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Physiological stress response in white suckers from agricultural drain waters containing pesticides and selenium
ARTICLE IN PRESS Ecotoxicology and Environmental Safety 72 (2009) 1249–1256
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Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Physiological stress response in white suckers from agricultural drain waters containing pesticides and selenium$, $$ L.L. Miller a,, J.B. Rasmussen a, V.P. Palace b, A. Hontela a a b
Department of Biological Sciences, University of Lethbridge, 4401 University Drive, Lethbridge, Alberta, Canada T1K 3M4 Department of Fisheries and Oceans Canada, Center for Environmental Research on Pesticides, 501 University Crescent, Winnipeg, Manitoba, Canada R3T 2N6
a r t i c l e in fo
abstract
Article history: Received 16 April 2008 Received in revised form 15 September 2008 Accepted 23 September 2008 Available online 18 November 2008
To assess the effect of agriculture drain water, a complex mixture containing pesticides and selenium (Se), on the physiological stress response, white suckers were collected from irrigation return flows in the summer and the fall and subjected to a stress challenge. Water (0.40–26.71 mg/L) and muscle Se (0.37–1.52 mg/g ww) levels were elevated at two sites and plasma acetylcholinesterase (AChE) activity (a marker of pesticide exposure) was lower in the fall (5.9770.45 mmol/min/mL) than the summer (10.7370.73 mmol/min/mL). Fish raised plasma cortisol levels in response to the stress challenge 11.8 times above basal levels (12.874.9 ng/mL). Multivariate statistics linked Se exposure to elevated plasma glucose levels, and pesticide exposure to elevated liver glycogen levels generating hypotheses for further testing. This study showed that white suckers accumulated Se from agricultural drain water and the complex mixtures present in the drain water influenced the physiological stress response. & 2008 Elsevier Inc. All rights reserved.
Keywords: Selenium Pesticides Acetylcholinesterase White sucker Cortisol Stress response Agricultural drain water
1. Introduction Irrigation return flows and agricultural drain waters often contain multiple contaminants including pesticides (Anderson, 2005), selenium (Se) (Ohlendorf and Santolo, 1994), veterinary pharmaceuticals (Forrest et al., 2006), salts (Schlenk et al., 2003), particulates (Velasco et al., 2006), and nutrients (Velasco et al., 2006) that may adversely affect fish. Carbamate and organophosphate pesticides, entering surface waters from direct applications, aerial drift or leaching from soils, affect fish behaviour, food consumption, and reproductive function (Dell’Omo et al., 1997; Begum and Vijayaraghavan, 1999). They also inhibit the activity of acetylcholinesterase (AChE); thus, decreased AChE activity is used as a biomarker of pesticide exposure in fish (Sturm et al., 1999). This is very useful in areas where pesticide exposure data are not available or are difficult to obtain due to the large number of pesticides present. While pesticides are xenobiotics, Se is an
$ Funding sources: We gratefully acknowledge the support of the Natural Science and Engineering Research Council (NSERC) Metals In The Human Environment (MITHE) Research Network. A full list of sponsors is available at: www.mithe-rn. org. We also acknowledge a NSERC PGS-M awarded to L.L. Miller. $$ Assurance of animal care: Animal-use protocols have been approved by the University of Lethbridge Animal Care Committee in accordance with national guidelines. Corresponding author. Fax: +1 403 329 2082. E-mail address: [email protected] (L.L. Miller).
0147-6513/$ - see front matter & 2008 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2008.09.018
essential element that bioconcentrates and can be toxic at concentrations slightly greater than needed to maintain homeostasis (Hilton et al., 1980). Se is maternally deposited into eggs and can cause teratogenic deformities above-threshold concentrations (Holm et al., 2005; Muscatello et al., 2006). At the Kesterson reservoir, a wetland that received agricultural drain water from the San Joaquin Valley, CA, USA, many aquatic birds and native fish species were adversely impacted by Se-induced teratogenesis (Ohlendorf and Santolo, 1994). It is not known if the Kesterson reservoir is a unique case of Se contamination or a wide spread phenomenon occurring in many agricultural drain systems. While the effect of Se on the reproduction of fish and many effects of pesticides have been documented, the effects of Se and the effects of Se–pesticide mixtures on the physiological stress response (PSR), are not well understood. The PSR is a series of responses, including mobilization of energy reserves, enabling organisms to maintain homeostasis during exposure to a stressor. When a fish perceives a chemical, physical, or biological stimulus as a stressor, cortisol, and catecholamines are released (Hontela, 1998). Elevated cortisol levels may induce metallothionein (Hyliner et al., 1989), stimulate protein catabolism (Freeman and Idler, 1973), and increase plasma glucose levels and Na+/K+-ATPase activity (Shrimpton and McCormick, 1999). Secondarily, cortisol can also suppress the immune system, gonadal maturation, and sex steroid secretion (Carragher and Sumpter, 1990; Hontela, 1997). Activation or impairment of the PSR by a contaminant
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(Hontela, 2005) depends on the duration of exposure and type of chemical. Acute and subchronic (30 day) laboratory exposures to waterborne sodium selenite activated the PSR in rainbow trout (Miller et al., 2007). In contrast, field exposure to agricultural chemicals, including pesticides impaired the PSR (Dorval et al., 2005), potentially interfering with the allocation of energy reserves. Pesticides may also activate the PSR, elevating plasma glucose levels and decreasing glycogen stores (Bretaud et al., 2002). The PSR integrates the response to all of the stressors into a series of responses that can be easily evaluated. For example, resource management agencies could use low condition, together with abnormal glycogen reserves and disrupted hormonal stress response, as indicators of a suboptimal physiological status. Similar endpoints were used in yellow perch from lakes impacted by mining and smelting in Sudbury and Rouyn-Noranda regions (Campbell et al., 2008; Rasmussen et al., 2008). Fish exposed to chemical stressors in the field often exhibit altered seasonal energy reserve patterns. Yellow perch from metal-contaminated lakes exhibited little or no seasonal cycling in glycogen reserves, while perch from reference lakes seasonally increased their glycogen (Levesque et al., 2002). Winter-stress syndrome is an extension of this phenomenon where chemical stressors, such as Se, and the natural reduction in feeding during the winter result in mortality due to depleted energy reserves (Lemly, 1993a). Contaminant mixtures in field studies add another layer of complexity as the specific contaminants may occur at various ratios at different sites influencing parameters in ways that require multivariate statistical analyses. These can be used to generate hypotheses to be tested in future controlled experiments. Fluctuations in biotic and chemical stressors may significantly alter energy metabolism, and subsequently, growth and survival of fish exposed to agricultural drain water. Most ecotoxicological studies focus on large game fish species, but non-game fish (e.g., Catostomus species) may also be useful bioindicators of contaminant exposure and effects as benthivors may be vulnerable to particulate-bound contaminants. The use of non-game fish as bioindicators or sentinel species is an emerging research area because of their abundance and presumed higher site fidelity ensuring exposure to effluents under investigation (Palace et al., 2005). The objectives of this study were to determine (1) if Se contamination is a concern for fish living in agricultural drain water from southern Alberta and (2) the effect of agricultural drain water on the PSR of white suckers.
2. Materials and methods 2.1. Chemicals Acetylthiocholine iodide, tricaine methanesulfonate (MS-222), KH2PO4, Na2 HPO4, OMPA, DNTB, Tris buffer, Bradford reagent, KOH, glacial acetic acid, sodium acetate trihydrate, and amyloglucosidase were purchased from Sigma-Aldrich (Oakville, Ontario, Canada). Ultra pure nitric acid was purchased from Fisher Scientific (Ottawa, Ontario, Canada). GOD-PAP reagent was purchased from Roche Diagnostic (Laval, Que´bec, Canada).
2.1. Study sites Fish were sampled from four different drains (canals receiving irrigation return flows from a canal network) in the Lethbridge Northern Irrigation District (LNID; 491380 N, 1121480 W) and from the Little Bow River, Alberta, Canada (Fig. 1). The LNID drainage canals (filled from April to October) receive irrigation drain water from 175,000 acres of land used for intensive grain and animal production. The Little Bow River (a permanent stream) receives drain water from the dry land agricultural areas. Fish and water samples were collected in the summer (late June to early August) and the fall (mid-September to mid-October) at all sites.
Fig. 1. Sample sites in the Lethbridge Northern Irrigation District (LNID) drainage canals and the Little Bow River, Southern Alberta, Canada. % indicates sampling sites, - indicates direction of water flow, solid lines indicate roads. Solid black represents the city of Lethbridge, and the grey represents the LNID.
2.3. Fish sampling Animal-use protocols have been approved by the University of Lethbridge Animal Care Committee in accordance with national guidelines. All fish were captured with a Smith Root LR-24 electroshocker in the morning (0900–1200 h) and held in enclosures (0.27 m3) in the stream until sampling (1300–1400 h) as a standardized stress challenge. White suckers (8.4–28.0 cm) were anaesthetized with MS-222 (0.1 g/L), weighed, measured and blood sampled from the caudal blood vessels. Fish were then sacrificed by spinal transection. Plasma for AChE, cortisol, and glucose analyses was recovered (blood centrifuged at 16,000g for 5 min) and flash frozen in liquid nitrogen. Fork length and weight were recorded and condition factor (K ¼ (wt(g) 100)/length(cm)3.09) and liver somatic index (LSI ¼ liver wt(g)/body wt(g) 100) were calculated. Livers for glycogen analysis and gills for Na+/K+-ATPase activity were removed and flash frozen in liquid nitrogen. Muscle samples for Se analysis were obtained from the left side of the fish under the dorsal fin. Additionally, some white suckers were lightly anaesthetized immediately after electroshocking (without confinement stress) and a blood sample was taken from the caudal blood vessels for basal plasma cortisol, glucose, and AChE analysis. These fish were released after they had recovered from the anaesthetic. Water temperatures were recorded during electrofishing. Conductivity, pH, nitrites, and nitrates were measured on archived water samples (Table 2).
2.4. Selenium analysis Unfiltered water samples collected from each site were acidified with 0.05% ultra pure nitric acid and analysed for total Se by Inductively Coupled Plasma–Mass Spectrometry (ICP–MS) on an Elan DRC-II ICP-MS with CH4 as the reaction gas. The method has a detection limit of 0.01 mg/L. NIST 1640 (NIST, USA) and TM-Rain 95 (Environment Canada) were used as the certified reference materials for the Se analysis. Further QA/QC was done through the Ecosystem Proficiency Testing QA Program of Environment Canada. Muscle samples were analysed for total Se by hydride generation–atomic absorption (detection limit: 0.05 mg/g). Samples were weighed, and digested (4 mL nitric acid, 0.5 mL sulphuric acid, and 1.0 mL perchloric acid) at two temperatures (140 1C for 5 h and 200 1C for 2 h). Samples were analysed using a nonlinear curve through zero calibration curve with pre-determined concentrations with a Flame Atomic Absorption–Hydride Generation (model: Varian AA-55, VGA-77 Accessory). Analysis of certified reference material Dorm 2, Tort 2, CRM 2976 was included.
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Table 1 Performance characteristics of biochemical assays Interassay variabilitya
Assay
Assay sensitivityb
High Glucose (mg/mL) Na+/K+ ATPase (mmol/mg protein) Glycogen (mg/g) Protein (mg/mL) AChE (mmol/min/mL) a b
0.885 7.94 1.99 0.683 17.50
Low (2.8%) (10.6%) (5.8%) (7.3%) (7.2%)
0.166 1.21 0.054 0.240 9.72
(1.5%) (13.7%) (12.8%) (12.8%) (4.9%)
0.005 0.010 0.040 0.030 0.960
Mean values for low and high concentrations or activities, and percent coefficient of variation (CV) measured in different assays (n ¼ 10). The lowest detectable dose defined as the dose resulting in a response 2 SD from the zero dose response (n ¼ 10).
Table 2 Water quality characteristics from LNID drainage canals and the Little Bow River, Southern Alberta, Canada in the summer (S) and the fall (F) Site
Water (1C)
Little Bow Monarch BASF Pyami Battersea a b
Nitritesa (mg/L)
Conductivity (mS)
pH
Nitratesb (mg/L)
S
F
S
F
S
F
S
F
S
F
18.0 17.0 17.3 21.3 24.5
8.5 10.3 8.7 10.5 9.5
9.91 9.06 9.14 9.24 9.73
9.56 9.26 9.72 8.71 9.58
392 6 1958 508 734
783 6 872 409 458
0.05 0.10 0.25 0.05 0.05
0.05 0.05 0.05 0.05 0.05
o0.1 3.0 1.0 o0.1 0.6
0.1 5.0 1.0 0.3 0.6
Measured using APHA Standard Methods, method 4500-NO 2 B (Clescieri et al., 1998). Measured using APHA Standard Methods, method 4500-NO 3 E (Clescieri et al., 1998).
2.5. Biochemical parameters Acetylcholinesterase activity was measured by the change in absorbance at 405 nm for 10 min (Ellman et al., 1961) in a 96-well microplate. The final reaction (142 mL) contained 5 mM OMPA (tetraisopropyl pyrophosphoramide), 0.07 mM DNTB (5, 50 -dithiobis-2-nitro-benzoic acid), and 4.2 mM acetylthiocholine iodide in Tris buffer and 2 mL of sample. Activity was expressed as mmol DNTB degradation per minute per mL plasma or mg protein. Assay characteristics are given in Table 1. Cortisol was measured using a commercially available radioimmunoassay (catalogue #07-221102, Medicorp, Montre´al, Que´bec). Assay characteristics, including intra- and interassay variability, were assessed with internal standards, as previously described (Levesque et al., 2003). Glucose was measured by incubating plasma samples (60 min, 23 1C) with the GOD-PAP reagent and measuring the absorbance at 510 nm. Na+/K+-ATPase activity, expressed as mmol PO4 liberated per mg of protein in a gill homogenate, was measured by liberating PO4 from a hydrolysis reaction with ATPase, as described previously (Miller et al., 2007). Protein concentrations were determined at 595 nm using the Bradford regent in a spectrophotometric assay. Assay characteristics are given in Table 1. Liver glycogen was measured by digesting glycogen with amyloglucosidase and determining the resulting glucose concentration (Bleau et al., 1996). Livers were initially digested with KOH (1 N) and acetic acid (1.5 N), and the samples centrifuged (16,000g, 10 min). The glycogen in the supernatant was digested in a 1 N acetate buffer with amyloglucosidase (30 min, 37 1C). Glycogen levels were determined by incubating samples with the GOD-PAP reagent (60 min, 23 1C) and measuring the absorbance at 510 nm. Results were expressed as mg glycogen per g wet liver weight. Assay characteristics are given in Table 1.
2.6. Statistical analyses Data were analysed using JMP IN 5.1.2. (1989–2002 SAS Institute Inc.) with
a ¼ 0.05 for all tests. To determine different exposure characteristics, site and season were compared with a two-way ANOVA and post hoc Tukey–Kramer HSD test (normal data) or a Kruskal–Wallis test followed by multiple Wilcoxon tests with the Bonferroni correction (non-parametric data). To determine the effects of multiple stressors on the PSR of white suckers, an analysis of covariance was used. Muscle Se levels (indicator of Se exposure), plasma AChE activities (indicator of pesticide exposure), site, season (to control for natural seasonal variation), and length were considered as co-variates. Sex and maturity (gonadal somatic index) were also included in initial model testing but they did not significantly influence any of the parameters measured. Final models were selected by successive removal of co-variates that were not significant (p40.05). The PSR parameters (plasma cortisol, plasma glucose, gill Na+/K+ ATPase activity, liver glycogen, condition
factor, and liver somatic index) were transformed (Box–Cox) to ensure residuals were normally distributed. The assumption of homogeneity of slopes was also tested by including an interaction term during model testing.
3. Results 3.1. Site and exposure characteristics In the summer, water temperatures ranged from 17.0 to 24.5 1C while in the fall, water temperatures ranged from 8.5 to 10.5 1C (Table 2). Additional water quality characteristics are also described in Table 2. Water Se levels were the highest at the Monarch site in both the summer and the fall, and muscle Se levels of white suckers were significantly (po0.05) elevated at both the Monarch and BASF (Fig. 2). All Se levels were above the detection limit. Muscle Se levels were significantly (po0.05) lower in the fall than the summer at Battersea and BASF (Fig. 2). All samples were above the detection limit and reference material read within the acceptable limits. The mean water content of all muscle samples was 78.470.23%. Concentrations of pesticides in irrigation drains are not routinely monitored by Alberta Environment, thus only limited data are available. Pesticides were detected in the Little Bow River (only 2001 data available) and the Battersea drain (2001 and 2005) by Alberta Environment (Table 3). Plasma AChE activity of white suckers was significantly lower in the fall than the summer (po0.05) at all sites (Fig. 3). In the summer, fish from the Little Bow River had lower (po0.05) plasma AChE activity than fish from Monarch, BASF, and Battersea (Fig. 3). There was also a significant (po0.05) negative relationship between the fork length of white sucker and plasma AChE activity in the summer (y ¼ 5.9353x+295.82; R2 ¼ 0.1327; data not shown) and the fall (y ¼ 3.7911x+183.15; R2 ¼ 0.1666; data not shown). This relationship was not evident in the suckers sampled without confinement stress (basal), however the seasonal differences in AChE activity (summer ¼ 164.078.1 mmol/min/mL;
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1.8
30
Muscle Se (µg/g wet weight)
1.6 1.4 1.2
25
20
1.0 15 0.8 0.6
10
Total Water Se (µg/L)
Muscle Summer Muscle Fall Water Summer Water Fall
0.4 5 0.2 0.0
0 Little Bow
Monarch
BASF Site
Pyami
Battersea
Fig. 2. White sucker muscle (n ¼ 10–17) and water (n ¼ 1) Se levels from LNID drainage canals and the Little Bow River in the summer and the fall. Different letters indicate a significant difference. The solid line identifies the Canadian Water Quality Guideline (1 mL) for Se.
Table 3 Pesticides detected by Alberta Environment in the Little Bow River and the Battersea drain in 2001 and 2005 Maximum concentrations of pesticides detected (mg/L)
Site
Little Bow 2001c Battersea 2001 Battersea 2005 a b c
Dichlorophenoxyacetic acid (2,4-D)
Atrazine
Bromoxynil
Clopyralid (Lontrel)
Dichlorprop (2,4-DP)
MCPAa
MCPP
Picloram (Tordon)
0.020 0.377 0.433
o0.005 o0.005 0.109
o0.005 0.007 0.077
o0.020 0.069 0.018
o0.005 0.046 0.058
0.023 7.279 0.219
o0.005 o0.005 0.012
0.016 o0.005 o0.005
Simazineb
0.024
Triallate (Avadex BW)
o0.005 0.036 0.037
Detection limit of MCPA is 0.005 mg/L. Simazine was not tested for in 2001. Alberta Environment did not collect pesticide data at this site in 2005.
AChE Activity (µmol/min/ml plasma)
Summer 300
Fall
a a
a
250 ab 200 150
bc
bc c
c
c c
100 50 0 Little Bow
Monarch
BASF Site
Pyami
Battersea
Fig. 3. Plasma AChE activities of confinement-stressed white suckers (n ¼ 10–17) from the LNID drainage canals and the Little Bow River. Different letters indicate a significant difference.
fall 89.679.3 mmol/min/mL) were still observed (data not shown). Plasma AChE activities were not correlated (r2 ¼ 0.04) to muscle Se levels. 3.2. Physiological stress response (PSR) Models including the covariates that accounted for a significant (po0.05) amount of variation in each PSR parameter for
white suckers are given in Table 4. Plasma cortisol levels in white suckers were not influenced by muscle Se levels or plasma AChE activity. They were, however, significantly influenced (po0.05) by site and season (Table 4), and they were higher in the fall than in the summer (Fig. 4A). Gill Na+/K+ ATPase activities of white suckers were also significantly influenced by site and season (Table 4), and lower levels were observed in the fall (Fig. 4B). Condition factor was only influenced by season (Table 4) and it was lower in the fall (1.2070.02) than in the summer (1.2870.02). The liver somatic index model did not have homogeneous slopes, thus it was removed from further analyses. Plasma glucose levels were significantly influenced by muscle Se levels, site and season, but not pesticide exposure, as estimated by plasma AChE activity (Table 4). Plasma glucose levels increased with increasing muscle Se levels at all sites, both in the summer and the fall (Fig. 5A). A significant amount of variation in liver glycogen levels was explained by plasma AChE activities, site and season, but not muscle Se levels (Table 4). In the summer, fish with higher plasma AChE activities had lower liver glycogen levels, but in the fall the glycogen levels were higher and no such relationship was evident (Fig. 5B). Plasma cortisol (confinement stressed ¼ 151.0715.7 ng/mL; basal ¼ 12.874.9 ng/mL) and glucose (confinement stressed ¼ 1.7070.09 mg/mL; basal ¼ 0.5470.07 mg/mL) levels were lower in white suckers sampled without confinement (basal, data not shown) than those that had been stressed.
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Table 4 Analysis of covariance models that describe a significant (po0.05) portion of the variation in the PSR parameters of white suckersa Parameter
Box–Cox transformation (l)
Adjusted r2
Model
Plasma cortisol Plasma glucose Gill Na+/K+ ATPase Liver glycogen Condition factor
0.6 0.8 0.2 0.4 2.0
0.1980 0.2692 0.3954 0.7185 0.1185
Site+season Selenium+site+season+site season Site+season+site season AChE+site+season+site season Season
a
Confinement-stressed fish.
300
Summer
Cortisol (ng/ml plasma).
250
Fall
200 150 100 50 0 Little Bow
Gill Na+/K+ ATPase Activity. (µmol/mg protein)
1.00
Monarch
BASF Site
Pyami
Battersea
BASF Site
Pyami
Battersea
Summer
0.80
Fall
0.60 0.40 0.20 0.00 Little Bow
Monarch
Fig. 4. Plasma cortisol levels (A), and gill Na+/K+ ATPase activity (B), in confinement-stressed white suckers from the LNID drainage canals and the Little Bow River.
4. Discussion Drainage of agricultural lands can increase Se runoff into adjacent streams, potentially increasing concentrations in aquatic foodwebs beyond thresholds for Se toxicity, including the appearance of deformities in the offspring of fish (Ohlendorf and Santolo, 1994). The water concentrations observed in Monarch and BASF were similar to those in exposed sites where deformities were observed in rainbow trout (Holm et al., 2005). The highest Se muscle concentrations were detected in white suckers from the BASF and Monarch drains (1.5270.12 and 1.2370.08 mg/g wet weight, respectively). When converted to mg Se/g dry weight whole body using the muscle moisture content and the USEPA’s equation (USEPA, 2004), these values (6.4770.67 and 5.3670.47 mg Se/g dry weight whole body, respectively) were well below those associated with deformities in adult white suckers (42 mg Se/g dry weight whole body) (Lemly, 1993b), but in the range that embryos exhibit deformities (1.8 mg Se/g wet weight) (de Rosemond et al., 2005). Se is efficiently deposited into eggs by female fish because it binds to the primary yolk precursor, vitellogenin (Holm et al., 2005). However, the relationship between concentrations of Se in muscle and those deposited into
the eggs differ between species, with some species approximating a 1:1 ratio between the two compartments (Palace et al., 2004, Muscatello et al., 2006). If white sucker embryos in this system had similar Se levels as the muscle of adult fish, larval deformities could occur. Deformed larvae have lower survival rates, thus persistence of deformities in adult fish is rare. Additional studies should focus on the link between larval survival and Se in contaminated systems. Together with Se, pesticides are also present in the agricultural drain water from Southern Alberta. Pesticides and muscle Se levels were not correlated, thus, sites with high Se did not necessarily have higher pesticide contamination. It is difficult to establish contaminant exposure-effects links in such complex systems. In white suckers, plasma AChE, a marker of pesticide exposure (Whitehead et al., 2005) activities did not vary greatly between sites, but the activities were consistently lower in fish sampled in the fall than the summer. This may indicate that fish were exposed to greater concentrations of pesticides in the fall or the cumulative exposure to pesticides throughout the summer progressively lowered AChE activity. However, other factors (e.g., water temperature, and body size) may also influence AChE activities. Temperature differences within 10 1C do not appear to significantly influence cholinesterase activities in fish (Beauvais et al., 2002), but a temperature decrease of more than 10 1C lowers AChE activity (Hazel, 1969). The seasonal temperature differences in this study ranged from 6.7 to 15.0 1C, thus the decrease observed in AChE activity in the fall may be due to temperature and/or pesticide exposure. Size may also influence AChE levels—a significant negative relationships between length or weight, and AChE activities have been documented in stickleback (Sturm et al., 1999), bluegill (Beauvais et al., 2002), and also in the present study. The effects of temperature and other seasonal factors on AChE activities in white suckers have not been characterized, yet such data are needed to validate the use of AChE activities in monitoring pesticide contamination. Similarly, use of plasma cortisol in assessing fish health is debated (Hontela, 2005). In white suckers, plasma cortisol levels were elevated 11.8 times the basal level with the stress challenge and this response was influenced by site and season, with higher cortisol levels in the fall. Cool fall temperatures (8–11 1C), well below white suckers’ thermal preference of 27 1C (Cincotta and Stauffer, 1984), may have mediated the season effect. Alternatively, the season effect may be due to differences in photoperiod, annual cortisol cycles, flow alterations (Flodmark et al., 2002), and contaminant exposure duration. The site effect may be caused by additional chemical and physical stressors that were not quantified. In the complex system investigated in the present study, the classic stress response was not observed (cortisol was not correlated with Se exposure). However, plasma glucose levels were significantly influenced by muscle Se levels, site, and season with glucose increasing with muscle Se levels in the summer and
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Summer
Fall 4
Pyami Battersea LB Monarch
3
2
Plasma Glucose (mg/ml)
Plasma Glucose (mg/ml)
4
BASF
1
0
3 Pyami BASF, LB, Battersea Monarch
2
1
0 0
0.5 1.0 1.5 2.0 Muscle Se (µg Se/ g wet weight)
2.5
0.5 1.0 1.5 2.0 Muscle Se (µg Se/ g wet weight)
0
2.5
150 150 Pyami
125 Liver Glycogen (mg/g)
Liver Glycogen (mg/g)
125 100 75 50 25 BASF, Battersea, Pyami Monarch, LB
0 0
100 200 300 Plasma AChE (µmol/min/ml)
400
100
BASF
75
Battersea, LB
50 Monarch
25 0 0
100 200 300 400 Plasma AChE (µmol/min/ml)
Fig. 5. The relationship between (A) muscle Se and plasma glucose levels, and (B) plasma AChE activity and liver glycogen levels in the summer and the fall in confinementstressed white suckers from LNID drainage canals (Monarch ( ), BASF (’), Pyami (n), and Battersea (K) and the Little Bow River (J)). Lines represent the ANCOVA model that describes r2 (refer to Table 4) of the variation in the parameter.
the fall. This suggests white suckers exposed to increased Se may mobilize more energy reserves. Similarly, higher plasma glucose levels have also been found in fish exposed to other contaminants including Se (Miller et al., 2007) and bleached kraft mill effluent (Jardine et al., 1996). To explain why the classic stress response was not observed and to confirm Se’s effect on the PSR of white suckers, controlled studies are needed. The activity of gill Na+/K+-ATPase, an enzyme involved in osmoregulation, was influenced by site and season, with lower activity in the fall than the summer at all sites, except Monarch. Fish may express different variants and levels of gill Na+/K+-ATPase in the summer and the winter (Packer and Garvin, 1998); moreover site-specific chemical stressors and temperature may also influence gill Na+/K+-ATPase activity. For example, exposure to silver (Morgan et al., 1997) and copper (Gagnon et al., 2006) decreased gill Na+/K+-ATPase activity in rainbow trout. More research is needed to determine what component of the agricultural drain water influenced gill Na+/K+-ATPase activity in white suckers, as muscle Se levels and plasma AChE activity did not significantly explain the variation observed. Plasma AChE activity, site and season explained a significant amount of the variation in liver glycogen levels, a secondary stress response. Unexpectedly, white suckers with higher AChE (lower pesticide exposures) had lower liver glycogen levels in the summer. The expected pattern after contaminant exposure is a decrease in glycogen reserves (Bleau et al., 1996), as fish metabolize glycogen to meet the increased energy demands of a
chemical stressor. However, if contaminants interrupt the pathway leading to glycogen mobilization, exposed fish may have greater glycogen reserves than unexposed fish, as has been reported in some impacted systems (Andersson et al., 1988; Hontela et al., 1995; Dorval et al., 2005). This difference may be further magnified by additional stressors that stimulate glycogen mobilization in the unimpaired fish. It has not been established if carbamate and organophosphate pesticides can interrupt glycogen mobilization in white suckers and how this influences their response to additional environmental stressors. In the fall, the relationship between liver glycogen and AChE activity is no longer evident and liver glycogen levels appear higher than in the summer. Fish normally increase their glycogen reserves in preparation for winter (Lemly, 1993a) and also after they have spawned (Rinchard and Kestemont, 2003). This seasonal increase could mask the effect of pesticide exposure on glycogen reserves by naturally elevating the low glycogen reserves of the unimpaired fish. Traditionally, an increase in cortisol and glucose levels would also be linked to pesticide exposure, thus the hypothesis that pesticides impair glycogen degradation should be tested in a controlled pesticide exposure with a standardized stress challenge. Condition factor was significantly influenced only by season. Condition factor of confinement-stressed white suckers was lower in the fall, suggesting that there is an energetic cost associated with living in agricultural drain water for an extended period of time. Yellow perch exposed to organic contaminants and metals
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also had lower condition factors than reference fish in various field exposure systems (Hontela et al., 1995; Couture and Kumar, 2003). Investigations of the effects of agricultural drain waters on fish health and aquatic ecosystems are important since these systems often drain large areas impacted by intensive agricultural activities and return some of the flow to rivers, lakes or reservoirs. Agricultural drain water contains a complex mixture of contaminants including Se and pesticides that may occur in different ratios. In this study, white suckers living in agricultural drain water accumulated Se at some sites, but levels were below those associated with deformities in adult fish. Suckers had elevated plasma cortisol levels and decreased condition factor in the fall suggesting that the stressors present in the agricultural drain water negatively impacted fish over time. Se exposure increased plasma glucose levels, suggesting white suckers mobilized energy reserves in response to this stressor. Liver glycogen levels in the summer were higher in fish exposed to pesticides, implying an impairment of the glycogen mobilization pathway. The hypotheses that Se elevates plasma glucose levels and pesticide exposure can impair glycogen mobilization should be tested in controlled studies.
Acknowledgments We thank R. Flitton, C. Friesen, I. Harper, W. Warnock, and T. Miller for assistance in the field and in the laboratory. We also thank Dr. F. Wang and D. Armstrong (Ultra-Clean Trace Elements Laboratory, Univ. of Manitoba) for water Se analysis, S. Kollar for muscle Se analysis (DFO, Winnipeg), and K. Forbes (University of Lethbridge) for the map. We gratefully acknowledge the support of the Natural Science and Engineering Research Council (NSERC) Metals In The Human Environment (MITHE) Research Network. A full list of sponsors is available at: www.mithe-rn.org. We also acknowledge a NSERC PGS-M awarded to L.L. Miller. References Anderson, A., 2005. Overview of Pesticide Data in Alberta Surface Waters since 1995. Alberta Environment, Edmonton. Andersson, T., Fo¨rlin, L., Ha¨rdig, J., Larsson, A., 1988. Physiological disturbances in fish living in coastal water polluted with bleached kraft pulp mill effluents. Can. J. Fish. Aquat. Sci. 45, 1525–1536. Beauvais, S.L., Cole, K.J., Atchison, G.J., Coffey, M., 2002. Factors affecting brain cholinesterase activity in bluegill (Lepomis macrochiurs). Water Air Soil Pollut. 135, 249–264. Begum, G., Vijayaraghavan, S., 1999. Effect of acute exposure of the organophosphate insecticide rogor on some biochemical aspects of Clarias batrachus (Linnaeus). Environ. Res. 80, 80–83. Bleau, H., Daniel, C., Chevalier, G., van Tra, H., Hontela, A., 1996. Effects of acute exposure to mercury chloride and methylmercury on plasma cortisol, T3, T4, glucose and liver glycogen in rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 34, 221–235. Bretaud, S., Saglio, P., Saligaut, C., Auperin, B., 2002. Biochemical and behavioural effects of carbofuran in goldfish (Carassius auratus). Environ. Toxicol. Chem. 21, 175–181. Campbell, P.G.C., Kraemer, L.D., Giguere, A., Hare, A., Hontela, A., 2008. Subcellular distribution of cadmium and nickel in chronically exposed wild fish: inferences regarding metal detoxification strategies and implications for setting water quality guidelines for dissolved metals. Hum. Ecol. Risk Assess. 14, 290–316. Carragher, J., Sumpter, J., 1990. The effect of cortisol on the secretion of sex steroids from cultured ovarian follicles of rainbow trout. Gen. Comp. Endocrinol. 77, 403–407. Cincotta, D.A., Stauffer, J.R., 1984. Temperature preference and avoidance studies of six North American freshwater fish species. Hydrobiologia 109, 173–177. Clescieri, L.S., Arnold, E., Deaton, A.D., 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, American Water Works Association, and Water Environment Federation, Baltimore. Couture, P., Kumar, P.R., 2003. Impairment of metabolic capacities in copper and cadmium contaminated wild yellow perch, Perca flavescens. Aquat. Toxicol. 64, 107–120.
1255
Dell’Omo, G., Bryenton, R., Shore, R.F., 1997. Effects of exposure to an organophosphate pesticide on behaviour and acetylcholinesterase activity in the common shrew, Sorex araneus. Environ. Toxicol. Chem. 16, 272–276. de Rosemond, S.C., Liber, K., Rosaasen, A., 2005. Relationship between embryo selenium concentration and early life stage development in white sucker (Catostomus commersoni) from a northern Canadian Lake. Bull. Environ. Contam. Toxicol. 74, 1134–1142. Dorval, J., Leblond, V., Deblois, C., Hontela, A., 2005. Oxidative stress and endocrine endpoints in white sucker (Catostomus commersoni) from a river impacted by agricultural chemicals. Environ. Toxicol. Chem. 24, 1273–1280. Ellman, G.L., Courtney, K.D., Andres, V., Featherstone, R.M., 1961. A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7, 88–95. Flodmark, L.E.W., Urke, H.A., Halleraker, J.H., Arnekleiv, J.V., Vøllestad, L.A., Pele´o, A.B.S., 2002. Cortisol and glucose responses in juvenile brown trout subjected to a fluctuating flow regime in an artificial stream. J. Fish Biol. 60, 238–248. Forrest, F., Keenliside, J., Kendall, J., Thompson, T., Noot, D., Wuite, J., 2006. Livestock Pharmaceuticals in Agricultural Streams: A Scoping Study for Alberta. Alberta Agriculture, Food, and Rural Development, Edmonton. Freeman, H.C., Idler, D.R., 1973. Effects of corticosteroids on liver transaminases in two salmonids, the rainbow trout (Salmo gairdnerii) and the brook trout (Salvelinus fontinalis). Gen. Comp. Endocrinol. 20, 69–75. Gagnon, A., Jumarie, C., Hontela, A., 2006. Effects of Cu on plasma cortisol and cortisol secretion by andrenocortical cells of rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 78, 59–65. Hazel, J., 1969. The effect of thermal acclimation upon brain acetlycholinesterase activity of Carassius auratus and Fundulus heterolitus. Life Sci. 8, 775–784. Hilton, J.W., Hodson, P.V., Slinger, S.J., 1980. The requirement and toxicity of selenium in rainbow trout (Salmo gairdneri). J. Nutr. 110, 2527–2535. Holm, J., Palace, V., Siwik, P., Sterling, G., Evans, R., Baron, C., Weerner, J., Wautier, K., 2005. Developmental effects of bioaccumulated selenium in eggs and larvae of two salmonid species. Environ. Toxicol. Chem. 24, 2373–2381. Hontela, A., 1997. Endocrine and physiological responses of fish to xenobiotics: role of glucocorticosteroid hormones. Rev. Toxicol. 1, 1–46. Hontela, A., 1998. Interrenal dysfunction in fish from contaminated sites: in vivo and in vitro assessment. Environ. Toxicol. Chem. 17, 44–48. Hontela, A., 2005. Adrenal toxicology: environmental pollutants and the HPI axis. In: Mommsen, T.P., Moon, T.W. (Eds.), Biochemistry and Molecular Biology of Fishes. Elsevier B.V., Amsterdam, pp. 331–363. Hontela, A., Dumont, J., Duclos, D., Fortin, R., 1995. Endocrine and metabolic dysfunction in yellow perch, Perca flavescens, exposed to organic contaminants and heavy metals in the St. Lawrence River. Environ. Toxicol. Chem. 14, 725–731. Hyliner, H., Andersson, T., Haux, C., Olsson, P.-E., 1989. Cortisol induction of metallothionein in primary culture of rainbow trout hepatocytes. J. Cell Physiol. 139, 24–28. Jardine, J.J., Van Der Kraak, G., Munkittrick, K.R., 1996. Capture and confinement stress in white sucker exposed to bleached kraft pulp mill effluent. Ecotoxicol. Environ. Saf. 33, 287–298. Lemly, A.D., 1993a. Metabolic stress during winter increases the toxicity of selenium to fish. Aquat. Toxicol. 27, 133–158. Lemly, A.D., 1993b. Teratogenic effects of selenium in natural populations of freshwater fish. Ecotoxicol. Environ. Saf. 26, 181–204. Levesque, H.M., Moon, T.W., Campbell, P.G.C., Hontela, A., 2002. Seasonal variation in carbohydrate and lipid metabolism of yellow perch (Perca flavescens) chronically exposed to metals in the field. Aquat. Toxicol. 60, 257–267. Levesque, H.M., Dorval, J., Hontela, A., Van Der Kraak, G.J., Campbell, P.G.C., 2003. Hormonal, morphological, and physiological responses of yellow perch (Perca flavescens) to chronic environmental metal exposures. J. Toxicol. Environ. Health A 66, 657–676. Miller, L.L., Wang, F., Palace, V.P., Hontela, A., 2007. Effects of acute and subchronic exposures to waterborne selenite on the physiological stress response and oxidative stress indicators in juvenile rainbow trout. Aquat. Toxicol. 83, 263–271. Morgan, I.J., Henry, R.P., Wood, C.M., 1997. The mechanism of acute silver nitrate toxicity in freshwater rainbow trout (Oncorhynchus mykiss) is inhibition of gill Na+ and Cl transport. Aquat. Toxicol. 38, 145–163. Muscatello, J.R., Bennett, P.M., Himbeault, K.T., Belknap, A.M., Janz, D.M., 2006. Larval deformities associated with selenium accumulation in northern pike (Esox lucius) exposed to metal mining effluent. Environ. Sci. Technol. 40, 6506–6512. Ohlendorf, H.M., Santolo, G.M., 1994. Kesterson reservoir—past, present, and future: an ecological risk assessment. In: Frankenberger, Jr., W.T., Benson, S. (Eds.), Selenium in the Environment. Marcel Dekker Inc., New York, pp. 69–117. Packer, R.K., Garvin, J.L., 1998. Seasonal differences in activity of perch (Perca flavescens) gill Na+/K+ ATPase. Comp. Biochem. Physiol. B 120, 777–783. Palace, V.P., Evans, R.E., Wautier, K., Baron, C.L., Kollar, S., Werner, J., Holm, J., Siwik, P., Sterling, G., Johnson, C.F., 2004. An assessment of the potential for selenium to impair reproduction in bull trout (Salvelinus confluentus) from an area of active coal mining. Environ. Biol. Fishes 70, 169–174. Palace, V.P., Doebel, C., Baron, C.L., Evans, R.E., Wauiter, K.G., Klaverkamp, J.F., Werner, J., Kollar, S., 2005. Caging small bodied fish as an alternative method for environmental effects monitoring (EEM). Can. Water Qual. Res. J. 40, 328–333. Rasmussen, J.B., Gunn, J.M., Sherwood, G.D., Iles, A., Gagnon, A., Campbell, P.G.C., Hontela, A., 2008. Direct and indirect (foodweb mediated) effects of metal
ARTICLE IN PRESS 1256
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exposure on the growth of yellow perch (Perca flavescens): implications for ecological risk assessment. Hum. Ecol. Risk Assess. 14, 317–350. Rinchard, J., Kestemont, P., 2003. Liver changes releated to oocyte growth in roach, a single spawner fish, and in bleak and white bream, two multiple spawner fish. Int. Rev. Hydrobiol. 88, 68–76. Schlenk, D., Zubcov, N., Zubcov, E., 2003. Effects of salinity on the uptake, biotransformation, and toxicity of dietary seleno-L-methionine to rainbow trout. Toxicol. Sci. 75, 309–313. Shrimpton, J.M., McCormick, S.D., 1999. Responsiveness of gill Na+/K+-ATPase to cortisol is related to gill corticosteroid receptor concentration in juvenile rainbow trout. J. Exp. Biol. 202, 987–995. Sturm, A., Wogram, J., Hansen, P.-D., Liess, M., 1999. Potential use of cholinesterase in monitoring low levels of organophosphates in small streams: natural
variability in three-spined stickleback (Gasterosteus aculeatus) and relation to pollution. Environ. Toxicol. Chem. 18, 194–200. United States Environnemental Protection Agency, 2004. Draft Aquatic Life Water Quality Criteria for Selenium—2004. Office of Water, Office of Science and Technology, US Environmental Protection Agency, Washington, DC, 334pp. Velasco, J., Lloret, J., Millan, L., Marin, L., Barahona, J., Abellan, P., SanchezFernandez, D., 2006. Nutrient and particulate inputs into the Mar Menor Lagoon (SE Spain) from an intensive agricultural watershed. Water Air Soil Pollut. 176, 37–56. Whitehead, A., Anserson, S.L., Ramirez, A., Wilson, B.W., 2005. Cholinesterases in aquatic biomonitoring: assay optimization and species-specific characterization for a California native fish. Ecotoxicology 14, 597–606.
Contents lists available at ScienceDirect
Ecotoxicology and Environmental Safety journal homepage: www.elsevier.com/locate/ecoenv
Physiological stress response in white suckers from agricultural drain waters containing pesticides and selenium$, $$ L.L. Miller a,, J.B. Rasmussen a, V.P. Palace b, A. Hontela a a b
Department of Biological Sciences, University of Lethbridge, 4401 University Drive, Lethbridge, Alberta, Canada T1K 3M4 Department of Fisheries and Oceans Canada, Center for Environmental Research on Pesticides, 501 University Crescent, Winnipeg, Manitoba, Canada R3T 2N6
a r t i c l e in fo
abstract
Article history: Received 16 April 2008 Received in revised form 15 September 2008 Accepted 23 September 2008 Available online 18 November 2008
To assess the effect of agriculture drain water, a complex mixture containing pesticides and selenium (Se), on the physiological stress response, white suckers were collected from irrigation return flows in the summer and the fall and subjected to a stress challenge. Water (0.40–26.71 mg/L) and muscle Se (0.37–1.52 mg/g ww) levels were elevated at two sites and plasma acetylcholinesterase (AChE) activity (a marker of pesticide exposure) was lower in the fall (5.9770.45 mmol/min/mL) than the summer (10.7370.73 mmol/min/mL). Fish raised plasma cortisol levels in response to the stress challenge 11.8 times above basal levels (12.874.9 ng/mL). Multivariate statistics linked Se exposure to elevated plasma glucose levels, and pesticide exposure to elevated liver glycogen levels generating hypotheses for further testing. This study showed that white suckers accumulated Se from agricultural drain water and the complex mixtures present in the drain water influenced the physiological stress response. & 2008 Elsevier Inc. All rights reserved.
Keywords: Selenium Pesticides Acetylcholinesterase White sucker Cortisol Stress response Agricultural drain water
1. Introduction Irrigation return flows and agricultural drain waters often contain multiple contaminants including pesticides (Anderson, 2005), selenium (Se) (Ohlendorf and Santolo, 1994), veterinary pharmaceuticals (Forrest et al., 2006), salts (Schlenk et al., 2003), particulates (Velasco et al., 2006), and nutrients (Velasco et al., 2006) that may adversely affect fish. Carbamate and organophosphate pesticides, entering surface waters from direct applications, aerial drift or leaching from soils, affect fish behaviour, food consumption, and reproductive function (Dell’Omo et al., 1997; Begum and Vijayaraghavan, 1999). They also inhibit the activity of acetylcholinesterase (AChE); thus, decreased AChE activity is used as a biomarker of pesticide exposure in fish (Sturm et al., 1999). This is very useful in areas where pesticide exposure data are not available or are difficult to obtain due to the large number of pesticides present. While pesticides are xenobiotics, Se is an
$ Funding sources: We gratefully acknowledge the support of the Natural Science and Engineering Research Council (NSERC) Metals In The Human Environment (MITHE) Research Network. A full list of sponsors is available at: www.mithe-rn. org. We also acknowledge a NSERC PGS-M awarded to L.L. Miller. $$ Assurance of animal care: Animal-use protocols have been approved by the University of Lethbridge Animal Care Committee in accordance with national guidelines. Corresponding author. Fax: +1 403 329 2082. E-mail address: [email protected] (L.L. Miller).
0147-6513/$ - see front matter & 2008 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2008.09.018
essential element that bioconcentrates and can be toxic at concentrations slightly greater than needed to maintain homeostasis (Hilton et al., 1980). Se is maternally deposited into eggs and can cause teratogenic deformities above-threshold concentrations (Holm et al., 2005; Muscatello et al., 2006). At the Kesterson reservoir, a wetland that received agricultural drain water from the San Joaquin Valley, CA, USA, many aquatic birds and native fish species were adversely impacted by Se-induced teratogenesis (Ohlendorf and Santolo, 1994). It is not known if the Kesterson reservoir is a unique case of Se contamination or a wide spread phenomenon occurring in many agricultural drain systems. While the effect of Se on the reproduction of fish and many effects of pesticides have been documented, the effects of Se and the effects of Se–pesticide mixtures on the physiological stress response (PSR), are not well understood. The PSR is a series of responses, including mobilization of energy reserves, enabling organisms to maintain homeostasis during exposure to a stressor. When a fish perceives a chemical, physical, or biological stimulus as a stressor, cortisol, and catecholamines are released (Hontela, 1998). Elevated cortisol levels may induce metallothionein (Hyliner et al., 1989), stimulate protein catabolism (Freeman and Idler, 1973), and increase plasma glucose levels and Na+/K+-ATPase activity (Shrimpton and McCormick, 1999). Secondarily, cortisol can also suppress the immune system, gonadal maturation, and sex steroid secretion (Carragher and Sumpter, 1990; Hontela, 1997). Activation or impairment of the PSR by a contaminant
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(Hontela, 2005) depends on the duration of exposure and type of chemical. Acute and subchronic (30 day) laboratory exposures to waterborne sodium selenite activated the PSR in rainbow trout (Miller et al., 2007). In contrast, field exposure to agricultural chemicals, including pesticides impaired the PSR (Dorval et al., 2005), potentially interfering with the allocation of energy reserves. Pesticides may also activate the PSR, elevating plasma glucose levels and decreasing glycogen stores (Bretaud et al., 2002). The PSR integrates the response to all of the stressors into a series of responses that can be easily evaluated. For example, resource management agencies could use low condition, together with abnormal glycogen reserves and disrupted hormonal stress response, as indicators of a suboptimal physiological status. Similar endpoints were used in yellow perch from lakes impacted by mining and smelting in Sudbury and Rouyn-Noranda regions (Campbell et al., 2008; Rasmussen et al., 2008). Fish exposed to chemical stressors in the field often exhibit altered seasonal energy reserve patterns. Yellow perch from metal-contaminated lakes exhibited little or no seasonal cycling in glycogen reserves, while perch from reference lakes seasonally increased their glycogen (Levesque et al., 2002). Winter-stress syndrome is an extension of this phenomenon where chemical stressors, such as Se, and the natural reduction in feeding during the winter result in mortality due to depleted energy reserves (Lemly, 1993a). Contaminant mixtures in field studies add another layer of complexity as the specific contaminants may occur at various ratios at different sites influencing parameters in ways that require multivariate statistical analyses. These can be used to generate hypotheses to be tested in future controlled experiments. Fluctuations in biotic and chemical stressors may significantly alter energy metabolism, and subsequently, growth and survival of fish exposed to agricultural drain water. Most ecotoxicological studies focus on large game fish species, but non-game fish (e.g., Catostomus species) may also be useful bioindicators of contaminant exposure and effects as benthivors may be vulnerable to particulate-bound contaminants. The use of non-game fish as bioindicators or sentinel species is an emerging research area because of their abundance and presumed higher site fidelity ensuring exposure to effluents under investigation (Palace et al., 2005). The objectives of this study were to determine (1) if Se contamination is a concern for fish living in agricultural drain water from southern Alberta and (2) the effect of agricultural drain water on the PSR of white suckers.
2. Materials and methods 2.1. Chemicals Acetylthiocholine iodide, tricaine methanesulfonate (MS-222), KH2PO4, Na2 HPO4, OMPA, DNTB, Tris buffer, Bradford reagent, KOH, glacial acetic acid, sodium acetate trihydrate, and amyloglucosidase were purchased from Sigma-Aldrich (Oakville, Ontario, Canada). Ultra pure nitric acid was purchased from Fisher Scientific (Ottawa, Ontario, Canada). GOD-PAP reagent was purchased from Roche Diagnostic (Laval, Que´bec, Canada).
2.1. Study sites Fish were sampled from four different drains (canals receiving irrigation return flows from a canal network) in the Lethbridge Northern Irrigation District (LNID; 491380 N, 1121480 W) and from the Little Bow River, Alberta, Canada (Fig. 1). The LNID drainage canals (filled from April to October) receive irrigation drain water from 175,000 acres of land used for intensive grain and animal production. The Little Bow River (a permanent stream) receives drain water from the dry land agricultural areas. Fish and water samples were collected in the summer (late June to early August) and the fall (mid-September to mid-October) at all sites.
Fig. 1. Sample sites in the Lethbridge Northern Irrigation District (LNID) drainage canals and the Little Bow River, Southern Alberta, Canada. % indicates sampling sites, - indicates direction of water flow, solid lines indicate roads. Solid black represents the city of Lethbridge, and the grey represents the LNID.
2.3. Fish sampling Animal-use protocols have been approved by the University of Lethbridge Animal Care Committee in accordance with national guidelines. All fish were captured with a Smith Root LR-24 electroshocker in the morning (0900–1200 h) and held in enclosures (0.27 m3) in the stream until sampling (1300–1400 h) as a standardized stress challenge. White suckers (8.4–28.0 cm) were anaesthetized with MS-222 (0.1 g/L), weighed, measured and blood sampled from the caudal blood vessels. Fish were then sacrificed by spinal transection. Plasma for AChE, cortisol, and glucose analyses was recovered (blood centrifuged at 16,000g for 5 min) and flash frozen in liquid nitrogen. Fork length and weight were recorded and condition factor (K ¼ (wt(g) 100)/length(cm)3.09) and liver somatic index (LSI ¼ liver wt(g)/body wt(g) 100) were calculated. Livers for glycogen analysis and gills for Na+/K+-ATPase activity were removed and flash frozen in liquid nitrogen. Muscle samples for Se analysis were obtained from the left side of the fish under the dorsal fin. Additionally, some white suckers were lightly anaesthetized immediately after electroshocking (without confinement stress) and a blood sample was taken from the caudal blood vessels for basal plasma cortisol, glucose, and AChE analysis. These fish were released after they had recovered from the anaesthetic. Water temperatures were recorded during electrofishing. Conductivity, pH, nitrites, and nitrates were measured on archived water samples (Table 2).
2.4. Selenium analysis Unfiltered water samples collected from each site were acidified with 0.05% ultra pure nitric acid and analysed for total Se by Inductively Coupled Plasma–Mass Spectrometry (ICP–MS) on an Elan DRC-II ICP-MS with CH4 as the reaction gas. The method has a detection limit of 0.01 mg/L. NIST 1640 (NIST, USA) and TM-Rain 95 (Environment Canada) were used as the certified reference materials for the Se analysis. Further QA/QC was done through the Ecosystem Proficiency Testing QA Program of Environment Canada. Muscle samples were analysed for total Se by hydride generation–atomic absorption (detection limit: 0.05 mg/g). Samples were weighed, and digested (4 mL nitric acid, 0.5 mL sulphuric acid, and 1.0 mL perchloric acid) at two temperatures (140 1C for 5 h and 200 1C for 2 h). Samples were analysed using a nonlinear curve through zero calibration curve with pre-determined concentrations with a Flame Atomic Absorption–Hydride Generation (model: Varian AA-55, VGA-77 Accessory). Analysis of certified reference material Dorm 2, Tort 2, CRM 2976 was included.
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Table 1 Performance characteristics of biochemical assays Interassay variabilitya
Assay
Assay sensitivityb
High Glucose (mg/mL) Na+/K+ ATPase (mmol/mg protein) Glycogen (mg/g) Protein (mg/mL) AChE (mmol/min/mL) a b
0.885 7.94 1.99 0.683 17.50
Low (2.8%) (10.6%) (5.8%) (7.3%) (7.2%)
0.166 1.21 0.054 0.240 9.72
(1.5%) (13.7%) (12.8%) (12.8%) (4.9%)
0.005 0.010 0.040 0.030 0.960
Mean values for low and high concentrations or activities, and percent coefficient of variation (CV) measured in different assays (n ¼ 10). The lowest detectable dose defined as the dose resulting in a response 2 SD from the zero dose response (n ¼ 10).
Table 2 Water quality characteristics from LNID drainage canals and the Little Bow River, Southern Alberta, Canada in the summer (S) and the fall (F) Site
Water (1C)
Little Bow Monarch BASF Pyami Battersea a b
Nitritesa (mg/L)
Conductivity (mS)
pH
Nitratesb (mg/L)
S
F
S
F
S
F
S
F
S
F
18.0 17.0 17.3 21.3 24.5
8.5 10.3 8.7 10.5 9.5
9.91 9.06 9.14 9.24 9.73
9.56 9.26 9.72 8.71 9.58
392 6 1958 508 734
783 6 872 409 458
0.05 0.10 0.25 0.05 0.05
0.05 0.05 0.05 0.05 0.05
o0.1 3.0 1.0 o0.1 0.6
0.1 5.0 1.0 0.3 0.6
Measured using APHA Standard Methods, method 4500-NO 2 B (Clescieri et al., 1998). Measured using APHA Standard Methods, method 4500-NO 3 E (Clescieri et al., 1998).
2.5. Biochemical parameters Acetylcholinesterase activity was measured by the change in absorbance at 405 nm for 10 min (Ellman et al., 1961) in a 96-well microplate. The final reaction (142 mL) contained 5 mM OMPA (tetraisopropyl pyrophosphoramide), 0.07 mM DNTB (5, 50 -dithiobis-2-nitro-benzoic acid), and 4.2 mM acetylthiocholine iodide in Tris buffer and 2 mL of sample. Activity was expressed as mmol DNTB degradation per minute per mL plasma or mg protein. Assay characteristics are given in Table 1. Cortisol was measured using a commercially available radioimmunoassay (catalogue #07-221102, Medicorp, Montre´al, Que´bec). Assay characteristics, including intra- and interassay variability, were assessed with internal standards, as previously described (Levesque et al., 2003). Glucose was measured by incubating plasma samples (60 min, 23 1C) with the GOD-PAP reagent and measuring the absorbance at 510 nm. Na+/K+-ATPase activity, expressed as mmol PO4 liberated per mg of protein in a gill homogenate, was measured by liberating PO4 from a hydrolysis reaction with ATPase, as described previously (Miller et al., 2007). Protein concentrations were determined at 595 nm using the Bradford regent in a spectrophotometric assay. Assay characteristics are given in Table 1. Liver glycogen was measured by digesting glycogen with amyloglucosidase and determining the resulting glucose concentration (Bleau et al., 1996). Livers were initially digested with KOH (1 N) and acetic acid (1.5 N), and the samples centrifuged (16,000g, 10 min). The glycogen in the supernatant was digested in a 1 N acetate buffer with amyloglucosidase (30 min, 37 1C). Glycogen levels were determined by incubating samples with the GOD-PAP reagent (60 min, 23 1C) and measuring the absorbance at 510 nm. Results were expressed as mg glycogen per g wet liver weight. Assay characteristics are given in Table 1.
2.6. Statistical analyses Data were analysed using JMP IN 5.1.2. (1989–2002 SAS Institute Inc.) with
a ¼ 0.05 for all tests. To determine different exposure characteristics, site and season were compared with a two-way ANOVA and post hoc Tukey–Kramer HSD test (normal data) or a Kruskal–Wallis test followed by multiple Wilcoxon tests with the Bonferroni correction (non-parametric data). To determine the effects of multiple stressors on the PSR of white suckers, an analysis of covariance was used. Muscle Se levels (indicator of Se exposure), plasma AChE activities (indicator of pesticide exposure), site, season (to control for natural seasonal variation), and length were considered as co-variates. Sex and maturity (gonadal somatic index) were also included in initial model testing but they did not significantly influence any of the parameters measured. Final models were selected by successive removal of co-variates that were not significant (p40.05). The PSR parameters (plasma cortisol, plasma glucose, gill Na+/K+ ATPase activity, liver glycogen, condition
factor, and liver somatic index) were transformed (Box–Cox) to ensure residuals were normally distributed. The assumption of homogeneity of slopes was also tested by including an interaction term during model testing.
3. Results 3.1. Site and exposure characteristics In the summer, water temperatures ranged from 17.0 to 24.5 1C while in the fall, water temperatures ranged from 8.5 to 10.5 1C (Table 2). Additional water quality characteristics are also described in Table 2. Water Se levels were the highest at the Monarch site in both the summer and the fall, and muscle Se levels of white suckers were significantly (po0.05) elevated at both the Monarch and BASF (Fig. 2). All Se levels were above the detection limit. Muscle Se levels were significantly (po0.05) lower in the fall than the summer at Battersea and BASF (Fig. 2). All samples were above the detection limit and reference material read within the acceptable limits. The mean water content of all muscle samples was 78.470.23%. Concentrations of pesticides in irrigation drains are not routinely monitored by Alberta Environment, thus only limited data are available. Pesticides were detected in the Little Bow River (only 2001 data available) and the Battersea drain (2001 and 2005) by Alberta Environment (Table 3). Plasma AChE activity of white suckers was significantly lower in the fall than the summer (po0.05) at all sites (Fig. 3). In the summer, fish from the Little Bow River had lower (po0.05) plasma AChE activity than fish from Monarch, BASF, and Battersea (Fig. 3). There was also a significant (po0.05) negative relationship between the fork length of white sucker and plasma AChE activity in the summer (y ¼ 5.9353x+295.82; R2 ¼ 0.1327; data not shown) and the fall (y ¼ 3.7911x+183.15; R2 ¼ 0.1666; data not shown). This relationship was not evident in the suckers sampled without confinement stress (basal), however the seasonal differences in AChE activity (summer ¼ 164.078.1 mmol/min/mL;
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1.8
30
Muscle Se (µg/g wet weight)
1.6 1.4 1.2
25
20
1.0 15 0.8 0.6
10
Total Water Se (µg/L)
Muscle Summer Muscle Fall Water Summer Water Fall
0.4 5 0.2 0.0
0 Little Bow
Monarch
BASF Site
Pyami
Battersea
Fig. 2. White sucker muscle (n ¼ 10–17) and water (n ¼ 1) Se levels from LNID drainage canals and the Little Bow River in the summer and the fall. Different letters indicate a significant difference. The solid line identifies the Canadian Water Quality Guideline (1 mL) for Se.
Table 3 Pesticides detected by Alberta Environment in the Little Bow River and the Battersea drain in 2001 and 2005 Maximum concentrations of pesticides detected (mg/L)
Site
Little Bow 2001c Battersea 2001 Battersea 2005 a b c
Dichlorophenoxyacetic acid (2,4-D)
Atrazine
Bromoxynil
Clopyralid (Lontrel)
Dichlorprop (2,4-DP)
MCPAa
MCPP
Picloram (Tordon)
0.020 0.377 0.433
o0.005 o0.005 0.109
o0.005 0.007 0.077
o0.020 0.069 0.018
o0.005 0.046 0.058
0.023 7.279 0.219
o0.005 o0.005 0.012
0.016 o0.005 o0.005
Simazineb
0.024
Triallate (Avadex BW)
o0.005 0.036 0.037
Detection limit of MCPA is 0.005 mg/L. Simazine was not tested for in 2001. Alberta Environment did not collect pesticide data at this site in 2005.
AChE Activity (µmol/min/ml plasma)
Summer 300
Fall
a a
a
250 ab 200 150
bc
bc c
c
c c
100 50 0 Little Bow
Monarch
BASF Site
Pyami
Battersea
Fig. 3. Plasma AChE activities of confinement-stressed white suckers (n ¼ 10–17) from the LNID drainage canals and the Little Bow River. Different letters indicate a significant difference.
fall 89.679.3 mmol/min/mL) were still observed (data not shown). Plasma AChE activities were not correlated (r2 ¼ 0.04) to muscle Se levels. 3.2. Physiological stress response (PSR) Models including the covariates that accounted for a significant (po0.05) amount of variation in each PSR parameter for
white suckers are given in Table 4. Plasma cortisol levels in white suckers were not influenced by muscle Se levels or plasma AChE activity. They were, however, significantly influenced (po0.05) by site and season (Table 4), and they were higher in the fall than in the summer (Fig. 4A). Gill Na+/K+ ATPase activities of white suckers were also significantly influenced by site and season (Table 4), and lower levels were observed in the fall (Fig. 4B). Condition factor was only influenced by season (Table 4) and it was lower in the fall (1.2070.02) than in the summer (1.2870.02). The liver somatic index model did not have homogeneous slopes, thus it was removed from further analyses. Plasma glucose levels were significantly influenced by muscle Se levels, site and season, but not pesticide exposure, as estimated by plasma AChE activity (Table 4). Plasma glucose levels increased with increasing muscle Se levels at all sites, both in the summer and the fall (Fig. 5A). A significant amount of variation in liver glycogen levels was explained by plasma AChE activities, site and season, but not muscle Se levels (Table 4). In the summer, fish with higher plasma AChE activities had lower liver glycogen levels, but in the fall the glycogen levels were higher and no such relationship was evident (Fig. 5B). Plasma cortisol (confinement stressed ¼ 151.0715.7 ng/mL; basal ¼ 12.874.9 ng/mL) and glucose (confinement stressed ¼ 1.7070.09 mg/mL; basal ¼ 0.5470.07 mg/mL) levels were lower in white suckers sampled without confinement (basal, data not shown) than those that had been stressed.
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Table 4 Analysis of covariance models that describe a significant (po0.05) portion of the variation in the PSR parameters of white suckersa Parameter
Box–Cox transformation (l)
Adjusted r2
Model
Plasma cortisol Plasma glucose Gill Na+/K+ ATPase Liver glycogen Condition factor
0.6 0.8 0.2 0.4 2.0
0.1980 0.2692 0.3954 0.7185 0.1185
Site+season Selenium+site+season+site season Site+season+site season AChE+site+season+site season Season
a
Confinement-stressed fish.
300
Summer
Cortisol (ng/ml plasma).
250
Fall
200 150 100 50 0 Little Bow
Gill Na+/K+ ATPase Activity. (µmol/mg protein)
1.00
Monarch
BASF Site
Pyami
Battersea
BASF Site
Pyami
Battersea
Summer
0.80
Fall
0.60 0.40 0.20 0.00 Little Bow
Monarch
Fig. 4. Plasma cortisol levels (A), and gill Na+/K+ ATPase activity (B), in confinement-stressed white suckers from the LNID drainage canals and the Little Bow River.
4. Discussion Drainage of agricultural lands can increase Se runoff into adjacent streams, potentially increasing concentrations in aquatic foodwebs beyond thresholds for Se toxicity, including the appearance of deformities in the offspring of fish (Ohlendorf and Santolo, 1994). The water concentrations observed in Monarch and BASF were similar to those in exposed sites where deformities were observed in rainbow trout (Holm et al., 2005). The highest Se muscle concentrations were detected in white suckers from the BASF and Monarch drains (1.5270.12 and 1.2370.08 mg/g wet weight, respectively). When converted to mg Se/g dry weight whole body using the muscle moisture content and the USEPA’s equation (USEPA, 2004), these values (6.4770.67 and 5.3670.47 mg Se/g dry weight whole body, respectively) were well below those associated with deformities in adult white suckers (42 mg Se/g dry weight whole body) (Lemly, 1993b), but in the range that embryos exhibit deformities (1.8 mg Se/g wet weight) (de Rosemond et al., 2005). Se is efficiently deposited into eggs by female fish because it binds to the primary yolk precursor, vitellogenin (Holm et al., 2005). However, the relationship between concentrations of Se in muscle and those deposited into
the eggs differ between species, with some species approximating a 1:1 ratio between the two compartments (Palace et al., 2004, Muscatello et al., 2006). If white sucker embryos in this system had similar Se levels as the muscle of adult fish, larval deformities could occur. Deformed larvae have lower survival rates, thus persistence of deformities in adult fish is rare. Additional studies should focus on the link between larval survival and Se in contaminated systems. Together with Se, pesticides are also present in the agricultural drain water from Southern Alberta. Pesticides and muscle Se levels were not correlated, thus, sites with high Se did not necessarily have higher pesticide contamination. It is difficult to establish contaminant exposure-effects links in such complex systems. In white suckers, plasma AChE, a marker of pesticide exposure (Whitehead et al., 2005) activities did not vary greatly between sites, but the activities were consistently lower in fish sampled in the fall than the summer. This may indicate that fish were exposed to greater concentrations of pesticides in the fall or the cumulative exposure to pesticides throughout the summer progressively lowered AChE activity. However, other factors (e.g., water temperature, and body size) may also influence AChE activities. Temperature differences within 10 1C do not appear to significantly influence cholinesterase activities in fish (Beauvais et al., 2002), but a temperature decrease of more than 10 1C lowers AChE activity (Hazel, 1969). The seasonal temperature differences in this study ranged from 6.7 to 15.0 1C, thus the decrease observed in AChE activity in the fall may be due to temperature and/or pesticide exposure. Size may also influence AChE levels—a significant negative relationships between length or weight, and AChE activities have been documented in stickleback (Sturm et al., 1999), bluegill (Beauvais et al., 2002), and also in the present study. The effects of temperature and other seasonal factors on AChE activities in white suckers have not been characterized, yet such data are needed to validate the use of AChE activities in monitoring pesticide contamination. Similarly, use of plasma cortisol in assessing fish health is debated (Hontela, 2005). In white suckers, plasma cortisol levels were elevated 11.8 times the basal level with the stress challenge and this response was influenced by site and season, with higher cortisol levels in the fall. Cool fall temperatures (8–11 1C), well below white suckers’ thermal preference of 27 1C (Cincotta and Stauffer, 1984), may have mediated the season effect. Alternatively, the season effect may be due to differences in photoperiod, annual cortisol cycles, flow alterations (Flodmark et al., 2002), and contaminant exposure duration. The site effect may be caused by additional chemical and physical stressors that were not quantified. In the complex system investigated in the present study, the classic stress response was not observed (cortisol was not correlated with Se exposure). However, plasma glucose levels were significantly influenced by muscle Se levels, site, and season with glucose increasing with muscle Se levels in the summer and
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Summer
Fall 4
Pyami Battersea LB Monarch
3
2
Plasma Glucose (mg/ml)
Plasma Glucose (mg/ml)
4
BASF
1
0
3 Pyami BASF, LB, Battersea Monarch
2
1
0 0
0.5 1.0 1.5 2.0 Muscle Se (µg Se/ g wet weight)
2.5
0.5 1.0 1.5 2.0 Muscle Se (µg Se/ g wet weight)
0
2.5
150 150 Pyami
125 Liver Glycogen (mg/g)
Liver Glycogen (mg/g)
125 100 75 50 25 BASF, Battersea, Pyami Monarch, LB
0 0
100 200 300 Plasma AChE (µmol/min/ml)
400
100
BASF
75
Battersea, LB
50 Monarch
25 0 0
100 200 300 400 Plasma AChE (µmol/min/ml)
Fig. 5. The relationship between (A) muscle Se and plasma glucose levels, and (B) plasma AChE activity and liver glycogen levels in the summer and the fall in confinementstressed white suckers from LNID drainage canals (Monarch ( ), BASF (’), Pyami (n), and Battersea (K) and the Little Bow River (J)). Lines represent the ANCOVA model that describes r2 (refer to Table 4) of the variation in the parameter.
the fall. This suggests white suckers exposed to increased Se may mobilize more energy reserves. Similarly, higher plasma glucose levels have also been found in fish exposed to other contaminants including Se (Miller et al., 2007) and bleached kraft mill effluent (Jardine et al., 1996). To explain why the classic stress response was not observed and to confirm Se’s effect on the PSR of white suckers, controlled studies are needed. The activity of gill Na+/K+-ATPase, an enzyme involved in osmoregulation, was influenced by site and season, with lower activity in the fall than the summer at all sites, except Monarch. Fish may express different variants and levels of gill Na+/K+-ATPase in the summer and the winter (Packer and Garvin, 1998); moreover site-specific chemical stressors and temperature may also influence gill Na+/K+-ATPase activity. For example, exposure to silver (Morgan et al., 1997) and copper (Gagnon et al., 2006) decreased gill Na+/K+-ATPase activity in rainbow trout. More research is needed to determine what component of the agricultural drain water influenced gill Na+/K+-ATPase activity in white suckers, as muscle Se levels and plasma AChE activity did not significantly explain the variation observed. Plasma AChE activity, site and season explained a significant amount of the variation in liver glycogen levels, a secondary stress response. Unexpectedly, white suckers with higher AChE (lower pesticide exposures) had lower liver glycogen levels in the summer. The expected pattern after contaminant exposure is a decrease in glycogen reserves (Bleau et al., 1996), as fish metabolize glycogen to meet the increased energy demands of a
chemical stressor. However, if contaminants interrupt the pathway leading to glycogen mobilization, exposed fish may have greater glycogen reserves than unexposed fish, as has been reported in some impacted systems (Andersson et al., 1988; Hontela et al., 1995; Dorval et al., 2005). This difference may be further magnified by additional stressors that stimulate glycogen mobilization in the unimpaired fish. It has not been established if carbamate and organophosphate pesticides can interrupt glycogen mobilization in white suckers and how this influences their response to additional environmental stressors. In the fall, the relationship between liver glycogen and AChE activity is no longer evident and liver glycogen levels appear higher than in the summer. Fish normally increase their glycogen reserves in preparation for winter (Lemly, 1993a) and also after they have spawned (Rinchard and Kestemont, 2003). This seasonal increase could mask the effect of pesticide exposure on glycogen reserves by naturally elevating the low glycogen reserves of the unimpaired fish. Traditionally, an increase in cortisol and glucose levels would also be linked to pesticide exposure, thus the hypothesis that pesticides impair glycogen degradation should be tested in a controlled pesticide exposure with a standardized stress challenge. Condition factor was significantly influenced only by season. Condition factor of confinement-stressed white suckers was lower in the fall, suggesting that there is an energetic cost associated with living in agricultural drain water for an extended period of time. Yellow perch exposed to organic contaminants and metals
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also had lower condition factors than reference fish in various field exposure systems (Hontela et al., 1995; Couture and Kumar, 2003). Investigations of the effects of agricultural drain waters on fish health and aquatic ecosystems are important since these systems often drain large areas impacted by intensive agricultural activities and return some of the flow to rivers, lakes or reservoirs. Agricultural drain water contains a complex mixture of contaminants including Se and pesticides that may occur in different ratios. In this study, white suckers living in agricultural drain water accumulated Se at some sites, but levels were below those associated with deformities in adult fish. Suckers had elevated plasma cortisol levels and decreased condition factor in the fall suggesting that the stressors present in the agricultural drain water negatively impacted fish over time. Se exposure increased plasma glucose levels, suggesting white suckers mobilized energy reserves in response to this stressor. Liver glycogen levels in the summer were higher in fish exposed to pesticides, implying an impairment of the glycogen mobilization pathway. The hypotheses that Se elevates plasma glucose levels and pesticide exposure can impair glycogen mobilization should be tested in controlled studies.
Acknowledgments We thank R. Flitton, C. Friesen, I. Harper, W. Warnock, and T. Miller for assistance in the field and in the laboratory. We also thank Dr. F. Wang and D. Armstrong (Ultra-Clean Trace Elements Laboratory, Univ. of Manitoba) for water Se analysis, S. Kollar for muscle Se analysis (DFO, Winnipeg), and K. Forbes (University of Lethbridge) for the map. We gratefully acknowledge the support of the Natural Science and Engineering Research Council (NSERC) Metals In The Human Environment (MITHE) Research Network. A full list of sponsors is available at: www.mithe-rn.org. We also acknowledge a NSERC PGS-M awarded to L.L. Miller. References Anderson, A., 2005. Overview of Pesticide Data in Alberta Surface Waters since 1995. Alberta Environment, Edmonton. Andersson, T., Fo¨rlin, L., Ha¨rdig, J., Larsson, A., 1988. Physiological disturbances in fish living in coastal water polluted with bleached kraft pulp mill effluents. Can. J. Fish. Aquat. Sci. 45, 1525–1536. Beauvais, S.L., Cole, K.J., Atchison, G.J., Coffey, M., 2002. Factors affecting brain cholinesterase activity in bluegill (Lepomis macrochiurs). Water Air Soil Pollut. 135, 249–264. Begum, G., Vijayaraghavan, S., 1999. Effect of acute exposure of the organophosphate insecticide rogor on some biochemical aspects of Clarias batrachus (Linnaeus). Environ. Res. 80, 80–83. Bleau, H., Daniel, C., Chevalier, G., van Tra, H., Hontela, A., 1996. Effects of acute exposure to mercury chloride and methylmercury on plasma cortisol, T3, T4, glucose and liver glycogen in rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 34, 221–235. Bretaud, S., Saglio, P., Saligaut, C., Auperin, B., 2002. Biochemical and behavioural effects of carbofuran in goldfish (Carassius auratus). Environ. Toxicol. Chem. 21, 175–181. Campbell, P.G.C., Kraemer, L.D., Giguere, A., Hare, A., Hontela, A., 2008. Subcellular distribution of cadmium and nickel in chronically exposed wild fish: inferences regarding metal detoxification strategies and implications for setting water quality guidelines for dissolved metals. Hum. Ecol. Risk Assess. 14, 290–316. Carragher, J., Sumpter, J., 1990. The effect of cortisol on the secretion of sex steroids from cultured ovarian follicles of rainbow trout. Gen. Comp. Endocrinol. 77, 403–407. Cincotta, D.A., Stauffer, J.R., 1984. Temperature preference and avoidance studies of six North American freshwater fish species. Hydrobiologia 109, 173–177. Clescieri, L.S., Arnold, E., Deaton, A.D., 1998. Standard Methods for the Examination of Water and Wastewater, 20th ed. American Public Health Association, American Water Works Association, and Water Environment Federation, Baltimore. Couture, P., Kumar, P.R., 2003. Impairment of metabolic capacities in copper and cadmium contaminated wild yellow perch, Perca flavescens. Aquat. Toxicol. 64, 107–120.
1255
Dell’Omo, G., Bryenton, R., Shore, R.F., 1997. Effects of exposure to an organophosphate pesticide on behaviour and acetylcholinesterase activity in the common shrew, Sorex araneus. Environ. Toxicol. Chem. 16, 272–276. de Rosemond, S.C., Liber, K., Rosaasen, A., 2005. Relationship between embryo selenium concentration and early life stage development in white sucker (Catostomus commersoni) from a northern Canadian Lake. Bull. Environ. Contam. Toxicol. 74, 1134–1142. Dorval, J., Leblond, V., Deblois, C., Hontela, A., 2005. Oxidative stress and endocrine endpoints in white sucker (Catostomus commersoni) from a river impacted by agricultural chemicals. Environ. Toxicol. Chem. 24, 1273–1280. Ellman, G.L., Courtney, K.D., Andres, V., Featherstone, R.M., 1961. A new and rapid colorimetric determination of acetylcholinesterase activity. Biochem. Pharmacol. 7, 88–95. Flodmark, L.E.W., Urke, H.A., Halleraker, J.H., Arnekleiv, J.V., Vøllestad, L.A., Pele´o, A.B.S., 2002. Cortisol and glucose responses in juvenile brown trout subjected to a fluctuating flow regime in an artificial stream. J. Fish Biol. 60, 238–248. Forrest, F., Keenliside, J., Kendall, J., Thompson, T., Noot, D., Wuite, J., 2006. Livestock Pharmaceuticals in Agricultural Streams: A Scoping Study for Alberta. Alberta Agriculture, Food, and Rural Development, Edmonton. Freeman, H.C., Idler, D.R., 1973. Effects of corticosteroids on liver transaminases in two salmonids, the rainbow trout (Salmo gairdnerii) and the brook trout (Salvelinus fontinalis). Gen. Comp. Endocrinol. 20, 69–75. Gagnon, A., Jumarie, C., Hontela, A., 2006. Effects of Cu on plasma cortisol and cortisol secretion by andrenocortical cells of rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 78, 59–65. Hazel, J., 1969. The effect of thermal acclimation upon brain acetlycholinesterase activity of Carassius auratus and Fundulus heterolitus. Life Sci. 8, 775–784. Hilton, J.W., Hodson, P.V., Slinger, S.J., 1980. The requirement and toxicity of selenium in rainbow trout (Salmo gairdneri). J. Nutr. 110, 2527–2535. Holm, J., Palace, V., Siwik, P., Sterling, G., Evans, R., Baron, C., Weerner, J., Wautier, K., 2005. Developmental effects of bioaccumulated selenium in eggs and larvae of two salmonid species. Environ. Toxicol. Chem. 24, 2373–2381. Hontela, A., 1997. Endocrine and physiological responses of fish to xenobiotics: role of glucocorticosteroid hormones. Rev. Toxicol. 1, 1–46. Hontela, A., 1998. Interrenal dysfunction in fish from contaminated sites: in vivo and in vitro assessment. Environ. Toxicol. Chem. 17, 44–48. Hontela, A., 2005. Adrenal toxicology: environmental pollutants and the HPI axis. In: Mommsen, T.P., Moon, T.W. (Eds.), Biochemistry and Molecular Biology of Fishes. Elsevier B.V., Amsterdam, pp. 331–363. Hontela, A., Dumont, J., Duclos, D., Fortin, R., 1995. Endocrine and metabolic dysfunction in yellow perch, Perca flavescens, exposed to organic contaminants and heavy metals in the St. Lawrence River. Environ. Toxicol. Chem. 14, 725–731. Hyliner, H., Andersson, T., Haux, C., Olsson, P.-E., 1989. Cortisol induction of metallothionein in primary culture of rainbow trout hepatocytes. J. Cell Physiol. 139, 24–28. Jardine, J.J., Van Der Kraak, G., Munkittrick, K.R., 1996. Capture and confinement stress in white sucker exposed to bleached kraft pulp mill effluent. Ecotoxicol. Environ. Saf. 33, 287–298. Lemly, A.D., 1993a. Metabolic stress during winter increases the toxicity of selenium to fish. Aquat. Toxicol. 27, 133–158. Lemly, A.D., 1993b. Teratogenic effects of selenium in natural populations of freshwater fish. Ecotoxicol. Environ. Saf. 26, 181–204. Levesque, H.M., Moon, T.W., Campbell, P.G.C., Hontela, A., 2002. Seasonal variation in carbohydrate and lipid metabolism of yellow perch (Perca flavescens) chronically exposed to metals in the field. Aquat. Toxicol. 60, 257–267. Levesque, H.M., Dorval, J., Hontela, A., Van Der Kraak, G.J., Campbell, P.G.C., 2003. Hormonal, morphological, and physiological responses of yellow perch (Perca flavescens) to chronic environmental metal exposures. J. Toxicol. Environ. Health A 66, 657–676. Miller, L.L., Wang, F., Palace, V.P., Hontela, A., 2007. Effects of acute and subchronic exposures to waterborne selenite on the physiological stress response and oxidative stress indicators in juvenile rainbow trout. Aquat. Toxicol. 83, 263–271. Morgan, I.J., Henry, R.P., Wood, C.M., 1997. The mechanism of acute silver nitrate toxicity in freshwater rainbow trout (Oncorhynchus mykiss) is inhibition of gill Na+ and Cl transport. Aquat. Toxicol. 38, 145–163. Muscatello, J.R., Bennett, P.M., Himbeault, K.T., Belknap, A.M., Janz, D.M., 2006. Larval deformities associated with selenium accumulation in northern pike (Esox lucius) exposed to metal mining effluent. Environ. Sci. Technol. 40, 6506–6512. Ohlendorf, H.M., Santolo, G.M., 1994. Kesterson reservoir—past, present, and future: an ecological risk assessment. In: Frankenberger, Jr., W.T., Benson, S. (Eds.), Selenium in the Environment. Marcel Dekker Inc., New York, pp. 69–117. Packer, R.K., Garvin, J.L., 1998. Seasonal differences in activity of perch (Perca flavescens) gill Na+/K+ ATPase. Comp. Biochem. Physiol. B 120, 777–783. Palace, V.P., Evans, R.E., Wautier, K., Baron, C.L., Kollar, S., Werner, J., Holm, J., Siwik, P., Sterling, G., Johnson, C.F., 2004. An assessment of the potential for selenium to impair reproduction in bull trout (Salvelinus confluentus) from an area of active coal mining. Environ. Biol. Fishes 70, 169–174. Palace, V.P., Doebel, C., Baron, C.L., Evans, R.E., Wauiter, K.G., Klaverkamp, J.F., Werner, J., Kollar, S., 2005. Caging small bodied fish as an alternative method for environmental effects monitoring (EEM). Can. Water Qual. Res. J. 40, 328–333. Rasmussen, J.B., Gunn, J.M., Sherwood, G.D., Iles, A., Gagnon, A., Campbell, P.G.C., Hontela, A., 2008. Direct and indirect (foodweb mediated) effects of metal
ARTICLE IN PRESS 1256
L.L. Miller et al. / Ecotoxicology and Environmental Safety 72 (2009) 1249–1256
exposure on the growth of yellow perch (Perca flavescens): implications for ecological risk assessment. Hum. Ecol. Risk Assess. 14, 317–350. Rinchard, J., Kestemont, P., 2003. Liver changes releated to oocyte growth in roach, a single spawner fish, and in bleak and white bream, two multiple spawner fish. Int. Rev. Hydrobiol. 88, 68–76. Schlenk, D., Zubcov, N., Zubcov, E., 2003. Effects of salinity on the uptake, biotransformation, and toxicity of dietary seleno-L-methionine to rainbow trout. Toxicol. Sci. 75, 309–313. Shrimpton, J.M., McCormick, S.D., 1999. Responsiveness of gill Na+/K+-ATPase to cortisol is related to gill corticosteroid receptor concentration in juvenile rainbow trout. J. Exp. Biol. 202, 987–995. Sturm, A., Wogram, J., Hansen, P.-D., Liess, M., 1999. Potential use of cholinesterase in monitoring low levels of organophosphates in small streams: natural
variability in three-spined stickleback (Gasterosteus aculeatus) and relation to pollution. Environ. Toxicol. Chem. 18, 194–200. United States Environnemental Protection Agency, 2004. Draft Aquatic Life Water Quality Criteria for Selenium—2004. Office of Water, Office of Science and Technology, US Environmental Protection Agency, Washington, DC, 334pp. Velasco, J., Lloret, J., Millan, L., Marin, L., Barahona, J., Abellan, P., SanchezFernandez, D., 2006. Nutrient and particulate inputs into the Mar Menor Lagoon (SE Spain) from an intensive agricultural watershed. Water Air Soil Pollut. 176, 37–56. Whitehead, A., Anserson, S.L., Ramirez, A., Wilson, B.W., 2005. Cholinesterases in aquatic biomonitoring: assay optimization and species-specific characterization for a California native fish. Ecotoxicology 14, 597–606.