Keywords
climate matching, heterogeneous landscape, invasive species, lag phase,P. muralis , range expansion
Introduction
The global rise in the number of species introduced to regions beyond their native range via human-mediated translocation shows no sign of reaching saturation point (Seebens et al., 2017). While many species fail to establish or have little negative effect following introduction, a subset of these do spread and can have significant impact on economies, human health, native biodiversity and ecosystem services (Kolar & Lodge, 2001; Vila et al., 2010; Keller, Geist, Jeschke, & Kühn, 2011). The severity of potential negative impacts (e.g., extirpation and extinction of native species) are such that invasive non-native species (INNS) are justifiably regarded as one of the most significant threats to biodiversity worldwide (Genovesi, 2009; Simberloff et al., 2013).
For non-native species to become widespread and potentially damaging following introduction to new regions, introduced populations must negotiate the three stages of an introduction–establishment–invasion continuum (Blackburn et al., 2011). Evaluation of the likelihood of a species to be transported, establish, and to spread, as well as the potential for having ecological, economical, and health impacts, forms the basis of ‘invasive’ risk assessment for alien species (Bacher et al., 2018; Roy et al., 2019). Although it has been argued that the term ‘invasive’ doesn’t always necessarily equate with a species’ negative impact (Ricciardi & Cohen, 2007), the potential for damaging effects inherently increases as introduced species increase in population size and spread across novel landscapes, thus affecting broader areas and more ecological communities (Crooks, 2005). As such, there is great interest in understanding patterns and rates of expansion of introduced species, and the environmental factors which limit their distributions (Gallien, Munkemuller, Albert, Boulangeat, & Thuiller, 2010; Roy et al., 2019).
Following introduction and successful establishment beyond native ranges, species can further expand their range through local dispersal processes and/or by jump dispersal events that may be human-mediated (i.e., deliberate or accidental movement of individuals between habitats) or natural dispersal (i.e., long distance flight in birds) (Suarez, Holway, & Case, 2001; Ingenloff et al., 2017). Invading species typically exhibit several phases in the rate of spread. Firstly, there is an initial establishment phase where rate of spread is slow. Secondly, an expansion phase typified by increasing rates of spread, and finally, a saturation phase when available space is occupied and expansion rates reach a plateau (Arim, Abades, Neill, Lima, & Marquet, 2006).
A suite of factors influence patterns and rates of range expansion during these phases: propagule size, dispersal mode, matching of physiological and ecological traits of invading species with environmental conditions at the receptor site, vital rates (births and deaths) species interactions, evolutionary processes, spatial heterogeneity and temporal variability (reviewed in Hastings et al., 2005). Furthermore, our ability to assess and predict the temporal dynamics of invasions is often complicated by the phenomenon of lag phases, wherein an introduced species remains at low population levels in the early stages of establishment for a protracted period of time before the sudden onset of rapid range expansion (see Crooks (2005) for review of the causes of temporal lags at all stages in the invasion process). Introduced populations of the northern Racoon (Procyon lotor ), for example, remained small for a number of years following introduction to Europe before a population explosion in the mid 1990’s (Salgado, 2018). Similarly, landscape complexity can result in temporal and spatial patterns of invasion dynamics that deviate from classic theory of symmetrical, radial expansion from a central point (diffusion theory) (Skellam, 1951; Shigesada, Kawasaki, & Takeda, 1995; Kinezaki, Kawasaki, & Shigesada, 2010). The effects of landscape heterogeneity on patterns and rates of expansion have been demonstrated in the quick colonization of areas of suitable habitat in the early stages of the American mink (Neovison vison ) invasion, compared to uptake in areas of low habitat suitability in Scotland (Fraser et al., 2015), and the fluctuating rates in range expansion of Cane toad (Rhinella marina ) in response to changing environmental conditions in newly invaded areas of Australia (Urban, Phillips, Skelly, & Shine, 2008). Consideration of dispersal processes across heterogeneous landscapes is therefore central to predicting potential for range expansion during the invasion process (Travis, Harris, Park, & Bullock, 2011; Bocedi, Zurell, Reineking, & Travis, 2014; Grayson & Johnson, 2018). The development of platforms for spatially explicit individual-based modelling (Bocedi, Zurell, et al., 2014; Samson et al., 2017) have enabled the nested interactions between dispersal, landscape properties, and population dynamics to be considered in predicting species distributions, increasing the ecological realism of range expansion models (Andrew & Ustin, 2010; Ferrari, Preisser, & Fitzpatrick, 2014; Mang, Essl, Moser, Kleinbauer, & Dullinger, 2018; Hunter-Ayad & Hassall, 2020).
In this study, we determine the potential for range expansion of the non-native common wall lizard (Podarcis muralis ) in the UK.Podarcis muralis has a long history of introductions beyond its native range which covers most of Western and Southern Europe (Gassert et al., 2013). Many of these introductions have extended its range throughout continental Europe (Schulte, Gassert, Geniez, Veith, & Hochkirch, 2012; Wirga & Majtyka, 2015; Šandera, 2017), but the species also has several populations established in the New World, both in the United States (R. M. Brown, Gist, & Taylor, 1995) and Canada (Allan, Prelypchan, & Gregory, 2006). Introduced to Vancouver Island, British Columbia, in 1970, the species persisted in isolated populations until 2006, but has since spread with alarming speed due to jump dispersal (human mediated) and natural radial dispersal of 40-70 meters a year in urban areas (Engelstoft, Robinson, Fraser, & Hanke, 2020){Engelstoft, 2020 #2101}.
To date there is no empirical evidence of negative ecological impacts ofP. muralis introductions in the UK, and there is mixed social perception and opinion towards the species’ presence (Williams, Dunn, Quinn, & Hassall, 2019). However, suspected declines in native lizards through interference and/or exploitation contest have been reported following introductions of P. muralis , to both Germany (Münch, 2001; Kühnis & Schmocker, 2008; Schulte, 2009) and the UK (Mole, 2010).
There have been multiple introduction events of P. muralis to the UK both as deliberate releases of captive animals and as cargo stowaways, with some extant populations having been established on the UK mainland as early as the 1970s (Michaelides, While, Bell, & Uller, 2013). More recent introductions (1980s onwards) have mostly arisen from movement of individuals from already established populations (secondary introduction) or captive-bred animals, rather than directly sourced from the native range (Michaelides, While, Zajac, & Uller, 2015). The UK populations represent the species at the northern extent of its range, with sites having markedly different climatic conditions compared with the native range. For example, air temperatures during the main activity season in populations in England are 5–10°C lower than their source regions in Tuscany and western France (While et al., 2015).
We investigated the potential for range expansion of P. muralisin the UK with models highlighting different (but complementary) parameters likely to influence spread at two spatial scales. Firstly, since long distance jump dispersal via translocation is important in facilitating spread of this species, we aimed to predict the national extent of the area potentially available for further colonisation by running a species distribution model (SDM) based on climatic suitability at these northern extremes. As has been speculated elsewhere, the ability to survive cold winters is likely limiting to the spread of introduced Podarcis populations (Burke, Hussain, Storey, & Storey, 2002). We therefore hypothesised that latitudinal clines in climate would restrict the area available for northward expansion ofP. muralis via long distance human-assisted translocation in the UK. Second, to make predictions of population growth and dispersal patterns, as well as identify environmental features important to range expansion at a local level, we took a hybrid model approach combining SDMs, informed by variables characterising 10 local landscapes (i.e., microclimate, proximity to geographic features, and habitat type), with a high resolution (15 x15 m) spatiotemporal individual based model (IBM) simulating local population and dispersal dynamics. We expected that landscape characteristics (i.e., configuration and connectivity of suitable habitat patches), would result in asymmetrical patterns of predicted dispersal within populations, which in turn, would result in spatial and temporal variance in patterns of population growth and range expansion between populations. These analyses allow us to investigate the proximate and ultimate barriers to spread, as well as simulating the potential for invasion lag in each population.