Sardiña, Paula, Beringer, Jason, Roche, Dylan and Thompson, Ross M. 2015, Temperature influences species interactions between a native and a globally invasive freshwater snail, Freshwater science, vol. 34, no. 3, pp. 933-941. This is the published version. ©2015, The Society for Freshwater Science Reproduced with the kind permission of the publisher, University of Chicago Press. Originally published online at: 10.1086/681639 Available from Deakin Research Online: http://hdl.handle.net/10536/DRO/DU:30084905 Temperature influences species interactions between a native and a globally invasive freshwater snail Paula Sardiña1,2,6, Jason Beringer3,7, Dylan Roche4,8, and Ross M. Thompson1,5,9 1School of Biological Sciences, Monash University, Victoria 3800 Australia 2Universidad Nacional del Sur-INBIOSUR, Consejo Nacional de Investigaciones Científicas y Técnicas, San Juan 670, 8000 Bahía Blanca, Argentina 3School of Geography and Environmental Science, Monash University, Victoria 3800 Australia 4School of Life and Environmental Sciences, Deakin University, Victoria 3125 Australia 5Institute for Applied Ecology, University of Canberra, Australian Capital Territory 2601 Australia Abstract: We experimentally assessed the interaction between a globally invasive snail (Potamopyrgus antipo- darum) and an Australian native snail (Austropyrgus angasi) under temperatures based on current (1990–2000, mean = 17.94–19.02°C) and future (2100, mean = 19.42–21.65°C) predicted conditions. Temperature treatments were scenarios identified from down-scaled global circulation models. Growth rates (mm/d) for juveniles and adults were measured at low (1000 individuals [ind]/m2) and high (20,000 ind/m2) densities in intraspecific and interspecific interaction trials under the 2 temperature regimes. Juveniles of both species grew at similar rates regardless of temperature and density. On the other hand, adults had dissimilar growth rates among treatments. Under current temperatures, P. antipodarum adults grew significantly faster than A. angasi adults when both species were kept at high densities in the interspecific treatment (interspecific-high) and faster than when they were kept at high densities but with conspecifics in the intraspecific treatment (intraspecific-high). However, we did not detect intra- or interspecific competition effects on either species. Thus, our results suggest that under current conditions, P. antipodarum gained from foraging with A. angasi (unidirectional facilitation effect). Under 2100 temperatures, the facilitative effect of A. angasi on P. antipodarum growth was not apparent, a result sug- gesting that the facilitation was directly related to the temperature conditions. Our research shows the impor- tance of considering future temperature conditions as a factor that could alter species interactions and poten- tially influence the ecological effects of invasive species. Key words: invasion, competition, climate change, Potamopyrgus antipodarum, down-scaled models, Austro- pyrgus angasi Biological invasions are a key cause of global biodiversity The recognized effects of invasion and climate change loss (Vermeij 1996, MEA 2005). Effects of invasions are on patterns of biodiversity make understanding the inter- highly variable but can be significant, and include extinc- actions between these 2 global threats critically impor- tions of native fauna, invasional ‘meltdowns’ caused by cas- tant. Will competition between invaders and native spe- cading effects on other native flora and fauna, and complex cies change under climate-change scenarios? Will some 2nd-order effects on disturbance regimes and ecosystem species that are not currently invasive become so? Will functioning (e.g., Vitousek et al. 1987, 1996, O’Dowd et al. some ecosystems become more or less robust to invaders 2003). Collectively, the effects of invasion on biodiversity when climate changes? The results of several studies have and ecosystem functioning are sufficient for biological in- suggested that climate change will increase invasion suc- vasion to be considered an element of global change (Vi- cess because invasive species are often favored by distur- tousek et al. 1997). Evidence of the negative effects of bio- bance (Dukes and Mooney 1999, Diez et al. 2012). Others logical invasions is already abundant, but how these effects have suggested that the general habitat requirements of will be manifested in the future is unclear. Predicting inva- many invaders predispose them to being successful in the sion dynamics entails significant uncertainty, and predict- face of changing climate (Qian and Ricklefs 2006). How- ing dynamics in an altered climate is likely to be even more ever, such generalizations may apply only to some groups uncertain (Dukes and Mooney 1999). of invaders. Responses of invaders to altered climates may E-mail addresses: [email protected]; [email protected]; [email protected]; [email protected] DOI: 10.1086/681639. Received 4 April 2014; Accepted 6 September 2014; Published online 6 April 2015. Freshwater Science. 2015. 34(3):933–941. © 2015 by The Society for Freshwater Science. 933 All use subject to University of Chicago Press Terms and Conditions (http://www.journals.uchicago.edu/t-and-c). 934 | Climate change and invasion P. Sardiña et al. depend, in part, on the origin of the invaders. Warming in history because of the prolonged aridification of the Aus- some regions may reduce the effects of some invaders of tralian continent over the last 10,000 y (Byrne et al. 2008). temperate origin, particularly if the invader is at its ther- Thus, the possibility exists that the outcome of competi- mal limit. For example, in freshwater systems, invasive sal- tion between P. antipodarum and Austropyrgus could dif- monids have known major effects on native populations, com- fer under predicted climate change. Climate models for munities, and ecosystem functions (e.g., Simon and Townsend southern Australia predict increases in mean temperatures 2003), but these invaders have relatively low upper lethal lim- and in the frequency of extreme summer temperatures its for temperature (Clark et al. 2001). (Hobday and Lough 2011). However, few investigators The expansion of invasive species into new habitats is have examined experimentally the effect of climate change broadly controlled by the physical environment (Moyle and on species interactions to predict invasion success. Light 1996) and species interactions (e.g., Hector et al. We experimentally assessed the effects of climate change 2001, Kennedy et al. 2002, Levine et al. 2004). Climate af- on the interaction between P. antipodarum and Austropyr- fects habitat occupancy, dispersal, and species interactions gus angasi using realistic temperature treatments based on (Dukes and Mooney 1999), so understanding the responses down-scaled global circulation models (Thompson et al. of invaders to climate change is important for predicting 2013a) for current (1990–2000) and future (2100) condi- and modeling invasion success and spread. Invasion mod- tions. We tested the effect of density-dependent intra- and eling has been carried out for many species (e.g., Loo et al. interspecific interactions on growth and mortality of both 2007, Steiner et al. 2008, Dullinger et al. 2009), but pre- species under each climate scenario. We hypothesized that dicted rates of spread and habitat occupancy to date have P. antipodarum would reach higher growth rates than A. been based on current climates. From a management per- angasi under current temperatures, which could result in a spective, understanding invasion probabilities and rates of competitive advantage of the invader over the native. We spread against a backdrop of changing climate remains a also predicted that A. angasi, which has evolved in an envi- challenging but necessary endeavor (Hellmann et al. 2008). ronment subjected to higher and more variable tempera- The New Zealand mud snail (Potamopyrgus antipoda- tures than those occurring in New Zealand, would perform rum) is a global invader. Invasive populations have been better than P. antipodarum under the 2100 scenario. established in the UK (since at least the 1850s), eastern Europe (1880s), southern Europe (1950s; Čejka et al. 2008), Australia (1970s; Ponder 1988), North America (1980s; Da- METHODS vidson et al. 2008), Asia (2000s), and the Middle East (2008; Study species and source populations Naser and Son 2009). In the western USA, P. antipodarum We used P. antipodarum and A. angasi (identities con- now dominates invertebrate stream secondary production firmed by W. F. Ponder, Australian Museum, Sydney). in many waterways at densities ranging from 10,000 to Both species are in the Tateidae family (formerly Hydro- 40,000/m2 (Hall et al. 2006). Invasion by this species has biidae) and are small (A. angasi: up to 4 mm, P. antipoda- been implicated in declines in the abundance and diversity rum: up to 6 mm), operculate snails that feed on bacterial, of native invertebrates in the USA (Kerans et al. 2005) and fungal, and algal films and are found primarily in fresh Europe (Strzelec 2005), in at least some cases via strong water. Potamopyrgus antipodarum can reproduce sexu- competition with other snails (Richards and Shinn 2004, ally and parthenogenically (Schreiber et al. 1998), whereas Strzelec 2005). Potamopyrgus antipodarum negatively af- only sexual reproduction has been reported for A. angasi fects the condition and productivity of native fish because (Miller et al. 1999). Both species are widespread in Victo-