Effects of herbicides and the chytrid fungus Batrachochytrium dendrobatidis on the health of post-metamorphic northern leopard frogs (Lithobates pipiens)

Ecotoxicology and Environmental Safety 80 (2012) 372–380

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Effects of herbicides and the chytrid fungus Batrachochytrium dendrobatidis on the health of post-metamorphic northern leopard frogs (Lithobates pipiens) Linda J. Paetow a, J. Daniel McLaughlin a, Roger I. Cue b, Bruce D. Pauli c, David J. Marcogliese d,n a

Department of Biology, Concordia University, 7141 Sherbrooke St. W., Montreal, Quebec, Canada H4B 1R6 Department of Animal Science, McGill University, 21111 Lakeshore Rd., Ste. Anne-de-Bellevue, Quebec, Canada H9X 3V9 c Environment Canada, National Wildlife Research Centre, Carleton University, 1125 Colonel By Drive, Ottawa, Ontario, Canada K1A 0H3 d Fluvial Ecosystem Research Section, Aquatic Ecosystem Protection Research Division, Water Science and Technology Directorate, Science and Technology Branch, Environment Canada, St. Lawrence Centre, 105 McGill, 7th Floor, Montreal, Quebec, Canada H2Y 2E7 b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 28 October 2011 Received in revised form 2 April 2012 Accepted 3 April 2012 Available online 19 April 2012

Effects of exposure to contaminants such as pesticides along with exposure to pathogens have been listed as two major contributors to the global crisis of declining amphibian populations. These two factors have also been linked in explanations of the causes of these population declines. We conducted a combined exposure experiment to test the hypothesis that exposure to two agricultural herbicides would increase the susceptibility of post-metamorphic northern leopard frogs (Lithobates pipiens) to the amphibian fungal pathogen Batrachochytrium dendrobatidis (Bd). We assessed the independent and interactive effects of these exposures on the health and survival of the frogs. Wild-caught frogs underwent a 21-day exposure to a nominal concentration of either 2.1 mg/L atrazine (Aatrexs Liquid 480) or 100 mg a.e./L glyphosate (Roundups Original), followed by Bd, and then were observed until 94 days post-initial exposure to the herbicides. Actual levels of atrazine were between 4.28 7 0.04 mg/L and 1.70 70.26 mg/L while glyphosate degraded from 100 mg a.e./L to approximately 7 mg a.e./L within 6 days of initial exposure to the herbicides. Compared to controls, the glyphosate formulation reduced the snout-vent length of frogs during the pesticide exposure (at Day 21), and the atrazine formulation reduced gain in mass up to Day 94. No treatment affected survival, splenosomatic or hepatosomatic indices, the densities and sizes of hepatic and splenic melanomacrophage aggregates, the density and size of hepatic granulomas, proportions of circulating leucocytes, the ratio of neutrophils to lymphocytes, or the ratio of leucocytes to erythrocytes. Histological assessment of samples collected at Day 94 revealed no evidence of Bd infection in any Bd-exposed frogs, while real-time PCR detected only one case of light infection in a single atrazine- and Bd-exposed frog. Frogs exposed to Bd shed their skin significantly more frequently than Bd-unexposed frogs, which may have helped them resist or clear infection, and could explain why no interaction between the herbicides and Bd was detected. The results suggest that these frogs were resistant to Bd infection and that pre-exposure to the herbicides did not alter this resistance. The effects seen on the growth following herbicide exposure is a concern, as reduced growth can lower the reproductive success and survival of the amphibians. Crown Copyright & 2012 Published by Elsevier Inc. All rights reserved.

Keywords: Atrazine Glyphosate Amphibians Batrachochytrium dendrobatidis Chytrid fungus Lithobates pipiens

1. Introduction Amphibian populations have experienced severe declines over the past three decades and more than one-third of amphibian species are now threatened with extinction (Stuart et al., 2004). The disease chytridiomycosis, caused by infection by the fungal pathogen Batrachochytrium dendrobatidis (hereafter ‘‘Bd’’), has been

n

Corresponding author. Fax þ 1 514 496 7398. E-mail address: [email protected] (D.J. Marcogliese).

identified as a proximate cause of both amphibian population declines and species extinctions, but the factors underlying its recent emergence and varying virulence remain under investigation (Fisher et al., 2009). With respect to virulence, one hypothesis states that an interaction between environmental contaminants, such as pesticides, and Bd may contribute to susceptibility to infection (Carey et al., 1999; Parris and Baud, 2004; Davidson et al., 2007). The prevalence of pesticides in aquatic environments combined with observations that amphibians exhibit immune suppression and increased parasitism following pesticide exposure (e.g. Taylor et al., 1999; Kiesecker,

0147-6513/$ - see front matter Crown Copyright & 2012 Published by Elsevier Inc. All rights reserved. http://dx.doi.org/10.1016/j.ecoenv.2012.04.006

L.J. Paetow et al. / Ecotoxicology and Environmental Safety 80 (2012) 372–380

2002; Christin et al., 2003, 2004) suggest that environmental pesticide exposure may influence the dynamics of amphibian disease, including chytridiomycosis. Northern leopard frogs (Lithobates pipiens; formerly Rana pipiens) are exposed to pesticides and Bd throughout their range (Ouellet et al., 1997; Smith and Keinath, 2007; Tennessen et al., 2009). Although widely distributed across North America (Voordouw et al., 2010), certain northern leopard frog populations are threatened and the species’ range is contracting (Smith and Keinath, 2007; Voordouw et al., 2010); in Canada, the western populations are listed as endangered (COSEWIC, 2009). However, eastern populations of leopard frogs appear to exhibit resistance to Bd and may serve as a reservoir of infection within amphibian communities (Ouellet et al., 2005; Longcore et al., 2007; Woodhams et al., 2008). Other leopard frog populations appear more susceptible to Bd, as chytridiomycosis was the cause or suspected cause of mass die-offs (Carey et al., 1999; Green et al., 2002; Voordouw et al., 2010) and individual mortality (Adama et al., 2004). The objective of the current study was to investigate whether the resistance that some amphibians exhibit toward Bd is compromised by exposure to environmental contaminants including pesticides. Therefore, we investigated whether post-metamorphic northern leopard frogs from eastern Canada could become susceptible to clinical Bd infections following exposure to pesticides. To accomplish this, we performed a combined exposure experiment, assessed the independent and interactive effects of common pesticides and Bd on individual frog health and survival, and tested the hypothesis that exposure to the herbicides would increase susceptibility to infection by Bd. The frogs were exposed to sublethal concentrations of the herbicides Aatrexs Liquid 480 or Roundups Original, then Bd, and subsequently observed until 94 days post-initial exposure to the herbicides. The Aatrexs formulation contains the active ingredient atrazine, which has been linked to depressed immune systems in amphibians (Christin et al., 2003, 2004; Forson and Storfer, 2006; Rohr et al., 2008a, 2008b). The Roundups formulation contains the active ingredient glyphosate and the surfactant polyethoxylated tallowamine (POEA) (Giesy et al., 2000), both of which have been widely investigated for their toxic effects in amphibians (e.g. see review in Govindarajulu, 2008) but less so for their potential immunomodulating effects (Rohr et al., 2008a). We evaluated survival, growth (gain in mass and snout-vent length), measures of condition (hepatosomatic index, densities and mean sizes of hepatic and splenic melanomacrophage aggregates, density and mean size of hepatic granulomas, and neutrophil/lymphocyte ratios), markers of immune function (splenosomatic index, differential leucocyte counts, ratio of leucocytes to erythrocytes), and susceptibility to Bd using standard molecular and histological techniques (Hyatt et al., 2007).

2. Materials and methods 2.1. Ethics statement This study was conducted in accordance with national and institutional guidelines for the protection of animal welfare, as outlined by the Canadian Council on Animal Care (CCAC, 1993) and the Aquatic Ecosystem Protection Research Division Animal Care Committee (St. Lawrence Centre, Environment Canada). 2.2. Animals In August 2007, recently-metamorphosed northern leopard frogs were captured on land, near a pond located in a wildlife reserve near Boucherville, Quebec (451380 0600 N 731260 0600 W). Immediately after their capture, a skin swab was taken from each of the 20 randomly selected individuals to test for infection with Bd using methods described by Kriger et al. (2006). Each swab was placed in a

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sterilised Eppendorf tube and sent within 24 h to the British Columbia Ministry of Agriculture’s Animal Health Centre (Abbotsford, BC, Canada), where all swabs were tested for Bd using a real-time PCR assay designed to target the ITS1 gene of the fungus (Julie Bidulka, personal communication). Water samples taken from the pond were analysed for 39 pesticides commonly used in eastern Canadian agriculture, including atrazine and glyphosate. Pesticide analyses were conducted at the Centre d’expertise en analyse environnementale du Que´bec (CEAEQ, Quebec City, QC, Canada). The frogs were transported to the laboratory where they were isolated in individual 1.84 L plastic containers. The containers held 50 mL of dechlorinated, UV-treated water, and two platforms that permitted the frogs to exit the water. The frogs were acclimated to a 16 h:8 h light:dark photoperiod and a room temperature of 217 1.3 1C over a period of eight weeks. During the experiment, each frog was housed in a 1 L mason jar containing 50 mL of the appropriate treatment liquid for the corresponding phase of the study (see below). The mouth of each jar was covered with fibreglass screening secured with the jar’s screw-on ring to prevent escape. With the exception of the period during which the frogs were exposed to Bd, the jars were placed on a slant to allow the frogs to enter and leave the liquid at will, given the semi-terrestrial nature of the species post-metamorphosis. During the acclimatisation and experimental periods, the water or exposure solution in each jar was changed daily, as it rapidly became fouled. The location of individual jars was randomly interchanged on a weekly basis by blindly drawing labels out of a container, to minimise effects of uneven lighting or temperature within the laboratory. The frogs were fed one to two crickets (Mirdo Importations Canada, Inc., Montreal, QC, Canada) every 2 days. Once a week, the crickets were pre-dusted with phosphorus-free CaCO3 powder (Mirdo Importations Canada, Inc., Montreal, QC, Canada) to prevent nutrient deficiencies in the frogs. At various stages of the experiment, frogs were killed by immersion in a solution of 0.8% buffered tricaine-methanylsulfonate (MS-222) in distilled water (Syndel Laboratories Ltd, Vancouver, BC, Canada).

2.3. Chemicals and reagents Aatrexs Liquid 480, containing the active ingredient atrazine, was purchased from Syngenta Crop Protection Canada Inc. (Guelph, ON, Canada). Roundups Original, containing the active ingredient glyphosate, was purchased from Monsanto Canada Ltd. (Winnipeg, MB, Canada). All reagents used in the study were purchased from Fisher Scientific (Ottawa, ON, Canada) unless otherwise noted.

2.4. Aatrexs Liquid 480 and Roundups Original exposures A time-frame of 21 days was selected for the chemical exposures to mimic realistic exposure times in southwestern Quebec (Giroux et al., 2006). A concentration of 2.1 mg/L was selected for atrazine because it reflected concentrations measured in summer surface waters within the same region (Gendron et al., 2003). A concentration of 100 mg a.e./L (a.e. ¼ acid equivalents of glyphosate) was selected for glyphosate, based on average and maximum concentrations (600 and 1800 mg a.e./L, respectively) reported in a widely-cited ecotoxicological risk assessment study (Giesy et al., 2000). Although this was well above concentrations measured in surface waters of the study area (o 0.8 mg a.e./L, Giroux et al., 2006), concentrations of glyphosate in surface waters, including vernal pools, streams and forest wetlands, which are important amphibian habitats, can reach levels as high as 1950 mg a.e./L (Feng et al., 1990; Giesy et al., 2000; Thompson et al., 2004; Scribner et al., 2007; Battaglin et al., 2009). One day prior to the start of the experiment, two chemical stock solutions were prepared, each at 1000  the target exposure concentrations for atrazine and glyphosate, by dissolving the water-soluble end-use herbicide formulations Aatrexs Liquid 480 and Roundups Original, respectively, in deionized water. Stock solutions were stored at 4 1C in sealed 2 L amber flasks that were held in light-proof boxes to prevent photodegradation. During the pesticide exposure phase of the experiment (Days 1–21), aliquots of the stock solutions were diluted daily with dechlorinated water, and used to replace the pesticide solutions in jars holding the frogs being exposed. At Day 20 of the experiment, two samples of freshly diluted stock solution of each herbicide were collected. These samples were later analysed to determine the initial herbicide concentration in the exposure solutions used for the daily exposures during the pesticide exposure phase. At Day 21 of the experiment, immediately following the final exposure of the frogs to the herbicides, water samples from six randomly-chosen exposure vessels per pesticide group were collected for analysis. The atrazine samples (the Day 20 freshly-diluted stock solution representing the exposure solution, along with the Day 21 water samples from the exposure vessels) were immediately shipped to the National Laboratory for Environmental Testing (NLET, Environment Canada, Burlington, ON, Canada) for analysis. The corresponding glyphosate samples were stored in sealed amber bottles and refrigerated at 4 1C for three months prior to being sent for analysis by AXYS Analytical Services Ltd. (Sidney, BC, Canada). The original concentrated stock solutions were not analysed.

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2.5. Maintenance of pathogen culture and pathogen exposures We used strain JEL423 of Bd, which was isolated during an amphibian die-off in Panama and provided by Dr. J.E. Longcore (University of Maine, Orono, Maine, USA). We confirmed its viability by successfully infecting an American bullfrog (Lithobates catesbeianus; formerly Rana catesbeiana) tadpole during a pilot experiment conducted during 2007. The culture was maintained in 1% tryptone broth at 4 1C and subcultured every three months. Streptomycin–penicillin (Sigma-Aldrich Canada Ltd., Oakville, ON, Canada) was added to each subculture to prevent bacterial growth (methods followed a protocol provided by Dr. L. Rollins-Smith, Vanderbilt University, Nashville, TN, USA). During the course of the experiment, the frogs were exposed to Bd on four separate occasions to increase the potential for infection. Bd inocula were prepared for each exposure by culturing Bd on 1% tryptone agar plates at 23 1C for seven days, flooding the plates with distilled water to release zoospores, and decanting the plates into a single beaker to obtain a suspension of distilled water and zoospores. The concentration of the active zoospores in suspension was then estimated using a hemocytometer. Following this, a 50 mL aliquot of the suspension was added to the container water of each frog selected for Bd exposure. Control frogs were exposed to sham inocula that consisted of distilled water rinses of Bd-free tryptone agar plates. Following inoculation, all jars were left upright for 24 h to ensure contact between the frogs and zoospores. Then, the liquid in each jar was replaced with clean dechlorinated water and the jars were returned to their tilted position so that the frogs could exit the water. The concentrations of zoospores for the four exposures were estimated to be 79,600 zoospores/mL, 86,400 zoospores/mL, 130,000 zoospores/mL and 67,200 zoospores/mL.

2.6. Experimental design We used recently metamorphosed post-metamorphic frogs for this experiment because the clinical symptoms associated with infection by Bd are more serious in amphibians at this life stage (Gahl et al., 2011) and because leopard frogs may be exposed to atrazine, glyphosate and Bd at this age (Ouellet et al., 2005; Giroux et al., 2006). There were two phases to the experiment. In the first phase, the frogs were exposed to one of the two herbicides. In the second phase, some of these frogs (see numbers below) were subsequently exposed to Bd zoospores. Initially, 190 frogs were weighed (mean7 SD¼ 4.41 71.13 g) and measured for snout-vent length (SVL¼ 38.05 7 2.98 mm). Individual frogs were then randomly assigned to one of three groups as follows: one group (n¼64) was exposed to a nominal concentration of 2.1 mg/L atrazine, a second group was exposed to a nominal concentration of 100 mg a.e./L glyphosate (n ¼62), and the third group (n ¼64) consisted of untreated controls that received plain dechlorinated water in the place of a chemical exposure solution. The pesticide-treated frogs were exposed to the herbicides for 21 days, as described in Section 2.4. On Day 21, at the termination of the pesticide exposures, the frogs were again weighed and measured. The frogs in each group were then randomly assigned to one of three subgroups consisting of 20–22 frogs per group, producing three subgroups of frogs that had been exposed to atrazine, three that had been exposed to glyphosate, and three control subgroups. Frogs in one atrazine-exposed subgroup, one glyphosate-exposed subgroup, and one control subgroup, selected at random, were killed and necropsied. Frogs in the remaining two subgroups were returned to their jars and held in dechlorinated water in preparation for the second phase of the study. On Day 22, one subgroup from each pair of the remaining treatment groups was selected at random for Bd exposure; the other served as the control. On Day 22 and every eighth day of the next 25-day period, the frogs were exposed to Bd for a 24 h period, as described in Section 2.5. Frogs were then held until 94 days post-initial exposure to the pesticides, which was also 74 days post-initial exposure to Bd, longer than the 60-day period required for all frogs of a more susceptible species (Mixophyes fasciolatus) to die when they were experimentally exposed to Bd under similar conditions (i.e. a room temperature of 17 or 23 1C) (Berger et al., 2004). At Day 94, the treated and control frogs were weighed, measured, killed and necropsied, and samples for effects measures were taken.

2.7. Analytical methods Data gathered from frogs necropsied immediately after the pesticide exposures (Day 21) were used to assess short term effects of the pesticides on survival, growth, hepatosomatic index (HSI) and splenosomatic index (SSI). Growth was measured both as a change in mass and in SVL. Data collected from frogs necropsied on Day 94 were used to assess longer term effects of the pesticides, the effects of exposure to Bd, and their interaction, on survival, growth and various biomarkers of frog health. The measured biomarkers of health included HSI, densities and mean sizes of hepatic and splenic melanomacrophage aggregates (HMAs and SMAs, respectively), density and mean size of hepatic granulomas (HGAs), the neutrophil to lymphocyte ratio (N/L ratio), and immune function measures including SSI, differential leucocyte counts, and the ratio of leucocytes to erythrocytes (WBC/RBC).

To obtain these measures, all frogs were first examined for gross pathology with a dissecting microscope. The spleen and liver were removed, rinsed with distilled water, blotted dry and weighed to the nearest 0.0001 g. Fresh organ weights were divided by the total body weight and used to obtain the HSI and SSI, which are commonly used to assess adverse impacts of contaminants in fish (Schlenk et al., 2008). Three blood smears prepared from each frog killed at Day 94 were air-dried, fixed and stained using the Fisher Diagnostics PROTOCOL Hema 3s Manual Staining System. A total of 100 leucocytes per smear were identified, based on descriptions in Rouf (1969). The differential leucocyte counts and ratio of WBC/ RBC were compared, as was the N/L ratio, which served as an indicator of stress (Davis et al., 2008). Spleens and livers from all the frogs killed at Day 94 were fixed in 10% neutral buffered formalin and submitted to the McGill Bone Centre (McGill University, Montreal, QC, Canada) for histological processing. They were sectioned at 5 mm, mounted on slides, and stained with hematoxylin and eosin. Slides from 60 randomly-selected frogs (10 in each of the six subgroups) were examined microscopically to count and measure HMAs, SMAs and HGAs, which can be affected by exposure to contaminants (Agius and Roberts, 2003; Linzey et al., 2003; Thomas, 2008; Froese et al., 2009). Counts were performed using 100  magnification with a 100 mm2 grid placed over one section of the slide while measurements were taken under 200  magnification. Because HMAs were common, only those observed within five randomly-selected grid squares were counted. Because HGAs were rare and spleens were small, HGAs and SMAs in every square of the grid that completely fit over part of the respective organ were counted. The counts of HMAs and SMAs were converted to density per mm2 while the data on HGA counts were treated as described below. The long toe from the right hind foot of each frog killed at Day 94 was fixed in 10% formalin and submitted to the McGill Bone Centre for histological processing. Following decalcification, all of these toes were sectioned at 5 mm along the longitudinal axis, stained with hematoxylin and eosin and examined microscopically for Bd and hyperkeratosis. Toe samples were used to confirm infection in Bd-exposed frogs and to ensure that frogs that were not experimentally exposed to Bd were not infected. The long toe from the left hind foot was severed, fixed in 70% ethanol and sent to Trent University (Peterborough, ON, Canada) for diagnosis and quantification of Bd infection by real-time PCR (Boyle et al., 2004). Toes from 45 randomly selected frogs were examined molecularly (10 Bd-exposed and 5 Bd-unexposed frogs from each of the two pesticide-exposed groups and the control group). Full body swabs were also taken at several time-points of the study to both confirm infection by Bd and monitor the evolution of infection, but problems at the outside laboratory prevented them from being analysed. The frequency with which frogs shed their skin (ecdysis) during the four-week Bd exposure period was recorded and used as an additional indicator of infection by Bd, because ecdysis may occur in frogs to prevent or clear infection by Bd (Woodhams et al., 2007a). As animals were housed individually, we were able to record the presence of shed skin in the water during daily water changes for each animal. Several shed skins of both Bd-exposed and Bd-unexposed frogs were examined microscopically to look for evidence of infection by Bd, but none was found in any frog using this method.

2.8. Statistics Analyses were performed using SASs version 9.1 (SAS Institute Inc., 2004). Kaplan–Meier analyses and the log-rank statistic were used to compare survival among treatment groups. One-way ANOVAs were used to compare the initial size of frogs assigned to the three pesticide groups (including controls) at the start of the experiment, and growth, HSI and SSI measured at Day 21, following the herbicide exposures. Two-way factorial ANOVAs (3  2) were used to compare growth from the onset of Bd exposure until the end of the experiment (Days 21–94); growth over the course of both exposures (Days 1–94); along with HSI, SSI, the densities and sizes of HMAs and SMAs, and the sizes of HGAs that were measured at Day 94. Where necessary, these data were transformed to meet the assumptions of ANOVAs. Scheffe´’s tests were used when multiple comparisons were performed. Between one and eight individuals were excluded from some analyses (growth, HSI and SSI) due to errors during data collection or because they (n ¼2) were intersex individuals (i.e. possessed both male and female gonads). Data describing the density of HGAs were zero-inflated, and were therefore analysed using the GENMOD procedure available in SAS. GENMOD fits a generalised linear model (e.g. logistic regression or other) with a link function to data, and the distribution can be other than normal (SAS Institute Inc., 2004). The negative binomial distribution with a log link function provided the best fit to the data set, based on the deviance/df ratio, which approaches 1 as goodness of fit improves. Due to variation among individuals, the number of grid squares in which hepatic granulomas were counted was set as an offset term in the analysis, allowing us to model the dependent variable as a rate model (count/grid square). Due to overdispersion in the data (i.e. deviance/df41), a scaling option was implemented (i.e. scale¼ deviance), thereby adjusting the covariance matrix and likelihood function by the scale parameter, which was fixed at a value of 1. GENMOD and the same scaling option were similarly used to compare the blood cell ratios and

L.J. Paetow et al. / Ecotoxicology and Environmental Safety 80 (2012) 372–380 frequency of epidermal shedding between Bd-exposed and Bd-unexposed frogs during the Bd exposures. However, because these data sets were recorded as proportions, the binomial distribution was fitted with a log link function, as above.

3. Results 3.1. Collection site Only glyphosate and its primary degradation product AMPA were detected in the water samples collected at the frogs’ natal pond on the wildlife preserve; the concentrations were 0.091 mg a.e./L and 1.2 mg/L, respectively. Atrazine was not detected. None of the 20 frogs swabbed upon capture tested positive for Bd infection.

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(at approximately 7 mg a.e./L) in the exposure vessels between Days 20 and 21. Finally, we can state that the mock exposure results indicate that the experimental animals were likely exposed to a glyphosate concentration near our 100 mg a.e./L nominal concentration for under six days, and then to approximately 7 mg a.e./L until Day 21. The glyphosate exposures were environmentally relevant because the actual glyphosate concentrations in the exposure vessels were within levels measured in the environment, and glyphosate reportedly dissipates relatively rapidly from natural aquatic environments as it binds to particles and sediment and is broken down by microbial activity (Giesy et al., 2000). In addition, the exposures reflected concentrations recently measured in surface waters in southwestern Quebec ( r0.8 mg a.e./L; Giroux et al., 2006).

3.2. Pesticide exposure concentrations

3.3. Aatrexs Liquid 480 and Roundups Original exposures

The concentrations of atrazine in the two samples of exposure solution (stock solution diluted 1:1000) prepared on Day 20 were 4.30 and 4.25 mg/L (mean7SD¼4.28 70.04 mg/L). The concentrations of atrazine in the water samples collected from the exposure vessels on Day 21 ranged from 1.30 to 2.03 mg/L (meanþSD¼ 1.7070.26 mg/L). These results demonstrate that the frogs were exposed to a level of atrazine that was close to our nominal concentration of 2.1 mg/L. They also indicate that the active ingredient was dissipating at an approximate rate of 50% over 24 h between water changes. The concentrations of glyphosate measured in the two samples of exposure solution prepared at Day 20 (and kept in refrigerated storage for 3 months before analysis) were 3.15 and 4.50 mg a.e./L (mean7SD ¼3.8370.95 mg a.e./L). The glyphosate concentrations in the samples collected from the exposure vessels on Day 21 (and similarly stored) ranged between 6.86 and 9.84 mg a.e./L (mean7SD ¼8.1371.27 mg a.e./L). Because our nominal glyphosate concentration was 100 mg a.e./L, we suspected that our solutions had degraded during the three-month interim preceding analysis due to improper storage (they should have been frozen rather than refrigerated). This prompted us to verify our protocols for solution preparation and storage. As in the original experiment, we prepared a new glyphosate stock solution for storage at 4 1C in an amber bottle. From this solution we immediately prepared one sample of exposure solution that was frozen at 20 1C (rather than refrigerated at 4 1C) before shipment to the original laboratory for analysis. Analysis of this sample revealed a concentration of 92.6 mg a.e./L glyphosate, confirming that our method for preparing the original stock solution had been adequate. The new stock solution was then refrigerated for a further six days and a second sample was taken and frozen. Finally, a third sample was held in a tilted mason jar (minus the frog) for 24 h to simulate the original exposure conditions prior to its being taken and frozen. Upon analysis the second and third samples of exposure solution contained 7.29 and 7.00 mg a.e./L of glyphosate, respectively. These results indicate that the glyphosate concentration in the original stock solution was reduced by one order of magnitude during six days of refrigerated storage, and this likely occurred because the water used to prepare the stock solutions had not been pre-sterilised. Further, the glyphosate contained in the samples of exposure solution that we collected on Days 20 and 21 of the original experiment likely underwent additional microbial degradation during the three-month period of refrigerated storage prior to analysis. Therefore, we consider the analyses obtained from Days 20 and 21 of the original experiment to be unreliable. Analysis of the samples collected during the mock exposure reliably demonstrate that the concentration of glyphosate was likely stable

There was no significant difference in mean mass or SVL among the treatment and control groups at Day 1 (Appendix A, Table A.1). Survival to Day 21 ranged from 96% to 100% in the three treatment groups (0–2 deaths per group) (Table A.2) and was unaffected by pesticide exposure (Kaplan–Meier test, logrank statistic: x2 ¼2.00, df ¼2, p ¼0.37). The mass gain of the surviving frogs in these groups was also unaffected by pesticide exposures. However, there was a significant overall effect of pesticides on change in SVL (ANOVA: n ¼184; F2,181 ¼4.63, p¼0.01) (Tables 1 and A.3). SVL in frogs exposed to glyphosate was significantly less than in control frogs (Scheffe´’s test: p¼0.02), but there was no difference in the SVL of atrazineexposed versus control frogs (Scheffe´’s test: p ¼0.84) (Fig. 1). Neither the HSI nor SSI of frogs was affected by pesticide exposure (Table A.3).

Table 1 Mean (7 SD) changes in snout-vent length (SVL) and body mass of northern leopard frogs (Lithobates pipiens) that underwent a 21-day exposure to either a nominal concentration of 2.1 mg/L atrazine contained in Aatrexs Liquid 480 (Group A) or 100 mg a.e./L glyphosate contained in Roundups Original (Group G)a, or that remained pesticide-free (Group C), and that were subsequently exposed or not to Batrachochytrium dendrobatidis (Bd). Treatment groups

n

Following exposure to the A 61 G 59 C 64

Change in SVL (mm)

n

pesticides (Days 1–21) 2.68 71.36ab 60 2.12 71.44b 57 a 2.83 71.26 62

Change in body mass (g)

0.927 0.47 0.957 0.46 1.047 0.38

Following exposure to Bd (Days 22–94) A 21 8.50 72.70 A þ Bd 20 8.81 71.56 G 19 8.63 71.81 Gþ Bd 18 8.38 72.37 C 20 8.14 72.33 C þ Bd 22 8.92 71.34

21 20 19 18 20 21

3.867 0.83 4.257 0.66 4.387 0.91 3.957 0.98 4.437 0.94 4.297 0.65

Following both exposures (Days 1–94) A 21 10.92 73.10 A þ Bd 19 11.35 71.76 G 19 10.85 72.93 Gþ Bd 18 10.42 73.04 C 20 11.26 72.73 C þ Bd 23 10.93 72.76

21 20 20 18 20 22

4.807 0.94n 5.027 0.63n 5.127 1.08 4.997 1.10 5.397 1.09 5.477 0.82

Different superscripted letters identify significantly different mean SVLs (one-way ANOVA, Scheffe´’s test, p o 0.05). a Actual levels of atrazine were between 4.28 70.04 mg/L and 1.707 0.26 mg/L while glyphosate degraded from 100 mg a.e./L to approximately 7 mg a.e./L within 6 days of initial exposure to the herbicides. n Indicates a significant effect of atrazine on mass gain (p o0.05); see text and Appendix A (Table A.4) for details.

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Fig. 1. Mean (7 SE) gain in snout-vent length (SVL) of northern leopard frogs (Lithobates pipiens) that underwent a 21-day exposure to a nominal concentration of either 2.1 mg/L atrazine (Aatrexs Liquid 480) or 100 mg a.e./L glyphosate (Roundups Original). Actual levels of atrazine were between 4.28 7 0.04 mg/L and 1.70 7 0.26 mg/L while glyphosate degraded from 100 mg a.e./L to approximately 7 mg a.e./L within 6 days of initial exposure to the herbicides. Different letters indicate significant differences among the treatment groups (one-way ANOVA, Scheffe´’s test, po 0.05).

Fig. 2. Mean (7 SE) gain in body mass over 94 days of northern leopard frogs (Lithobates pipiens) that underwent a 21-day exposure to a nominal concentration of either 2.1 mg/L atrazine (Aatrexs Liquid 480) or 100 mg a.e./L glyphosate (Roundups Original), followed by Batrachochytrium dendrobatidis (Bd). Actual levels of atrazine were between 4.28 7 0.04 mg/L and 1.70 7 0.26 mg/L while glyphosate degraded from 100 mg a.e./L to approximately 7 mg a.e./L within 6 days of initial exposure to the herbicides. Different letters indicate significant differences among treatment groups (two-way ANOVA, Scheffe´’s test, po 0.05). Frogs are pooled across Bd exposure treatments.

3.4. Combined pesticide and pathogen exposures 3.4.1. Survival and growth Survival following Bd exposure (Days 22–94) and over the course of the entire experiment (Days 1–94) ranged from 90% to 100% (i.e. mortality of 0–2 individuals per group) (Table A.2). Survival was not significantly different among the treatment groups (Kaplan–Meier test, log-rank statistic: Bd exposure x2 ¼ 3.79, df ¼5, p ¼0.58; overall x2 ¼2.66, df ¼5, p¼0.75). There was no significant effect of pesticides, Bd, or their interaction on growth following Bd exposure (Days 22–94). However, pesticide exposure had a significant effect on mass gain, but not SVL, over the course of both exposures (Days 1–94) (ANOVA: n ¼121; pesticide F2,115 ¼3.91, p ¼0.02; Bd F1,115 ¼0.02, p ¼0.90; pesticidenBd F2,115 ¼0.25, p ¼0.78) (Tables 1 and A.4). When pooled across the Bd exposure treatments, frogs exposed to atrazine gained significantly less mass than did the control frogs between Days 1 and 94 (Scheffe´’s test: p¼0.03) (Fig. 2). Frogs exposed to glyphosate also gained less mass than control frogs, but the difference was not significant (Scheffe´’s test: p ¼0.21) (Fig. 2). 3.4.2. Biomarkers of animal health There was no significant effect of pesticides, Bd, or their interaction on HSI (Tables A.5 and A.11), density and mean size of HMAs (Tables A.6 and A.11), density and mean size of HGAs (Tables A.7, A.11 and A.12), density and mean size of SMAs (Tables A.8 and A.11), or the N/L ratio (Tables A.10 and A.12) in frogs that survived to the end of the experiment (Day 94). Among the 60 frogs examined for HGAs, 18 (approximately 30%) had none, and were therefore excluded from the analysis comparing mean size of HGAs (Tables A.7 and A.11); another frog was excluded from both analyses comparing HGAs as an obvious outlier (Tables A.7, A.11 and A.12). 3.4.3. Biomarkers of immune function There was no significant effect of pesticides, Bd, or their interaction on SSI (Tables A.5 and A.13). Blood cell ratios were almost always affected by the pesticide exposures, however, and were occasionally affected by Bd exposure (Tables A.9, A.10 and A.14). Minor to high overdispersion occurred among all haematological data, as indicated by the deviance/df ratios (Table A.14).

After implementation of the scaling option in GENMOD, none of the experimental treatments showed a significant effect on measured haematological values (Tables A.9, A.10 and A.14).

3.4.4. Indicators of infection and disease None of the toe samples examined microscopically showed evidence of zoosporangia or hyperkeratosis, which are indicative of Bd infection. Real-time PCR analysis of one toe sample obtained from a frog exposed to both atrazine and Bd tested positive for Bd infection, although with a low DNA copy number (1.6 DNA copies). The rate of ecdysis during the Bd exposure period was not affected by earlier exposure to the pesticides, or by an interaction between the pesticide and Bd exposures. However, ecdysis was significantly affected by Bd exposure alone (logistic regression: pesticide x22,117 ¼2.01, p¼0.37; Bd x21,117 ¼33.00, po0.0001; pesticidenBd x22,117 ¼0.86, p¼0.65). Bd-exposed frogs shed on average 1.0371.03 times during the four-week Bd exposure period, while those not exposed to Bd shed an average of 0.1970.54 times (Fig. 3a). Most of the frogs that failed to shed skin had not been exposed to Bd, whereas most of the frogs that shed their skin had been exposed (Fig. 3b).

4. Discussion There is evidence that exposure to pesticides, including atrazine, can influence infection by certain types of pathogens (e.g. viruses or parasites) in amphibians (Kiesecker, 2002; Forson and Storfer, 2006; Koprivnikar et al., 2007; Rohr et al., 2008a; Koprivnikar, 2010). However, the effect of pesticide exposure on infection by Bd in amphibians is not well known. Our results suggest that the leopard frogs we studied exhibit resistance to Bd infection and that this resistance was not diminished by exposure to atrazine- or glyphosate-based herbicides at the concentrations tested. We also found no evidence that the independent or combined pesticide and Bd exposures influenced survival, differential leucocyte counts, the WBC/RBC ratio, the density and number of hepatic and splenic melanomacrophage aggregates and granulomas, or hepatosomatic and splenosomatic indices in metamorphosed northern leopard frogs. A low susceptibility to infection by Bd in the population of frogs we sampled could

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Fig. 3. (a) Mean (7 SE) number of times northern leopard frogs (Lithobates pipiens) pre-exposed to a nominal concentration of either 2.1 mg/L atrazine (Aatrexs Liquid 480) or 100 mg a.e./L glyphosate (Roundups Original), or untreated controls, shed their skin during the time (4 X 24 h over 25 days) they were subsequently exposed to Batrachochytrium dendrobatidis (Bd) (p o 0.0001). (b) Number of northern leopard frogs that shed their skin 0, 1, 2 or 3 times during exposure to Bd. Frogs shown in (a) and (b) are pooled across pesticide treatments. Actual levels of atrazine were between 4.28 7 0.04 mg/L and 1.70 7 0.26 mg/L while glyphosate degraded from 100 mg a.e./L to approximately 7 mg a.e./L within 6 days of initial exposure to the herbicides.

explain why we did not measure significant interactive effects between the herbicide and fungal exposures. Three earlier studies have examined whether exposure to pesticides can increase the susceptibility of post-metamorphic frogs to chytrid infection (Davidson et al., 2007; Gahl et al., 2011; Buck et al., 2012). In one, recently-metamorphosed yellow-legged frogs (Rana boylii) were exposed to carbaryl (an insecticide) followed by Bd (Davidson et al., 2007). In another, three assemblages of pacific treefrogs (Pseudacris regilla) and Cascades frogs (Rana cascadae) were simultaneously exposed to carbaryl and Bd during larval development and then monitored beyond metamorphosis (Buck et al., 2012). In the third study, which employed moderate to high exposure concentrations of glyphosate (210 and 2890 mg a.e./L) (Giesy et al., 2000; Struger et al., 2008), wood frogs (Lithobates sylvaticus) were simultaneously exposed to glyphosate and Bd during larval development and then monitored beyond metamorphosis (Gahl et al., 2011). Based on infection and/or survival rates, none of these studies found any indication that the pesticides altered susceptibility to Bd in post-metamorphic frogs. Taken together, these earlier studies and the present one do not provide support for a hypothesis that environmental exposure to pesticides is contributing to the chytridiomycosis epidemic (Gahl et al., 2011). On the other hand, the pesticides used in the four

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exposure studies may not interfere with immune defenses that protect the species of frogs tested against Bd. The use of different pesticides and/or a different amphibian species in similar experiments may produce different results. Alternatively, the pesticide doses used in the four studies may not have been sufficient to suppress host immune defenses below a critical threshold that increases susceptibility to Bd (Davidson et al., 2007). Exposure of amphibians to higher concentrations of atrazine in particular (between 30 and 400 mg/L) has been linked with increased susceptibility to infection by trematodes and a down-regulation of genes associated with immunity (Kiesecker, 2002; Koprivnikar et al., 2007; Rohr et al., 2008a; Langerveld et al., 2009). It would therefore be important to test whether higher doses of atrazine also alter susceptibility to Bd. Moreover, it would be prudent to test a range of doses of atrazine, as its immunosuppressive effects in amphibians are not always dose-dependent (Forson and Storfer, 2006). We detected significant independent effects of the pesticide exposures on the growth of individual animals. Numerous studies have shown that sublethal levels of atrazine- and glyphosatebased herbicides can affect growth in larval amphibians (Diana et al., 2000; Boone and James, 2003; Brown Sullivan and Spence, 2003; Howe et al., 2004; Rohr et al., 2004; Hayes et al., 2006; Floyd et al., 2008; Relyea, 2009; Koprivnikar, 2010; Rohr and McCoy, 2010). Atrazine, for instance, has been found to reduce size at metamorphosis in most amphibian species tested (Rohr and McCoy, 2010). Proposed mechanisms for this effect include endocrine disruption, a reduction in food intake, an increase in metabolic rate, or a decrease in food conversion efficiency following a reallocation of energy towards processes such as detoxification (Langerveld et al., 2009; Koprivnikar, 2010). The results of the present study suggest that exposure to atrazine- and glyphosate-based herbicides slow growth rates in post-metamorphic frogs. However, given that the mass gain in particular of the atrazine-exposed frogs was reduced over the longer term while their gain in SVL was unaffected, it is possible that atrazine caused the experimental frogs to become dehydrated. This was observed in an earlier exposure study involving atrazine and streamside salamanders (Ambystoma barbouri) (Rohr and Palmer, 2005). In either case, these findings are worrisome as smaller metamorphs have a lower survival rate, reduced reproductive success and a compromised immune system (Rohr and McCoy, 2010), while amphibians in general are especially susceptible to desiccation (Wells, 2007; Rohr and McCoy, 2010). Growth of frogs exposed to Roundups Original was reduced compared to control frogs at the end of the pesticide exposure. Roundups Original, like most glyphosate-based herbicides, contains a surfactant to facilitate its penetration into plant tissues (Mann et al., 2009). In the case of Roundups Original, this surfactant is the polyethoxylated tallowamine POEA. Exposure studies on a variety of amphibian species have found that POEA rather than the glyphosate active ingredient was the compound responsible for a range of observed toxic effects that included delayed growth (Mann et al., 2009). We cannot rule out POEA as the causative agent for delayed growth in the present case, as we did not include separate exposures to the individual components of the herbicide in our experimental design. Future exposure studies should test effects of both glyphosate and POEA separately (Mann et al., 2009). Traces of glyphosate and its primary breakdown product AMPA were found in samples of the natal pond water. Therefore the leopard frogs may have been exposed to formulated products containing glyphosate, and perhaps POEA, at the tadpole stage, although at lower concentrations. We have no data to determine whether the effects of Roundups Original that we observed on the growth of metamorphs in the laboratory were in addition to effects of potential field exposures at the tadpole stage, prior to animal collection. However, our control post-metamorphic animals would

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also have been exposed to glyphosate as tadpoles in the field, so the growth difference between control and herbicide-exposed postmetamorphic frogs remains a significant finding. None of the frogs tested positive for Bd prior to commencement of the experiment. With these results, we could not confirm that the source population was uninfected. Assuming a large population of leopard frogs in the area (e.g. 10,000), and given the fact that only 20 frogs were initially tested for Bd, a prevalence of infection as high as 13.9% could have gone undetected in this case (Cannon and Roe, 1982). In southwestern Quebec, prevalence of infection by Bd may fall below this level in amphibian populations during the summer (Ouellet et al., 2005). Thus, although our results may reflect an absence of Bd infection in this population, it may also reflect insufficient sampling effort or seasonal infection patterns. As with post-metamorphic foothill yellow-legged frogs experimentally exposed to Bd (Davidson et al., 2007), here laboratory exposures to the fungus did not reduce frog survival, and only a single Bd-exposed frog subsequently tested positive for chytrid infection. In contrast to the earlier study, however, we did not find that exposure to Bd slowed the growth of frogs. The different responses might be explained by a more energy-costly immune response among the yellow-legged frogs compared to the northern leopard frogs. The other indicators of general and immune health that we examined were not affected by the frogs’ exposure to the pesticides, the fungus, or by their interaction. In fish, melanomacrophage aggregates (MAs) typically increase in number and size following exposure to contaminants and infectious agents (viruses, bacteria and parasites), and this generalised tissue response to environmental perturbations is often used as a biomarker of degrading health (Thilakaratne et al., 2007; Borucinska et al., 2009). Amphibian tissues may respond similarly, as the livers of marine toads (Bufo marinus) collected in sites contaminated with pesticides and heavy metals also exhibited an increased number of melanomacrophages compared to tissues of toads collected at control sites (Linzey et al., 2003). However, Rohr et al. (2008a) measured a reduced abundance of hepatic melanomacrophage aggregates in northern leopard frogs exposed to atrazine. Moreover, this response was correlated with increased infections by larval trematodes, prompting the suggestion that MAs may form part of an important immune response to trematode infections, and that atrazine may suppress this response (Rohr et al., 2008a). Correlations between MA responses and environmental perturbations need to be more fully evaluated in amphibians, as it is unclear why exposure to atrazine did not affect MAs in the present study. Possibly, it was because MAs were only measured more than two months after the pesticide exposures were terminated, allowing sufficient time to regain homoeostasis following any changes. We recorded more frequent ecdysis in the northern leopard frogs exposed to Bd compared to those that were not exposed to the fungus. Given that Bd infections are localised in the skin and other keratinised surfaces of post-metamorphic amphibians, accelerated ecdysis may be a mechanism by which hosts reduce or clear infection (Davidson et al., 2003; Woodhams et al., 2007a). To our knowledge, this is the first report of increased epidermal shedding in leopard frogs. We cannot rule out that unmonitored aspects of immunity (e.g., antimicrobial peptides; Rollins-Smith and Conlon, 2005) contributed to the lack of chytridiomycosis among our frogs. However, whatever aspects of immunity were involved, our results indicate that preexposure to the pesticide formulations at the doses used did not significantly interfere with ecdysis. Altered leucocyte profiles (i.e. the relative proportions of different circulating lymphocytes) following infection by Bd suggests that certain leucocytes migrate to foci of epidermal degeneration (Woodhams et al., 2007b), and/or that there is significant stress associated with the fungal exposures (Blaustein et al., 2009; Davis

et al., 2010). Exposure to atrazine can alter leucocyte profiles (Kiesecker, 2002) through direct immunosuppressive action and/or via an associated stress response (Forson and Storfer, 2006). In the present study, we did not detect significant changes in leucocyte profiles following exposure to the pesticides, Bd or their combination, suggesting that there was no direct suppression or activation of this immune component, and no associated stress response at the time of measurement. Alterations in leucocyte profiles of wildcaught leopard frog metamorphs could not be attributed to atrazine exposure in another recent study (Shutler and Marcogliese, 2011).

5. Conclusion Our results suggest that exposure to two common herbicides prior to Bd exposures in post-metamorphic northern leopard frogs did not alter the frogs’ resistance to Bd infection and chytridiomycosis. However, exposure to the herbicide formulations did reduce growth, which ultimately can have adverse effects on amphibian populations in agricultural landscapes. Given the increasing understanding of immune function in leopard frogs exposed to pesticides (Rouf, 1969; Christin et al., 2003, 2004; Gilbertson et al., 2003; Hayes et al., 2006; Houck and Sessions, 2006; Albert et al., 2007; Brodkin et al., 2007; Rohr et al., 2008a), and the apparent resistance of certain leopard frog populations, but not others, to chytridiomycosis (Ouellet et al., 2005; Longcore et al., 2007; Woodhams et al., 2008), this species appears to be an excellent model for the study of immune response against Bd. This study challenges the hypothesis that the prevalence of pesticide residues in aquatic environments is contributing to the emergence of chytridiomycosis. However, a low susceptibility to infection by Bd in the population of leopard frogs that we sampled could explain why we did not measure significant interactive effects of the herbicide and fungal exposures. A follow-up study examining whether pesticide-Bd interactions take place within host species that initially demonstrate a greater susceptibility to Bd would be a logical direction for future research.

Acknowledgments We thank Andre´e Gendron, Malorie Ge´linas, Sophie Tre´panier, Jean-Franc- ois Lafond, Ste´phanie Arseneault, Hubert De´silets, Violaine Peltier, Coralie Beaudry, E´milie Lessard, Maxime Guerard and Ariane Laurence for their assistance in the laboratory, Louise Rollins-Smith, Douglas Woodhams, Todd Smith, David M. Green, Colin Rousseaux, Sylvia Ruby, Chris Blanar and Marc Champagne for their advice on protocols, and Joyce E. Longcore for providing the fungal culture. Partial funding was provided by a Natural Sciences and Engineering Research Council of Canada Postgraduate Scholarship (PGS M) to LJP, and by Environment Canada’s Pesticide Science Fund (PSF) and STAGE funding programme to DJM and BDP. Permission to collect frog specimens was issued by the Quebec provincial government (Ministe re des Ressources Naturelles et de la Faune) (permit number 2007-05-17-789-16-SF).

Appendix A. Supplementary materials Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.ecoenv.2012.04.006.

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