The Distribution and Abundance of Parasites in Aquatic Ecosystems in a Changing Climate: More than Just Temperature

Synopsis

Evaluation of the potential response of parasites of aquatic organisms to climate change illustrates the complexity of host–parasite relationships and the difficulty of making accurate predictions for these biological systems. In recent years, trematodes have proven to be a useful model to evaluate potential effects of climate change on host–parasite systems. In the first part of this article, I review and summarize results from the recent use of trematodes and specifically their early life cycle stages in testing effects of temperature and other climate-driven variables on life history traits and host–parasite interactions. However, metazoan parasites in aquatic systems respond directly to changes in temperature and also to changes in other climate-driven abiotic parameters that are mediated directly on the parasite or indirectly through changes in the distribution and abundance of their hosts. In addition, though most research to date has focused on the effects of temperature, it is imperative to explore effects of precipitation, eutrophication, acidification, water levels and flow rates, habitat loss and fragmentation, extreme weather, and other forms of anthropogenic interference on the distribution of both hosts and parasites, as these biotic and abiotic factors and stressors do not operate independently of climate. In the second part of this article, the effects of some of these factors derived from our own field studies, as well as other investigations both in the laboratory and the field, on the distribution, abundance, and community structure of parasites in aquatic ecosystems will be reviewed and discussed.

Introduction

Since 2007, there has been an upsurge in studies exploring the linkages between infectious disease and climate change ( Altizer et al. 2013 ). In the numerous reviews on the subject, several emphasize aquatic ecosystems ( Harvell et al. 1999 , 2002 ; Hayes et al. 2001 ; Marcogliese 2001 , 2008 ; Marcos-López et al. 2010 ; Lõhmus and Björklund 2015 ). The majority of studies have focused on chytridiomycosis in amphibians ( Lips et al. 2008 ; Rohr et al. 2008 , 2011 ; Blaustein et al. 2010 ; Rohr and Raffel 2010 ; Hof et al. 2011 ) and infectious diseases in corals ( Rosenberg and Ben-Haim 2002 ; Bruno et al. 2007 ; Harvell et al. 2007 ; Sokolow 2009 ; Ruiz-Moreno et al. 2012 ), both of which are the focus of conservation concerns. However, in recent years trematodes and particularly their cercariae have proven to be useful experimental models to evaluate potential effects of temperature and other climate-driven variables on host–parasite systems. While the effects of climate change on disease are usually framed primarily in terms of increases in temperature and changes in precipitation, Marcogliese (2001) emphasized the indirect effects of climate change on eutrophication, stratification, ice cover, acidification, water levels, flow rates, oceanic currents, ultraviolet-light penetration, and extreme weather. Furthermore, other forms of anthropogenic interference, such as habitat loss and fragmentation, pollution, species introductions, and hydrological modifications interact with climate change ( Marcogliese 2008 ). Herein, I follow up on these general themes, with a focus on metazoan parasites in aquatic ecosystems, concentrating on developments during the last 10 years. More specifically, this article discusses first, the frequent and increasing use of freshwater and marine trematodes as experimental systems to test effects of temperature increases due to climate change, emphasizing the need to account for life cycle complexity, and second, recent studies from our laboratory and others on parasite populations and communities related to potential indirect effects of climate change.

Trematodes of medical significance have been studied in the laboratory for decades, and new host–parasite experimental systems appear on a regular basis as different host–parasite interactions become the subject of study and manipulation. In recent years, trematodes have been very popular models for parasitological studies investigating climate change and its effects on host–parasite interactions. Typically, they have an obligate molluscan first intermediate host, often a snail, in which the larval parasite asexually produces tens, hundreds, or even thousands of cercariae. Molluscan first intermediate hosts release cercariae into the environment, where they infect the next host in the life cycle. This portion of the life cycle has proven relatively amenable to experimentation and manipulation. The high number of cercariae released, referred to as productivity, permits sufficient sample sizes and replicability in experimental design. The effects of temperature have been, and continue to be, the primary focus of studies on potential impacts of climate change on these parasites and others.

Empirical studies on temperature

In recent years, numerous experimental studies have examined the effects of temperature on cercarial release, survival and activity of different trematode species, often with unexpected results. Among infected estuarine snails, cercarial shedding (productivity) increased with increasing temperature in certain trematode species and decreased in others even though some infected the same host species ( Koprivnikar and Poulin 2009a , 2009b ). In a meta-analyses of published experimental results, Poulin (2006) found that cercarial emergence increased up to 200-fold following a 10°C rise in temperature. The increases in cercarial productivity were much higher than would be predicted from those estimated from the temperature sensitivity of most physiological processes (Q10 of 2–3). In another meta-analysis, while accounting for a minimum emergence temperature threshold and acclimation, emergence was unaffected by temperature over relatively wide optimal ranges (20–25 °C) for mid- and low-latitude trematode species, but declined at higher temperatures ( Morley and Lewis 2013 ). These results contradict those of Poulin (2006) , who did not account for acclimation of infected snails to experimental temperatures used in laboratory studies. Shedding also varied differentially among sites, salinities, and water levels, depending on the parasite species ( Koprivnikar and Poulin 2009a , 2009b ). In these studies, however, snails were not acclimated and results should be interpreted cautiously ( Morley and Lewis 2013 ).

Cercarial survival and activity also were differentially affected by increasing temperature and salinity among trematode species that infected the same species of molluscan host ( Koprivnikar et al. 2010 ). Given that climate change will impact salinity and water levels as well as temperature ( Marcogliese 2001 ), and that different parasites respond differentially to various environmental parameters, it becomes difficult to predict effects of a warmer climate on parasite transmission based solely on temperature, even within the same host species or among host populations. To complicate climate change scenarios even further, other studies have suggested that cercarial production and activity can vary with the strain or genotype of molluscan host ( Morley 2011 ; Morley and Lewis 2013 ; Berkhout et al. 2014 ).

However, physiological tolerances and thermal optima vary not only among parasite species but among life stages within species ( Marcogliese 2001 ; Barber et al. 2016 ). For trematode miracidia, which hatch from eggs and infect molluscs, a meta-analysis of published experimental results showed that all trematode species examined exhibited distinct zones of thermostability and temperature had only limited effect on survival and metabolism. Miracidia (which hatch from eggs and infect snails) proved more thermally resistant than cercariae, with climate change apparently having limited impact on this segment of a trematode’s life cycle ( Morley 2012 ). Furthermore, another meta-analysis on published experimental data did not detect an effect of temperature on parasite development within the molluscan host ( Morley and Lewis 2013 ).

Thus, in order to better predict effects of climate change, it is important to consider the net effect of temperature on a parasite’s entire life cycle. Logistically, this is a daunting task, but a few pioneering studies have examined the effects of temperature on different life history traits and parasite–host interactions at different stages of a life cycle in trematodes. One series of studies was carried out on an estuarine trematode ( Maritrema novaezealandensis ) that infects snails, amphipods, and birds ( Studer et al. 2010 ; Studer and Poulin 2013 ), and another examined a freshwater trematode ( Ribeiroia ondatrae ) that infects snails, frogs, and birds ( Paull and Johnson 2011 ; Paull et al. 2012 ; Paull et al. 2015 ; Altman et al. 2016 ). In these cases, the authors focused on parasite–host interactions within ectothermic hosts in which, presumably, temperature effects should be most pronounced ( Marcogliese 2008 ; but see Morley and Lewis 2014a ). A summary of experimental results is presented in Table 1 and Table 2 for the two different host–parasite systems. Productivity and rate processes of many life history traits (e.g., host growth and fecundity, parasite development and activity) simply increased with temperature ( Paull and Johnson 2011 ; Paull et al. 2012 ), while others (e.g., parasite survival and infectivity) decreased ( Paull et al. 2012 ). A common pattern observed in different host–parasite systems was that cercarial survival decreased with increasing temperature ( Studer et al. 2010 ; Morley 2011 ; Paull et al. 2012 ; Studer and Poulin 2014 ). However, Morley’s (2011) meta-analysis of published laboratory studies demonstrated that cercariae tended to display zones of thermostability, resulting in minimal effects of climate change on survival under the predicted temperature-change scenarios on the order of 2–4 °C ( Morley 2011 ).

Table 1.

Relative strength of temperature and temperature variability on different parasite and host traits in the life cycle of the intertidal trematode Maritrema novaezealandensis in its gastropod first intermediate host ( Zeacumantus subcarinatus ) and its amphipod second intermediate host ( Paracalliope novizealandiae )

Trait	Temperature (°C)
	16	20	25	30	34
Cercarial productivity	+	++	+++	++
Cercarial survival	+++	++	++	+	+
Cercarial infection rate	+	++	+++	++	+
Amphipod susceptibility	+	+	+	+	+
Amphipod survival	+++	++	++	–	–
Metacercarial development	+	++	+++	–	–
	Temperature variability (°C)
	15	15 + 5	15 + 10	15 + 15	15 + hw
Cercarial productivity	+	++	++	++	+++
Cercarial infection rate	+	++	+	+	++
Amphipod survival	+	+	+	+	+

Notes : Modified from Studer et al. (2010 ) with information from Studer and Poulin (2013 ). hw = simulated heatwave. A + sign indicates an increase in the measured parameter, with the number of + signs indicating the relative strength of the effect. A – sign indicates a negative effect.

Table 2.

Relative strength of temperature on different parasite and host traits in the life cycle of the freshwater trematode Ribeiroia ondatrae in its gastropod first intermediate host ( Planorbella trivolvis ) and its anuran second intermediate host ( Pseudacris regilla ).

Trait	Temperature (°C)
	13	20	26
Parasite egg development	–	+	++
Snail growth	+	++	+++
Infected snail growth	+	++	++ ++
Snail fecundity	+	++	+++
Infected snail fecundity	+	++	+
Snail survival	+	–	–
Cercarial development	=	+	+++
Cercarial survival	+++	++	+
Cercarial penetration of frog	+	++	+++
Cercarial establishment in frog	+++	++	+
Metacercarial numbers in frog	++	++	+
Frog growth	+	++	+++
Frog survival	+	+	+
Frog malformations	++	+++	+

Notes : Information from Paull and Johnson (2011) and Paull et al. (2012) . A + sign indicates an increase in the measured parameter, with the number of + signs indicating the relative strength of the effect. A – sign indicates a negative effect.

Some parasite life history traits peaked then decreased with increasing temperatures, again a pattern observed among many trematode species ( Studer et al. 2010 ; Morley 2011 ; Studer and Poulin 2014 ; Morley and Lewis 2015 ). Life history traits of the first intermediate molluscan hosts and the second intermediate hosts also were affected differentially by temperature and depended on whether they were infected or not, with effects also differing between first and second intermediate hosts for the same parasite ( Studer et al. 2010 ; Paull and Johnson 2011 ; Paull et al. 2012 ; Morley and Lewis 2015 ). Indeed, meta-analyses derived from published experimental studies of infectivity of trematode miracidia, cercariae, and metacercariae demonstrated that miracidial and cercarial infectivity increases to a plateau then decreases with higher temperature, while that of metacercaria decreases ( Studer and Poulin 2014 for cercaria only; Morley and Lewis 2015 ).

Pathology in some invertebrate intermediate hosts increases with temperature while in other hosts it may peak at intermediate temperatures ( Studer et al. 2010 ; Paull and Johnson 2011 ; Paull et al. 2012 ). Warming due to climate change can alter the timing of infection at key and possibly vulnerable points in the hosts’ susceptibility to infection ( Paull and Johnson 2011 ; Paull et al. 2012 ; see also discussion on host–parasite synchronicity in Marcogliese 2001 ).

Experiments repeated under conditions of varying temperatures, which better simulate natural conditions in wetlands and intertidal habitats, differentially affected both trematode and host life history traits compared with those carried out under constant temperature conditions ( Studer and Poulin 2013 ; Paull et al. 2015 ). Importantly, parasites and hosts responded differently to the same acclimation temperature ( Altman et al. 2016 ). Thermal acclimation led to different temperature effects on the resistance of the anuran host to infection by R. ondatrae cercariae and the clearance of established metacercariae by the host. Thus, the host appears to use separate immune mechanisms to combat different parasite stages of infection (cercarial establishment vs. metacercarial persistence) that respond differently to temperature variability ( Altman et al. 2016 ). Furthermore, it is important to consider the magnitude and the frequency of temperature variability in predicting parasite dynamics under a changing climate ( Paull et al. 2015 ). Taken as a whole, these studies on trematodes demonstrate that temperature has differential effects on parasites, their hosts and host–parasite interactions that vary among host and parasite species and the life history stages of the parasite itself.

Empirical studies on other factors

Extreme weather

Extreme weather provides the opportunity to examine disease and parasite dynamics in novel conditions that may mimic future climate change ( Studer and Poulin 2014 ). Prolonged heat waves can disrupt parasite transmission and increase parasite-induced mortality, a realistic possibility under proposed climate change scenarios ( Studer et al. 2010 ; Studer and Poulin 2013 ; see also Marcogliese 2001 for a review of effects of thermal effluents on host–parasite systems). Effects of interactions between parasitism and temperature can extend beyond the host–parasite system in question to the entire community of which they are a part. Parasite-induced mortality of the amphipod Corophium volutator , an important intermediate host for microphallid trematodes and a known ecosystem engineer, influences the structure of mud flat invertebrate communities in the Wadden Sea ( Poulin and Mouritsen 2006 ; Larsen and Mouritsen 2014 ). A heatwave led to mass mortalities of mud snails ( Hydrobia ulvae ) and amphipods ( C. volutator ), apparently caused by infection with two microphallid trematodes ( Microphallus claviformis , Maritrema subdolum ) ( Poulin and Mouritsen 2006 ; Larsen and Mouritsen 2014 ). Removal of the amphipod caused wholesale physical transformation of the mud flats, affecting primary productivity and benthic community structure, and depleting the major food resource for migrating shorebirds (reviewed in Marcogliese 2008 ). Subsequent laboratory experiments (reviewed in Poulin and Mouritsen 2006 ; Marcogliese 2008 ) and mesocosm studies ( Larsen and Mouritsen 2014 ) supported the links between high temperature, cercarial productivity, parasite-induced amphipod mortality and changes in benthic community structure.

Other examples include studies of parasite communities of the horn snail Cerithidea pliculosa , an important first intermediate host for a number of trematodes in the Yucantan before and after Hurricane Isidore in September 2001 ( Aguirre-Macedo et al. 2011 ), fish parasite populations along the Mississippi coast following Hurricane Katrina in August 2005 ( Overstreet 2007 ), and the impacts on infections of wild and domestic animals in terrestrial, freshwater, and coastal marine ecosystems in the UK of a major 16-month drought that occurred in 1976 (reviewed in Morley and Lewis 2014b ). A general result of the hurricanes mentioned above is that parasite populations and communities crashed for a period of time ranging from months to years before becoming re-established ( Overstreet 2007 ; Aguirre-Macedo et al. 2011 ). Given that most metazoan parasites have complex life cycles that include numerous hosts occurring at different trophic levels and often rely on trophic interactions for transmission, their presence suggests they are good indicators of food web structure, ecosystem degradation, and recovery ( Marcogliese and Cone 1997 ; Marcogliese 2004 , 2005 ; Overstreet 2007 ; Aguirre-Macedo et al. 2011 ; see also Okamura 2016 ).

Ocean acidification

Despite ocean acidification and disease in corals and other organisms fueling increasing concerns in recent years, little information exists on potential effects on metazoan parasites, unlike studies on those in freshwater fishes (reviewed in Marcogliese 2001 , 2005 ). However, given that molluscs are sensitive to low pH ( Marcogliese 2001 ), it is reasonable to presume that trematodes can be concomitantly affected. A series of studies has examined the effects of pH on different trematode and host life history characteristics using different species of trematodes, some of which infect the same gastropod intermediate host ( Hartland et al. 2015 ; MacLeod and Poulin 2015a ). Cercariae of all trematode species displayed reduced longevity under acidified conditions, as did metacercariae of one that encysts on the surface of invertebrate transport hosts ( MacLeod and Poulin 2015a ). However effects of acidification and parasitism on growth and length of the mud snail ( Zeacumantus subcarinatus ) varied differentially with trematode species ( MacLeod and Poulin 2015b ). In a parallel experiment, net transmission success of cercariae of M. novaezealandensis to its second intermediate host, the amphipod Paracalliope novizealandiae , was highest at the lowest pH tested through differential effects on hosts and parasites ( Hartland et al. 2015 ). Again, the complexity of direct and indirect effects of climate change on hosts, parasites, and their interactions renders predictions difficult.

Eutrophication

Many freshwater systems are expected to experience eutrophication as a result of climate change, thus affecting parasite populations and communities ( Marcogliese 2001 ). For example, outbreaks of salmonid proliferative kidney disease, caused by the myxozoan Tetracapsuloides bryosalmonae , known to occur with increasing temperature, similarly may be promoted by eutrophication ( Okamura et al. 2011 ; Okamura 2016 ). Both the parasite and its bryozoan intermediate host proliferated at higher temperature and nutrient levels ( Okamura et al. 2011 ). In studies in the St. Lawrence River on effects of sewage effluents on the myxozoan communities in spottail shiners ( Notropis hudsonius ), prevalence and mean infracommunity richness (mean number of species per host fish) increased downstream of the effluent outflow ( Marcogliese and Cone 2001 ; Marcogliese et al. 2009 ). The authors proposed that this was not a toxicological, but rather an ecological effect that results from the huge organic input from the effluent outflow into the St. Lawrence River, in turn promoting oligochaete population growth in the riverine effluent waters, oligochaetes being obligate alternate hosts for numerous myxozoan species ( Marcogliese and Cone 2001 ). They subsequently found oligochaete density to be higher downstream of Montreal than it was upstream, supporting the hypothesis that the increase in infection was due to greater oligochaete densities in waters receiving effluents ( Marcogliese et al. 2009 ). Thus, the local distribution of myxozoan parasites in spottail shiners in the St. Lawrence River appeared to be affected by eutrophication.

Blanar et al. (2011) examined the helminth and arthropod parasite fauna of mummichogs ( Fundulus heteroclitus ) in the Miramichi and Bouctouche rivers, New Brunswick as a follow-up to a prior series of ecotoxicological studies on effects of pulp and paper mill effluents on the health of these fish in the Miramichi. Analyses of parasite community structure incorporating water quality, contaminants in the fish, and host physiological data showed that the parasite community structure was partially influenced by upstream–downstream differences in eutrophication as indicated by fecal coliforms, and not chemical contaminants, from municipal and pulp mill effluents ( Blanar et al. 2011 ). While this study and those above illustrate the effects of eutrophication from anthropogenic sources, they also demonstrate potential indirect consequences of climate change on fish parasite populations and communities in a diverse array of parasites ( Marcogliese 2001 ).

Water levels and precipitation

Under conditions of high water flow, parasite diversity is seen to decrease (reviewed in Marcogliese 2001 ). Strong flow rates may transport small, motile, free-living stages of parasites out of a system. Indeed, high current flow reduced oligochaete abundance, transmission to fish, and subsequent transmission to oligochaetes in an experimental study on Myxobolus cerebralis , the causative agent of whirling disease ( Hallett and Bartholomew 2008 ). Faster currents also reduced infections of Ceratomyxa shasta in polychaete invertebrate hosts ( Manayunkia speciosa ) and rainbow trout ( Oncorhynchus mykiss ) ( Bjork and Bartholomew 2009 ).

Water levels in many rivers dictate water flow, with high levels meaning high flow rates. Following up on the St. Lawrence River study above, when samples within the river were pooled by year over the entire 5-year sampling period, mean infracommunity species richness of myxozoans in the spottail shiner was negatively correlated with mean water levels measured during the month prior to fish sampling ( Marcogliese et al. 2009 ). Water levels in the St. Lawrence River are determined by precipitation and snow melt. Thus, myxozoan diversity over the length of the river is driven by climate at a regional scale, an effect which is superimposed over the eutrophication effect discussed above, fueled by municipal pollution at a local scale ( Marcogliese et al. 2009 ).

In a 2-year study on spottial shiners in the Richelieu River, a tributary of the St. Lawrence River receiving municipal and agricultural effluents, the helminth and arthropod parasite communities in fish were drastically different between the years, annual variation being much greater than site variation within years ( Marcogliese et al. 2016 ). Indeed, overall species richness at each site in 2003 was 15–18 species, declining to 9–13 species at the four sites in 2004. In contrast, on the scale of an individual fish, total parasite abundance and the mean infracommunity richness were lower in 2003 at each site. These results could not be linked to any measurements of water quality or pollution. However, precipitation in the month prior to sampling was 40% higher in 2003 compared with 2004, and was the only significant parameter consistently explaining community structure ( Marcogliese et al. 2016 ). The Richelieu River is artificially regulated in the section sampled, with water levels completely stable. Thus, increased precipitation means higher flow rates. The high flow rates in 2003 resulted in reduced diversity and abundance of parasites in individual fish (see Marcogliese 2001 ), although the reasons for the higher overall species richness in 2003 remain unclear.

The St. Lawrence River downstream from Montreal consists of two water masses, one of high conductivity (green water) emanating from the Great Lakes to the west, and the other of low conductivity (brown water) draining from the Ottawa River to the north. The relative extent of the two water masses depends on the respective water volume from each drainage, which is a function of temperature and precipitation. Conductivity, which differentiates the green and brown waters, was a major determinant of helminth parasite community structure ( Marcogliese et al. 2006 ). As with the myxozoan communities discussed above, hydrology of the St. Lawrence River, which is hugely impacted by climate change, exerts a major influence on parasite community structure in fish at a large spatial scale.

These studies provide examples illustrating further potential indirect consequences of climate change on parasites of aquatic organisms. It should be noted that further anthropogenic regulation and intercession to mitigate effects of climate change may amplify other potential indirect consequences of climate change, with concomitant implications for flow rates, eutrophication, and other physicochemical parameters in aquatic ecosystems, thus further affecting parasitism in aquatic ecosystems ( Marcogliese 2001 ).

Salinity

Sea level rise due to climate change will lead to salt water intrusion in coastal rivers, with consequences for the parasite communities of fishes and other aquatic organisms in coastal and low-lying areas ( Marcogliese 2008 ). Mummichogs are euryhaline and extremely common along the eastern seaboard of North America between the Gulf of Mexico and the Gulf of St. Lawrence, Canada. In the Blanar et al. (2011) study discussed above, the downstream sites in both the Miramichi and Bouctouche rivers were brackish, while upstream sites were fresh, establishing a salinity gradient in both rivers. The most prominent pattern in community structure was a distinct difference between upstream and downstream sites. Mean infracommunity richness was significantly lower at both downstream sites ( Blanar et al. 2011 ). In their study, salinity appears to have a stronger effect on parasite community structure than eutrophication ( Blanar et al. 2011 ), even though both are indirect effects of a changing climate. As with the myxozoan study in spottail shiners, large-scale patterns in water quality over large sections of the rivers, in this case a salinity gradient, appear to superimpose their effects on parasite communities in fishes over local effects such as eutrophication due to pollution.

Landscape changes

While habitat fragmentation typically is a indirect consequence of human actions, it likely will be exacerbated by a changing climate ( Moore et al. 1997 ). Recent studies have demonstrated that parasite community richness in frogs is linked to landscape structure. Prevalence of the trematode parasite Alaria sp. in leopard frogs ( Lithobates pipiens ) was positively associated with surrounding forest cover surrounding ponds in southern Ontario ( Koprivnikar et al. 2006 ). Alaria sp. matures in mammals such as foxes, commonly found in forested areas. Mean infracommunity richness was negatively correlated with urban and agricultural area, also in leopard frogs, in southwestern Quebec ( King et al. 2007 ). In bullfrogs ( Lithobates catesbeianus ) from southwestern Quebec, the number of parasite species found at a locality was negatively correlated with agricultural area, while the mean infracommunity richness was positively related to forest area surrounding ponds ( King et al. 2010 ). In these studies, parasites with avian and mammalian definitive hosts tended to be either lower in abundance or absent in agricultural and urban landscapes. King et al. (2007, 2010 ) suggested that habitat fragmentation due to anthropogenic landscape development restricts access to wetlands by these hosts. In another large-scale study, the abundance, richness, and diversity of helminths in leopard frogs were associated with greater amounts of forested and woody wetland habitats, shorter distances between woody wetlands, and smaller-sized open water patches in surrounding landscapes to anuran habitats ( Schotthoefer et al. 2011 ). In terms of landscape structure, local scale (1 km) was considered the most important for larval trematode abundances, whereas both local and regional (10 km) landscape variables appeared most significant for adult helminths ( Schotthoefer et al. 2011 ). With habitat fragmentation, wetland food webs and aquatic–terrestrial interactions become disrupted, followed by the reduction or elimination of parasite species. Wetland area may also be compromised by human adaptation to climate change. Furthermore, the studies presented above illustrate the importance of spatial scale in determining effects of a changing climate on aquatic ecosystems.

Concluding remarks

Studies solely focusing on a single parasite species and even single stages within parasite life cycles can lead to erroneous generalizations and predictions on the effects of climate change on parasite transmission. Future research should attempt to encompass a variety of parasite species experimentally and in situ to observe a range of potential effects. Furthermore, transmission must be examined using as many hosts as possible in a parasite’s life cycle to determine overall net effects of global warming ( Barber et al 2016 ). Acclimation effects and temperature variability also must be considered ( Morley and Lewis 2013 ; Raffel et al. 2013 , 2015 ; Rohr et al. 2013 ; Studer and Poulin 2013 , 2014 ; Paull et al. 2015 ; Altman et al. 2016 ).

The effects of climate change, however, are not limited to increasing temperature and changes in precipitation. Indirect effects will impact numerous biotic and abiotic variables in aquatic ecosystems. Effects often are nonlinear due to species interactions, confounding variables, temperature thresholds, and context dependencies ( Marcogliese 2008 ; Rohr et al. 2011 ; Lõhmus and Björklund 2015 ). Indeed, some environmental factors, especially those that are anthropogenic, will overshadow or mask effects of climate change ( Marcogliese 2008 ; Lafferty 2009 ). Studies of changes in parasite populations and communities in response to environmental change should consider the broad picture at the ecosystem level, and also account for the scale of any observed impacts. Indeed, there are increasing appeals for the interdisciplinary integration of community ecology, biodiversity, and physiology into disease and climate change research ( Martin et al. 2010 ; Altizer et al. 2013 ; Molnár et al. 2013 ; Rohr et al. 2013 ).

Acknowledgments

I thank Chris Boyko and Jason Williams for organizing the symposium “Parasites and Pests in Motion: Biology, Biodiversity and Climate Change” and extending an invitation to participate as a guest speaker. Special thanks go to Drs Bruno Pernet, Art Woods, Jane Cook, and one anonymous reviewer whose comments substantially improved the article. Many of the studies mentioned in Canada and done by the paper’s author have been supported via Environment and Climate Change Canada’s St. Lawrence Action Plan.

Funding

Studies on the St. Lawrence and Richelieu rivers were funded by Environment and Climate Change Canada's St. Lawrence Action Plan.

Author notes

From the symposium “Parasites and Pests in Motion: Biology, Biodiversity, and Climate Change” presented at the annual meeting of the Society for Integrative and Comparative Biology, January 3–7, 2016 in Portland, Oregon.