INTRODUCTION
The ability of a non-native plant species to invade depends on resource
availability, natural enemies and the physical environment (Shea &
Chesson 2002). Accordingly, three hypotheses have been proposed to
explain invasion success (Jeschke 2014): (1) The
diversity-invasibility hypothesis (Elton 1958; Kennedy et al. 2002;
Levine et al. 2004) posits that species- and functionally-rich
communities limit establishment opportunities for invaders by reducing
access to resources; (2) The environmental-matching hypothesis(Mack et al. 2000; Shwartz et al. 2009) postulates that,
irrespective of local native biota, plant species will have higher
invasion success if introduced to environments similar to those in their
natural optimal range; (3) The invasion paradox (Levine &
D’Antonio 1999; Fridley et al. 2007) proposes that invasive species
spread over time through native metacommunities (communities linked by
dispersal) in a heterogeneous landscape, leading to the regional
coexistence of both native and non-native plant species. We subsequently
refer to these as invasion hypotheses 1, 2 and 3 ,
respectively.
Processes affecting biotic interactions and species distributions are
scale-dependent (McGill 2010). At the landscape scale, both habitat
quality and heterogeneity (defined here as the spatial and temporal
variation in biotic and abiotic conditions) within and between lakes may
facilitate species coexistence, regardless of their native status (Shea
& Chesson 2002; Davies et al. 2005). Habitat heterogeneity
therefore has the potential to concurrently increase invasion
probability whilst reducing invasion impact by promoting coexistence in
space and time that might be precluded in a more homogenous setting
(Melbourne et al. 2007; Clark et al. 2013). Thus,
biological assemblages in a given landscape may include species that are
extirpated at some sites but present at others through spatial and/or
temporal storage effects, provided there is sufficient regional
connectivity and spatiotemporal habitat heterogeneity (Melbourne et al.
2007).
In lake landscapes, macrophyte species distributions may be determined
by the hydrological network (Salgado et al. 2018a). For example,
communities in hydrologically-connected lakes may be particularly
influenced by repeated colonisation events; i.e. mass effects (Capers et
al. 2010) while local environmental factors may dictate community
structure through species sorting according to habitat optima in more
isolated lakes (Salgado et al. 2019). Habitat heterogeneity may
therefore be spatially autocorrelated with hydrological connectivity.
Although, there have been attempts to quantify how spatial
autocorrelation affects invasibility dynamics (e.g. Davies et al. 2005),
disentangling the simultaneous effects of abiotic factors, biodiversity
and spatially-structured dynamics on invasion processes in nature has
proved challenging (von Holle 2013; Nunez-Mir et al. 2017). There
is evidence that temporal variation in environmental stress factors and
dispersal-related mechanisms may promote the co-existence of both native
and non-native species in lake landscapes (Clark et al. 2013), and that
this may explain a positive correlation between native diversity and
exotic macrophyte abundance at any one time (Capers et al. 2007).
Systems comprising multiple sites that span environmental gradients
offer the possibility of more explicitly quantifying how spatial
autocorrelation affects invasibility dynamics and hence scope to
disentangle factors that contribute to regional coexistence.
Canadian pondweed (Elodea canadensis ) is considered amongst the
most widespread non-native plant species in Europe (Hussner 2012;
Nentwig et al. 2018). It was first recorded in Great Britain in
1836 (Simpson 1984). Thereafter it spread rapidly, reaching the maximum
extent of its distribution in Great Britain and Ireland by the middle of
the twentieth century (Simpson 1984). The rapid colonization and spread
of this species is commonly attributed to a high capacity for vegetative
propagation (Barrat-Segretain 1996) and tolerance of a broad range of
physical-chemical conditions, including low illumination, enabling
growth at a wide range of water depths and under eutrophication-induced
shade (Abernethy et al. 1996; Zehnsdorf et al. 2015). Once
established E. canadensis can quickly replace native submerged
macrophytes by forming a dense, closed canopy (Howard-Williams et al.
1987; de Winton & Clayton 1996; Zehnsdorf et al. 2015). Indeed, the
propensity for encountering E. canadensis in eutrophic isolated
temperate lakes has promoted the view that its spread and colonisation
across Britain and Ireland is attributable to environmental–matching;
i.e. hypothesis 2 above (O’Hare et al. 2012). However, few
studies have investigated the patterns of spread of E. canadensisover space and time [in a multivariate environmental context].
Here we examine the drivers of E. canadensis abundance in space
and time by focusing on the macrophyte-rich (> 40 submerged
and floating-leaved water plant species), and
environmentally-heterogeneous lowland shallow lake complex of the Upper
Lough Erne (ULE) system in Northern Ireland. Present-day and historical
data derived from surveys and sediment core analyses enable us to
address the following questions:
To what extent do diversity-related factors (invasion hypothesis
1 ), abiotic factors (invasion hypothesis 2 ) and a regional
gradient in lake habitat quality (invasion hypothesis 3 )
contribute to variation in E. canadensis abundance?
Does habitat heterogeneity (including spatial autocorrelation)per se promote both E. canadensis invasibility and
coexistence with native macrophyte communities in space and time
(decades-centuries)?
Answering these questions provides novel demonstration that both habitat
heterogeneity and habitat quality influence the coexistence of native
and non-native plants at landscape scales. We also provide evidence that
invasion resistance is linked to stressful environments and when high
native plant cover limits opportunities for E. canadensis to
proliferate.