Sub-lethal effects of a neonicotinoid, clothianidin, on wild early life stage sockeye salmon (Oncorhynchus nerka)

Journal Pre-proof Sub-Lethal Effects of a Neonicotinoid, Clothianidin, on Wild Early Life Stage Sockeye Salmon (Oncorhynchus nerka) Vicki Lee Marlatt, Tsz Yin Ginny Leung, Sarah Calbick, Chris Metcalfe, Christopher Kennedy

PII:

S0166-445X(19)30499-0

DOI:

https://doi.org/10.1016/j.aquatox.2019.105335

Reference:

AQTOX 105335

To appear in:

Aquatic Toxicology

Received Date:

20 June 2019

Revised Date:

19 September 2019

Accepted Date:

13 October 2019

Please cite this article as: Marlatt VL, Ginny Leung TY, Calbick S, Metcalfe C, Kennedy C, Sub-Lethal Effects of a Neonicotinoid, Clothianidin, on Wild Early Life Stage Sockeye Salmon (Oncorhynchus nerka), Aquatic Toxicology (2019), doi: https://doi.org/10.1016/j.aquatox.2019.105335

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Sub-Lethal Effects of a Neonicotinoid, Clothianidin, on Wild Early Life Stage Sockeye

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Salmon (Oncorhynchus nerka)

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Vicki Lee Marlatta, Tsz Yin Ginny Leunga, Sarah Calbicka, Chris

of Biological Sciences, Simon Fraser University, 8888 University

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aDepartment

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Metcalfeb,c, Christopher Kennedya

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Drive, Burnaby, British Columbia, Canada

bWater

Quality Centre, Trent University, Peterborough, ON, Canada

for Watershed Science, Trent University, ON, Canada

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cInstitute

*Corresponding author: Vicki L. Marlatt, Department of Biological Sciences, Simon Fraser University, 8888 University Drive, Burnaby, BC, Canada, V5A 1S6, Ph: 778-782-4107, Email: [email protected]

Highlights -chronic, waterborne exposure to a neonicotinoid, clothianidin, did not decrease survival, hatching, growth or deformities in sockeye salmon at 0.15, 1.5, 15 and 150 μg/L 

low level clothianidin (0.15 µg/L) increased 17β-estradiol levels in sockeye salmon swim-up fry, but testosterone was unaffected liver glucocorticoid gene expression was reduced after 150 µg/L clothianidin

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exposure in sockeye salmon swim-up fry

Abstract One of the categories of environmental contaminants possibly contributing to declining sockeye salmon (Oncorhynchus nerka) in the Fraser River, British Columbia, Canada is pesticides. In this 4-month study, the effects of environmentally relevant concentrations of a waterborne neonicotinoid, clothianidin (0.15, 1.5, 15 and 150 μg/L), on embryonic, alevin and early swim-up fry sockeye salmon derived from four unique genetic crosses of the Pitt River, BC stock were investigated. There were no significant effects of

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clothianidin on survival, hatching, growth or deformities, although genetic variation

significantly affected these endpoints. Clothianidin caused a significant 4.7-fold increase

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in whole body 17β-estradiol levels in swim-up fry after exposure to 0.15 µg/L, but no

effects were observed on testosterone levels. In addition, hepatic expression of the gene

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encoding glucocorticoid receptor 2 was also impacted at the highest concentration of clothianidin tested, and was found to be ~4-fold lower compared to the sockeye reared in control water. These results indicate additional examination of clothianidin and its

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effects on salmonid gonad development and the reproductive and stress endocrine axes

growth; development; estradiol, testosterone; gene expression; glucocorticoid receptor

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Keywords:

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in general, is warranted.

Introduction Currently, one of the most frequently used groups of synthetic pesticides of the 21st century is the neonicotinoids, which are broad-spectrum, systemic insecticides (Jeschke et al., 2008; Jeschke et al. 2011; Tomizawa et al., 2004; Casida et al., 2013). There are seven neonicotinoids available and registered in over 120 countries, including clothianidin, imidacloprid and thiamethoxam which account for more than 26% of the total global insecticide sales (Jeschke et al, 2011; Simon-Delso et al., 2015).

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Neonicotinoids are mainly used as seed treatments, soil applications, and foliar sprays on a wide variety of agricultural crops such as oilseeds, grains, fruits, vegetables,

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greenhouse crops and ornamental plants (Canadian Council of Ministers of the

Environment (CCME), 2007; Pest Management Regulatory Agency (PMRA), 2004;

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Health Canada, 2017; Simon-Delso et al. 2015). Due to the structural similarity with nicotine, neonicotinoids share the same mode of action as nicotine in invertebrates and vertebrates, binding agonistically to the nicotinic acetylcholine receptor (nAChRs) in the

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nervous system interfering with acetylcholine neurotransmitter signalling (Simon-Delso et al. 2015; Sattelle, 2009; Crossthwaite, 2017; Taylor, 2012). Neonicotinoids are more

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toxic to invertebrates because these insecticides have greater affinity for invertebrate nAChRs compared to vertebrate nAChRs (Matsuda et al. 2001; Matsuda et al. 2005). When an invertebrate such as an insect is exposed to neonicotinoids, the neonicotinoid

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molecules irreversibly and selectively bind to nAChRs keeping ion channels open resulting in overstimulation, and eventually paralysis and death within minutes (SimonDelso et al. 2015; Gibbons et al., 2014). Therefore, this group of insecticides is efficient for controlling a wide range of economically important pests including aphids, leafhoppers and phytophagous mites, and as such, were rapidly adopted by both

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agricultural and urban pesticide users globally (Simon-Delso et al. 2015). In general, neonicotinoids are quite water soluble (0.33 g/L, 4.1 g/L and 0.61 g/L

for clothianidin, thiamethoxam and imidacloprid respectively, at 1 atm and 25 °C), allowing them to be readily absorbed into roots and leaves of plants and transported systemically (Bonmatin et al., 2015; Fossen, 2006). However, this also points towards a concern for the potential leaching and movement of neonicotinoids into surface waters and groundwater (Bonmatin et al., 2015; Fossen, 2006; United States Environmental Protection Agency (US EPA), 2010). Indeed, neonicotinoids have been frequently

detected in a variety of water bodies, typically at concentrations in the low μg/L range. In a review of 29 studies from nine countries, neonicotinoids were listed as one of the more common contaminants of surface water (Morrissey et al., 2015). In particular, there are numerous reports of clothianidin contaminating various water bodies including ponds, river, groundwater, puddled water, soil water and run-off and are detected most frequently in proximity to agricultural areas in North America, Asia, Australia and Europe (de Perre et al., 2015; Morrissey et al., 2015; Yamamoto et al., 2012). A study by Main et al. (2014) collected water samples from 136 wetlands across four rural municipalities

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in Saskatchewan, Canada during the spring, summer and fall of 2012 as well as in the

spring of 2013. This study revealed that clothianidin and thiamethoxam were present in

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the majority of samples and were detected in all 4 sampling periods/seasons, with

clothianidin being the most commonly detected neonicotinoid in the water samples (Main

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et al., 2014). Clothianidin also had the highest maximum (max) and mean concentrations during three of the sampling periods: spring 2012 (max: 144 ng/L), summer 2012 (max: 3110 ng/L), and spring 2013 (max: 173 ng/L). In another study conducted in

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southwestern Ontario, Canada, 76 agricultural surface water samples were collected within commercial maize farms, all of which had clothianidin at a mean concentration of

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2.28 µg/L and a maximum concentration of 43.60 µg/L (Schaafsma et al., 2015). This study also reported that the total neonicotinoid concentration increased 6-fold after the planting season, suggesting the main source of pesticide pollution was from agricultural

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settings (Schaafsma et al., 2015). Currently low-level, chronic exposure to many neonicotinoids, including clothianidin, appears to be a potentially important environmentally relevant exposure scenario for aquatic wildlife. Adverse population level effects in terrestrial and aquatic invertebrates after low

level, environmentally relevant exposure to neonicotinoids has been demonstrated in

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numerous studies (Beketov et al., 2008; Gibbons et al., 2015; Miles et al., 2017; Morrissey et al., 2015; Roessink et al., 2013; Van Dijk et al., 2013). For example, clothianidin has been shown to suppress the honey bee immune system resulting in an increased vulnerability to disease and parasites, which can cause lethal effects on larvae and reduction in queen survival resulting in colony collapse disorder in honey bees (Di Prisco et al., 2013). Recent experiments in a model aquatic benthic invertebrate (Chironomus dilutus) demonstrated clothianidin exhibits similar acute and chronic toxicity compared to another neonicotinoid, imidacloprid, thus it is likely that the current

environmental quality guidelines for imidacloprid in Canada and the US would apply to clothianidin as well (Cavallaro et al., 2017). In mammals, clothianidin has been shown to cause detrimental effects on reproductive organs and gametes. For example, exposing rats to 32 mg/kg/day of clothianidin (oral administration; 90-day) caused a significant decrease in the weight of epididymis and seminal vesicles (Bal et al., 2012). This chronic exposure to this sub-lethal concentration also induced oxidative stress via enhanced reactive oxygen species production causing sperm DNA fragmentation, and reduced serum testosterone levels in these male rats (Bal et al., 2012). A chronic toxicity study

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reported by U.S. Environmental Protection Agency (EPA) revealed significant

differences in the dry body weight and body length of fathead minnows after exposure to

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20 mg/L clothianidin in a flow-through system (US EPA, 2010). However, in general, the chronic, low-level, sub-lethal effects of clothianidin in various wild aquatic vertebrates is

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still largely unknown, thus, environmentally relevant studies in such taxa are warranted especially those inhabiting surface waters where contamination is prevalent.

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Historically, the Fraser River in British Columbia, Canada was the river that produced the highest numbers of sockeye salmon in the world, but the population of

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Fraser River sockeye has been declining since 1990s (Cohen, 2012; Noakes, 2011). Several possible causes of this decline identified to date include: climate change (i.e. increase in temperature); fishing pressures caused by the growth in fishery industries;

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habitat destruction; disease or parasites; and, environmental contaminants (Cohen, 2012). To date, no studies examining the adverse effects of low-level, chronic exposure of a sockeye salmon to a neonicotinoid have been reported. Therefore, the main objective of this study was to examine the effects of a neonicotinoid insecticide, clothianidin, during chronic exposure in early life stages of a wild salmon species, sockeye salmon (Oncorhynchus nerka). In this study, 4 concentrations of clothianidin

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(0.15, 1.5, 15 and 150 μg/L) plus a water control were tested in chronic exposure experiments that were initiated ~1 hour post-fertilization and continued through to the swim-up fry developmental stage. The concentrations were selected based on the CCME water quality guideline for imidacloprid of 0.23 µg/L, the USEPA aquatic life benchmarks of 35 (acute) and 1.05 µg/L (chronic) for imidacloprid for invertebrates, and the range of neonicotinoid concentrations reported in various surface waters in North American, Asian and Australian studies (0.0035 - 320 μg/L) (Schaafsma et al., 2015; Main et al., 2014; Morrissey et al., 2015; de Perre et al., 2015; Hladik et al., 2014; Van

Dijk et al., 2013; Sánchez-bayo and Hyne, 2014; Miles et al., 2017; Starner and Goh, 2012). Since wild salmon are not routinely studied and to better understand the influence of parentage on toxicant responses in this species, four unique offspring sets (crosses) were incorporated into the experimental design. The endpoints measured to assess the adverse effects of clothianidin in developing wild sockeye salmon included survival and several sub-lethal endpoints, specifically, growth, hatching, emergence, various gene

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expression changes and sex steroid hormone levels.

Methods 1.1. Chemicals The clothianidin stock used for this study was prepared using ≥ 98.0 % pure clothianidin (CAS#: 210880-92-5, Sigma-Aldrich, Oakville, Ontario, Canada). Tricaine methanesulfonate (MS-222) and Ovadine for fish euthanization and disinfection

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procedures, respectively, were obtained from Syndel Laboratories Ltd., Nanaimo, B.C., Canada. All other chemicals used in this study were of analytical grade from commercial

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sources and are specified upon mention below.

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1.2. Sockeye Salmon Gamete Collection and Fertilization Four sexually mature mating pairs of wild sockeye salmon were captured in the

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Pitt River, BC, Canada during the fall 2015 spawning season. Fish were caught by Fisheries and Oceans Canada staff of the Inch Creek Hatchery (Dewdney, BC), and

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were generously donated for this project. Approximately 2,000 to 3000 eggs were collected from each female and 1 to 3 ml of milt were collected from each male. Gamete collections were performed by applying gentle pressure to the body in an anterior to

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posterior direction on the ventral surface into dry food grade plastic containers, and containers were transported at 6 - 10 °C to Simon Fraser University, Burnaby, BC on September 8, 2015.

Dry fertilizations were performed within 6 hours of the collection of gametes to

create four unique offspring sets, referred to as cross A, B, C and D. As such, eggs and

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milt from each mating pair were fertilized independently and kept separate throughout all procedures and subsequent exposures to evaluate differences between different offspring sets/genetic crosses. Dry fertilizations were performed according to Patterson et al. (2004) whereby the eggs from one female were combined with the milt from one male in a 4 L food grade plastic container, followed by the addition of 1.5 L of dechlorinated water (10 ± 1 °C) and gentle swirling (Patterson et al. 2004). The fertilized eggs remained in these containers for a minimum of 60 minutes to allow for water hardening and then transferred into separate netted cylindrical egg containers (food

grade polyvinyl chloride (PVC)) and placed in exposure vessels. Fertilization success was determined by the proportion of eyed embryos on 28 post-fertilization day (dpf) to the total number of eggs placed in the control or exposure vessels at the onset of the experiment.

1.3. Aquatic Exposure Systems

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This study was divided into two different aquatic exposure systems: 1) glass tank flow-through test vessels conducted in duplicate; and, 2) gravel-bed flume incubators

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that simulate a streambed environment conducted in duplicate. The fish in the glass tanks allowed for monitoring survival and development throughout the experiment.

However, the fish in the gravel-bed flume incubators mimicked a more natural incubation

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system whereby fish were buried in the gravel upon reaching the eyed embryo stage and allowed to emerge naturally from the gravel. As such, survival and development

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were monitored only at the end of the experiment in the gravel-bed flume incubators. Both aquatic exposures were conducted from 1-3 hours post-fertilization through to the swim-up fry developmental stage (duration: 118 days for glass tank exposures and 110

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days for gravel-bed flume exposures).

Glass tank exposure systems were monitored daily throughout the experiment for

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survival, health and development and dead embryos/alevins were also removed daily. For each mating pair, approximately 100 fertilized eggs were divided between three netted cylindrical PVC containers and placed in each of the duplicate tanks (i.e. ~100 eggs/glass tank divided among three netted cylindrical PVC containers). The total volume of each tank was 28 L and the dimensions of each glass tank were as follows:

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22 cm height (with a 20 cm high drainage hole) by 26 cm wide and 52 cm long. Duplicate tanks for each test concentration and the control were placed in random order at two opposite sides of the temperature-controlled room. In order to maintain a uniform water pressure, an overhead tank was constructed to ensure consistency of water flow into the glass tanks. Dechlorinated municipal water was dispensed through food grade Tygon tubing and flow rates were monitored and adjusted every 48 hours throughout the exposures. Embryos were maintained in darkness until 90-100% hatching was achieved in the control glass tanks, and alevins were then reared under a 16 h light: 8 h dark

photoperiod until termination of experiment. Termination was performed when 87-98% of surviving alevins reached the swim-up fry developmental stage (100% yolk sac absorption) in the control glass tanks. The endpoints examined in the glass tank exposure experiment included the following for each individual fish in all 4 genetic crosses: survival; hatching; body morphometrics and condition factor; % of deformities; hepatic gene expression, and, whole body 17β-estradiol and testosterone concentrations in swim-up fry (hormones analyzed in 1 genetic cross only).

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The gravel-bed flume exposure system was employed to simulate the natural

environment and allowed the fish to naturally ‘swim-up’ or emerge from the gravel once

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their yolk sacs were fully resorbed. This is a key process and behavioral endpoint

integral to the survival of a developing salmonid (Pilgrim et al. 2013), thus monitoring the

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emergence of swim-up fry was included as a sub-lethal endpoint to examine adverse impacts of clothianidin on swim-up success. The design of gravel-bed flumes was adapted from Pilgrim et al. (2013). The dimension of each flume was 250 cm in length by

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40 cm wide and 32 cm deep, and the flume was divided into five isolated sections with a total volume of 64 L each. Each of the five sections were sub-divided by stainless steel

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mesh to create five 10 cm long x 40 cm wide x 32 cm deep sub-compartments. Of these five sub-compartments that shared a single inflow and outflow, the middle compartment was used for drainage (outflow) while the other four compartments were used to house

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the developing sockeye salmon from each of the four genetic crosses. Similar to the glass tanks, approximately one hundred fertilized eggs from each of the four crosses were divided into three netted PVC cylindrical containers and placed in each of the four sub-compartments (i.e. one sub-compartment contained 3 PVC containers from a single cross). This entire gravel-bed flume was assembled in duplicate, and these two replicates were located on opposite sides of a temperature-controlled room. The

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arrangement of each test concentration and the control was randomly assigned within each replicate of gravel-bed flume incubator. Dechlorinated municipal water was delivered to the flumes through food grade Tygon tubing and flow rates were monitored and adjusted every 48 hours throughout the exposures. Two sizes of gravel rock, 10 mm and 25 mm, were selected as substrate and to rear developing sockeye salmon in the gravel-bed flume system (Kondolf and Wolman 1993). The two gravel rocks were mixed in a 1:1 ratio and disinfected using 1% Ovadine

solution followed by a dechlorinated municipal water rinse before placement into the gravel-bed flumes. The flumes were filled with the gravel to a depth of 5 cm and were flushed for a minimum of 24 hours with dechlorinated municipal water to remove any residual ovadine. The eyed embryos were housed in netted cylindrical egg containers on top of the gravel until 92-100% of the embryos developed eyes in the control. On 28 dpf, eyed embryos were gently deposited on top of the gravel and then buried with additional gravel to a height of 15 cm. The test volume of 28 L was maintained throughout the exposure by adjusting the height of the outflow pipe in each of the 5 sections in both of

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the replicate gravel-bed flumes. The photoperiod and termination date for the gravel-bed flume incubators was identical to that in the glass tank exposure system described

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above. The endpoints examined in the fish in the gravel-bed flume incubator exposure experiment included the following for each individual fish in all 4 genetic crosses:

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1.4. Pesticide Exposures

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survival; hatching; body morphometrics and condition factor, and % of deformities.

Clothianidin stock solutions were prepared fresh every 48 hours to prevent any

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degradation of the chemical. This was selected based on research indicating an aqueous photolysis half-life ranging from 0.35 to 3.31 days in freshwater mesocosm studies during 4 different seasons in Winnipeg, Canada (Lu et al., 2015), and 25.1 to

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27.7 hour half-life in nonsterile river water under 9 h light: 15 h dark (Rexrode et al. 2003). In the present study, fish were exposed from 1 hour post-fertilization through to the swim-up fry developmental stage in a water control and four concentrations of clothianidin: 0.15, 1.5, 15 and 150 μg/L. The stock was prepared using ≥ 98.0 % pure clothianidin (CAS#: 210880-92-5). Since clothianidin is soluble in water (0.327 g/L at 20 o

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C; US EPA, 2003), the clothianidin stock solution was prepared by adding 0.200 g of

clothianidin to 4 L of dechlorinated municipal water. This solution was allowed to mix for one hour until clothianidin fully dissolved. This solution was then further diluted with dechlorinated water and distributed into glass stock solution containers and delivered into glass tanks and gravel-bed flumes housing the fish by a Masterflex peristaltic pump using Masterflex silicone tubing at 2.0 ml/minute. The nominal concentrations of clothianidin in the treatment tanks/gravel-bed flumes were achieved by each tank/flume receiving a water flow rate of 95 ml/minute and a pesticide stock solution flow rate of 2.0

ml/minute. The pesticide and water inflow rates were monitored every 48 hours and adjusted if necessary throughout the entire duration of the exposure experiments. The actual clothianidin water concentrations were measured in 10 mL grab samples collected from one replicate glass tank and one gravel-bed flume replicate per test concentration on 70 dpf according to the methods described for grab water samples described in Sultana et al. (2018). Water temperature, dissolved oxygen concentration, pH and conductivity were measured every 48 hours using an HQd Portable Meter (Hach Company, Loveland, CO, USA). Ammonia concentrations were monitored every two

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weeks using Seachem MultiTest Ammonia Test Kit (Seachem Laboratories, Madison,

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USA; detection limit 0.05 mg/L).

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1.5. Measurement of Survival, Hatching and Emergence The survival of embryos/alevins in both exposure systems was monitored every

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48 hours post-fertilization until eyed embryo developmental stage. On 28 dpf, eyed eggs were counted to assess the fertilization rates and embryonic survival in both systems. The survival monitoring in the glass tank exposure system was then daily until

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termination of the experiment 119 dpf, and only upon termination of the experiment in the gravel-bed flumes. Daily survival in glass tanks was determined by calculating the

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proportion of surviving swim-up fry to the total eyed embryos observed on 28 dpf. Hatching was monitored daily and started on 49 dpf in glass tanks. The daily

percent hatched was calculated by the number of hatched alevins divided by the total eyed embryos on 28 dpf. In addition, the 10th percentile (H10), the 50th percentile or median (H50) and 90th percentile (H90) was determined for each treatment according to

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Sternecker and Geist 2010. The average time required for alevins in replicates to reach the H50 and H90 and the duration between the H10 and H90 were examined in order to compare the timing of hatching across treatments and genetic crosses. Emergence from the gravel started on 88 dpf in gravel-bed flumes and was monitored daily. The number of emerged fry was recorded in each sub-compartment, and the order of observations was randomized daily based on treatment to randomly distribute human disturbance during netting in adjacent tanks. Emerged fry were

captured daily using a net and placed in covered/netted cylindrical containers and maintained on top of the gravel in corresponding sub-compartments until termination 119 dpf. The daily emergence and survival rate was calculated by the number of swim-up fry captured that day divided by the total eyed embryos buried on 28 dpf in a single subcompartment.

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1.6. Morphometric and Deformity Analyses Termination was conducted when the yolk sac was absorbed in 87-98% alevins

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in the control glass duplicate tanks. Fish were placed in an observation container to

examine skeletal and swim abnormalities and then individuals were euthanized with an overdose of tricaine methanesulfonate buffered with sodium bicarbonate to pH 7.0-7.5

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on 118-119 dpf and 110-119 dpf in the glass tanks and gravel-bed flumes, respectively. Fish were then weighed, snout to fork length was measured, external deformities were

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assessed followed by archiving livers or whole bodies by snap freezing on dry ice. The condition of the fry was determined by calculating Fulton’s condition factor (K): K=100W/L3 where W= Wet weight of fry (g) and L=Length of fry (cm) (Datta et al. 2013).

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Some of the individual severely deformed fry showing significantly lower body length and body weight (outliers) were removed in the statistical analysis. Four main categories of deformities (skeletal, craniofacial, finfold and edema) were assessed using a graduated

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severity index (GSI) under a dissecting microscope immediately after euthanization for each individual fish according to the methods described by Rudolph et al. (2008).

1.7. Biochemical Analyses

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Whole body concentrations of 17β-estradiol and testosterone per gram of body

weight were quantified in fish from the glass tank exposure system in one genetic cross (cross A). Five fish from the two replicate glass tanks from each of the five treatments were included in this analysis (n = 2; 5 individual swim-up fry per tank). Homogenization and extraction of hormones from swim-up fry whole bodies for both 17β-estradiol and testosterone hormone assays were performed according to Arukwe et al. (2008). Estradiol and testosterone levels of the five cross A whole body hormone extracts from

two replicate glass tanks were measured using enzyme immunoassay kits: Estradiol ELISA Kit and Testosterone ELISA Kit (Cayman Chemical Company, Michigan, USA, Item No. 582251 and 582701, respectively) according to the manufacture’s protocol. Briefly, multiple 96-well assay plates were used on the same day using all of the same reagents and standards, and each plate included the following: duplicate blank wells; duplicate non-specific binding wells; triplicate maximum binding wells; duplicate 8-point standard curve concentrations (6.6 pg/ml to 4,000 pg/ml estradiol or 3.9 to 500 pg/ml testosterone); and, a unique set of duplicate whole body samples. An EPOCH2

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microplate reader (BioTek Instruments Inc., Winooski, Vermont, USA) and Gen 5.02 Software (BioTek Instruments Inc., Winooski, Vermont, USA) were used to read the

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absorbance of the assay in 96-well microplates at 70 and 65 min for estradiol and

testosterone, respectively. Hormone concentrations in samples were quantified using the

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standard curve for estradiol and testosterone which was linearized by logit transformation (logit (Sample binding/maximum binding)) according to the

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manufacturer’s protocol.

Several quality assurance/control measures were undertaken during the

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hormone ELISA procedures. The degree of difference between measurements was expressed by coefficient of variability (CV), which was calculated by the standard deviation divided by the mean. The intra-assay CV, the variability of sample

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measurements on different wells within the same plate, was the average of all individual CVs of duplicates on a microplate. In addition, a set of two known concentrations prepared from the hormone stock and another set of 2-fold dilutions from the same batch of whole body samples of each tank were tested in duplicate on the same plate to analyze intra-assay variation. The mean intra-assay variation was 11.2% for estradiol and 11.0% for testosterone. As samples were run on multiple microplates and each

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microplate had its own calibration for the standard curve, plate-to-plate consistency was calculated by averaging the CVs of the same sample with known concentration (102.4 and 256 pg/ml for estradiol and 102.4 and 256 pg/ml for testosterone) on different plates. The overall inter-assay CV for estradiol and testosterone was 16.1% and 14.5%. The lower limit of detection was approximately 20 pg estradiol/ml and 6 pg testosterone/ml. The extraction efficiency was determined by homogenizing, extracting and resuspending 1,600 pg/ml of testosterone in an identical procedure except no whole fish samples were added. Extraction of spiked samples was performed in triplicate and the concentration of

spiked extracts was quantified in duplicate on the same microplate tested for whole body testosterone. The average recovery efficiency was 76 ± 5 % (mean ± standard error of 3 spiked samples).

1.8. Liver Collection, RNA Isolation and cDNA Synthesis Livers were collected at random from individual swim-up fry from the glass tank exposure system for gene expression analyses. Specifically, for each genetic cross, two

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livers from each of the five treatment groups (n = 2 per treatment per cross) were collected at the swim-up fry stage. Thus, all four genetic crosses tested in the

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clothianidin waterborne exposures (water control and 0.15, 1.5, 15, 150 μg/L) were

included for gene expression analyses in this study. Livers were excised and placed in

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1.5 ml safe-lock Eppendorf tubes (DNase/RNase free) and snap frozen immediately on dry ice, followed by transfer to -80°C for long term storage. All instruments used in liver tissue collection were cleaned in between each animal using 10% hydrogen peroxide

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and double rinsed with ultrapure water to ensure it was RNase/DNase free. All RNA isolation and cDNA synthesis procedures for samples used in subsequent quantitative

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real-time PCR (qPCR) experiments adhered to the guidelines in Bustin et al. (2010) for accurate design, documentation and reporting of qPCR experiments. Total RNA was isolated from the swim-up fry livers using TRIzol® Reagent as

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described by the manufacture (Invitrogen, Burlington, ON, Canada). Homogenization of one liver was performed utilizing 1 mL TRIzol reagent, and two 1 mm tungsten-carbide beads in a safe-lock Eppendorf 1.5 mL microcentrifuge tubes in a Retsch Mixer Mill MM 400 (Fisher Scientific, Ottawa, ON, Canada) at 30 Hz for a total of 8 min. Total RNA obtained from the TRIzol® RNA isolation procedure was subsequently reconstituted in

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30 μL DNase/RNase-free water, and stored at -80°C. The quantity (ng/µL) of RNA was determined by measuring the optical density unit (OD260) with an Epoch 2 Microplate Spectrophotometer (BioTek, Winooski, VT, USA), and the RNA purity was assessed measuring OD260/280 and OD260/230 ratios. Following determination of RNA concentrations, total RNA samples were DNase treated using TURBO DNA-free kits™ (Ambion, Austin, TX) to remove any co-extracted DNA according to the manufacturer’s instructions. RNA integrity of DNase-treated RNA samples was evaluated using a Bio-Rad Experion™ Automated Electrophoresis System and Experion software (version 3.20; Bio-Rad,

Mississauga, ON, CAN). RNA integrity values (RIN’s) for all samples used in this study ranged from 7.9 to 9.7 (RIN average ± 9.0; standard deviation = ± 0.50, n = 38). The DNase-treated RNA samples were stored at -80°C for long-term storage until subsequent cDNA synthesis. Reverse transcription of 1 μg of DNase-treated total RNA into cDNA for each liver sample was performed using qScript™ cDNA SuperMix (Quanta Biosciences, Beverly, MA, USA) following supplier’s instructions, and these cDNA samples were maintained at -20°C until subsequent quantitative real-time PCR

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experiments were performed.

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1.9. Quantitative Real-Time PCR (qPCR) for Hepatic Gene Expression

The relative quantification of each gene target/gene of interest in an individual

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fish liver cDNA sample was measured in qPCR experiments using the ΔΔCq method via the Bio-Rad CFX384™ Real-Time PCR Detection System and CFX Manager™

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Software (Bio-Rad Laboratories Ltd, Mississauga, ON, Canada), including normalizing the data to 3 reference genes and following the guidelines outlined for MIQE (Bustin et

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al., 2010). The 3 reference genes were those encoding cytoplasmic Beta-actin, (CBA), Glyceraldehyde 3-phosphate dehydrogenase (GAPDH) and Elongation factor 1α (EF1α). The gene expression targets examined in this study were relevant to various biological

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processes including reproduction, growth, stress responses and immune system function. A complete list of these genes is presented in Tables 1. Primers with optimal annealing temperature between 55-58°C were designed using Integrated DNA Technologies (IDT) OligoAnalyzer 3.1 (www.idtdna.com/calc/analyzer), and sequences obtained from GenBank National Center for Biotechnology Information (https://www.ncbi.nlm.nih.gov/) database for sockeye salmon when available, or rainbow

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trout (Oncorhynchus mykiss) and/or zebrafish (Danio rerio). Primer sets used to measure gene expression levels were tested for efficiency

using a 4 to 8 point standard curve generated by a 4-fold dilution of a 50 ng cDNA/μL of water (comprised of pooled liver samples from sockeye swim-up fry and/or liver from a 10 month old sockeye salmon). The criteria for acceptance of the standard curve included in single peak melt curve, efficiencies between 90-110%, amplification in at least 4 concentrations of the standard curve, and an R2 of the standard curve > 0.900

(Bustin et al., 2010). Primers that satisfied these criteria are shown in Table 1 and include product size, PCR efficiency, and correlation coefficient (R2). All primers sets were also evaluated for specificity by conducting PCR with each primer set on sockeye salmon fry cDNA and verification of each amplicon via sequencing. All sequencing was performed by Sanger Sequencing at the University of British Columbia Sequencing and Bioinformatics Consortium (Vancouver, BC, Canada). Sequencing results were aligned using the European Bioinformatics Institute EMBOSS Needle nucleotide alignment tool (https://www.ebi.ac.uk) with the respective rainbow trout nucleotide

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sequence obtained from GenBank National Center for Biotechnology Information for

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each gene.

All qPCR experiments were carried out on a Bio-Rad CFX384™ Real-Time PCR Detection System, following the guidelines outlined for MIQE (Bustin et al., 2010). All

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qPCR reactions for each gene were performed in triplicate on Hard-Shell 384 well PCR plates (Bio-Rad) according to the manufacturer’s instructions. Each qPCR reaction

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contained the following master mix for each target gene or reference gene in total reaction volume of 10 μL: 0.29-0.44 μL forward and reverse primers (0.23-0.35

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μMol/reaction), 6.25 μl of SsoFast™ EvaGreen® Supermix (Bio-Rad) and RNase-free water for the remaining volume. To the 10 µL of the qPCR master mix, a volume of a 2.5 µL of cDNA template (50 ng/μL) diluted to either 1:20 or 1:80 was added to make up a

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12.5 μL total reaction volume. A standard curve was performed in triplicate, for each gene on each plate to confirm amplification efficiency for each target and reference gene. Each plate contained three technical replicates per sample, and biological replicates (i.e. cDNA of an individual fish liver) were as follows; control (n = 5-8), 0.15 μg/L (n = 5-8), 1.5 μg/L (n = 5-8), 15 μg/L (n = 5-7), and 150 μg/L (n = 5-7). A no template control (NTC) was prepared for each primer set and tested in triplicate,

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whereby 2 μL of RNase-free water was used in place of cDNA. A no reverse transcriptase control (NoRT) was also prepared for each primer set and tested in triplicate, whereby 2 μL of DNase treated RNA was used in place of cDNA. Amplification reactions were performed using the following instrument settings: initial cycle 1 activation at 95 °C for 30 s, followed by 45 cycles at 95 °C for 5 s, and primer annealing at 55-58 °C for 5 s. After 45 cycles, a melt curve analysis was performed to confirm a single amplicon was amplified for each set of primers as indicated by a single peak. The melt

curve analysis was conducted using the following instrument settings: initial temperature of 65.0°C and was increased by increment of 0.5°C for 5 s to a maximum of 95.0°C.

1.10. Statistical Analyses All statistical analyses were performed with JMP®, Version 13 (SAS Institute Inc., Cary, North Carolina, United States). A split-plot design with two replicate test vessels

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was employed in a randomized complete block design to test the effect of pesticide concentration and genetic cross on the lethal and sub-lethal endpoints on sockeye

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salmon in this study. The test concentration was the main plot, and genetic cross was the second factor applied to sub-plots within the whole main plots within each block. Each block (one replicate of glass tanks and gravel-bed flumes) contained all five

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clothianidin concentrations (0, 0.15, 1.5, 15 and 150 μg/L) assigned at random while the five main plots within a block (glass tanks or sections of gravel-bed flumes) for each test

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concentration were further divided to four sub-plots housing all four genetic crosses (A, B, C and D). A 2-factor-split-plot-analysis of variance (ANOVA) followed by a Tukey’s post hoc test (p< 0.05) on the two replicate glass tanks/gravel bed flumes was performed

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on the following endpoints to analyze the interaction between the factors and two main fixed factors of the pesticide concentration and genetic cross: % survival of eyed embryos; % survival to swim-up; % emergence; timing and duration of hatching (the time

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at H50 and H90 and the duration between H10 and H90); average body weight and mass; average condition factor; and, proportion of deformities. Only one of the genetic crosses was tested for whole body hormone

concentrations, thus the difference in mean hormone levels was analyzed by a

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randomized complete block (RCB) ANOVA with the fixed factor of clothianidin concentration and blocking effect followed by a Tukey’s post hoc test (p< 0.05). For hepatic gene expression, statistical comparisons were conducted on the normalized hepatic gene expression values of 4 to 8 biological replicates (i.e. individual fish livers from all four genetic crosses; actual n is indicated in each figure caption for each gene) per treatment. A test for normality was conducted on Log10 transformed normalized expression values (ΔΔCq values) using a Sharpo-Wilk test, and a Levene’s test was employed to evaluate homogeneity of variance. Significant differences in normalized

gene expression between treatments were determined by one-way analysis of variance (ANOVA; p < 0.05) followed by a Tukey's post-hoc, except for ERα and ERβ2 that displayed non-normally distributed data and were therefore analyzed using a Kruskal-

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Wallis test.

Results 2.1. Clothianidin Exposures and Water Quality Nominal concentrations of clothianidin were 0, 0.15, 1.5, 15 and 150 µg/L during this 119-day waterborne clothianidin exposure experiment, and the measured and predicted concentrations were similar to these nominal values (Table 2). Throughout the

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entire experiment the water flow rate and pesticide flow rates were relatively consistent with an average of 91.5 (standard deviation (SD) 2.14) ml/min and 2.0 (SD 0.03) mL/min

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in each glass tank, and 92.0 (SD 0.594) ml/min and 2.0 (SD 0.036) ml/min in each section of gravel-bed flume, respectively (Table 2). The predicted exposure

concentrations were relatively close to the nominal concentrations (Table 2). Likewise,

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the measured pesticide concentrations from a single sampling event from one replicate were close to the nominal concentrations, and were within the range of the minimum and

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maximum predicted concentrations. Water quality parameters were within the desired range (Table 3), and ammonia concentrations were below 0.05 mg/L in all glass tanks

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and gravel-bed flumes throughout the experiment.

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2.2. Fertilization and Survival to Eyed Embryo Stage The control treatment demonstrated high survival (mean= 88.4 ±SE 5.7% and

93.5 ±SE 1.6 % in glass tanks and gravel-bed flume, respectively) up to the eyed embryo developmental stage, indicating a high fertilization success rate for all crosses prior to the exposure period. Clothianidin exposures were initiated initiated ~1 hour postfertilization and did not affect survival to the eyed embryo stage in any treatments,

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regardless of the glass tank exposure or gravel bed rearing conditions (glass tank, p = 0.521; gravel-bed flume, p = 0.573; Figure 1). However, there were differences in survival between genetic crosses. The mean survival of cross A was significantly lower than the cross C and D in glass tanks (p = 0.0003; data not shown). Interestingly, in the gravel-bed flume system, cross B exhibited decreased survival compared to cross C and D but not compared to A in all treatments (p = 0.0125; Figure 1). The accumulated

thermal units (ATU) on 28 dpf (92-100% eyes developed in the control) was 385.7 and 386.0 oC in glass tank and gravel-bed flume exposure system respectively (Table 4).

2.3. Effects of Clothianidin on Survival from the Eyed Embryo to Swim-up Fry Stage Upon termination of the experiment 118 dpf at the swim-up fry developmental

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stage cross D had an extremely low survival rate of 11 and 41% in two replicate control glass tanks, and thus was excluded from further analyses. There were no significant

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differences in survival between any treatments (p = 0.327) or genetic crosses A, B and C (p = 0.470) upon termination at the swim-up fry developmental stage in the glass tanks

(standard error (SE)) 5.1% in control (Figure 2).

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(Figure 2). Mean survival from eyed ebmryo to the swim-up fry stage was 73.8% ±

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In the gravel bed flume systems only 34% of the swim-up fry at most were recovered at the end of the experiment due to the difficulty catching the fish in these systems (data not shown). The swim-up fry were hiding in the gravel and at the back of

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the flume most of the time during swim-up fry counts and collection attempts rendering netting unsuccessful. Therefore, with this low recovery rate, the survival rate was not

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reliably obtained and cannot be calculated.

2.4. Hatch and Emergence Success and Timing Overall, the average hatching success was 74.7 ± SE 5.5 % in control fish in

glass tanks compared to 57.9 ± SE 10.2 % in the highest (150 µg/L) clothianidin

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treatment, but no statistically significant effect of clothianidin treatments (p = 0.341) or genetic crosses (p = 0.412) on hatching success were detected (Figure 3). The ATU when the first hatch was observed in the glass tank system was 49 dpf, and this equated to 658.9 and 661.4 degree days in the glass tanks and gravel-bed flume systems, respectively (Table 4). The average timing of hatching across treatments and genetic crosses is presented in Figure 4 to indicate the first hatch, H10, H50, H90 and the last hatch. There

was no significant effect of treatment on H50 (p = 0.886), but there was a significant difference in H50 between genetic crosses (p = 0.0014; Figure 5A). Specifically, fish in cross A and B reached H50 on 53 ± SE 0.45 dpf and 54 ± SE 0.62 dpf, while cross C was delayed 2 to 3 days before 50% of the total fish had hatched (H50 = 56 ± SE 0.88 dpf; Figure 5A). Similarly, no significant difference in the mean time to H90 was observed between treatments (p = 0.440) (Figure 5B); however, cross C required 67 ± SE 1.8 days which was 9 and 6 days longer than cross A and B, respectively, to reach 90% of total hatched (p = 0.0003). On average, no significant effect of treatments on duration

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from 10% of total hatched to 90% of total hatched was observed (Figure 5C; p = 0.498).

However, significantly longer mean duration between H10 and H90 was detected in cross

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C (p = 0.0008). Cross A and B fry took an average of 7.1 ± SE 0.67 and 10 ± SE 1.5

days to reach H90 from H10, while cross C required an average of 15-days in duration (SE

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= 1.8 days).

The swim-up behavior was first observed on 88 dpf in gravel-bed flume (1088.1 o

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C-days; Table 4). However, this experiment failed to reliably assess the effect of

clothianidin on the emergence of swim-up fry from the gravel due to only capturing 34%

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of the total fish originally added to the gravel-bed flumes at the end of the experiment (data not shown). Therefore, with this low recovery rate, the emergence success and

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timing of emergence endpoints were not obtained.

2.5. Morphometric Analysis There was no effect of clothianidin concentration on the mean body weight in

swim-up fry reared in the glass tanks and gravel-bed flume systems (0.172 ± SE

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0.00221 g, p = 0.0700 and 0.170 ± SE 0.00207 g, p = 0.320, respectively; Figure 6). However, genetic cross B swim-up fry exhibited a significantly higher body weight compared to cross A and C in both the glass tank and gravel-bed flume systems (p < 0.0001 in both systems). Similarly, there was no significant difference in mean body length between treatments in glass tank or gravel-bed systems (29.8 ± SE 0.132 mm, p = 0.355 and 29.8 ± SE 0.127 mm, p = 0.230, respectively), but the average body length for cross B was significantly higher than the other genetic crosses (p = 0.0069 and p< 0.0001; Figure 6). There was no significant effect of clothianidin concentration (p = 0.115

in glass tanks; p = 0.242 in gravel-bed incubators) or genetic cross (p = 0.143 in glass tanks; p = 0.765 in gravel-bed incubators) on K in sockeye salmon in both systems (data not shown). The average condition factor in swim-up fry in the control glass tanks and gravel-bed flumes was 0.630 ± SE 0.00690 and 0.652 ± SE 0.00413 (data not shown).

2.6. Analysis of Deformities

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There were no fin or edema related deformities at the swim-up fry developmental stage in the glass tank or gravel-bed flume systems, and swim-up fry in gravel-bed

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flumes had no deformities. Only skeletal and craniofacial deformities were observed in

the glass tank exposure system. Out of 2382 swim-up fry in all glass tank experimental groups, a total of 85 swim-up fry exhibited skeletal deformities and out of the 85

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deformed fry the following deformities were observed: 39.3% kyphosis; 14.3% lordosis; 39.3% scoliosis; and, 7.14% 2-headed fish with a single body. The mean skeletal

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deformity rate in the glass tanks was 2.65 ± SE 1.18% in the control. There was no significant effect of clothianidin treatments (p = 0.1243) on the overall deformity rate, but the skeletal deformity rate in cross C was significantly lower than the other two crosses

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(p = 0.0149; data not shown). The mean skeletal deformity rate in all groups in cross A and B was 4.23% ± SE 0.75% and 4.67 ± SE 0.78%, respectively, while cross C was 2 times lower than other crosses (2.12% ± SE 0.64%). The severity ranged from mild,

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moderate to severe at a similar ratio and fish that exhibited severe spinal deformities had obvious mobility impairment. For craniofacial deformity, a total of 103 deformed swim-up fry were recorded with 86.2% of these fry having a reduced eye to head ratio, 9.23% had reduced pupil to eye ratio and 4.62% with a malformed head. The mean percent of craniofacial deformities in the control glass tanks was 5.02% ± SE 1.32%. There was no

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significant effect of clothianidin or cross on craniofacial deformities and all were assigned a severity score of 1, which is a minor malformation (data not shown).

2.7. Biochemical Analyses To determine if sub-chronic clothianidin exposure could adversely affect hormones associated with the reproductive endocrine axis in early life stages of sockeye

salmon, whole body 17β-estradiol and testosterone were measured in swim-up fry reared in the glass tank exposure system for cross A. Swim-up fry exposed to 0.15 µg/L clothianidin had significantly higher concentrations of 17β-estradiol than any other treatments (p < 0.0001; n= 2; 5 fry/tank; Figure 7A). The mean 17β-estradiol level in the control was 1536 ± SE 368 pg/ml/g bw and it was 7312 ± SE 743 pg/ml/g bw for the clothianidin 0.15 µg/L exposed fish, which was 3 to 4.7-fold higher than other clothianidin treatments and the control. There was no significant difference in the testosterone levels (p = 0.117) in swim-up fry between treatments with a mean concentration of 153 ± SE 30

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pg/ml/g body weight (bw) in the control and 289 ± SE 36 pg/ml/g bw in all treatments

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(0.15 - 150 µg/L clothianidin; Figure 7B).

2.8. Hepatic Gene Expression

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The target stability function using CFX Manager™ Software Gene Expression Analysis demonstrated that the combined M-value for the three reference genes (CBA,

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GAPDH, and EF1α) was 0.56 with a coefficient of variation (CV) = 0.22, thus all 3 of these reference genes were used for normalization purposes for all target genes tested

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in this study. No significant difference was detected between all treatments for CYP1A, ERα, ERβ2, and SOCS3 at any concentration (p > 0.05; data not shown). However, a significant 4-fold decrease for GR2 was detected between the normalized expression in

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the liver of swim-up fry exposed to 150 μg/L of waterborne clothianidin compared to liver of the swim-up fry in the water control treatment group (Figure 8). The average normalized expression ± SE for the 150 μg/L clothianidin was 1.01 ± 0.49, and for the

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water control was 4.29 ± 1.64 (p = 0.0281).

Discussion Controlled laboratory studies examining the effects of environmentally relevant concentrations of pollutants on wild sockeye salmon are limited in the literature, and this is the first study to report the effects of the neonicotinoid clothianidin in this species. In this study, four unique genetic crosses were exposed to the clothianidin at concentrations of 0.15, 1.5, 15 and 150 μg/L initiated 1 hour post-fertilization through to

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the swim-up fry stage. No significant effects of clothianidin on survival, hatch success/duration/timing, growth and deformity rates were observed at any of the

concentrations tested. However, the 0.15 μg/L clothianidin treatment significantly

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increased whole body 17β-estradiol levels in one of the genetic crosses, resulting in a non-monotonic concentration response curve. It is well established that elevated

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circulating levels of 17β-estradiol feminize undifferentiated gonads in developing salmonids and many other teleosts. These results in addition to decreased hepatic

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glucocorticoid receptor 2 transcript levels in the highest concentration tested indicate additional examination of clothianidin and its effects on salmonid gonad development, and the endocrine system in general, is warranted. Furthermore, the significant

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differences observed in growth and development of the four unique genetic crosses of wild caught sockeye in this study also underscores the influence of genetics on variation

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in apical endpoints, and should be considered in future toxicity testing in this species.

Genetic Cross Differences in Wild Sockeye Salmon Several endpoints were significantly different between crosses, specifically,

embryonic survival for all four crosses, timing and duration of hatching, body length,

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body weight and skeletal deformity rate were significantly different between 3 of the crosses. Previous studies have demonstrated parental influence is a significant driver of offspring variation in fish populations (Burt et al., 2012; Marshall et al., 2008; Green 2008), including in a BC sockeye salmon population (Burt et al., 2012). Burt et al. (2012) reported the survival to hatch of Weaver Creek, BC sockeye salmon with an early incubation temperature of 16 oC caused substantial variation in responses in different genetic crosses of early life stage salmon to thermal stress (Burt et al., 2012). In particular, the embryonic survival significantly decreased to an average of 60 ± SD 23%

with a huge variation ranging from 31.2 to 92.3% in different genetic crosses, compared to those same genetic crosses subjected to an early 12 oC pulse with average survival 95% ± SD 5% (Burt et al., 2012). The early high temperature incubation in the Weaver Creek study showed a persistent high temperature effect on fry survival even if the thermal stress was removed after hatching, and the genetic cross variation in survival tended to increase as temperature increased from the optimal incubation temperature of 4 - 12.5 oC (Burt et al., 2012). The mean fry survival showed a similar trend compared to the embryo survival, whereby fry survival at 14 oC between the 4 families varied from ≤

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75% to ≥ 90% while at 16 oC survival ranged from ≤ 25% to 80% (Burt et al., 2012). This high variation may coincide with the present study since one of the four crosses

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exhibited lower survival (≤26.5% average survival in 2 replicate control tanks) than the

other three crosses (>75% average survival in respective 2 replicate control tanks). This

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variability in survival in the present study may be natural variation due to parentage or perhaps a combination of parentage and transient thermal stress. The latter is hypothesized because two instances during embryonic development at 8 dpf (48 hours

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at 14.0 oC) and 17 dpf (~24 hours at 14.5 oC) the exposure water exceeded 12.5 oC in the present study. Although no significant differences were observed in average survival

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between clothianidin treatments in this study, there was a large range in survival for all four crosses within a treatment. It is hypothesized that the different genetic compliments of the four crosses combined with the chemical stressor (i.e. clothianidin) and a transient

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thermal stress may have had additive effects on variability with respect to survival in the four genetic crosses in the clothianidin treatments. In addition to effect on the embryonic survival, parentage appears to influence

the timing and duration of hatching in sockeye salmon. In the present study, one of the three crosses (cross C) exhibited significantly delayed hatching based on the time to

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achieve 50% and 90% hatched and an increased duration of hatching compared to the two other crosses (cross A and B). Weaver Creek, BC sockeye salmon also exhibited genetic differences with respect to time 50% hatch and duration of hatching from 5% to 95% of hatch at both 14 and 16 oC (Burt et al., 2012). Furthermore, variation in fry wet mass and length between different families was also observed in the Weaver Creek study (Burt et al., 2012). This is in line with the present study using Pitt River, B.C. sockeye salmon crosses whereby cross B exhibited a significantly larger body size based on average body length and weight compared to cross A and C, and this may

indicate different genetics underlying body size or a growth rate differences between crosses. Collectively, the numerous cross-specific differences in development, size and survival in the three crosses tested in the present study strongly support the hypothesis that there is considerable genetic variation due to parentage in wild sockeye salmon. In addition, it is hypothesized that this genetic variation due to parentage also influences sockeye salmon’s response to environmental stressors (i.e. thermal or chemical), and

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this should be considered when testing the toxicity of contaminants in this species.

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Adverse Effects of Neonicotinoids in Sockeye Salmon During Early Development

These results coincide with shorter duration acute toxicity studies indicating that

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lethal concentrations in several fish species are 3 orders of magnitude higher, with LC50 values of ~100 mg/L. Additional studies in the literature also support no lethality due to

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sub-chronic or chronic exposure to another neonicotinoid, imidacloprid, at low mg/L concentrations in fish species tested to date. For example, in a 60-day toxicity test reported no effect on survival in rainbow trout (Oncorhynchus mykiss) at 19,000 µg/L

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imidacloprid (Canadian Council of Ministers of the Environment, 2007). Similarly, no significant reduction in survival at low level of exposure was reported in a 98-day imidacloprid exposure (1,300 – 20,000 µg/L) on newly fertilized rainbow trout embryos in

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a flow-through system (Anatra-Cordone and Durkin, 2005). Furthermore, exposure to even a higher concentration, 320,000 µg/L of imidacloprid, did not induce any toxicity in zebrafish during larval development (Tišler et al., 2009). Similar to the findings for imidacloprid in other teleosts, the results of the present study suggest that clothianidin at concentrations equal to or below 150 μg/L during early life stage development of

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sockeye salmon would be unlikely to cause high direct mortality in sockeye salmon. Growth parameters including body mass, length and condition factor reflect

overall fish health, and are the culmination of several complicated and not fully understood molecular and biochemical processes that can be influenced by environmental contaminants (Morado et al., 2017). For example, a xenobiotic can induce reactive oxygen species causing oxidative imbalance, and this triggers metabolically expensive detoxification processes in fish, ultimately depleting the limited energy reserve

in the yolk sac and results in growth inhibition (Zhu et al., 2008). Although the present study did not show any clothianidin concentration related effects on salmon growth, a reduction in growth has been observed in studies with the neonicotinoid imidacloprid in other fish species. For instance, a 98-day imidacloprid exposure to newly fertilized rainbow trout eggs (Oncorhynchus mykiss) in a flow-through system caused a significant decrease in body length at 36 and 60 days post-hatch, while significant reduction of body weight occurred at 60 days post-hatch (Anatra-Cordone and Durkin, 2005). The LOAEC and NOAEC were determined to be 19,000 µg/L and 9,800 µg/L imidacloprid,

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respectively (Anatra-Cordone and Durkin, 2005). Also, a 60-day study reported a LOAEC of 2,300 µg/L imidacloprid for growth inhibition for rainbow trout from the

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fertilized egg developmental stage to the juvenile stage (Canadian Council of Ministers of the Environment, 2007). Another 7-day toxicity test, conducted for derivation of

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Canadian water quality guidelines for the protection of aquatic life determined the LOAEC for growth inhibition in larval inland silverside (Menidia beryllina) to be 34,000 µg/L (Canadian Council of Ministers of the Environment, 2007). In contrast to

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these aforementioned higher concentration imidacloprid studies in rainbow trout and silverside, in a 3-month field study on Japanese medaka effects on fish growth occurred

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at a lower concentration of imidacloprid, and within the range of clothianidin tested in the present sockeye salmon study (Sánchez-Bayo and Goka, 2005). Specifically, Japanese medaka were exposed to imidacloprid treated rice paddy fields, whereby water

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concentrations measured in the first week in field water was 33-240 µg/L and an average of 0.75 µg/L imidacloprid was measured in the following months. Adult Japanese medaka were released and reproduced in the water collected from the rice paddy fields, hence developing embryos were exposed throughout spawning and development (Sánchez-Bayo and Goka, 2005). Imidacloprid was shown to significantly reduce the

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weight/length ratio (i.e. condition factor) in Japanese medaka fry compared to the control (Sánchez-Bayo and Goka, 2005). This reduction in condition factor after imidacloprid exposure may reflect decreased energy reserves which has been shown to be associated with this endpoint along with an inhibition of fish growth (Morado et al., 2017; Zhu et al. 2008; Sánchez-Bayo and Goka 2005). In contrast to these findings in Japanese medaka fry with imidacloprid exposure, no significant effects of clothianidin on morphometrics was detected in the present study. Whether this is due to unique modes

of action for these two different neonicotinoids, species specific effects, high variation in body size between genetic crosses in the present study requires further study. This study revealed that a concentration of 0.15 µg/L clothianidin significantly elevated whole body 17β-estradiol concentrations up to 4.7-fold in sockeye salmon swim-up fry, with no significant concomitant increase in testosterone. This is the first study to report a change in sex steroid hormones in a fish after a neonicotinoid exposure and these data show a non-monotonic concentration-response curve, which is

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commonly reported for chemicals impacting endocrine system endpoints. Although no

studies in non-mammalian vertebrates on sex steroid levels after neonicotinoid exposure

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have been reported, studies in mammals in vitro and in vivo thus far show effects on

reproductive endocrine axis hormones, but a consistent pattern is not evident and may

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be due to various factors (i.e. tissue type, duration, exposure route, species, etc.). For example, a 24-hour exposure of imidacloprid (LOAEL= 10 µM), thiamethoxam (LOAEL= 0.1 µM) and thiacloprid (LOAEL= 0.1 µM) significantly increased aromatase activity and

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estradiol biosynthesis in co-culture of human adrenocortical carcinoma (H295R) cells and BeWo human choriocarcinoma cells (Caron-beaudoin et al., 2017). Likewise,

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another 24-hour in vitro exposure of thiacloprid and thiamethoxam induced aromatase activity at concentrations of 0.1-1.0 µM with decreasing catalytic activity at a higher concentration in H295R cells and exhibited a non-monotonic dose-response curve

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(Caron-Beaudoin et al., 2016). One in vivo study showed that imidacloprid reduced testosterone levels and epididymis mass in male rats exposed to 0.5 mg/kg for 90 days (Bal et al., 2012). Bal et al. (2012) hypothesized that in these rats imidacloprid, an AchR agonist, somehow interacts with the gonadotropin-releasing hormone (GnRH) from the hypothalamus to reduce the release of luteinizing hormone (LH) and follicle-stimulating hormone (FSH) from anterior pituitary causing the inhibition of testosterone secretion in

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Leydig cells and spermatogenesis in testes (Bal et al., 2012). In a female rat study sex steroids were not measured, however, oral administration of 20 mg/kg/day imidacloprid for 90 days significantly induced hormone imbalance (LH, FSH and progesterone), ovarian oxidative damage, decreased weight and patho-morphological changes (Kapoor et al., 2011). Additional studies are required to determine if clothianidin is directly or indirectly influencing 17β-estradiol levels at the conversion of testosterone to 17βestradiol via the enzyme aromatase, decreased degradation pathways of this sex steroid hormone in fish or at higher levels within the hypothalamic-pituitary-gonad axis in

sockeye and other teleosts. Regardless of the mechanism of action resulting in increased 17β-estradiol observed in the present study, these results combined with the evidence from mammalian studies suggests further inquiry is warranted to examine the adverse effects of clothianidin on the reproductive endocrine axis. Indeed, xenoestrogens in particular that mimic endogenous estrogens have been shown to cause population level outcomes in fish. One of the most definitive studies on a xenoestrogen involved chronic exposure of a fathead minnow population to 5-6 ng/L 17α-ethynylestradiol (a synthetic estrogenic hormone used in birth control pills) in an

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experimental lake and caused significantly increased intersex in males, ovarian follicle

degeneration in females and ultimately a population level collapse and near extinction of

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this species from the lake (Kidd et al., 2007). Recently it has been reported that at least

105 pesticides, including atrazine (a herbicide in maize and sugarcane crops), and DDT

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(a banned insecticide used mainly in 1950s), are identified as endocrine disruptors or chemicals that interfere with the normal function of the endocrine system (Mnif et al., 2011). To more fully assess clothianidin for endocrine disrupting activity, future studies

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should include testing clothianidin in the suite of the standardized test methods available for the Testing and Assessment of Endocrine Disrupters listed in the Organization for

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Economic Cooperation and Development Conceptual Framework (OECD, 2018). This is the first study to report the effects of environmentally relevant levels of a neonicotinoid pesticide, clothianidin, on gene expression in wild sockeye salmon. This

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study evaluated five target genes relevant to various biological processes/systems, including the stress endocrine axis (glucocorticoid receptor 2), reproductive endocrine axis (estrogen receptor α and β), xenobiotic detoxification (cytochrome P450 1A) and immune system (suppressor of cytokine signaling 3) in the liver of early life stage sockeye salmon after chronic waterborne clothianidin exposures. One of the five genes

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of interest, glucocorticoid receptor 2, decreased ~4 fold in the highest clothianidin exposure concentration tested. The glucocorticoid receptor 2 (GR2) is a key steroid hormone receptor that binds to glucocorticoids (i.e. cortisol) to regulate the stress response through modulation of metabolic, cardiovascular and immune function. Interestingly, several previous studies have demonstrated hypothalamic-pituitaryinterrenal axis (HPI) down-regulation resulting from negative feedback of cortisol mediated by GRs following either physical or chemical chronic stress (crowding, handling stress, water quality), or intraperitoneal implants of cortisol to mimic chronic

stress in salmonids (Barton, 2002; Madaro et al., 2015; Moltesen et al., 2016; Vijayan et al., 1990). Recently, Madaro et al. (2015) conducted a study that investigated the effects of exposing Atlantic salmon smolts to an unpredictable chronic stressor (UCS) for three weeks (Madaro et al., 2015). Smolts in the UCS group were stressed three times per day using a total of eight types of stressors given in a random and unpredictable order (Madaro et al., 2015). These stressors included both physical and chemical changes such as temperature shock, brief hypoxia, chasing, and noise disturbance (Madaro et al., 2015), and all stressors induced a down-regulation of both GR1 and GR2 mRNA

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levels in the preoptic area of the hypothalamus. This reduced expression of both GRs

also coincides with previous observations for UCS in zebrafish (Piato et al., 2011) and

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common carp (Cyprinus carpio) (Stolte et al., 2008). While both GRs are involved in the stress response pathway, the current dogma is that these two receptors have unique

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physiological functions that are yet to be fully elucidated (Bury et al., 2003; Bury et al., 2007; Stolte et al., 2006; Vijayan et al., 2003). Interestingly, in zebrafish and some salmonid species low levels of GRs may protect neurons from high levels of cortisol

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during chronic stress, as cells containing GRs are more prone to apoptosis by elevated levels of glucocorticoids (Moltesen et al., 2016; Piato et al., 2011; Sapolsky et al., 2000).

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The suppressive effect of chronic stress in the form of toxicant exposure (i.e. clothiandin) in the present study on GR2 mRNA levels in sockeye swim-up fry is hypothesized to be due to negative feedback of cortisol on the HPI axis. However, cortisol levels were not

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measured in this study, and therefore future studies measuring cortisol levels are necessary to test this hypothesis. Given that clothianidin can potentially interfere with neuron function, more specifically nAChRs in the nervous system and at the neuromuscular junction in vertebrates, and in light of studies linking reduced GR levels to reduced neuronal apoptotis, more focused studies on evaluating effects of clothianidin

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on the central and peripheral nervous system are recommended.

Conclusion The present study examined the adverse effects of clothianidin at low level, environmentally relevant concentrations on wild sockeye salmon and has several key findings. The first is that there is significant variation in size and development in wild caught salmon genetic crosses that should be taken into account when investigating

adverse effects of contaminants in this species. One of the limitations of the present study was a duplicate test vessel experimental design, and increasing replicates to improve statistical power in light of this significant natural variation between genetic crosses is recommended in the future. Nonetheless, exposures up to 150 μg/L clothianidin did not affect the survival, growth or external development of sockeye salmon in early life stages. However, this study did demonstrate an increase in whole body 17β-estradiol levels and decreased hepatic GR2 gene expression in a developing wild salmonid after chronic, low-level clothianidin exposure. Thus, future studies

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examining the potential disruption of endocrine system function or development in

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teleosts are warranted.

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Funding Source: National Contaminants Advisory Group, Fisheries and Oceans Canada

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Funding: This research was supported by the National Contaminants Advisory Group, Fisheries and Oceans Canada.

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Conflict of Interest Statement for following manuscript:

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There are no conflicts of interest.

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Figure 1 The average percent embryonic survival of sockeye salmon after chronic clothianidin exposure in A) glass tanks and B) gravel-bed incubation system. Salmon embryos were exposed to waterborne clothianidin from 1 hour post-fertilization to the eyed embryo stage. Percent survival to eyed embryonic developmental stage in duplicate test vessels was calculated based on taking the average number of surviving salmon at 28 dpf/total number of eggs seeded in two replicate test vessels. Means ± standard error are presented (n=2 test vessels per treatment with ~100 fish/genetic cross in each test vessel). There is no significant difference between treatments and different letters indicate differences between genetic crosses (2-factor-split-plot-analysis of variance followed by a Tukey’s post hoc test, p< 0.05).

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The average percent survival of sockeye salmon during waterborne clothianidin exposures initiated 1 hour post-fertilization to the swim-up fry developmental stage in flow through glass tanks. Percent of survival of swim-up fry in each glass tank was calculated based on number of surviving salmon 119 dpf /total number of eyed embryos added to each tank (~100 fish/genetic cross in each tank). Means of 2 replicate glass tanks ± standard error are presented (n=2). No significant difference in mean survival between any treatments or genetic crosses was observed (2-factor-split-plot-analysis of variance followed by a Tukey’s post hoc test, p< 0.05).

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Figure 2

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The average percent of sockeye salmon that hatched during clothianidin waterborne exposures initiated 1 hour post-fertilization to the swim-up fry developmental stage in flow-through glass tanks. Percent hatched in each glass tank was calculated based on number of eyed embryos hatched by 119 dpf/total number of eyed embryos added to each test vessel (~100 fish/genetic cross in each tank). Means of 2 replicate glass tanks ± standard error are presented (n=2). No significant effect of clothianidin concentration or genetic cross was observed (2-factor-split-plot-analysis of variance followed by a Tukey’s post hoc test, p< 0.05).

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Figure 3

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The percent of hatched of sockeye salmon over time during waterborne clothianidin exposures initiated 1 hour post-fertilization to the swim-up fry developmental stage. Boxes represent the average H10 (day indicating 10% of total number of embryos hatched), and H90 (day indicating 90% of the total embryos hatched) in duplicate glass tanks (n=2 tanks; ~100 fish/genetic cross in each tank). Horizontal lines within each box represent the median (H50, day indicating 50% of embryos hatched). The whiskers indicate the first (H0) and last hatch (H100) observed.

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Figure 4

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The average time required for sockeye salmon to reach A) 50% hatched, H50, B) 90% hatched, H90 and, C) H90 from H10 during clothianidin exposures initiated 1 hour post-fertilization to the swim-up fry developmental stage in glass tanks. Means ± standard errors of two replicate glass tanks are presented (n=2 tanks; ~100 fish/genetic cross in each tank). No effect due to clothianidin treatment was observed, but different letters indicate significant differences between genetic crosses (2-factor-split-plot-analysis of variance followed by a Tukey’s post hoc test, p< 0.05).

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Figure 5

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The average body length and weight of swim-up fry sockeye salmon after clothianidin waterborne exposures initiated 1 hour post-fertilization to the swim-up fry developmental stage in glass tanks and gravel-bed incubators. Means ± standard error of two test vessls are presented (n=2 tanks or gravel-bed incubators; ~100 fish/genetic cross in each test vessel). No significant differences due to clothianidin were observed, but different letters indicate significant differences between genetic crosses (2-factor-split-plot-analysis of variance followed by a Tukey’s post hoc test, p< 0.05).

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Figure 6

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Average whole body A) 17β-estradiol and B) testosterone concentrations in swim-up sockeye salmon fry from a single genetic cross after chronic clothianidin exposures in flow through glass tanks initiated1 hour postfertilization to the swim-up fry developmental stage. Hormones were measured via enzyme-linked immunosorbant assays. Means ± standard errors of two replicate glass tanks are presented (5 fish per tank). Different letters indicate significant differences between treatments (randomized complete block analysis of variance, followed by Tukey’s post-hoc test, p<0.05).

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Effects of waterborne clothianidin on sockeye salmon hepatic glucocorticoid receptor 2 gene expression after clothianidin exposures initiated 1 hour post-fertilization to the swim-up fry developmental stage. Box plots demonstrating the upper and lower quartiles (25%; whiskers) and the median values (horizontal black line within box) for each treatment group are presented. The number of individual fish livers analyzed per treatment ranged from 6 to 8 as follows: water control (n = 6), 0.15 μg/L (n = 8), 1.5 μg/L (n = 7), 15 μg/L (n = 6), and 150 μg/L (n = 7). Normalized gene expression was calculated based on the ΔΔCq method for relative quantitation of a target gene using three reference genes (CBA, GAPDH and EF1α). Different letters indicate significant differences between treatments (one-way analysis of variance followed by Tukey’s post-hoc test, P < 0.05).

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Zhu, X., Zhu, L., Lang, Y. and Chen, Y. 2008. Oxidative Stress and Growth Inhibition in the Freshwater Fish Carassius Auratus Induced by Chronic Exposure to Sublethal Fullerene Aggregates. Environmental Toxicology and Chemistry, 27(9): 1979–85. doi:10.1897/07-573.1.

Table 1: Primer sets used to examine the effects of clothianidin on various target genes in sockeye salmon swim-up fry. National Center for Biotechnology and Information accession identifiers, primer sequences (5’ to 3’), annealing temperature (Tm), efficiency of primer pair (%E), and goodness of fit of linear regression for the relative standard curve (R2) are provided for reference genes CBA, GAPDH, and EF1α, and target genes GR2, CYP1A, ERα, ERβ2, and SOCS3 Primer Sequence (5’-3’)

Tm oC

Product size

%E

R2

CBA (cytoplasmic Betaactin)

FJ226383

Forward: GGACTTTGAGCAGGAGATGG Reverse: TCGTTGCCGATTGTGATG

57

96

99.8

0.984

GAPDH (Glyceraldehyde 3phosphate dehydrogenase)

FJ226385

Forward: CCATCACAGCCACACAGAAG Reverse: CCATTCAGCTCGGGGATAA

58

EF1α (Elongation factor 1α)

NP001117811.1

Forward: GGTGGTGTGGGTGAGTTTGA Reverse: CCAGAGTGTAGGCGAGGAGA

58

GR2 (Glucocorticoid receptor 2)

NM001124482

Forward: CTGGCTGATGACACTCTCCTG Reverse: CTGGCTTGGAGGTGGAGTTG

CYP1A (Cytochrome P450 1A)

FJ226380

ERα (Estrogen receptor α)

AJ242741

ERβ2 (Estrogen receptor β2)

DQ248229

139

0.900

79

99.5

0.976

56

70

91.4

0.781

Forward: GCCTGTGGTTGTTCTGAGTG Reverse: TCTTGCCGTCGTTGATGA

55

116

104.1

0.718

Forward: CCT GGA GAT GCT GGA CGG T Reverse: CCT GTG GAG GTG GTA GTG GT Forward: TTCTCCTCCACTATGTCCAGCCT Reverse: TCCAGGTGTCCGTTGACTGTT

57

102

112.6

0.997

57

62

107.0

0.996

Forward: GAG CAT CCA AGG TCA CAA TG Reverse: CAC TTT GTC ATG CCC ACT TC

57

126

94.2

0.988

lP

re

-p

ro

105.0

ur na NM001146168. 1

Jo

SOCS3 (Suppressor of Cytokine Signaling 3)

of

Accession

Target Gene

Table 2 Predicted and measured concentrations of clothianidin in glass tanks and gravel-bed flumes in flow through systems during chronic clothianidin exposures of sockeye salmon from 1 hour post-fertilization to the swim-up fry developmental stage. Predicted values were based on measured pesticide and water flows into the flow through systems every 48 hours throughout the exposure period. Measured concentrations were based on 1 sampling event collected from 1 of 2 replicate test vessels. Predicted concentration (μg/L)

*GBF, gravel-bed flume; GT, glass tank

Jo

re

of

N 43 43 43 43 43 43 43 43 43 43 44 43 44 44 44 44 44 44 44 44

Measured concentration (μg/L) 0 114 0.230 28.9 1.08

ro

Maximum 0.00 227 0.198 20.8 1.78 18.7 0.00 185 0.172 1.90 0.226 189 24.9 3.21 0.00 19.8 0.00 245 2.04 0.269

-p

Minimum 0.00 136 0.124 0.766 1.30 11.3 0.00 132 0.125 1.17 0.122 110 9.83 0.950 0.00 11.2 0.00 99.7 0.787 0.0712

lP

Average±SE 0.00±0.00 158±2.25 0.157±0.002 15.8±0.42 1.57±0.02 15.6±0.27 0.00±0.00 154±1.75 0.151±0.002 1.53±0.023 0.156±0.0033 146±2.46 16.3±0.47 1.65±0.06 0.00±0.00 15.2±0.29 0.00±0.00 161±4.35 1.53±0.037 0.159±0.0056

ur na

Nominal Exposure Concentration System* (μg/L) 0 150 0.15 15 1.5 GBF 15 0 150 0.15 1.50 0.15 150 15 1.5 0 GT 15 0 150 1.5 0.15

15.0

5.36 0.00 165 0.750 0.140

Min*

Max *

Max

SD *

Temp (°C) * pH

48 40

12.1 7.14

9.6 6.90

15.2 8.42

1.33 0.13

45 39

12.3 7.14

10.2 7.00

15.0 8.16

1.23 0.14

Dissolved Oxygen mg/L) Conductivity (μS/cm)

48

9.9

8.00

11.08

0.70

45

10.10

8.83

10.74

0.42

47

26.2

20.9

31.4

4.14

44

Ammonia (mg/L)

16

0.00

0.00

0.00

0.00

16

26.2

20.8

31.5

3.94

0.00

0.00

0.00

0.00

-p

Glass tanks n Mean

of

SD *

Gravel-bed flumes n Mean* Min

ro

Table 3 Water quality monitoring summary from 4-month waterborne clothianidin (0, 0.15, 1.5, 15 and 150 μg/L) exposure experiment of sockeye salmon 1 hour postfertilization to the swim-up fry developmental stage. Two flow through systems using either glass tanks or flumes filled with gravel to mimic natural salmonid rearing substrate were used as test vessels. Water quality was monitored every 48 hours for all parameters, except ammonia (monitored every 2 weeks).

re

*SD, standard deviation; Min, miniumum; Max, maximium; Temp, temperature

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Table 4 The accumulated thermal units (ATUs) for sockeye salmon at the eyed embryo, hatching and swim-up developmental stages during chronic exposures to 0, 0.15, 1.5, 15 and 150 μg/L clothianidin initiated 1-hour post-fertilization to the swim-up fry stage in glass tanks and gravel-bed flumes. Temperature was recorded every 48 hours, and the average of this temperature was used to calculate ATU (average daily temperature x days post fertilization) in both systems. Developmental Stage

Day post-fertilization

Eyed embryo stage (92-100% developed eyes in the control/ Day of gravel burial)

ATU (oC-days) Gravel-bed Flume

28

340.8

346.2

Alevin (Day of 1st hatched observed in glass tanks)

49

596.4

605.8

Swim-up fry (Day of 1st swimup observed in gravel-bed flume)

88

1071.2

1088.1

Jo

Glass tank