted from POCIS following methodsdescribed by Li et al. [6]. The compounds monitored includedpharmaceuticals (carbamazepine, gemfibrozil, and ibuprofen), 2antibiotics (sulfamethoxazole and trimethoprim), an estrogen(estrone), the antibacterial compound, triclosan, and 2 syntheticmusk compounds, HHCB (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-g-2-benzopyran; Galaxolide) and AHTN(6-Acetyl-1,1,2,4,4,7-hexamethyltetraline; Tonalide). Several ofthese indicator compoundswere selected according to the criteriaprovided in a recent publication on indicator compounds [21],which identified the value of monitoring a small number ofpharmaceutical and personal care products (PPCPs) in waste-water treatment facilities in the United States. The advantage ofusing these indicator compounds is that they are almost alwaysdetected in wastewater and surface waters impacted byWWTPs,and they show variations in removals by microbial, oxidative,and photolytic processes, as well as by partitioning to sedimentsor air, and thus provide an indication of the most importantremoval processes within the river.

Data on the sampling rates (L/d) for the target compounds inPOCIS [6] and in SPMDs [20] were used to estimateconcentrations in river water (i.e., ng/L) integrated over the 2-wk deployment period. While sampling rates in SPMDs andPOCIS will vary with temperature, flow rate, and biofoulingthese data give an estimate of the time-weighted averageconcentrations of PPCPs in the river, independent of the spatialand temporal variations that occur with repeated grab samples.

The concentrations of total and fecal coliform at the studysites were determined by a private company (Exova Accutest,Ottawa, ON, Canada). Coliform levels were quantified in asurface water grab sample taken after 2 wk of deployment, atthe time the passive samplers were retrieved.

Biomarkers of exposure

Lipid peroxidation (LPO), GST, P4503A, and vitellogenin-like proteins (vtg) were selected as indictors of contaminant-induced physiological stress. Lipid peroxidation was examinedin the digestive gland. Glutathione S-transferase levels wereassessed in digestive gland as well as gonad tissue. Vitellogenin-like proteins were measured in gonad tissue. All biomarkermeasurements were normalized against total protein content inthe supernatant or homogenate [22].

Gonad and digestive gland tissues for GST and P450-3Awere dissected out over ice, and homogenized with a Teflonpestle tissue grinder (5 passes) in a 10mM N-2-hydroxyethylpi-perazine-N-2-ethanesulfonic acid (HEPES)-NaOH buffer, pH7.4, containing 125mM NaCl, 0.1mM ethylenediamine tetra-acetic acid (EDTA), and 0.1mM dithiothreitol, at a 1:5 (w/v)ratio at 4 8C. The homogenate was centrifuged (15 000 g, 20min,2 8C). The supernatant (S1) was separated from the pellet andutilized for determination of enzyme activities, specifically,dibenzyl fluorescein dealkylase (DBF) or P450-3A and GSTactivity.

Glutathione S-transferase activity was measured followingthe methodology developed by Boryslawskyj et al. [23]. Briefly,a sample of 50mL of S1 was added to 200mL of 1mMglutathione and 1mM 1-chloro-2,4-dinitrobenzene in 10mMHEPES-NaOH, pH 6.5, containing 125mM NaCl. Absorbance(340 nm) was measured at 0min, 5min, 10min, 20min, and30min to determine the appearance of the glutathione conjugate.Homogenization buffer was used as a sample control. Resultswere expressed as absorbance change per min per mg protein.

Cytochrome P450-3A activity was measured following theDBF substrate methodology [24]. Briefly, 25mL of S1 wasmixed with 25mM DBF and 100mM reduced nicotinamideadenine dinucleotide phosphate (NADPH) in 125mM NaClmedia buffered with 10mM HEPES-NaOH, pH 7.4. Thesamples were kept at 30 8C, and the liberation of fluoresceinwas measured at 0min, 15min, 30min, and 60min. Fluoresceinwas determined by fluorometry using 485-nm (excitation) and532-nm (emission) filters. The amount of fluorescein wasdetermined with an external fluorescein standard calibrationcurve between 1mM and 20mM. Results are expressed asnanomole per minute per milligram total protein.

The relative levels of vtg-like proteins were determinedfollowing the principle of alkali labile phosphate (ALP) assayperformed on high-molecular-weight proteins [10]. Gonadtissues for vtg analysis were homogenized at 4 8C with apolytron in 25mM HEPEs buffer at pH 7.5, containing 125mMNaCl, 1mM dithiothreitol, and 1mM EDTA. The homogenatewas then centrifuged (12 000 g, 30min, 4 8C). The supernatantwas carefully removed, and the poorly soluble high-molecular-weight proteins were precipitated in 35% acetone as described.The levels of ALP in the protein pellet were determined byadding 200mL of 1M NaOH and heating at 60 8C for 30min.Levels of ALP were measured following the molybdenumreagent for phosphate determination [25] with potassiumphosphate as a standard. The level of vtg-like proteins isexpressed as mg of ALP/mg total gonad protein.

Lipid peroxidation was determined by the thiobarbituric acidmethod, which measures the production of malonaldehyde [26].Standard solutions of tetramethoxy-propane, a stabilized form ofmalonaldehyde, were prepared in the presence of the homogeni-zation buffer for calibration. Fluorescence was measured at520 nm excitation and 590 nm emission with a fluorescencemicroplate reader (Bioscan). Because the reagent could reactwith other aldehydes, results were expressed as mg ofthiobarbituric acid reactants (TBARS) per mg of total protein.

Immune function

Hemocyte viability was analyzed in triplicate 100-mLhemolymph samples following the method of Blaise et al.[27]. Briefly, viability was quantified by determining the amountof hemocyte-retained fluorescamine, normalized with totalprotein content. Hemocyte protein concentration was determinedin lyzed cells with fluorescamine (Sigma). Standards of

136 Environ Toxicol Chem 33, 2014 P.L. Gillis et al.

fluorescein (hemocyte viability) and bovine serum albumin(protein) were used for calibration. Fluorescence was measuredat 485 nm excitation with 520 nm emission for the viabilityassay, and 400 nm excitation with 450 nm emission for theprotein assay, using a microplate reader (BioTek). Cell viabilityis expressed as micromoles of cell fluorescein per milligram ofcell proteins.

Lysozyme activity was measured in plasma according to themethod described by Lee and Yang [28]. Briefly, 100mL ofhemolymph were added to 100mL of a suspension ofMicrococcus lysodeikticus prepared at 0.4mg/mL in 0.1Mphosphate buffer, pH 6.2. The reaction was carried out and readwith a microplate reader (BioTek) at 450 nm and 25 8C. Theactivity of lysozyme was calculated from the slope of the timecourse by linear regression of data points. A unit of lysozymeactivity is defined as the quantity of enzyme that causes adecrease in absorbance of 0.001 per minute at pH 6.2 at 25 8C.Lysozyme activity was normalized to protein concentration inplasma, following the protein–dye binding method [22] withbovine serum albumin (Sigma) for calibration.

Statistical analysis

Results of biomarkers and immune function are presented asmeans with standard deviation (SD) as mean� SD. Data werecompared across deployment sites (UP, D1, D2) with analysis ofvariance (ANOVA) followed by Tukey’s test (p< 0.05; SigmaStat 3.5). For data that did not meet the normality and/or equalvariance assumptions, differences were identified with Kruskal-Wallis ANOVA on ranks (p< 0.05) followed by Dunn’s test(Sigma Stat 3.5). Results from contaminant analysis fromSPMDs (n¼ 3) are presented as means� standard deviation.The biomarker data were also analyzed by discriminant functionanalysis (Statistica 8.0) to seek out differences among the studysites and determine which biomarkers contributed the most forsite classification.

RESULTS

Mussel deployments

No equipment vandalism or mussel mortality was experi-enced at any of the deployment sites. A heavy rainstorm on theday of the deployment resulted in a large increase in river flow(from 13 m3/s to 33 m3/s) and an increase in water level of over25 cm. The hydrological flow profile for the Grand River overthe 4-wk deployment period is illustrated in Figure 2.

Dissection of deployed mussels on retrieval revealed that45% (UP), 65% (D1), and 40% (D2) of the mussels deployed for2 wk were female, and that 60% (UP), 67% (D1), and 47% (D2)of the mussels deployed for 4 wk were female.

Contaminants

The POCIS and SPMD passive samplers accumulated alltarget analytes (PCPPs) in amounts that were detectable byliquid chromatography–tandem mass spectrometry analysis,except for estrone, which was not detected in POCIS deployed atthe upstream site. Table 1 shows the estimated concentrations inwater (ng/L) calculated from the amounts accumulated in thepassive samplers over the 2-wk deployment period using thesampling rates (L/d) for the target analytes determined inprevious laboratory experiments with POCIS [6] and SPMDs[20]. The sampling rates selected for these estimates weredetermined in laboratory experiments at 15 8C. Overall, theconcentrations of PCPPs were higher at both downstream sitescompared with levels upstream of the WWTP.

The pattern of contamination by coliform bacteria at thedeployment sites also confirmed that the downstream cages weredeployed in the MWWE plume. Fecal coliform concentrationswere 76 coliform units/100mL at the upstream site, 680 coliformunits/100mL at the D1 downstream site, and 468 coliform units/100mL at the D2 downstream site (Table 2).

Biomarkers of exposure

Mussels exposed to theMWWE for 4 wk at both downstreamsites exhibited significantly more cellular membrane damage(lipid peroxidation) than mussels caged upstream of the outfall,although levels were similar across all sites after 2 wk ofexposure (Figure 3A). Similarly, analysis of 4-wk deployedmussels from both the D1 and D2 exposure sites revealedsignificantly elevated levels of GST activity in the digestivegland compared with upstream mussels, while no difference wasobserved in GST activity in 2-wk deployed mussels, regardlessof exposure site (Figure 3B). Although GST was also measuredin gonad tissue, levels did not differ across study sites (data notshown). Although MWWE-exposed mussels tended to haveslightly lower levels, there was no significant difference in P450-3A activity among deployment sites after either 2 wk or 4 wk ofexposure (Figure 3C).

No significant differences in mean levels of vtg-like proteinsacross the study sites were observed whether comparisons weremade with data from male and female mussels combined orseparated (Figure 4). However, the mean levels (expressed asalkali-labile phosphate mg/mg protein) in male mussels (Figure4B) deployed at the near-field site (D1) for 4 wk were slightly(p¼ 0.1) higher (28.5� 14.8) than those caged upstream(19.8� 3.8) or further downstream (D2) of the effluent(22.9� 5.7). As with the other biomarkers examined, therewas no significant response in vtg levels in the mussels deployedfor 2 wk.

Immune function

Mussels exposed to MWWE had higher levels of plasmalysozyme activity than those deployed upstream of the treatmentplant (Figure 5A). Lysozyme activity (expressed as D slope/mgprotein) in mussels from both downstream sites (D1,

Figure 2. Mean daily river flow rate (m3/s) in the study area as recorded bythe Grand River Conservation Authority (Gauge 2GAC05; 43.4001,–80.4268). The exposure period for the 2-wk and 4-wk deployments areindicated. Figure produced using information under License with the GrandRiver Conservation Authority. Copyright© Grand River ConservationAuthority, 2012.

Effect of municipal wastewater on caged freshwater mussels Environ Toxicol Chem 33, 2014 137

14.1� 10.6; D2, 12.1� 7.5) was significantly higher after 4 wkof exposure than the upstream-deployed mussels (UP,5.4� 4.6). The level of lysozyme activity at the nearestdownstream site (D1) was also elevated after 2 wk of exposure,but the difference was not significant.

MWWE exposure also had a significant effect on theviability (expressed cell fluorescein/mg of cell proteins) ofmussel hemocytes (Figure 5B). Mussels deployed at bothdownstream sites (D1, 3.7� 10�3� 1.1� 10�3; D2,2.4� 10�3� 8.0� 10�4) for 2 wk had significantly higherhemocyte viability compared with upstream mussels(1.4� 10�3� 3.5� 10�4), while those deployed in the effluentplume for 4 wk only showed a significant increase in viability atthe second downstream site (D2).

Relationships among biomarker responses

In an attempt to gain a global understanding of the observedresponses after 4 wk of exposure, we performed a discriminantfunction analysis. This analysis can be used to determine howwell the treatments, in this case caging sites, are separated fromeach other and also which biomarkers among the batteryemployed contribute the most to site classification or treatmentidentification (principal component analysis). The analysisrevealed that all sites (UP, D1, and D2) were discriminated bymore than 70% and hence, the sites could be considered differentfrom each other (Figure 6). The upstream site was readilydiscriminated from the 2 downstream sites, based on theobserved responses in LPO, GST, and lysozyme activity(component 1). Male vtg-like proteins were equally weightedwith lysozyme activity. Site D1 was somewhat discriminatedfrom site D2 based on hemocyte viability, vtg-like proteins infemales, and P450-3A activity (component 2).

DISCUSSION

Contaminants

A large number of PPCPs and endocrine-disruptingsubstances have been detected in both wastewater and surface

waters and could be considered for this research. However, tominimize analytical effort and cost, while monitoring a series ofcompounds that will provide valuable information on the variousmechanisms of removal, we analyzed a small group of indicatorcompounds that are typically detected in wastewater and forwhich the processes governing their fate in the aquaticenvironment are reasonably well understood. Nine indicatorcompounds were analyzed, including several PPCPs and anatural estrogen. The objective of the passive sampling approachwas to demonstrate that there are differences in the concen-trations of contaminants of wastewater origin at upstream anddownstream locations in the Grand River and not necessarily tolink biomarker responses to specific compounds.

The POCIS and SPMD passive samplers accumulated all ofthe target analytes. The data show that the target analytes werepresent at all deployment sites including the upstream location,probably as a result of effluents from the 6 municipal WWTPsthat discharge into the Grand River upstream of the study area.However, the concentrations of all analytes increased at the firstdeployment site (D1) immediately downstream of the KitchenerWWTP discharge. While all target compounds were still presentat the second downstream deployment site (D2), the concen-trations decreased with the increased distance downstream of thedischarge point (Table 1). The concentrations of the indicatorcompounds estimated from the passive sampler data areconsistent with concentrations of these compounds in surfacewaters and in MWWE at locations in Ontario, Canada [1,3,6].

Overall, the chemistry data indicate that the freshwatermussels deployed downstream of the Kitchener WWTP wereexposed to PPCPs and to estrone at ng/L concentrations over atleast the initial 2-wk deployment period and there is no reason tobelieve that exposures were not approximately the same orhigher over the remainder of the 4-wk deployment.

Biomarkers of exposure

Significant responses in GST activity and cellular membranedamage were observed in freshwater mussels deployeddownstream of a municipal wastewater treatment plant for 4wk, but were less evident in mussels exposed to the effluent for 2wk. While some biomarkers may indeed require a longer time torespond, indicators of xenobiotic transformation such as GSTand induction of vtg-like proteins have been induced after 2 wkin comparable caging studies with other species of freshwatermussels [29,30]. In the present study, although the magnitude ofthe response (4 wk vs 2 wk) appears to be the result of thedifference in exposure duration, this was not the onlycontributing factor. We believe that the reduced stress responseobserved in the 2-wk deployed mussels was the result at least inpart of higher river flow during the first 2 wk of deployment,which resulted in additional dilution of the MWWE plume.Indeed, there was a major rainstorm the day the mussels and

Table 1. Estimated mean (with standard deviation) concentrations of selected pharmaceutical and personal care products at deployment sites, upstream (UP) anddownstream (D1, D2) of a municipal wastewater treatment plant for 2 wka

Site

Estimated mean concentration in water (ng/L)

CBZ IBU TPM SMX GMZ EST TCS HHCB AHTN

UP 2.47 (0.41) 2.30 (0.54) 1.89 (0.83) 0.23 (0.09) 0.05 (0.01) ND 0.6 18.3 14.5D1 6.44 (0.33) 11.80 (1.13) 6.42 (1.14) 1.30 (0.43) 0.12 (0.02) 0.21 (0.02) 3.8 91.4 48.9D2 6.71 (0.27) 12.97 (0.87) 5.75 (1.24) 0.77 (0.13) 0.10 (0.03) 0.07 (0.04) 3.4 73.7 22.5

aTCS, HHCB, and AHTN were sampled using semipermeable membrane devices (n¼ 2). All other compounds were sampled using polar organic contaminantsintegrated samplers (n¼ 3). Mean water temperature during passive sampler deployment was 14.0 8C (range, 12.0 8C–15.8 8C).CBZ¼ carbamazepine; IBU¼ ibuprofen; TPM¼ trimethoprim; SMX¼ sulfamethoxazole; GMZ¼ gemfibrozil; EST¼ estrone, TCS¼ triclosan; HHCB¼Galaxolide® (synthetic musk); AHTN¼Tonalide® (synthetic musk); ND¼ not detected.

Table 2. Concentrations of fecal coliform and total coliform at musseldeployment sitesa

Field siteTotal coliforms(CFU/100 mL)

Fecal coliforms(CFU/100 mL)

UP 1500 76D1 7000 680D2 3400 468

aOf the 3 sample sites, one was upstream (UP) of the municipal wastewatertreatment plant, and two were downstream (D1 [0.5 km] and D2 [1.5 km]) ofthe municipal wastewater treatment plant. Coliform levels (coliform formingunits [CFU]/100 mL) were measured in surface water grab samples.

138 Environ Toxicol Chem 33, 2014 P.L. Gillis et al.

passive samplers were deployed, which was followed by asubstantial increase in river flow for the following 14 d (Figure2). Provisional flow data from the Grand River ConservationAuthority (GRCA; Cambridge, ON, Canada) demonstrate thatthe flow rate peaked 2 d after deployment and took 14 d to returnto typical fall levels (GRCA, unpublished data). During the first2 wk of deployment, on average, 3.2% of the river flow (range,2.3–5.0%) at the WWTP outfall was comprised of MWWE(GRCA, unpublished data; [31]), whereas MWWE accountedfor 4.8% (3.8–5.8%) of the river flow at the WWTP outfallduring the latter 2 wk of the deployment (GRCA, unpublisheddata; [31]). The additional dilution of MWWE and its associatedcontaminants during the first 2 wk of deployment likelytranslated into the marginal responses observed after 2 wk of

exposure. Unfortunately, this exposure dilution theory cannot beconfirmed by chemical data from the passive samplers becausethey were only deployed for the first 2 wk of the study.

Biomarkers. Mussels located at the deployment sites down-stream of the MWWE had increased GST activity and elevatedLPO, which indicates the presence of bioavailable xenobiotics inthe plume. Elevation of GST was also reported in a previousstudy with Elliptio complanata caged downstream of amunicipal WWTP [30] and in another laboratory study inwhich freshwater mussels injected with the arthritis medicationmetotrexate and a C8 solid-phase methanol extract from aprimary-treated MWWE displayed an increase in GST activityafter 48 h [32]. Both of these studies suggest that mussels may beresponsive to PPCPs in municipal wastewater effluents. Inaddition, in the marine mussel Mytilus galloprovincialis,

Figure 3. Biomarkers of xenobiotic exposure and oxidative stress. (A) Lipidperoxidation (LPO), (B) glutathione-S transferase (GST), and (C)cytochrome P450-3A (CytoP450) activity in the digestive gland tissue offreshwater mussels (Lasmigona costata) deployed upstream (UP) anddownstream (D1 [0.5 km], D2 [1.5 km]) of a municipal wastewater treatmentplant for either 2 wk or 4 wk. Error bars indicate standard deviation (n¼ 12).Bars identified with different letters of the same case are significantlydifferent (p< 0.05). TBARS¼ thiobarbituric acid reactants; ABS¼ absor-bance; FL¼fluorescence.

Figure 4. Vitellogenin-like proteins (referred to as alkali labile phosphate[ALP]) in gonad tissue of freshwater mussels (Lasmigona costata) deployedupstream (UP) and downstream (D1 [0.5 km], D2 [1.5 km]) of a municipalwastewater treatment plant for either 2 wk or 4 wk. (A) Mean ALP for allmussels; (B) only female mussels; and (C) only male mussels. Error barsindicate standard deviation. There were no significant differences (p> 0.05)between groups.

Effect of municipal wastewater on caged freshwater mussels Environ Toxicol Chem 33, 2014 139

exposure to 11mg/L propranolol, an antihypertensive drug, ledto increased GST activity in gills, and acetaminophen exposureincreased LPO in digestive gland [33]. While increased GSTactivity and LPO in MWWE-exposed mussels in the presentstudy are consistent with the induction of these biomarkers bysome PPCPs, further study is necessary to establish a causal linkbecause of the complex chemical composition of MWWE.

Pharmaceuticals induce CYP3A4 in mammals [34] andCYP3 homologues in fish [35], thus, it was anticipated that thesedetoxification enzymesmight be induced inmussels as a result ofexposure to pharmaceuticals in the wastewater. However, therewas no significant induction of P450-3A in the mussels cageddownstream of the WWTP.

Vitellogenin-like proteins. Although there were no statisticallysignificant differences between MWWE-exposed and unex-posed mussels, there was a trend toward higher levels of vtg-likeproteins in the downstream male mussels deployed for 4 wk.While separation of male and female mussels was necessary toreveal sex-specific trends, separating the sexes reduced thesample size and thus the power of the statistical analysis for thissomewhat variable endpoint. The lack of vtg induction was notexpected because the estrogenic effects of MWWE have beenwell documented, in mussels (other locations), and in fish fromthe Grand River system. Specifically, Tetreault et al. [12]reported 100% feminization of wild male darters (Etheostomasp.) in this same MWWE-impacted area of the Grand River, andthere have been reports of MWWE-induced changes in vtg-likeproteins in other species of freshwater mussels [10,11,36,37].Indeed Gagné et al. [11] reported that wild male Elliptiocomplanata collected downstream of the MWWE from 2Quebec cities had significantly increased vtg-like proteins. In thepresent study, because the slight elevation in male vtg levels wasonly evident after 4 wk, and not 2 wk of exposure, one couldspeculate that the exposure duration or strength (see river

Figure 5. (A) Lysozyme activity and (B) hemocyte viability in freshwatermussels (Lasmigona costata) deployed upstream (UP) and downstream (D1[0.5 km], D2 [1.5 km]) of a municipal wastewater treatment plant for either2wk or 4wk. Error bars indicate standard deviation (2wkn¼ 20; 4wkn¼15).Bars identified with different letters of the same case are significantly different(p< 0.05). FL¼fluorescence.

Figure 6. Discriminant function analysis of the biomarker data. Mussel deployment sites are referred to as UP (upstream of effluent outfall), D1 (0.5 kmdownstream), and D2 (1.5 km downstream) of effluent outfall. Biomarkers are referred to as HVia (hemocyte viability), vtgfemale (vitellogenin-like proteins infemales), P4503A (cytochrome P450-3A), LPO (lipid peroxidation), GST (glutathione-S transferase), and vtgmale (vitellogenin-like proteins in males). Percentage(%) refers to the correctness of treatment classification (e.g., at 100% all samples are correctly ascribed to their site). The grouping circles are a delineation ofsamples from the 3 sites (not statistically derived).

140 Environ Toxicol Chem 33, 2014 P.L. Gillis et al.

dilution discussion above) was not adequate to stimulatesignificant production of vtg. However, it is also important tonote the timing of the present study with respect to thereproductive cycle of the mussels employed. Many of themussels deployed in the cages were gravid, meaning they werebrooding glochidia (larvae) in their gills and not in an activestage of gametogenesis. Long-term brooders (i.e., bradytictic)such as L. costata hold their glochidia over the winter beforereleasing them the following spring and initiating the nextreproductive cycle [38]. Although there are no published reportsof endocrine disruption (reproductive cycle-influenced orotherwise) in L. costata, marine mussels (Mytilus) have beenreported to be more sensitive to estrogens during the early stagesof gonad development, but were impervious to estrogens atmaturity [39]. Perhaps if mussels had been deployed during anactive stage in their reproductive cycle, rather than early in thefall during reproductive quiescence, stronger induction of vtgwould have been observed. In addition, a larger sample size,especially of male mussels, would have been required to confirmwhether MWWE exposure induces changes in vtg-like proteinsin mussels in this ecosystem.

Immune function

The significant increase in lysozyme activity in the musselsdeployed in the effluent plume can be attributed to the 2-fold(D2) to 4-fold (D1) increase in waterborne bacteria levels at thedownstream sites compared with the upstream site (Table 2).Lysozymes are a key component of the bivalve’s innate immuneresponse against bacteria [40] that are secreted by hemocytesafter pathogen recognition or physiological stress [41]. This isconsistent with the observation that lysozyme activity followedthe level of total coliforms in the water. In addition, hemocyteviability generally (with 1 exception, D1 at 4-wk) followed thecoliform concentration at the deployment sites. Typically,pathogenic bacteria or sustained inflammation reduce hemocyteviability, although in some cases bacteria can increase lipaseactivity with lysozyme activity [42,43], which can result in anincrease in hemocyte viability, because the cell viability assayemployed quantifies a nonspecific esterase substrate that couldbe influenced by lipases (lipase is an esterase).

Although mussel energy stores were not assessed, we suspectthat the chronic stimulation of the immune system from elevatedlevels of waterborne bacteria could have associated energy costsfor wild MWWE-exposed mussels. In addition, the ongoingexposure to MWWE and its associated pathogens couldeventually lead to immunosuppression and render the musselsmore susceptible to diseases [44]. Such chronic effects ofMWWE exposure may be reflected in the finding that musselsliving downstream of urban effluents in the Grand River havelower condition factors and do not live as long as those from theupstream reference area [14].

Relationships among biomarker responses

Based on discriminant function analysis, the biomarkerresponses inmussels caged at the upstream and downstream sitescould be discriminated based on xenobiotic conjugation (GST),oxidative stress (LPO), lysozyme activity, and levels of vtg-likeproteins in males. These biomarkers are well known to respondin organisms exposed to municipal wastewater effluent. Forexample, lysozymes are produced as a defense mechanismagainst bacteria, which corroborates the reported levels ofbacteria in the surface waters of the study sites. In mussels cageddownstream of municipal effluents in the Mille-Isle River (QC,Canada), lysozyme activity was also induced in the hemolymph

and followed total heterotrophic bacteria [10]. Increasedoxidative stress (LPO) and xenobiotic conjugating activity(GST) were also reported in bivalves exposed to municipaleffluents [11,45]. Changes in vtg-like proteins in male musselshave also been reported in wild mussel populations collected at 2sites downstream of the discharge of a municipal wastewatereffluent in Quebec (Canada) [11]. The discriminant functionanalysis is a useful approach for evaluating whether responsesacross several biomarkers in mussels are different at sitesupstream and downstream of MWWE discharges. However,more work is required to evaluate whether these responses arethe result of exposure to contaminants of emerging concern orare more generalized responses to the complex wastewatermatrix.

Implications for resident freshwater mussel populations

Although the Grand River has historically contained adiverse population of freshwater mussels, 30% to 50% of thespecies once found there are believed to have been lost, andsewage pollution may have contributed to the decline [15]. Therange of waterborne contaminants measured in the presentstudy (PCPPs) and previous studies (agricultural chemicals [46]and PCPPs [7]) and in long-term water quality monitoring(metals and nutrients [17]) demonstrates that water qualitydownstream of municipal WWTPs in the Grand River isimpaired. In addition to the chemical data, there is a growingbody of evidence that the complex mixture of anthropogeniccompounds in this river negatively impact resident biota [12–14]. By isolating the exposure of a single municipal WWTP, wehave demonstrated that the compounds associated withMWWE induce physiological stress and likely contributed tothe negative effects observed in wild mussels downstream ofmultiple WWTP [14]. While the mussels examined in thepresent study displayed effects after a 4-wk exposure toMWWE, because of their relatively sessile nature, wild musselscould be exposed to MWWE throughout their lifespan. Suchchronic exposure to MWWE has been shown to negativelyimpact freshwater mussel populations. Goudreau et al. [47]investigated the impact of MWWE on wild mussels in Virginia(USA) and found a strong correlation between mussel lossesand exposure to wastewater effluents. These authors conductedextensive surveys upstream and downstream of WWTPs andfound that the areas downstream of the outfalls were nearlydevoid of unionids, in 1 case declining from 6 mussels/m2

upstream of the WWTP to less than 1 mussel/m2 below theoutfall [47]. Findings in Goudreau et al. [47] suggest either thatat some point over their lifecycle, mussels exposed to MWWEexperience acutely toxic conditions, perhaps as a result of theelevated levels of chloride [48] and ammonia [49] that areassociated with MWWE [17], or that the cumulative damage ofchronic MWWE exposure translates into negative impacts atthe whole-organism level [14] and eventually the populationlevel. In the present study, although no mortality of the adultmussels deployed in the MWWE plume was observed after a 4-wk exposure, it is unknown whether the sensitive early lifestages of glochidia (larvae) and juvenile mussels would havesurvived the deployment. The reduced condition factor of adultL. costata living downstream of 11 municipal WWTPs suggeststhat long-term exposure to MWWE leads to organism-levelimpacts [14]; however, to our knowledge neither the short-termnor the long-term effects of MWWE on the more sensitive lifestages of freshwater mussels are understood.

Based on the fact that populations of L. costata are found inmost areas of the Grand River [15], including those downstream

Effect of municipal wastewater on caged freshwater mussels Environ Toxicol Chem 33, 2014 141

of multiple municipal WWTPs [14], this species may berelatively pollution tolerant. Therefore, we speculate that if amore sensitive species had been employed in the presentMWWE caging study, the impacts of the exposure would in facthave been heightened. Of course, it is neither reasonable norpossible to conduct similar studies with rarer (i.e., species at risk)and potentially more sensitive species.

CONCLUSIONS

Freshwater mussels deployed in the effluent plume of amunicipal WWTP exhibited evidence of oxidative stress andstimulated immune response, as well as possible estrogeniceffects in male mussels. Passive sampling devices demonstratedthat the mussels were exposed to a wide range of PPCPs. Theestrogenic effects of the MWWE on mussels were not as strikingas those recently reported in wild fish in this watershed;therefore, further investigation is needed to determine whethermussel feminization is occurring. Both the MWWE-inducedphysiological stress reported in the present study and previouslyreported whole-organism impacts corroborate long-standingassumptions that anthropogenic-derived contaminants contrib-uted to the significant decline in freshwater mussel density anddiversity in this ecosystem. The municipal WWTP examined inthe present study (Kitchener) is currently undergoing significantfacility upgrades, which will hopefully translate into improvedwater quality and in turn contribute to the recovery of thisimperiled group of organisms.

Acknowledgment—The authors thank K. Oakes, M. Servos, and M. Jorge fortheir contributions to the project, as well as the Grand River ConservationAuthority (S. Shifflett and M. Anderson) for providing river flow andtemperature data. This project was funded by Environment Canada (to P.L.Gillis) and the Canadian Water Network (to C.D. Metcalfe and F. Gagné).

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