This study provides important information on the effects of historically and presently used therapeutants, drugs, and pesticides on an economically and ecologically important salmonid species that can migrate past aquaculture sites during or following treatment of sea lice in Atlantic salmon in areas such as the Broughton Archipelago, BC, an area of active aquaculture (Krkošek et al., 2006; Jones and Hargreaves 2007) and may be at risk from exposure to chemotherapeutants.
All 5 chemotherapeutants were lethal to pink salmon fry in the concentration range tested; lethal concentrations varyied by orders of magnitude. The chemicals in order of most to least toxic were DM > CP > AZ > EB > HP in both water and sediment exposures. Pink fry avoided HP, EB and AZ and were not attracted to any of the chemicals at the concentrations tested. AZ, CP and DM altered the olfactory ability that was both concentration and time-dependent. Swimming performance was also affected by all chemicals except EB, and again, concentrations resulting in effects thresholds varied by orders of magnitude.
Organisms rely on constituent chemical defense mechanisms to avoid the potential toxic effects of foreign compound exposure (Tierney et al., 2016). Ideally, behavioural avoidance acts to limit exposures to toxic substances by sensing the substance and moving into a cleaner environment. The avoidance/attraction responses to each of the 5 chemotherapeutants were chemical- and concentration-dependent. No attraction behaviour was exhibited for any of the chemicals, and pink salmon did not avoid either CP or DM at any concentration tested. Hydrogen peroxide initiated limited avoidance in fish at concentrations in the 50–80 mg/L range, emamectin benzoate resulted in more avoidance at concentrations > 300 µg/L, but avoidance was most pronounced to AZ occurring at water concentrations as low as 50 µg/L.
There are many examples of responses to pesticides including avoidance (e.g. acrolein; [Folmar, 1976]; aschlorpyrifos [Hansen et al., 1972]; 2,4-D [Tierney et al., 2011]), and attraction (e.g. bentazone; (Saglio et al., 2001). Avoidance behaviour cannot be extrapolated for all compounds within a class or for all fish species; for example, fenitrothion was avoided by goldfish (Scherer, 1975) and medaka (Hidaka and Tatsukawa, 1989) but sheepshead minnow (Cyprinodon variegatus) did not avoid malathion or carbaryl formulations (Hansen, 1969). Avoidance has been shown for other organophosphates like AZ including malathion (in G. affinis; Hansen et al., 1972) and parathion (in G. affinis; Kynard et al., 1974). There have been limited studies on the avoidance of these compounds in water or when associated with sediments; some evidence exists showing that AZ and EB provoke some level of avoidance behaviour in marine organisms. Under continuous exposure to AZ, juvenile American lobsters (Homarus americanus) exhibited an avoidance response (exiting shelters) with increasing water AZ concentrations, however, at concentrations used by aquaculture operations (100 µg/L and short exposure times), avoidance responses and effects in this species were not seen (Abgrall et al., 1999). Naïve and chronically pre-exposed E. estuarius (marine amphipod) or Neresis virens (marine polychaete) placed into sediment containing 0.5 to 200 µg/kg of EB in avoidance assay chambers, showed no significant differences in the proportions found on the non-seeded/uncontaminated side of test chambers (Woof and Kennedy 2021).
The calculated toxicological parameters for acute lethality in the 48-h water and 10-d sediment exposure tests for pink salmon show that toxicity trends were similar regardless of the exposure media and that toxicity occurred within the range tested for each chemical. DM was consistently the most toxic to pinks (LC50 values 1 µg/L and 1 µg/kg in water and sediment, respectively) with values which are similar to those reported for fish (juvenile starry flounder [Platichthys stellatus], adult threespine stickleback [Gasterosteus aculeatus], and adult tidepool sculpin [Oligocottus maculosus]) (500–870 ng/L and 510 ng/kg [in Strachan and Kennedy 2021] and approximately 20-fold less sensitive than crustaceans (Burridge et al., 2014b; Fairchild et al., 2010). Few other comparable studies with DM exist; a study with Atlantic salmon supports this sensitivity range (Sievers et al., 1995). CP was the next most acutely lethal chemical (LC50 values of 5 µg/L and 97 µg/kg in water and sediment, respectively) with values in the range of those previously reported (summarized in Clark et al., 1989 and Haya 1989; Ernst et al., 2001; Strachan and Kennedy 2021). Comparatively, crustaceans are approximately 3–10 fold more sensitive to CP than fish species; LC50 values for crustaceans range from 0.005 µg/L (Clark et al., 1989) to 0.82 µg/L (Strachan and Kennedy 2021). AZ was the next most toxic chemotherapeutant to pink salmon in water (LC50 value 80 µg/L), this is approximately 10-fold less toxic than reported for other marine fish species (Strachan and Kennedy 2021). This was within the range of toxicity values have been reported for a variety of crustaceans (1.03 µg/L to 191 µg/L) (Burridge et al., 1999; Burridge et al., 2014b). EB was the second least toxic compound to pink fry (LC50 values 1090 µg/L and 2100 µg/kg in water and sediment, respectively). Comparable values are 96 h LC50s in the range of 200–1300 µg/L for rainbow trout, bluegill sunfish (Lepomis macrochirus), and the Sheepshead minnow (Cyprinodon variegatus) (McHenery and Mackie 1999; Lumaret et al., 2012; Chukwedebe et al., 1996) and marine fish species (Strachan and Kennedy 2021). HP was the least toxic of all chemicals to both pink salmon fry (227 mg/L) with similar values to other marine species (Strachan and Kennedy 2021). Similar relative insensitivity of fish to HP has been widely reported (Burridge et al., 2014b; Kiemer and Black 1997).
The results for survival in the sediment exposure tests (with EB, CP and DM) for pink salmon fry exhibited similar toxicity trends with DM being the most toxic, followed by CP and then EB and were similar to those reported for adult tidepool sculpin [Oligocottus maculosus]) (Strachan and Kennedy 2021). 10-d LC50 values in the present study indicate that these 3 compounds, which due to low log Kow values, partition to sediments relative to the water phase, and appear to be bioavailable to pink fry which are considered a pelagic species. Pink salmon fry feed mainly on planktonic and epibenthic prey; in one study, between 38 and 51 % of the diet comprised epibenthic prey (Godin 1981). Kaczynski et al., 1973 also reported the predominant occurrence of epibenthic prey in the diets of pink salmon fry in littoral areas of the marine environment; this feeding strategy suggests that there exists an exposure pathway and potential bioavailability of sediment-associated contaminants for pink salmon in their early marine life stages.
Chemical information from the environment is received by the olfactory and gustatory systems in fishes and the relayed information can be critical to many activities including food location, predator avoidance, mating, kin discrimination, and particular to salmonids, migration and homing behaviours. Although the underlying mechanisms may vary, xenobiotics can impair olfactory function by gross anatomical alteration or by inhibiting key specific molecules, resulting is aberrant or dysfunctional behaviours to naturally occurring chemical stimuli.
The olfactory responsiveness of pink fry to a food extract was examined following exposure to sublethal concentrations of chemotherapeutants in water or sediments for varying time periods. Control pink salmon showed a typical positive response to the food odorant used in the test system. Olfactory systems function as important screening systems for both the respiratory and the gastrointestinal systems, and the classification of odors into the food or non-food category is of eminent survival value (Boesveldt et al., 2010). Concentration-dependent decreases in olfactory responsiveness was seen after 48-h AZ, CP and DM exposures at values lower than those causing acute mortality (20 µg/L v. 80 µg/L, 3 µg/L v. 5.1 µg/L, and 500 ng/L v. 980 µg/L [LOEC for olfactory effects v. LC50 value). Interestingly, pinks avoided AZ without prior exposure that may be protective, however, following a 48-h exposure to AZ and olfactory inhibition, it is unclear whether this avoidance response would still occur. Exposure to chemotherapeutants in sediments only resulted in olfactory inhibition with CP. Longer exposures to low concentrations of chemotherapeutants (that did not result in olfactory inhibition at 48 h) increased the potential for olfactory dysfunction suggesting that the threshold for inhibition is likely much lower and a function of exposure duration. Longer exposure durations are not likely to occur with water exposures, as models predict rapid dilution of chemotherapeutants released following treatment (ref) even though half lives in seawater can be long (e.g. AZ 13 d, CP 20 d, DM 18 d [Strachan and Kennedy 2021]); however, sediment-bound chemicals can be available for uptake for long periods due to their long half-lives (e.g. CP 560 d, DM 45 d, EB 230 d [Strachan and Kennedy 2021]) and limited dilution under farms.
Several classes of pesticides affect fish olfactory responses including carbamates, organophosphates and triazine herbicides (Tierney et al., 2010). Other organophosphates like AZ that have shown olfactory inhibition through behavioural measures include diazinon (O. tshawytscha; Scholtz et al., 2000), parathion (C. auratus; Rand et al., 1975) or olfactory sensory neuron impairment including diazinon (Moore and Waring 1998) and chlorpyrifos (O. kisutch; Sandahl et al., 2004). As in this study, alterations in olfaction were seen with CP exposure in Salmo salar (Moore and Waring 2001).
In most species of fish, swimming performance is a main determinant of survival and strongly influences the ability of a fish to obtain food, find mates, avoid unfavourable conditions, and migrate (Plaut 2001), and the Ucrit test has been used as an ecologically relevant definitive test for rover–predator teleosts with direct application to assessing their Darwinian fitness.
In control fish, Ucrit values ranged from 5.0 to 6.3 BL/s which are similar to the average swimming speeds determined in a study comparing critical swimming speed and maximal swimming speed in pink salmon fry (2.8 g), where mean swimming speeds ranged from 4.54 to 5.2 BL/s (Nendick et al., 2009). Concentration-dependent decreases in Ucrit were seen following exposure to HP [threshold 100 mg/L], AZ [10 µg/L], CP [2 µg/L] and DM [200 ng/L].
It is well established that exposure to a variety of contaminants including metals (Waiwood and Beamish 1978; Beaumont et al., 1995; Taylor et al., 2000; Rajotte and Couture 2002), petroleum (Kennedy and Farrell 2006; Alderman et al., 2020), pesticides (Little et al., 1990; Mackinnon and Farrell 1992; Nikl and Farrell 1993), and other contaminants (Howard 1975; Wood et al., 1996) can alter swimming performance in teleosts. This includes effects on critical swimming speed following exposure to organophosphate (OP) pesticides ([Cyprinodon variegatus] Cripe et al., 1984, [Salvelinus fontinalis] Peterson, 1974, [O. kisutch] Tierney et al., 2007; [Oreochromis niloticus] McKenzie et al., 2017)) such as AZ. The outcomes of OP exposure on locomotor activity appears to be two-fold; hyperactivity or decreases in swimming activity or ability (Tierney et al., 2007). For example, increased swimming activity was seen in eastern rainbow fish (Melanotaenia duboulayi) exposed to profenofos (Kumar and Chapman 1998) and in goldfish (Carassius auratus) following carbofuran exposure (Bretaud et al., 2001) In contrast, Coho salmon displayed a concentration-dependant decrease in swimming activity rather than hyperactivity following exposure to chlorpyrifos (Tierney et al., 2007). (Little et al., 1990) reported an incremental decrease in swimming activity in rainbow trout exposed to methyl parathion. The OP pesticide trichlorfon caused prolonged impairments to swimming performance in O. niloticus, but individuals varied widely in their relative sensitivity to the pesticide (McKenzie et al., 2017).
Exposure to pyrethroids has shown contradictory results with respect to swim performance. Goulding et al. (2013) showed no effects of permethrin exposure on the swimming performance in juvenile rainbow trout, results contrary to another study showing Ucrit declines following permethrin exposure in the same species (Kumaraguru and Beamish 1983). The discrepancy was attributed to differences in the size of fish used as it has been shown that permethrin toxicity in trout is inversely related to mass (Kumaraguru and Beamish 1981,1983,1986). Reductions in Ucrit were seen following deltamethrin exposure under the same conditions (Goulding et al., 2013). Effects on swimming performance were explained by considering the sublethal toxicity generally described for pyrethroids that include muscle tremors, and rapid and erratic swimming (Glickman and Lech, 1982; Haya, 1989; Velíšek et al., 2007; Werner and Moran, 2008). It was suggested that these manifestations of toxicity interrupted the constant gait required for prolonged swimming tests, forcing fish to transition to burst swimming earlier, which is unsustainable at the step durations used in Ucrit protocols (Farrell, 2008). In a study with zebrafish, no effects on swimming performance were seen following exposure to DM at 2 µg/L (Strungaru et al., 2019). DM caused behavioural effects including rapid swimming, loss of balance, aggressiveness and increases in the surface activity in brown trout (Salmo trutta) at concentrations as low as 1 µg/L (Karatas et al., 2019). Killifish (Jenynsia multidentata) exposed to cypermethrin at (4 µg/L) showed a decrease in swimming activity and an increase in the time spent at the bottom of test tanks (Bonansea et al., 2016).
Although not directly applicable to water exposures, an i.p dose of EB administration (3 d prior) caused dose-dependent decreases in swimming performance in juvenile rainbow trout at 5 mg/kg, an unlikely internal dose to be achieved in the present study. Three different swimming outcomes were examined and at these higher doses, Ucrit, burst swimming and schooling were affected (Kennedy and Tierney 2014). Although the contribution of glutamate-gated chloride ion channels and GABA-neurosynaptic transmission in the CNS to fish swimming performance is unknown, signs of avermectin toxicity (Katharios et al., 2001) suggests that increased accumulation of this neurotoxin in the brain of fish leads to altered parameters related to swimming. Avermectins inhibit signal transmission at GABA-gated and glutamate-gated chloride channels by binding GABA receptors, which leads to hyperpolarization of the neuronal cells (MSD, 1988). Data from the present study indicates that altering GABA transmission affects swimming performance and behaviour.
The postulated toxic modes of action of each chemotherapeutant provide clarity in the effects observed. EB, CP, DM and AZ are all neurotoxic agents and target specific biomolecules, however HPs mechanism of action is general with multiple targets.
Organophosphates (e.g. AZ) inhibit acetylcholinesterase (AChE) which hydrolyzes acetylcholine (ACh) in cholinergic neuropathways leading to ACh accumulation and repeated post-synaptic action potentials and insensitivity to further signalling; this translates as convulsions, twitching, agitation, and eventual partial or complete paralysis (Couillard and Burridge, 2014; Fulton and Key, 2001; Xuereb et al., 2009). AChE-impairing pesticides inhibit both olfaction (Sandahl et al., 2004) and muscle performance (Tierney et al., 2007).
In coho salmon, chlorpyrifos-induced declines in Ucrit performance are associated with reduced AChE activity in slow-twitch aerobic muscle and compromised neuromuscular coordination (Tierney et al., 2007). OPs are also reported to influence both the metabolism and cardiorespiratory physiology of fishes, reducing metabolic rate, heart rate, ventilatory activity and spontaneous swimming activity (da Silva et al., 1993, De Aguiar et al., 2004, Gehrke, 1988, Tryfonos et al., 2009). For example, trichlorfon exposure decreased the ability of Nile tilapia (Oreochromis niloticus) to regulate aerobic metabolism due to an impaired capacity to hyperventilate (Thomaz et al., 2009). Reduced exercise performance following OP exposure in fishes may therefore reflect direct effects on swimming muscles but also on the ability of the cardiorespiratory system to meet the oxygen demands of activity (Mackenzie et al., 2017). Impaired neuromuscular coordination presumably due to AChE inhibition (Tierney et al., 2007) suggests that higher oxygen consumption is required to power swimming at any given speed. (Guimarães et al., 2007) showed that trichlorfon exposure had no direct effect on respiratory metabolism, suggesting that the mechanism underlying Ucrit declines was a decline in swimming efficiency in O. niloticus
Pyrethroids such as cypermethrin and deltamethrin interact with voltage-gated Na+ channels, and with other ion channels including voltage-gated Cl- channels (Burr and Ray, 2004) leading to repetitive neuronal firing (Vijverberg and van den Bercken, 1990). Acute pyrethroid poisoning in fish manifests with symptoms that include muscle tremors, rapid and erratic swimming, loss of equilibrium, jaw spasms, gulping respiration, and lethargy (Werner and Moran, 2008).
The effects of individual pyrethroid exposure on the swimming performance of fish depends on compound-specific interactions with Na+ channels. For example, exposure of rainbow trout to 2 pyrethroids (permethrin and DM) resulted in reduced swim performance only with deltamethrin (Goulding et al., 2013). The divergent effects seen between these pyrethroids was attributed to their differing effects on peripheral motor neurons (Vijverberg et al., 1982), where permethrin causes repetitive action potential firing in response to stimulus (Vijverberg et al., 1982), while deltamethrin causes a frequency-dependent depression of action potentials due to the gradual depolarization of the cell membrane (Vijverberg and van den Bercken, 1979). Gradual depolarization occurs more rapidly at higher action potential frequencies such as those needed for elevated tail beat frequencies at faster swimming speeds (Goulding et al., 2013).
Another contributing factor to pyrethroid-associated reductions in Ucrit may be due to an increase in resting metabolic rate and aerobic capacity through increased energy requirements associated with physiological stress, tissue repair, and detoxification (Kumaraguru and Beamish, 1983, 1986; Balint et al., 1995; Philip and Anuradha, 1996; Velíšek et al., 2007). Histological studies have shown that acute DM exposure causes damage to gill, liver, and kidney in the common carp (Cyprinus caprio: Cengiz, 2006), and gill, liver, and gut tissue of the mosquito fish (Gambusia affinis: Cengiz and Unlu, 2006).
The precise mechanism(s) of action of EB is not fully understood. In invertebrates, avermectins are thought to interfere with GABA- and glutamate-gated Cl- channel receptors in nerve and muscle cells by stimulating the influx of chloride ions (Burridge et al., 2010; Lumaret et al., 2012; Benchaoui and McKellar 1996), leading to hyperpolarization of the neuronal cells (MSD, 1988) and subsequent paralysis (Reddy, 2012; Lumaret et al., 2012; Benchaoui and McKellar 1996). Avermectins can affect fish swimming and perhaps other systems that rely on glutamate-gated chloride ion channels and GABA-neurosynaptic transmission in the CNS; signs of avermectin toxicity (Katharios et al., 2001; Kennedy and Tierney 2014).
The mechanism of toxicity of HP is non-specific and not fully understood. As with other reactive oxygen species (ROS), high concentrations have been attributed to cell damage (Cabiscol et al., 2000), cell death (Saito et al., 2006) and carcinogenesis (Liou and Storz, 2010). In sea lice HP is believed to invoke mechanical paralysis through the formation of gas bubbles in the haemolymph (Burka et al., 1997; Bruno and Raynard, 1994; Grant 2002).
Each of the chemotherapeutants examined have different environmental fates and resulting water and sediment concentrations near farms due to modes of application, chemical characteristics, and environmental conditions. The highest exposure concentrations will occur in the water phase at the time of application and immediate release: Interox® Paramove 50 (AI HP), is applied in Canada as a bath treatment at 1500 mg L-1 for 20–30 min (PMRA 2014), and elsewhere at 1200–1800 mg L-1 for 30 min (Burridge, 2013; Burridge and Van Geest, 2014; Grant, 2002); Salmosan® (AI AZ) is applied as a bath treatment at 100 µg AZ L-1 for 30–60 min in well boats and tarps and at 150 µg AZ L-1 in skirt treatments (Burka et al., 1997; Van Geest et al., 2014; Burridge et al., 1999; Burridge et al., 2010; Grant, 2002; Haya et al., 2001; Haya et al., 2005); Excis® (AI CP) is applied as a bath treatment at 5 µg CP L-1 for 60 min (Burridge and Van Geest 2014); Alphamax® (AI DM) is applied as a bath treatment at 2–3 µg DM L-1 for 40 min (Burridge and Van Geest 2014). Following application, tides and currents strongly dictate the dilution and distribution of the chemical in the water column. For example, A field study in Atlantic Canada analyzed marine concentrations following the release of Salmosan®-treated baths using rhodamine dye as a tracer in an effort to characterise contaminant plume distribution (Ernst et al., 2014). Azamethiphos concentrations ranged from 1.1–11 µg/L and 0.2–1 µg/L approximately 1 m and 1000 m from application release areas, respectively, 2–3 h after treatment. A dispersion study utilizing simulated bath treatments with the pyrethroid cypermethrin found that the pesticide remained detectable for up to 5.5 h, and at distances up to 3000 m from the site of release; however, concentrations quickly diluted to levels 10–1000 times lower than the treatment concentration (Ernst et al., 2001). The following half-lives of the AIs in formulation have been reported in the literature: Salmosan (AI: AZ), 9–50 d (Mayor et al., 2008; Strachan and Kennedy 2021); Excis® (AI:CP), 19.8–80 d (Mayor et al., 2008; Strachan and Kennedy 2021); Alphamax® (AI:DM), 17.9–285 d (Benskin et al., 2016; Strachan and Kennedy 2021); Paramove® 50 (AI: HP), 8–19 d (Lyons et al., 2014; Strachan and Kennedy 2021); and Slice® (AI: EB), 164–175 d (Mayor et al., 2008) and > 400 d (Benskin et al., 2016). The Slice® formulation (AI EB) has an optimal prescribed dose of 50 µg kg-1 d-1 in feed for 7 d (Stone et al., 1999). The highest EB accumulation in sediment is generally within 25–60 m of the net pens but can be detected greater than 300 m away (Weston 1990, Schendel et al., 2004) as a result of seawater hydrodynamics (Tefler et al., 2006; DFO 2012). EMB concentrations in the water column in the vicinity of a salmon farm undergoing treatment have been found between 0.006–0.635 ng/L in Canada (DFO 2012). Some sampling events found concentrations as high as 140 and 366 µg/kg (McHenery and Mackie 1999, Boxall et al., 2002, Lalonde et al., 2012).
In areas where wild salmon migration routes co-exist with aquaculture, such as the coastal waters of Canada, juvenile wild salmon as small as 0.2 g (O. gorbuscha) may be exposed (Heard 1991). Exposure will be largely site-specific, influenced by local currents and tides. Each of the toxicity endpoints used here have different effects levels and can be compared to initial application rates (for HP, AZ, CP and DM) or highest measured sediment levels for EB (or CP and DM if available) to determine the potential for effects. For example, at initial chemotherapeutant application concentrations with bath applications, lethality of pink salmon would occur, however rapid dilution would reduce concentrations to non-lethal levels. The potential for sublethal effects on olfaction and swimming ability in pink salmon is a distinct possibility near farms. Dilution models in conjunction with toxicity effects levels should be used to make such risk determinations; this research highlights the importance of concentration-response data for regulators in this regard.