Biological invasions represent significant economic and conservation challenges, though it is widely acknowledged that their impacts are often poorly documented and difficult to predict. In the Antarctic, one non-native vascular plant species is widespread and studies have shown negative impacts on native flora. Using field “common garden” experiments, we evaluate the competitive impact of the increasingly widespread invasive grass Poa annua on the only two native vascular species of Antarctica, the forb Colobanthus quitensis and the grass Deschampsia antarctica. We focus on interactions between these three plant species under current and a future, wetter, climate scenario, in terms of density of individuals. Our analysis demonstrates Poa annua has the potential to have negative impacts on the survival and growth of the native Antarctic vascular species. Under predicted future wetter conditions, C. quitensis communities will become more resistant to invasion, while those dominated by D. antarctica will become less resistant. Under a recently developed unified scheme for non-native species impacts, P. annua can be considered a species that can cause potentially moderate to major impacts in Antarctica. If current patterns of increased human pressure and regional climate change persist and mitigation action is not taken (i.e. reduction of propagule pressure and eradication or control measures), P. annua is likely to spread in Antarctica, especially in the Antarctic Peninsula region, with significant negative consequences for some of the most remote and pristine ecosystems worldwide. Tighter biosecurity across all operators in the region, improved surveillance for the species, and prompt, effective control actions will reduce these risks.

Invasions, Poa annua, Climate change, Competition, Antarctic ecosystems

Biological invasions represent significant conservation challenges. A focus on their early stages, such as the pathways of, and barriers to, invasion is valuable given a cost-efficacy continuum exists from prevention, through early detection and rapid response, to eradication (Simberloff et al. 2013). The latter is usually the most costly management option and may have unanticipated consequences (Zavaleta et al. 2001). Such efforts are predicated on the assumption many invasive non-native species either have or will have substantial impacts (Catford et al. 2012; Richardson and Ricciardi 2013). However, recent reviews have argued that our understanding of impacts often remains poor, and: (1) a more substantive evidence base is required to improve management of non-native species given criticisms that impacts are unproven or cannot be predicted (Hulme et al. 2013); (2) despite considerable advances in understanding the early stages of the invasion pathway, forecasts of the conditions under which substantial impacts will be realized remain weak (Ricciardi et al. 2013); (3) generalizations regarding the groups most likely to cause impact, the suites of traits associated with impact, and the environments most sensitive to impacts, remain uncommon (Pyšek et al. 2012).

A consistent theme across these reviews is that predictions of impact are needed because impact is often used to assess the need for early intervention, and specifically which species or groups of species, and under what conditions, should be the subject of such intervention. Much uncertainty remains, however, about the species that will have most impact and the conditions under which such impact will be realized (McGeoch et al. 2010; Ricciardi et al. 2013). Although data mining approaches and meta-analyses are beginning to reduce this uncertainty (Vilà et al. 2011; Greenslade and Convey 2012; Pyšek et al. 2012), further quantification and forecasts of impacts are essential to improve management efficacy and reduce the impacts of biological invasions (Simberloff et al. 2013; Richardson and Ricciardi 2013). Although these priorities apply to invasions generally, they are particularly significant in the context of the conservation challenges faced by protected areas. Understanding the correlates or determinants of non-native species richness variation can assist with managing risk in the early stages of the invasion process (Wilson et al. 2009; Foxcroft et al. 2011). However, for management decision-making about eradication or control, either after initial detection or later in the invasion process, understanding the potential for negative effects on the ecophysiological performance of native species and the community functioning is essential.

Antarctica (including the sub-Antarctic islands) is considered to include many examples of the world’s last remaining wilderness areas (Convey and Lebouvier 2009; Shaw et al. 2014). The continent itself is protected under the Protocol on Environmental Protection to the Antarctic Treaty (Tin et al. 2009), and parts of most of the surrounding sub-Antarctic islands are either formally protected under national legislation, have World Heritage status, or both (Chown et al. 2001; de Villiers et al. 2005). Biological invasions (along with climate change and interactions among these two change drivers) are the most significant terrestrial conservation challenges facing the region (Frenot et al. 2005; Chown et al. 2015a; Hughes et al. 2015). In consequence, an increasing amount is known about the correlates of invasion, the pathways for and vectors of non-native species, and the management strategies required to limit inadvertent introductions, especially given deliberate introductions are, for the most part, not permitted to the continent and its surrounding islands (Hughes and Convey 2012; Molina-Montenegro et al. 2014; McGeoch et al. 2015; Hughes and Pertierra 2016). Nonetheless, as is more broadly the case, the extent of information on impacts is surprisingly limited, particularly for plants (Gremmen et al. 1998; Frenot et al. 2001; Le Roux et al. 2013, Molina-Montenegro et al. 2012a), which is remarkable given that plants together with invertebrates are the most speciose groups of non-native species across the region (Frenot et al. 2005; Hughes et al. 2015), and are showing propensity for establishment on the continent itself (Molina-Montenegro et al. 2014). In consequence, the evidence base for decisions about management actions to be taken for either established species or new arrivals is currently small (Hughes and Convey 2012). Nevertheless, some efforts have been conducted to eradicate non-native plants from Antarctica (Galera et al. 2017; Pertierra et al. 2017a). For example, Poa pratensis was successfully eradicated in January 2015, providing pivotal information about eradication actions, allowing for the generation of a management protocol with high cost-efficacy, likely applicable to another non-native plant species in Antarctica (Pertierra et al. 2017a).

Here we begin to address some aspects about the impacts and management for the most widespread non-native vascular plant species in the Antarctic, Poa annua, which currently is the only non-native species of flowering plant that has successfully established a reproducing population on the Antarctic Peninsula (Frenot et al. 2005; Chwedorzewska 2009). This species is commonly associated with anthropogenically-modified habitats worldwide, but currently can also be found as an introduced species in natural habitats on most sub-Antarctic and some maritime Antarctic islands as well as a number of locations on the north-west coast of the Antarctic Peninsula (Molina-Montenegro et al. 2012a; Chwedorzewska et al. 2015; Hughes et al. 2015; Atala et al. 2019), and has been forecast to become more widespread (Chown et al. 2012; Pertierra et al. 2017b). Experimental laboratory studies have shown this species can potentially outcompete the only two flowering plant species indigenous to the Antarctic continent, the grass Deschampsia antarctica and the pearlwort Colobanthus quitensis (Molina-Montenegro et al. 2012a, 2016).

Here, using field ‘common garden’ experiments on King George Island (South Shetland Islands), we examined interactions between these three plant species with regard to variation in relative density of each, as density has been identified as an important factor influencing the invasion process, since higher densities enhance the competitive ability of a given species in a community (Lockwood et al. 2005; Arii and Parrott 2006). We also focus on water availability, as it is the primary limiting component in most Antarctic terrestrial ecosystems (Convey et al. 2014), with D. antarctica occupying a wider range of habitats compared with C. quitensis in the context of the water regime (Torres-Mellado et al. 2011). Furthermore, coastal parts of the Antarctic Peninsula region have experienced increased precipitation over the last century, a pattern which is forecast to continue (Bromwich et al. 2011; Turner et al. 2014; Thomas et al. 2015; Lee et al. 2017). As well as the indicated increase of temperatures for this area, an increase of precipitations is expected, generating higher soil water availability and hence, enhancing the physiological performance of vascular plants (Torres-Diaz et al. 2016). Combining these two elements (relative competitive ability and increased water availability), this study tested the following hypotheses: i) the most widespread non-native plant in Antarctica, P. annua, will exert a stronger competitive effect than the native vascular plants C. quitensis and D. antarctica, as assessed by survival and growth, and ii) these competitive effects currently are greater at higher relative densities of individuals of P. annua and under higher resource availability, as predicted under a simulated future climate change scenario. The ultimate goal was to provide initial predictions of how P. annua will affect native plants in both the short and long term, to aid in evidence-based conservation decision making within the region.

The combined outcomes of this field study demonstrate explicitly the negative potential impacts of an invasive plant on the native Antarctic vascular flora, and can inform models of how invasion scenarios are likely to play out given current and predicted future climatic conditions. Previous investigations have identified a range of Antarctic areas most susceptible to colonization (Chown et al. 2012; Pertierra et al. 2017b) and several of these areas are already being colonized (Olech and Chwedorzewska 2011; Molina-Montenegro et al. 2012a; Hughes et al. 2015). This study advances current understanding by seeking to identify impacts in the field, providing evidence for the relative density required and climatic conditions it may take for a non-native species to invade and then to displace its native competitors.

The marked asymmetry of competitive effects identified, based on the field experiment with P. annua and the two native species, suggests that the future spread of P. annua may result in the local displacement of both native species. In addition, the data and analyses indicate that knowledge of the relative local frequency dependence of performance between species in competition is important when evaluating the potential for invasion of non-native species in Antarctica. Although P. annua performed better than C. quitensis or D. antarctica at all densities of competitors tested, in general, even low densities of P. annua individuals would be sufficient to outcompete and invade the local vegetated areas, both under current climatic conditions and the future, wetter, scenario examined. In addition, other key aspects of potential for invasion, as propagule pressure should be assessed (Colautti et al. 2006; Simberloff 2009), under the specific conditions found at King George Island, in order to know the potential impacts of P. annua on the community structure and functioning. Habitats on this island are representative of much of the maritime Antarctic. In the context of the large numbers of propagules estimated to be entering the Antarctic annually (> 70 000 – Chown et al. 2012), including those of P. annua and other species that are pre-adapted to the environmental conditions of the region, this finding is of particular concern.

Based on the observation that P. annua currently grows associated with other plant species as well as on bare ground on King George Island, we also demonstrate that the probability of invasion depends on an interaction between the native plant species and the specific wetter climate scenario. Thus, invasion of P. annua in any new area will depend on whether the area is currently dominated by C. quitensis or D. antarctica. Under current climate conditions the competitive effect of P. annua on C. quitensis is greater than on D. antarctica. This may be due to D. antarctica having a set of functional traits that enables higher performance than C. quitensis (see Smith 2003), or because invaders that are functionally dissimilar from native species are often favored (see Richardson and Pyšek 2006; Mayfield and Levine 2010; Gallien et al. 2015). However, under a future scenario of higher soil moisture availability, P. annua exerted a weaker competitive effect on C. quitensis than on D. antarctica. This switching in the competitive effect exerted by the invasive P. annua on native species appears to be the result of an increase in the competitive ability of C. quitensis under moister conditions. Alternatively, there may also be an increase in niche overlap of P. annua and D. antarctica (sensu Hutchinson 1957).

Numerous studies have shown relationships between competitive effects and phylogenetic or functional structure in plant communities (Kraft and Ackerly 2010; Kunstler et al. 2012). These studies are based on the assumption that ecological similarity tends to lead to more intense resource competition than ecological dissimilarity (Kunstler et al. 2012). Ecological similarity can be quantified by using functional traits on the basis that these traits are linked to competitive ability such as rapid resource acquisition or biotic tolerance (see Chave et al. 2009). Nevertheless, it has been suggested that processes other than phylogenetic or trait similarity could drive competition between plants, generating a hierarchy in the competitive ability of species (Chesson 2000; Mayfield and Levine 2010). Our results suggest that the hypothesis of phylogenetic or trait similarity must be qualified as a generalized driver of competitive outcomes among Antarctic plants, because the competitive effect induced by P. annua on native plants was altered under different abiotic conditions. Previous studies have shown that P. annua exerts higher competitive effects on D. antarctica under two simulated climate change scenarios (well-watered condition or higher nutrient availability) compared with current climate conditions, due to higher resource use efficiency (Molina-Montenegro et al. 2016). In addition, it has been shown that C. quitensis possesses high phenotypic plasticity, improving its resource acquisition and ecophysiological performance under well-watered soil conditions (Molina-Montenegro et al. 2012b). On the other hand, Casanova-Katny and Cavieres (2012) showed that D. antarctica performs better when grown in moister microhabitats such as those provided by mosses compared with those in the bare ground, suggesting that this vascular species can be negatively affected in its physiological performance and growth when faced with low water availability. Thus, we suggest that competitive ability in these Antarctic plant communities could be governed by hierarchical differences driven primarily by climate conditions and secondarily by phylogenetic similarity. Such outcomes would also be in keeping with studies illustrating the importance of abiotic conditions altering the outcome of competitive interactions (see Keddy 1989). Although these hypotheses cannot be differentiated in the current study, further field experiments can be designed to unravel the mechanisms’ underlying interactions between P. annua and native plants under current and future climate scenarios.

There are indications that the well-documented trend of rapid regional warming in the Antarctic Peninsula region over the second half of the Twentieth Century has temporarily ceased (Turner et al. 2016). However, it is clear that over the last several decades the patterns of precipitation and temperature have changed in this region of Antarctica, along with nutrient input to the soil (Vaughan et al. 2003; Convey et al. 2009; IPCC 2013; Turner et al. 2014), with significant impacts on plant populations and communities (Parnikoza et al. 2009; Cannone et al. 2016, 2017). In a complex global change scenario, with simultaneous variation in different factors such as nutrients, temperature and water availability (see also Convey et al. 2014), formerly excluded areas may become available for colonization by those species with higher capacity to acquire the resources or improve performance, such as many invasive species (Dawson et al. 2012). Scenarios which then include the complexities of community interactions (Grime 2006), including native and alien species with varying functional similarity, in the context of the abiotic variation suggest that the trajectory of influence will differ over time, as the hierarchy of competitive ability is altered, and more complex communities potentially facilitate colonization (Bruno et al. 2003). Thus, studies such as that described here provide the basis for further investigation of how invasive plant species respond to multiple changing abiotic factors in a natural setting in Antarctica. In so doing, the work will also extend understanding of how impacts are realized more generally (see discussion in Catford et al. 2012; Richardson and Ricciardi 2013), and contribute to understanding of the role of ecological similarity in determining competitive outcomes in the context of invasion success, especially under changing climates (Chown et al. 2015b; Gallien et al. 2015; Hulme 2016).

Overall, this study indicates that the substantial concerns already expressed about invasive plant species for the Antarctic continent (Shaw et al. 2014; Hughes et al. 2015) are warranted, and particularly so for P. annua which is already spreading in the region (Olech and Chwedorzewska 2011; Molina-Montenegro et al. 2012a, 2014; Hughes et al. 2015; Atala et al. 2019). Moreover, it provides additional evidence supporting general concerns about the impacts of grasses (Pyšek et al. 2012), and for a region where few investigations have been made of the impacts of invasive alien species on local populations (McGeoch et al. 2015).

These findings underpin the growing number of biosecurity actions in the region and the importance of adherence to mitigation recommendations in the Antarctic Treaty’s Non-Native Species Manual (CEP 2011). Clearly, interventions at an early stage in the invasion pathway will be most efficient and cost effective (Simberloff et al. 2013), but substantial investment in their implementation is only likely if it can be demonstrated that negative effects will ensue from colonization by non-indigenous species (Hulme et al. 2013; Ricciardi et al. 2013). This study shows that, without such interventions, impacts will not only take place, but are also likely to change as water availability changes in the future along the Antarctic Peninsula. Indeed, P. annua is clearly a competitor with moderate (MO) to major impact (MR), or at least potentially so for the continent, as defined under the recently developed unified classification for alien species based on the magnitude of their environmental impacts (Blackburn et al. 2014). Such species cause local population decline (MO) or extinction (MR) of at least one native species, and in the case of MR species lead to changes in the structure of communities and the abiotic or biotic composition of ecosystems. In consequence, much impetus exists to improve biosecurity for the region, especially given that its implementation is currently inconsistent among different operators in Antarctica, with improvements required from many operators (Braun et al. 2001; Hughes and Pertierra 2016).

Nonetheless, the impact of P. annua is being realized on a continent that is considered a natural reserve, and one of the planet’s last wilderness areas and one with expanding ice-free areas (Shaw et al. 2014; Lee et al. 2017). Recent reports confirming that the species has colonized areas away from stations and is expanding along the Antarctic Peninsula are concerning (Olech and Chwedorzewska 2011; Molina-Montenegro et al. 2012a, 2014; Atala et al. 2019). Given that spread can be relatively fast (see Wilson et al. 2007), and we have demonstrated here that P. annua is capable of negatively impacting Antarctica’s two native vascular species, we encourage the development of a program of eradication that also will enable an effective evidence-based conservation decision-making protocol to be developed and applied in the region.

We are grateful for assistance from N. Ricote-Martínez and Fernando Carrasco-Urra in the field and for logistic support from the Polish Antarctic station “Arctowski”. We acknowledge the financial, permitting and logistic support of the Chilean Antarctic Institute (INACH). M.A. Molina-Montenegro is supported by project FONDECYT 11140607 and PII 20150126. P. Convey is supported by NERC core funding to the British Antarctic Survey’s “Biodiversity, Evolution and Adaptation” Team. S.L. Chown is supported by Project 4307 from the Australian Antarctic Division. This article contributes to the SCAR biological research programs “Antarctic Thresholds - Ecosystem Resilience and Adaptation” (AnT-ERA) and “State of the Antarctic Ecosystem” (AntEco).