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Keywords:

  • Freshwater mussels;
  • Municipal wastewater;
  • Biomarkers;
  • Semipermeable membrane devices;
  • polar organic contaminants integrated samplers

Abstract

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

To examine effects of municipal wastewater effluent (MWWE) on sentinel organisms, the authors deployed caged freshwater mussels (Lasmigona costata) in the Grand River (ON, Canada) upstream and downstream of an MWWE outfall. Passive sampling devices were 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 a natural estrogen (estrone) were found at the downstream sites. Mussels caged downstream of the effluent for 2 wk showed minimal evidence 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 and oxidative stress. With respect to immune responses, a significant increase in plasma lysozyme activity and hemocyte viability was observed in MWWE-exposed mussels. Vitellogenin (vtg)-like protein in male mussels showed a trend toward induction after 4 wk of deployment at the first downstream site, but mean levels were not significantly different. Discriminant function analysis indicated that mussels deployed for 4 wk upstream and downstream of the MWWE discharge could be discriminated on the basis of LPO, GST, plasma lysozyme, and vtg responses. The physiological stress observed in caged mussels indicates that wild mussels chronically exposed to MWWE in this ecosystem would also be negatively impacted. Environ Toxicol Chem 2014;33:134–143. © 2013 SETAC


INTRODUCTION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

Contaminants of emerging concern detected in the effluents from Canadian municipal wastewater treatment plants (WWTPs) and in receiving waters impacted by these discharges include pharmaceuticals, personal care products, and estrogens [1-7]. However, monitoring of receiving waters using analytical techniques does not provide information on the bioavailability and the resulting biological impacts of exposure to these contaminants [8]. Monitoring of aquatic organisms can be used to evaluate whether contaminants in wastewater are having ecological impacts. Organisms of particular utility from a monitoring perspective include bivalves [9], which can be caged (i.e., field-deployed) at varying distances from the source of contamination and compared with mussels caged at upstream and/or reference locations [10, 11].

The effects of municipal wastewater effluents (MWWEs) discharged into aquatic environments are especially evident in areas with high population density. The Grand River in southern Ontario, Canada, is an example of an urban-impacted watershed that receives inputs from 30 municipal WWTPs that treat sewage from the surrounding population of nearly 1 million people. Recently, feminization of wild fish [12] and alterations in fish community structure [13] were observed in regions of the Grand River watershed that are impacted by MWWEs. In addition, wild freshwater mussels living downstream of a large urban area on the Grand River have reduced condition factor and do not live as long as mussels living upstream of the cities [14]. Because the Grand River also receives inputs from agriculture, road runoff, and industry, it was not clear whether MWWE exposure contributed to the impacts observed in wild Grand River mussels [14].

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

The goal of the present study was to determine the direct impact of MWWE on freshwater mussels in the Grand River. Therefore, mussels were deployed at locations upstream and downstream of a municipal WWTP effluent outfall. Exposure to MWWE-associated compounds was characterized by deploying passive sampling devices alongside mussel cages. At retrieval, mussel tissues were removed to assess physiological indicators of exposure by using a series of biomarkers and immunological measurements. Because MWWE contains a complex mixture of chemicals, detoxification functions can be induced in various organs, and the effects of exposure may be evident across multiple biological systems, including the immune, reproductive, and digestive systems. Immune responses were characterized at both the cellular and humoral levels through hemocyte viability (cellular) and lysozyme activity (humoral) assessments. Estrogenic effects were investigated by quantifying levels of vitellogenin-like proteins, and oxidative stress was examined by quantifying cellular membrane damage (lipid peroxidation). Finally, phase 1 and 2 biotransformation activities were determined by quantifying the induction of cytochrome P450-3A (P4503A) and glutathione S-transferase (GST) activities, respectively.

MATERIALS AND METHODS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

Mussel deployments

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

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

image

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

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Mussels were deployed on 28 September 2010 upstream and downstream of the Kitchener (ON, Canada) municipal WWTP (Figure 1). The first site (UP) was immediately upstream of the WWTP discharge. The 2 downstream sites, referred to as D1 and D2, were located approximately 0.5 km and 1.5 km downstream of the WWTP outfall, respectively. Mussel cage design 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, and submerged to a depth of 40 cm to 70 cm. Passive sampling devices were deployed alongside mussel cages at each site, as described below. Deployed equipment was monitored at least once a week to confirm submersion and to remove any accumulated debris.

Mussels were removed from the cages after 2 wk (20 mussels) and 4 wk (15 mussels) of deployment (12 October and 25 October, respectively) and transported for 1 h in a cooler containing dechlorinated tap water (condition of collection permit) 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 were determined with digital calipers as well as total (whole animal) wet weight. Hemolymph was collected from the sinus of the posterior abductor muscle with a 22-gauge needle and immediately distributed into either 96-well black polystyrene microplates (six 100 µL replicates per mussel) for viability and protein assessment (details below) or centrifuged (1 mL; 200 g for 5 min at 4 °C) to separate plasma (supernatant) from the cell pellet. Plasma was preserved at –80 °C for lysozyme activity assessment (details below). Mussels were dissected to obtain digestive gland and gonad tissues for biomarker analysis. Gravidity (presence of brooding glochidia) of the mussels was noted during dissection and was used as a surrogate for gender because this species is not externally sexually dimorphic. The reliability of this assumption (i.e., gravid = female) was confirmed by microscopic examination of gonad tissue. Tissue samples were stored at –80 °C until analysis for biomarkers.

Contaminants

Semipermeable membrane devices (SPMDs) for monitoring nonpolar (water-insoluble) compounds and polar organic contaminants integrated samplers (POCIS) for monitoring polar (water-soluble) compounds were deployed alongside mussel cages and retrieved after 2 wk. The SPMDs were prepared in a clean lab at Trent University (Peterborough, ON, Canada) as described previously [19], and POCIS containing Oasis Hydrophilic–Lipophilic Balance sorbent were purchased from Environmental Sampling Technology. The SPMDs (n = 3) and POCIS (n = 2) were mounted in perforated stainless steel cages as described previously [20] and anchored underwater to the iron rods that held the mussel cages.

Nonpolar contaminants were extracted from the SPMDs following methods described by O'Toole et al. [19], and polar contaminants were extracted from POCIS following methods described by Li et al. [6]. The compounds monitored included pharmaceuticals (carbamazepine, gemfibrozil, and ibuprofen), 2 antibiotics (sulfamethoxazole and trimethoprim), an estrogen (estrone), the antibacterial compound, triclosan, and 2 synthetic musk compounds, HHCB (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-benzopyran; Galaxolide) and AHTN (6-Acetyl-1,1,2,4,4,7-hexamethyltetraline; Tonalide). Several of these indicator compounds were selected according to the criteria provided in a recent publication on indicator compounds [21], which identified the value of monitoring a small number of pharmaceutical and personal care products (PPCPs) in wastewater treatment facilities in the United States. The advantage of using these indicator compounds is that they are almost always detected in wastewater and surface waters impacted by WWTPs, and they show variations in removals by microbial, oxidative, and photolytic processes, as well as by partitioning to sediments or air, and thus provide an indication of the most important removal processes within the river.

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

The concentrations of total and fecal coliform at the study sites were determined by a private company (Exova Accutest, Ottawa, ON, Canada). Coliform levels were quantified in a surface water grab sample taken after 2 wk of deployment, at the 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 examined in the digestive gland. Glutathione S-transferase levels were assessed in digestive gland as well as gonad tissue. Vitellogenin-like proteins were measured in gonad tissue. All biomarker measurements were normalized against total protein content in the supernatant or homogenate [22].

Gonad and digestive gland tissues for GST and P450-3A were dissected out over ice, and homogenized with a Teflon pestle tissue grinder (5 passes) in a 10 mM N-2-hydroxyethylpiperazine-N-2-ethanesulfonic acid (HEPES)-NaOH buffer, pH 7.4, containing 125 mM NaCl, 0.1 mM ethylenediamine tetraacetic acid (EDTA), and 0.1 mM dithiothreitol, at a 1:5 (w/v) ratio at 4 °C. The homogenate was centrifuged (15 000 g, 20 min, 2 °C). The supernatant (S1) was separated from the pellet and utilized for determination of enzyme activities, specifically, dibenzyl fluorescein dealkylase (DBF) or P450-3A and GST activity.

Glutathione S-transferase activity was measured following the methodology developed by Boryslawskyj et al. [23]. Briefly, a sample of 50 µL of S1 was added to 200 µL of 1 mM glutathione and 1 mM 1-chloro-2,4-dinitrobenzene in 10 mM HEPES-NaOH, pH 6.5, containing 125 mM NaCl. Absorbance (340 nm) was measured at 0 min, 5 min, 10 min, 20 min, and 30 min to determine the appearance of the glutathione conjugate. Homogenization buffer was used as a sample control. Results were expressed as absorbance change per min per mg protein.

Cytochrome P450-3A activity was measured following the DBF substrate methodology [24]. Briefly, 25 µL of S1 was mixed with 25 µM DBF and 100 µM reduced nicotinamide adenine dinucleotide phosphate (NADPH) in 125 mM NaCl media buffered with 10 mM HEPES-NaOH, pH 7.4. The samples were kept at 30 °C, and the liberation of fluorescein was measured at 0 min, 15 min, 30 min, and 60 min. Fluorescein was determined by fluorometry using 485-nm (excitation) and 532-nm (emission) filters. The amount of fluorescein was determined with an external fluorescein standard calibration curve between 1 µM and 20 µM. Results are expressed as nanomole per minute per milligram total protein.

The relative levels of vtg-like proteins were determined following the principle of alkali labile phosphate (ALP) assay performed on high-molecular-weight proteins [10]. Gonad tissues for vtg analysis were homogenized at 4 °C with a polytron in 25 mM HEPEs buffer at pH 7.5, containing 125 mM NaCl, 1 mM dithiothreitol, and 1 mM EDTA. The homogenate was then centrifuged (12 000 g, 30 min, 4 °C). The supernatant was 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 by adding 200 µL of 1 M NaOH and heating at 60 °C for 30 min. Levels of ALP were measured following the molybdenum reagent for phosphate determination [25] with potassium phosphate as a standard. The level of vtg-like proteins is expressed as µg of ALP/mg total gonad protein.

Lipid peroxidation was determined by the thiobarbituric acid method, which measures the production of malonaldehyde [26]. Standard solutions of tetramethoxy-propane, a stabilized form of malonaldehyde, were prepared in the presence of the homogenization buffer for calibration. Fluorescence was measured at 520 nm excitation and 590 nm emission with a fluorescence microplate reader (Bioscan). Because the reagent could react with other aldehydes, results were expressed as µg of thiobarbituric acid reactants (TBARS) per mg of total protein.

Immune function

Hemocyte viability was analyzed in triplicate 100-µL hemolymph samples following the method of Blaise et al. [27]. Briefly, viability was quantified by determining the amount of hemocyte-retained fluorescamine, normalized with total protein content. Hemocyte protein concentration was determined in lyzed cells with fluorescamine (Sigma). Standards of fluorescein (hemocyte viability) and bovine serum albumin (protein) were used for calibration. Fluorescence was measured at 485 nm excitation with 520 nm emission for the viability assay, and 400 nm excitation with 450 nm emission for the protein assay, using a microplate reader (BioTek). Cell viability is expressed as micromoles of cell fluorescein per milligram of cell proteins.

Lysozyme activity was measured in plasma according to the method described by Lee and Yang [28]. Briefly, 100 µL of hemolymph were added to 100 µL of a suspension of Micrococcus lysodeikticus prepared at 0.4 mg/mL in 0.1 M phosphate buffer, pH 6.2. The reaction was carried out and read with a microplate reader (BioTek) at 450 nm and 25 °C. The activity of lysozyme was calculated from the slope of the time course by linear regression of data points. A unit of lysozyme activity is defined as the quantity of enzyme that causes a decrease in absorbance of 0.001 per minute at pH 6.2 at 25 °C. Lysozyme activity was normalized to protein concentration in plasma, following the protein–dye binding method [22] with bovine serum albumin (Sigma) for calibration.

Statistical analysis

Results of biomarkers and immune function are presented as means with standard deviation (SD) as mean ± SD. Data were compared across deployment sites (UP, D1, D2) with analysis of variance (ANOVA) followed by Tukey's test (p < 0.05; Sigma Stat 3.5). For data that did not meet the normality and/or equal variance 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 from SPMDs (n = 3) are presented as means ± standard deviation. The biomarker data were also analyzed by discriminant function analysis (Statistica 8.0) to seek out differences among the study sites and determine which biomarkers contributed the most for site classification.

RESULTS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

Mussel deployments

No equipment vandalism or mussel mortality was experienced at any of the deployment sites. A heavy rainstorm on the day 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 over 25 cm. The hydrological flow profile for the Grand River over the 4-wk deployment period is illustrated in Figure 2.

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

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Dissection of deployed mussels on retrieval revealed that 45% (UP), 65% (D1), and 40% (D2) of the mussels deployed for 2 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 all target analytes (PCPPs) in amounts that were detectable by liquid chromatography–tandem mass spectrometry analysis, except for estrone, which was not detected in POCIS deployed at the upstream site. Table 1 shows the estimated concentrations in water (ng/L) calculated from the amounts accumulated in the passive samplers over the 2-wk deployment period using the sampling rates (L/d) for the target analytes determined in previous laboratory experiments with POCIS [6] and SPMDs [20]. The sampling rates selected for these estimates were determined in laboratory experiments at 15 °C. Overall, the concentrations of PCPPs were higher at both downstream sites compared with levels upstream of the WWTP.

Table 1. Estimated mean (with standard deviation) concentrations of selected pharmaceutical and personal care products at deployment sites, upstream (UP) and downstream (D1, D2) of a municipal wastewater treatment plant for 2 wka
SiteEstimated mean concentration in water (ng/L)
CBZIBUTPMSMXGMZESTTCSHHCBAHTN
  • a

    TCS, HHCB, and AHTN were sampled using semipermeable membrane devices (n = 2). All other compounds were sampled using polar organic contaminants integrated samplers (n = 3). Mean water temperature during passive sampler deployment was 14.0 °C (range, 12.0 °C–15.8 °C).

  • CBZ = carbamazepine; IBU = ibuprofen; TPM = trimethoprim; SMX = sulfamethoxazole; GMZ = gemfibrozil; EST = estrone, TCS = triclosan; HHCB = Galaxolide® (synthetic musk); AHTN = Tonalide® (synthetic musk); ND = not detected.

UP2.47 (0.41)2.30 (0.54)1.89 (0.83)0.23 (0.09)0.05 (0.01)ND0.618.314.5
D16.44 (0.33)11.80 (1.13)6.42 (1.14)1.30 (0.43)0.12 (0.02)0.21 (0.02)3.891.448.9
D26.71 (0.27)12.97 (0.87)5.75 (1.24)0.77 (0.13)0.10 (0.03)0.07 (0.04)3.473.722.5

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

Table 2. Concentrations of fecal coliform and total coliform at mussel deployment sitesa
Field siteTotal coliforms (CFU/100 mL)Fecal coliforms (CFU/100 mL)
  • a

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

UP150076
D17000680
D23400468

Biomarkers of exposure

Mussels exposed to the MWWE for 4 wk at both downstream sites 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 of exposure (Figure 3A). Similarly, analysis of 4-wk deployed mussels from both the D1 and D2 exposure sites revealed significantly elevated levels of GST activity in the digestive gland compared with upstream mussels, while no difference was observed in GST activity in 2-wk deployed mussels, regardless of exposure site (Figure 3B). Although GST was also measured in gonad tissue, levels did not differ across study sites (data not shown). Although MWWE-exposed mussels tended to have slightly lower levels, there was no significant difference in P450-3A activity among deployment sites after either 2 wk or 4 wk of exposure (Figure 3C).

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Figure 3. Biomarkers of xenobiotic exposure and oxidative stress. (A) Lipid peroxidation (LPO), (B) glutathione-S transferase (GST), and (C) cytochrome P450-3A (CytoP450) activity in the digestive gland tissue of freshwater mussels (Lasmigona costata) deployed upstream (UP) and downstream (D1 [0.5 km], D2 [1.5 km]) of a municipal wastewater treatment plant for either 2 wk or 4 wk. Error bars indicate standard deviation (n = 12). Bars identified with different letters of the same case are significantly different (p < 0.05). TBARS = thiobarbituric acid reactants; ABS = absorbance; FL = fluorescence.

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No significant differences in mean levels of vtg-like proteins across the study sites were observed whether comparisons were made with data from male and female mussels combined or separated (Figure 4). However, the mean levels (expressed as alkali-labile phosphate µg/mg protein) in male mussels (Figure 4B) deployed at the near-field site (D1) for 4 wk were slightly (p = 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, there was no significant response in vtg levels in the mussels deployed for 2 wk.

image

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

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Immune function

Mussels exposed to MWWE had higher levels of plasma lysozyme activity than those deployed upstream of the treatment plant (Figure 5A). Lysozyme activity (expressed as Δ slope/mg protein) in mussels from both downstream sites (D1, 14.1 ± 10.6; D2, 12.1 ± 7.5) was significantly higher after 4 wk of exposure than the upstream-deployed mussels (UP, 5.4 ± 4.6). The level of lysozyme activity at the nearest downstream site (D1) was also elevated after 2 wk of exposure, but the difference was not significant.

image

Figure 5. (A) Lysozyme activity and (B) hemocyte viability in freshwater mussels (Lasmigona costata) deployed upstream (UP) and downstream (D1 [0.5 km], D2 [1.5 km]) of a municipal wastewater treatment plant for either 2 wk or 4 wk. Error bars indicate standard deviation (2 wk n = 20; 4 wk n =15). Bars identified with different letters of the same case are significantly different (p < 0.05). FL = fluorescence.

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MWWE exposure also had a significant effect on the viability (expressed cell fluorescein/mg of cell proteins) of mussel hemocytes (Figure 5B). Mussels deployed at both downstream sites (D1, 3.7 × 10−3 ± 1.1 × 10−3; D2, 2.4 × 10−3 ± 8.0 × 10−4) for 2 wk had significantly higher hemocyte viability compared with upstream mussels (1.4 × 10−3 ± 3.5 × 10−4), while those deployed in the effluent plume for 4 wk only showed a significant increase in viability at the second downstream site (D2).

Relationships among biomarker responses

In an attempt to gain a global understanding of the observed responses after 4 wk of exposure, we performed a discriminant function analysis. This analysis can be used to determine how well the treatments, in this case caging sites, are separated from each other and also which biomarkers among the battery employed contribute the most to site classification or treatment identification (principal component analysis). The analysis revealed that all sites (UP, D1, and D2) were discriminated by more than 70% and hence, the sites could be considered different from each other (Figure 6). The upstream site was readily discriminated from the 2 downstream sites, based on the observed responses in LPO, GST, and lysozyme activity (component 1). Male vtg-like proteins were equally weighted with lysozyme activity. Site D1 was somewhat discriminated from site D2 based on hemocyte viability, vtg-like proteins in females, and P450-3A activity (component 2).

image

Figure 6. Discriminant function analysis of the biomarker data. Mussel deployment sites are referred to as UP (upstream of effluent outfall), D1 (0.5 km downstream), and D2 (1.5 km downstream) of effluent outfall. Biomarkers are referred to as HVia (hemocyte viability), vtgfemale (vitellogenin-like proteins in females), 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 of samples from the 3 sites (not statistically derived).

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DISCUSSION

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

Contaminants

A large number of PPCPs and endocrine-disrupting substances have been detected in both wastewater and surface waters and could be considered for this research. However, to minimize analytical effort and cost, while monitoring a series of compounds that will provide valuable information on the various mechanisms of removal, we analyzed a small group of indicator compounds that are typically detected in wastewater and for which the processes governing their fate in the aquatic environment are reasonably well understood. Nine indicator compounds were analyzed, including several PPCPs and a natural estrogen. The objective of the passive sampling approach was to demonstrate that there are differences in the concentrations of contaminants of wastewater origin at upstream and downstream locations in the Grand River and not necessarily to link biomarker responses to specific compounds.

The POCIS and SPMD passive samplers accumulated all of the target analytes. The data show that the target analytes were present at all deployment sites including the upstream location, probably as a result of effluents from the 6 municipal WWTPs that discharge into the Grand River upstream of the study area. However, the concentrations of all analytes increased at the first deployment site (D1) immediately downstream of the Kitchener WWTP discharge. While all target compounds were still present at the second downstream deployment site (D2), the concentrations decreased with the increased distance downstream of the discharge point (Table 1). The concentrations of the indicator compounds estimated from the passive sampler data are consistent with concentrations of these compounds in surface waters and in MWWE at locations in Ontario, Canada [1, 3, 6].

Overall, the chemistry data indicate that the freshwater mussels deployed downstream of the Kitchener WWTP were exposed to PPCPs and to estrone at ng/L concentrations over at least the initial 2-wk deployment period and there is no reason to believe that exposures were not approximately the same or higher over the remainder of the 4-wk deployment.

Biomarkers of exposure

Significant responses in GST activity and cellular membrane damage were observed in freshwater mussels deployed downstream of a municipal wastewater treatment plant for 4 wk, but were less evident in mussels exposed to the effluent for 2 wk. While some biomarkers may indeed require a longer time to respond, indicators of xenobiotic transformation such as GST and induction of vtg-like proteins have been induced after 2 wk in comparable caging studies with other species of freshwater mussels [29, 30]. In the present study, although the magnitude of the response (4 wk vs 2 wk) appears to be the result of the difference in exposure duration, this was not the only contributing factor. We believe that the reduced stress response observed in the 2-wk deployed mussels was the result at least in part 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 passive samplers were deployed, which was followed by a substantial increase in river flow for the following 14 d (Figure 2). Provisional flow data from the Grand River Conservation Authority (GRCA; Cambridge, ON, Canada) demonstrate that the flow rate peaked 2 d after deployment and took 14 d to return to typical fall levels (GRCA, unpublished data). During the first 2 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 accounted for 4.8% (3.8–5.8%) of the river flow at the WWTP outfall during the latter 2 wk of the deployment (GRCA, unpublished data; [31]). The additional dilution of MWWE and its associated contaminants during the first 2 wk of deployment likely translated into the marginal responses observed after 2 wk of exposure. Unfortunately, this exposure dilution theory cannot be confirmed by chemical data from the passive samplers because they were only deployed for the first 2 wk of the study.

Biomarkers

Mussels located at the deployment sites downstream of the MWWE had increased GST activity and elevated LPO, which indicates the presence of bioavailable xenobiotics in the plume. Elevation of GST was also reported in a previous study with Elliptio complanata caged downstream of a municipal WWTP [30] and in another laboratory study in which freshwater mussels injected with the arthritis medication metotrexate and a C8 solid-phase methanol extract from a primary-treated MWWE displayed an increase in GST activity after 48 h [32]. Both of these studies suggest that mussels may be responsive to PPCPs in municipal wastewater effluents. In addition, in the marine mussel Mytilus galloprovincialis, exposure to 11 µg/L propranolol, an antihypertensive drug, led to increased GST activity in gills, and acetaminophen exposure increased LPO in digestive gland [33]. While increased GST activity and LPO in MWWE-exposed mussels in the present study are consistent with the induction of these biomarkers by some PPCPs, further study is necessary to establish a causal link because of the complex chemical composition of MWWE.

Pharmaceuticals induce CYP3A4 in mammals [34] and CYP3 homologues in fish [35], thus, it was anticipated that these detoxification enzymes might be induced in mussels as a result of exposure to pharmaceuticals in the wastewater. However, there was no significant induction of P450-3A in the mussels caged downstream of the WWTP.

Vitellogenin-like proteins

Although there were no statistically significant differences between MWWE-exposed and unexposed mussels, there was a trend toward higher levels of vtg-like proteins in the downstream male mussels deployed for 4 wk. While separation of male and female mussels was necessary to reveal sex-specific trends, separating the sexes reduced the sample size and thus the power of the statistical analysis for this somewhat variable endpoint. The lack of vtg induction was not expected because the estrogenic effects of MWWE have been well documented, in mussels (other locations), and in fish from the Grand River system. Specifically, Tetreault et al. [12] reported 100% feminization of wild male darters (Etheostoma sp.) in this same MWWE-impacted area of the Grand River, and there have been reports of MWWE-induced changes in vtg-like proteins in other species of freshwater mussels [10, 11, 36, 37]. Indeed Gagné et al. [11] reported that wild male Elliptio complanata collected downstream of the MWWE from 2 Quebec cities had significantly increased vtg-like proteins. In the present study, because the slight elevation in male vtg levels was only evident after 4 wk, and not 2 wk of exposure, one could speculate that the exposure duration or strength (see river dilution discussion above) was not adequate to stimulate significant production of vtg. However, it is also important to note the timing of the present study with respect to the reproductive cycle of the mussels employed. Many of the mussels deployed in the cages were gravid, meaning they were brooding glochidia (larvae) in their gills and not in an active stage of gametogenesis. Long-term brooders (i.e., bradytictic) such as L. costata hold their glochidia over the winter before releasing them the following spring and initiating the next reproductive cycle [38]. Although there are no published reports of endocrine disruption (reproductive cycle-influenced or otherwise) in L. costata, marine mussels (Mytilus) have been reported to be more sensitive to estrogens during the early stages of gonad development, but were impervious to estrogens at maturity [39]. Perhaps if mussels had been deployed during an active stage in their reproductive cycle, rather than early in the fall during reproductive quiescence, stronger induction of vtg would have been observed. In addition, a larger sample size, especially of male mussels, would have been required to confirm whether MWWE exposure induces changes in vtg-like proteins in mussels in this ecosystem.

Immune function

The significant increase in lysozyme activity in the mussels deployed in the effluent plume can be attributed to the 2-fold (D2) to 4-fold (D1) increase in waterborne bacteria levels at the downstream sites compared with the upstream site (Table 2). Lysozymes are a key component of the bivalve's innate immune response against bacteria [40] that are secreted by hemocytes after pathogen recognition or physiological stress [41]. This is consistent with the observation that lysozyme activity followed the level of total coliforms in the water. In addition, hemocyte viability generally (with 1 exception, D1 at 4-wk) followed the coliform concentration at the deployment sites. Typically, pathogenic bacteria or sustained inflammation reduce hemocyte viability, although in some cases bacteria can increase lipase activity with lysozyme activity [42, 43], which can result in an increase in hemocyte viability, because the cell viability assay employed quantifies a nonspecific esterase substrate that could be influenced by lipases (lipase is an esterase).

Although mussel energy stores were not assessed, we suspect that the chronic stimulation of the immune system from elevated levels of waterborne bacteria could have associated energy costs for wild MWWE-exposed mussels. In addition, the ongoing exposure to MWWE and its associated pathogens could eventually lead to immunosuppression and render the mussels more susceptible to diseases [44]. Such chronic effects of MWWE exposure may be reflected in the finding that mussels living downstream of urban effluents in the Grand River have lower condition factors and do not live as long as those from the upstream reference area [14].

Relationships among biomarker responses

Based on discriminant function analysis, the biomarker responses in mussels caged at the upstream and downstream sites could be discriminated based on xenobiotic conjugation (GST), oxidative stress (LPO), lysozyme activity, and levels of vtg-like proteins in males. These biomarkers are well known to respond in organisms exposed to municipal wastewater effluent. For example, lysozymes are produced as a defense mechanism against bacteria, which corroborates the reported levels of bacteria in the surface waters of the study sites. In mussels caged downstream of municipal effluents in the Mille-Isle River (QC, Canada), lysozyme activity was also induced in the hemolymph and followed total heterotrophic bacteria [10]. Increased oxidative stress (LPO) and xenobiotic conjugating activity (GST) were also reported in bivalves exposed to municipal effluents [11, 45]. Changes in vtg-like proteins in male mussels have also been reported in wild mussel populations collected at 2 sites downstream of the discharge of a municipal wastewater effluent in Quebec (Canada) [11]. The discriminant function analysis is a useful approach for evaluating whether responses across several biomarkers in mussels are different at sites upstream and downstream of MWWE discharges. However, more work is required to evaluate whether these responses are the result of exposure to contaminants of emerging concern or are more generalized responses to the complex wastewater matrix.

Implications for resident freshwater mussel populations

Although the Grand River has historically contained a diverse population of freshwater mussels, 30% to 50% of the species once found there are believed to have been lost, and sewage pollution may have contributed to the decline [15]. The range of waterborne contaminants measured in the present study (PCPPs) and previous studies (agricultural chemicals [46] and PCPPs [7]) and in long-term water quality monitoring (metals and nutrients [17]) demonstrates that water quality downstream of municipal WWTPs in the Grand River is impaired. In addition to the chemical data, there is a growing body of evidence that the complex mixture of anthropogenic compounds in this river negatively impact resident biota [12-14]. By isolating the exposure of a single municipal WWTP, we have demonstrated that the compounds associated with MWWE induce physiological stress and likely contributed to the negative effects observed in wild mussels downstream of multiple WWTP [14]. While the mussels examined in the present study displayed effects after a 4-wk exposure to MWWE, because of their relatively sessile nature, wild mussels could be exposed to MWWE throughout their lifespan. Such chronic exposure to MWWE has been shown to negatively impact 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 losses and exposure to wastewater effluents. These authors conducted extensive surveys upstream and downstream of WWTPs and found that the areas downstream of the outfalls were nearly devoid of unionids, in 1 case declining from 6 mussels/m2 upstream of the WWTP to less than 1 mussel/m2 below the outfall [47]. Findings in Goudreau et al. [47] suggest either that at some point over their lifecycle, mussels exposed to MWWE experience acutely toxic conditions, perhaps as a result of the elevated levels of chloride [48] and ammonia [49] that are associated with MWWE [17], or that the cumulative damage of chronic MWWE exposure translates into negative impacts at the whole-organism level [14] and eventually the population level. In the present study, although no mortality of the adult mussels deployed in the MWWE plume was observed after a 4-wk exposure, it is unknown whether the sensitive early life stages of glochidia (larvae) and juvenile mussels would have survived the deployment. The reduced condition factor of adult L. costata living downstream of 11 municipal WWTPs suggests that long-term exposure to MWWE leads to organism-level impacts [14]; however, to our knowledge neither the short-term nor the long-term effects of MWWE on the more sensitive life stages of freshwater mussels are understood.

Based on the fact that populations of L. costata are found in most areas of the Grand River [15], including those downstream of multiple municipal WWTPs [14], this species may be relatively pollution tolerant. Therefore, we speculate that if a more sensitive species had been employed in the present MWWE caging study, the impacts of the exposure would in fact have been heightened. Of course, it is neither reasonable nor possible to conduct similar studies with rarer (i.e., species at risk) and potentially more sensitive species.

CONCLUSIONS

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

Freshwater mussels deployed in the effluent plume of a municipal WWTP exhibited evidence of oxidative stress and stimulated immune response, as well as possible estrogenic effects in male mussels. Passive sampling devices demonstrated that the mussels were exposed to a wide range of PPCPs. The estrogenic effects of the MWWE on mussels were not as striking as those recently reported in wild fish in this watershed; therefore, further investigation is needed to determine whether mussel feminization is occurring. Both the MWWE-induced physiological stress reported in the present study and previously reported whole-organism impacts corroborate long-standing assumptions that anthropogenic-derived contaminants contributed to the significant decline in freshwater mussel density and diversity in this ecosystem. The municipal WWTP examined in the present study (Kitchener) is currently undergoing significant facility upgrades, which will hopefully translate into improved water quality and in turn contribute to the recovery of this imperiled group of organisms.

Acknowledgment

  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES

The authors thank K. Oakes, M. Servos, and M. Jorge for their contributions to the project, as well as the Grand River Conservation Authority (S. Shifflett and M. Anderson) for providing river flow and temperature 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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  1. Top of page
  2. Abstract
  3. INTRODUCTION
  4. MATERIALS AND METHODS
  5. RESULTS
  6. DISCUSSION
  7. CONCLUSIONS
  8. Acknowledgment
  9. REFERENCES
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