Invasive alien species introduced by humans to areas beyond their native distributions (Richardson et al. 2000; Blackburn et al. 2011) are a major threat to the world’s biodiversity and economy (McGeoch et al. 2010; Blackburn et al. 2014; Brondizio et al. 2019; Pyšek et al. 2020). The numbers of alien species (and the subset of them that are invasive) are increasing rapidly world-wide and there is no sign of deceleration (Hulme et al. 2009; Seebens et al. 2017, 2018). Ongoing globalisation (Perrings et al. 2010), increasing levels of ecosystem modification, and climate warming (Walther et al. 2009) are expected further to accelerate alien species introductions, naturalisations and impacts (Essl et al. 2011a; Hulme 2017; Haeuser et al. 2018).

Research in invasion science over the last 30 years has focussed on questions aimed at improving predictions about which species will form invasive populations, and where these will occur (Drake et al. 1989; Rejmánek 2000; Kolar and Lodge 2002; Pyšek and Richardson 2007). These questions were motivated by the desire to prevent and mitigate the multiple environmental and socioeconomic impacts of alien species. This body of research has given us a better understanding of the importance of context dependence in biological invasions (Sapsford et al. 2020) and of the interactions among the multiple key drivers that influence the outcome of invasion (e.g. Higgins and Richardson 1998; Simberloff and von Holle 1999; Blumenthal 2006; Sol et al. 2008b; Pyšek et al. 2009a, 2015). This complexity is now fully appreciated and has been addressed by the development of numerous hypotheses and concepts (Catford et al. 2009; Enders et al. 2018, 2020; Jeschke and Heger 2018), theoretical frameworks (e.g. van Kleunen et al. 2010a; Gurevitch et al. 2011; Strayer 2012; Hulme et al. 2020; Wilson et al. 2020) and statistical models of macroecological patterns (e.g. Rouget and Richardson 2003; Thuiller et al. 2006; Wilson et al. 2007; Küster et al. 2008, 2010; Pyšek et al. 2009a, b, 2015; Castro-Díez et al. 2011; Schmidt and Drake 2011; Dawson et al. 2017; Essl et al. 2019). Since multiple factors determine invasion success and impacts, invasions can only be understood in the specific context in which they occur (Novoa et al. 2020; Sapsford et al. 2020). For this reason, studies need to be designed to consider the roles of these multiple factors to ensure that meaningful interpretations of outcomes can be made.

Given that thousands of alien species have established populations and spread across previously unoccupied environments, we are now in a position to (and indeed urgently need to) develop an understanding of the macroecological processes that underpin biological invasions. Macroecology is the study of large-scale (i.e. from hundreds of square kilometres to global in terms of space; from decades to centuries in time; and for large numbers of species or a broad range of taxonomic groups) patterns in the distribution and abundance of species, and the processes that determine those patterns (Gaston and Blackburn 2000; McGill 2019). To qualify as macroecological, a study needs to meet the scale requirement in at least one dimension; in invasion science, it is rare that studies conform to this definition in all three dimensions (but see Seebens et al. 2017, 2018) as can be inferred from the overview of studies presented in Appendix I.

Macroecology seeks to identify generality in complex ecological systems through comparative study of their properties, such as species assemblages or geographic ranges; it therefore addresses issues such as spatial and temporal variation in species richness, interspecific variation in abundance and range size, and how biological and environmental properties influence these aggregate entities (McGill 2019). For biological invasions, exploring macroecological patterns in the invaded range is a natural extension of research aiming to understand why some aliens become abundant and widespread while others do not, and why some sites accrue more alien species than others.

Attempts to associate biological traits and environmental characteristics with broad-scale patterns in the distribution, abundance, and richness of alien species have built on decades of macroecological research on native species. The assumption underlying this approach is that the ecologies of alien and native populations will be determined by the same drivers, albeit not necessarily in exactly the same way. For example, physiological tolerances of individuals to temperature or precipitation in the native range can be retained for many species in the alien range and climatic niche shifts are quite rare among terrestrial plant invaders (Petitpierre et al. 2012, but see Hulme and Barrett 2013; Early and Sax 2014; Atwater et al. 2018; Datta et al. 2019). Similarly, unless species’ life histories change when they move to a new range, effects of these traits on macroecological patterns in the native range should be maintained in the alien range. Plant species that are good competitors should retain this ability in the invaded range; some will become even better competitors due to enemy release (e.g. Keane and Crawley 2002), and some will become invaders by behaving in the same way as in their native range (Firn et al. 2011; Parker et al. 2013; Colautti et al. 2014).

The assumption that the ecologies of alien and native populations will be determined by the same drivers might not hold if the traits of conspecific individuals in the alien and native populations differ, e.g. due to founder effects, or evolution, or if resource limitation differs, e.g. when species move from an N-limited to a light-limited system. However, and more fundamentally, the identity and location of alien populations are determined by human activities, in a manner that is of a different order and type to that for native populations (Wilson et al. 2009). Thus, while human activities undoubtedly profoundly affect macroecological patterns in native populations (e.g., Gaston and Blackburn 2003; Faurby and Araújo 2017), the macroecological patterns and processes of alien populations are more strongly mediated by anthropogenic influences (Richardson et al. 2000; Blackburn et al. 2011). For example, similar factors seem to influence the native and alien range sizes of pine species (Richardson and Bond 1991), but alien range sizes are additionally profoundly influenced by anthropogenic factors (McGregor et al. 2012; Procheş et al. 2012).

With respect to alien abundance and distribution, a growing literature shows that some species traits are generally associated with the capacity to form self-sustaining populations that spread from points of introduction (i.e. invasive sensu Pyšek and Richardson 2007; van Kleunen et al. 2010b). For example, Pyšek et al. (2009a, 2015) used a source-area approach (as defined by Pyšek et al. 2004b) to show that the success of Central-European plant species introduced to other areas of the world results from the interaction of their distribution in the native range, habitats they occupy there, their biological traits, propagule pressure as a consequence of human use, and residence time. Jeschke and Strayer (2006) showed that invasiveness was related to native range size for mammals, birds and freshwater fish alien to Europe and North America. Recent studies revealed that fast life-history strategies, that allow for rapid increase in population size, characterise successful alien mammals (Capellini et al. 2015), reptiles (Allen et al. 2017) and plants (Richardson and Rejmánek 2004; van Kleunen et al. 2010b), while alien birds rather adopt slow strategies (Sol et al. 2012). In birds and mammals, a generalist life-style characterised by behavioural flexibility and larger trait variation is associated with successful establishment (Sol et al. 2008a, 2012; González-Suárez et al. 2015), while in insects specialised species seem to be more successful (Rossinelli and Bacher 2015). At the global scale, Dyer et al. (2016) showed that variation in the alien geographic range size of birds was positively associated with native geographic range size, while there was no effect of either body mass or ecological specialisation controlling for other variables. Environmental factors, including climate and habitat match between source and target regions (Thuiller et al. 2005; Hejda et al. 2009; Kalusová et al. 2013) are also likely to be important for invasiveness. For example, Duncan et al. (2001) showed that alien bird species with larger geographic ranges in Australia had a larger area of climatically suitable habitat on the continent.

For plants, several studies have addressed the role of traits in invasions in concert with other factors codetermining invasiveness (e.g. Herron et al. 2007; van Kleunen and Johnson 2007; Gravuer et al. 2008; Küster et al. 2008), but none of them simultaneously: (i) used a global dataset, (ii) analysed different stages of invasion process, (iii) took characteristics of the native and introduced ranges, such as its size, climate or habitat affiliation, into account together with species traits, and (iv) included the effect of residence time and propagule pressure (Table 1, Appendix I). Thuiller et al. (2006) studied how species traits, characteristics of the native and introduced ranges, residence time, and human usage shape the distribution of invasive alien plant species, but they based their analysis on the invading species pool in the target region of South Africa. Hamilton et al. (2005) analysed the role of several species traits in invasions at different spatial scales but, while they accounted for phylogenetic effects, they did not address different stages of the invasion process, and nor did they consider distributional characteristics in native ranges. Van Kleunen et al. (2007) studied different invasion stages by analysing introduction through horticultural trade and subsequent naturalisation separately, and employed distributional characteristics together with species traits, but only for species within the family Iridaceae. Gravuer et al. (2008) considered human and biogeographic factors as well as traits and three invasion stages, but only for a single genus (i.e. Trifolium). Küster et al. (2008) considered distributional characteristics and focused on important interactions among ecological characteristics for one invasion step. Dawson et al. (2009) addressed multiple stages of alien plant invasions for multiple genera in concert with a number of traits, but only for invasions in the tropics. Essl et al. (2011b) explored interactions among native range size, climate match, habitat affiliations, colonisation pressure and propagule pressure, but only for conifer naturalisations. McGregor et al. (2012) examined the role of species traits, biogeographic attributes (including native range size) and human factors on the likelihood of introduction and naturalisation of pine species in separate regions in the Northern and Southern hemispheres.

Despite advances in our understanding of invasion dynamics as discussed above, models in the literature that seek to elucidate the determinants of naturalisation and invasion success of alien species from a macroecological perspective (regional to global) rarely include a complete suite of factors that have been acknowledged as key elements in the process (Table 1, Appendix I). Yet, the application of models that analyse multiple factors in concert to determine their relative importance is crucial to address properly the role of biological traits promoting species invasiveness. Importantly, because of the context dependence of invasions, the real effect of a particular trait can be confounded, for example, if a species possessing a trait is introduced more frequently, or has had a longer time to adapt to, or take advantage of, conditions in the invaded area. Similarly, studies that ignore effects of, for example, habitats in which the species occurs either in the native and/or invaded range may overestimate the role of biological traits, which in turn may result in spurious predictions (Pyšek et al. 2015; Duncan et al. 2019). At the same time, factors interacting with the species traits themselves, such as propagule pressure and residence time, play important roles in determining the outcome of particular invasions.

Here, we develop a formal framework to explore the context dependence of invasions at broad geographical scales, and to increase awareness that macroecological analyses can yield biased results if these issues are ignored. We discuss different aspects of the framework by using examples of previous macroecological studies mostly based on plants and birds, as these two groups have been studied in most detail from this perspective. However, we believe that the framework is applicable to a broad range of taxa, and we hope that it will stimulate comparative research in other groups and environments.

Not all species have alien populations but, in principle, the size of the alien species pool (i.e. alien species richness) can to a large degree be attributed to the size of the donor species pool, dispersal success (incl. human transport, human commensalism and perceived utility) and the fit to the new environment in terms of environmental matching between donor and recipient regions (Karger et al. 2016). It therefore follows that, at the global level, observed aliens are a subsample of the world’s native species pool (though exceptions could occur where alien species hybridise and speciate in their new ranges; Ellstrand and Schierenbeck 2000; Levin 2003; Flores-Moreno et al. 2015; Brandenburger et al. 2019). Which species from this pool get entrained on the invasion pathway depends on the interaction of the socioeconomic motivations or determinants for translocation, and the distribution and characteristics of the species (Hulme et al. 2008; Essl et al. 2015; Sinclair et al. 2020). These latter features affect the probability that a species is selected (deliberately or otherwise) for transport. For example, a large native geographic range has been suggested to be among the best determinants of invasion success in seed plants (Rejmánek 1996; Goodwin et al. 1999; Hui et al. 2011), but this factor may affect invasiveness in several ways. First, having a large native range increases the probability of a species being selected for transport (Blackburn and Duncan 2001a) and therefore experiencing high propagule pressure (Cassey et al. 2004c). Second, the traits that allowed the species to achieve a large native range might also allow it to have a large alien range (Booth et al. 2003; Pyšek et al. 2009a; Dyer et al. 2016). Further, a large native range has been proposed to increase the probability that a species will sample a broader range of habitats and becomes better equipped for competition and novel interactions with species in the introduced ranges (Sax and Brown 2000). Nevertheless, this is not true for all taxa. For example, for parrots it has been shown that large geographic range size is a strong predictor of which species are transported outside their native ranges, and which transported species are subsequently introduced, but not which introduced species succeed in establishing (Cassey et al. 2004b); the net result of this, however, is that alien parrots tend to be those with large native ranges.

The biogeographic location of the native range also matters, as not all species pools are equally likely to be sampled for potential aliens. For example, bird species introduced in the 19^th and early 20^th centuries came primarily from Europe, were more likely to be introduced to regions of the British Empire, and were more likely to concern species in families of game birds (e.g. pheasants, ducks, and pigeons). These patterns arise because introductions in this period were largely driven by the deliberate activities of Acclimatisation Societies – organisations specifically aimed at promoting introductions of beneficial species, such as game animals, and which were especially active in British colonies (di Castri 1989; Pipek et al. 2015; Dyer et al. 2017).

The relative size and age of species pools in species’ native versus alien range also helps to indicate potential evolutionary imbalances (Fridley and Sax 2014). Alien species that have evolved over a longer period of time and in a more competitive and stable environment (e.g. mainland vs islands) tend to have higher competitive ability than co-occurring native species. As plant invasions in the Czech Republic, New Zealand, and eastern North America demonstrate (Fridley and Sax 2014), species from regions with highly diverse evolutionary lineages are more likely to become successful invaders in less diverse regions.

Disentangling the relative roles of species traits and properties of native geographic ranges in the context of anthropogenic effects is thus a fundamental task for invasion science. Knowing the extent to which the characteristics of the native range of a species can explain and predict its invasion, and under what contexts, would improve the precision of prediction systems used in weed-risk assessment (e.g. Pheloung et al. 1999; Weber et al. 2009).

There are at least three important consequences of the intersection of the socioeconomic motivations for introduction of aliens from the native species pool. First, the identities of introduced species are a non-random subset of all species that could have been introduced (see also Karger et al. 2016; Maurel et al. 2016). This can have significant consequences for our perceptions of the kinds of species that become invasive, and for our interpretation of the resulting macroecological patterns. For example, introduced wildfowl species are larger-bodied, on average, than those wildfowl that have not been introduced (Blackburn and Duncan 2001a). It follows that established wildfowl species are likely also to be large-bodied, and that the macroecological patterns expressed by alien wildfowl will be a consequence of how body size might influence the distribution and abundance of these species. It is important to factor such non-randomness into any analysis of later stages of the invasion process, including macroecological analyses, or incorrect conclusions about processes are likely to be reached (Cassey et al. 2004a; Pyšek et al. 2009a; Hui et al. 2014).

Second, sites to which species are introduced also depend on interactions between introduction pathways and the donor species pool. Again, incorrect conclusions about processes are likely to be reached without factoring in this context, especially as native species are not distributed randomly with respect to evolutionary history or associated traits, and hence pathway locations and species-pool composition interact. For example, socioeconomic changes in societies around the world have driven changes in the reasons for, and the geographical dimensions of, human-induced movement of bird species (Blackburn et al. 2009; Dyer et al. 2017); the source regions, destinations and identities of introduced species have shifted significantly in recent decades. Bird introductions are now driven largely by the pet trade, especially in rapidly developing economies in the Middle and Far East. This may explain why alien bird species follow Bergmann’s rule (Fig. 3), such that the average body mass exhibited by alien bird assemblages decreases toward the equator (Blackburn et al. 2019). Alien bird species appear to follow closely the relationship exhibited by native birds (Olson et al. 2009), but this is to a large extent a consequence of the fact that large-bodied species have been introduced at higher latitudes, on average, than small-bodied species, followed by latitudinal variation in establishment success that is independent of body mass (Blackburn et al. 2019). Historical introductions driven by Acclimatisation Societies tended to prefer large-bodied species and higher latitudes than recent introductions, which tend to be cage bird species such as parrots and estrildid finches, and to occur at lower latitudes (Dyer et al. 2017).

Figure 3.

Latitudinal variation in body mass for introduced (black, unfilled circles) and established (blue, filled circles) alien bird species worldwide, together with the mean (thick line) and range (thin line) of the relationship for native bird species. See text for details. Data from Blackburn et al. (2019) and Olson et al. (2009).

Third, patterns of selection from native species pools along different introduction pathways will affect the numbers of species (colonisation pressure; Lockwood et al. 2009) and individuals (propagule pressure; Lockwood et al. 2005; Simberloff 2009) that are introduced to different locations around the world. Models have shown repeatedly that the random selection of individuals from a species pool with realistic population structure will result in more species, and more individuals per species, in larger samples, as may occur for example in species transported in ballast water (Lockwood et al. 2009). More abundant species are more likely to be transported in this way. The same patterns hold for planned introductions (Cassey et al. 2004c). Variations in the levels of invasion among recipient communities, habitats or regions could be, in some cases, simply due to differences in the numbers of arriving aliens (Williamson 1996).

Lonsdale (1999) and Duncan et al. (2019) showed for plants and birds, respectively, that alien species richness at a location is a function of the number of species introduced to the location and the probability that any given introduced species establishes a viable population. Duncan et al. (2019) further showed that, for a closed system such as an island, establishment in turn is a function of the number of individuals introduced, and the probability that any one of those individuals leaves a surviving lineage (lineage survival probability; Fig. 2). Thus, alien species richness is primarily a consequence of the introduction process, and specifically colonisation and propagule pressures. These anthropogenic effects are fundamental to understanding the invasion process, and must be explicitly considered if the alien macroecological patterns that result are to be interpreted correctly (this is particularly notable early on in the invasion process, e.g. when looking at factors that determine the site of first detections; Huang et al. 2012). As an analogy, attempting to understand the drivers of alien species richness by performing a manipulative experiment in which the number of species added to each treatment was unknown would be unwise. It is similarly difficult to unravel the drivers of alien species richness in natural experiments where colonisation pressure is unknown. Duncan et al. (2019) carried out simple sensitivity analyses to show that by far the strongest determinant of alien species richness in their model was colonisation pressure; they show that increasing propagule pressure or lineage survival probability will increase alien species richness, but only up to an asymptote imposed by colonisation pressure. All else being equal, increasing colonisation pressure allows alien species richness to continue to grow as a linear function. While this model technically applies to closed systems, and it is not clear whether it applies to all taxa, most alien bird species at least do not spread far from points of introduction (Dyer et al. 2016). The implication is that for birds in most broad locations, colonisation pressure is a much more influential driver of incursion than spread. For many plant invasions, however, new population foci create potent propagule pressure sources that drive invasions much more quickly than the size and other dimensions of the source population, as demonstrated, for example, by the invasion of Opuntia stricta in Kruger National Park, South Africa (Foxcroft et al. 2004).

Data on colonisation pressure are rarely available for taxa other than vertebrates (i.e. alien species that were intentionally released outside of captivity, but see also insects released for biocontrol; Rossinelli and Bacher 2015). Quantification of colonisation pressure requires data on the number of species introduced in total, but data on failed invasions are generally scarce (but see Diez et al. 2009). Propagule pressure is also extremely difficult to measure at a large scale for plants (Fig. 4). Therefore, various quantitative surrogates have been used to attempt to capture variation in these key parameters. For example, the number of visitors to nature reserves (Lonsdale 1999; McKinney 2002), human population size or density (McKinney 2001, 2002; Pyšek et al. 2002; Taylor and Irwin 2004), the amount of trade and economic activity (Taylor and Irwin 2004; Pyšek et al. 2010; Essl et al. 2011a), species availability on the market (Dehnen-Schmutz et al. 2007a, b), the number of cultivars developed (Canavan et al. 2017), the type of land use such as the proportion of agricultural land and pastures (Chytrý et al. 2008b), or the number and distribution of botanic gardens (Hanspach et al. 2008; Hulme 2011) have all been used as proxies for propagule pressure in plants.

Figure 4.

Overview of the frequency of factors included in 92 macroecological studies of plants and vertebrates. The figure shows that the majority of studies in all taxonomic groups focus on traits, but that there is a difference among plants and animals in the frequency of studies addressing propagule and colonisation pressure, that is greater in the latter. On the contrary, plant studies more commonly address the role of residence time. Based on studies listed in Appendix I ; note that studies on invertebrates, fungi, and cross-taxonomic studies are not shown here (n = 10).

Despite the difficulty in accounting accurately for propagule pressure, it has been convincingly demonstrated that this factor, both over space (by widespread dissemination, abundant plantings, extensive release) and time (by long history of cultivation or captivity) fundamentally influences the probability of invasions by alien plant species (Rouget and Richardson 2003; Chytrý et al. 2008b). Models incorporating propagule pressure typically prove superior to those invoking only environmental parameters for explaining distribution patterns and abundance of invaders at a regional scale (Rouget and Richardson 2003) and only once propagule pressure of invaders is factored out, can the real effects of diverse physical and biotic factors on the outcome of plant invasions be identified (Chaneton et al. 2002).

Anthropogenic factors in the transport and introduction stages of the invasion influence the identities and numbers of species available for establishment at different locations, and the composition of the founding populations of those species (event-related effects). In general, propagule pressure needs to be sufficiently high to allow the founding population to escape the stochastic effects of demography, environment, genetics, and Allee effects, although the inherently random nature of these effects means that some very small founding populations avoid them. Following introduction, features of the new environment (including resource availability, disturbance regimes, environmental conditions, and native biota), and the ways that these features interact with the biological traits of the alien species, come into play in determining which species establish viable and persistent populations. Effectively, these features and traits determine lineage survival probability (Fig. 2). Populations that establish can then go on to spread across the new environment, by an ongoing sequence of establishment events realised through (and depending on) both their life history traits and further human-mediated dispersal. The spatial and temporal patterns in the distribution, abundance, richness and traits of the alien species that result, and the relationships between these population- and community ecology processes, are the fodder of the macroecological patterns and large-scale biological invasions (Fig. 2).

Even at this ‘terminal’ point in the macroecological study of biological invasions, however, it is important to remember that observed relationships bear the imprint of previous stages in the invasion process (Leung et al. 2012; Donaldson et al. 2014). For example, the right-hand (‘Invasion’) part in Fig. 2 presents a cartoon of the distributional extent and abundance of four hypothetical established alien species, plus the relative spatial positions of those populations in an oval region. A naïve assessment of these patterns might conclude that species represented by the triangle and star are naturally more invasive, being more abundant and having wider distributional (and latitudinal, if we assume the figure maps to the cardinal points) extents than the species represented by the cross and crescent. Species richness appears to decrease from the top (north) to the bottom (south) of the region. Species in the north tend to have pointed edges, with that in the south having more curves (although sample size is low to make inferences about traits). However, all these conclusions need to be tempered by information on which species were introduced, where and when, and in what numbers. In Fig. 2, we see that more species were introduced to the north than the south; we see that introduced species in the south were more likely to have had curved edges, while those in the north were more likely to have had points. Those species that established were generally those introduced in larger numbers. The star and triangle species were introduced more widely than the cross and crescent. The crescent species was only introduced in the south. All of this context modifies our conclusions, and demonstrates that we cannot reliably make the conclusions if we analysed only the current distribution pattern.

Field data for assemblages of alien species show that the effects depicted in Fig. 2 are real and complex. For example, the extent of the distribution ranges of established alien bird species increases with latitude poleward of the tropics, consistent with the well-known ecological pattern known as Rapoport’s rule, but ranges are smaller in the tropics (Stevens 1989). However, this pattern is largely a consequence of the latitudinal distributions of where bird species have been introduced, which is only modified slightly by latitudinal variation in establishment (Dyer et al. 2020). Hence, while alien and native bird species both follow Rapoport’s rule, the mechanisms underlying the similar patterns are unlikely to be the same (Dyer et al. 2020). The same is true for Bergmann’s rule in alien and native bird species (Blackburn et al. 2019), as noted earlier.

Various elements of introduction context may also interact. For example, individual pathways can deliver species with different levels of invasiveness (Thellung 1912; Pyšek et al. 2011), and species arriving via different pathways may differ in the impacts they cause (Pergl et al. 2017). The way in which species are introduced and spread around by humans within the new range can also have long-lasting impacts on invasion patterns. For example, trees used for forestry tend to be introduced to a few rural sites in large numbers, whereas ornamental trees tend to be introduced to many urban sites in low numbers, leading to profound differences in the pattern of the occurrence of invasions across spatial scales (Donaldson et al. 2014).

An important human-related effect on macroecological patterns of alien species that manifests most strongly in the naturalisation and invasion stages is residence time (Rejmánek 2000; Castro et al. 2005, Pyšek and Jarošík 2005, Williamson et al. 2009, Pyšek et al. 2011). For plants, residence time relates to species’ geographic alien range sizes but also their invasion status – in the Czech Republic casual species have significantly shorter mean residence times than naturalised and invasive aliens (Pyšek and Jarošík 2005), and in south-east Australia, alien graminoids with longer minimum residence times are more likely to be classified as invasive than non-invasive (Catford et al. 2016). Many regions contain species that have not been present long enough for them to naturalise and become invasive – yet, the importance of any particular plant trait in determining the success or failure of invasion is discernible only after the species has either established or failed in a new region. The longer a species is present, the more it is provided with opportunities for adaptation and spread, i.e. the more windows of opportunity it will encounter (Johnstone 1986). Another example of interaction with residence time is the lack of natural enemies in the new region following introduction, such as pathogens, herbivores or parasites. This process can operate on the scale of centuries, as shown for the accumulation of pathogens by alien plant species in North America (Mitchell et al. 2010).

Residence time interacts also with propagule pressure: the longer the species is present in a region, the greater the size of the propagule bank, and the greater the probability of dispersal, establishment, and founding of new populations (Rejmánek et al. 2005; Richardson and Pyšek 2006). In Europe, the effect of residence time is very long-term, and is still obvious after several millennia of plant invasions, as demonstrated for archaeophytes in the Czech Republic and UK (species introduced since the beginning of Neolithic agriculture until the end of Medieval; Pyšek et al. 2004a). Those archaeophytes that invaded soon after the beginning of Neolithic agriculture are still more common and have wider distribution ranges than those that arrived later (Pyšek and Jarošík 2005). Likewise, alien birds with longer residence times have larger alien range sizes worldwide (Dyer et al. 2016). However, the effect in birds is largely a consequence of species with longer residence times having been introduced to more locations, and only the effect of number of locations is significant in multivariate analysis (Dyer et al. 2016). Positive relationships between residence time and distributional extent have also been documented for many regional alien floras (Forcella and Harvey 1983; Crawley et al. 1996; Wilson et al. 2007; La Sorte and Pyšek 2009; see Rejmánek et al. 2005 and Pyšek and Jarošík 2005 for a review), although the influence of colonisation and propagule pressures here remain unexplored. Thus, failure to incorporate information on residence time may lead to spurious conclusions as, for example, we would expect species with different residence times to have different alien range sizes by chance alone (Wilson et al. 2007; Pyšek et al. 2009b, 2015).

To date, most invasion studies have attempted to explain the macroecological determinants of invasion by alien species and their assemblages by focusing on factors related to species traits and environmental characteristics, thus the interaction ‘Species biological traits × Geographic attributes × Habitats × Climate’. Few studies have explicitly considered event-related factors and their interactions with other factors. Searching for traits associated with invasiveness is partly practically motivated, and there is growing evidence that some species are inherently better equipped, i.e. have a more suitable suite of traits, to become invasive after translocation to new areas by humans (Pyšek and Richardson 2007; van Kleunen et al. 2010b). Identifying species with the potential to become weedy or pests based on their traits should provide information on the likely mechanisms by which a species becomes invasive, and the likely impacts it will have. It therefore provides a template for assessing the likely success of management options (Novoa et al. 2020). To achieve this, however, we need to identify the “real” and direct effects of the respective traits that can be then included into risk-assessment schemes, because often traits are associated with biases (e.g. resulting from variation in propagule pressure, residence time, pathways, habitats or other factors that are not explicitly addressed in analyses). Indeed, the few available studies that do account for this complexity suggest that the role of species traits is strongly context dependent, and that traits interact with other factors – there is a complex interplay of species’ traits, habitats occupied in both the native and invaded range (Hejda et al. 2009, 2015), characteristics of recipient ecosystems and native communities (Catford et al. 2019), and human activities (which influence propagule pressure and residence time in the new region) in determining invasion in novel environments (Bacon et al. 2014). Using multivariate approaches to examine suites of species traits linked with invasiveness may help to account for some of this context dependence (Kimmel et al. 2019).

Recent research on alien plants has shown that some of the species traits that were not commonly considered in the past due to the lack of information for large numbers of species forming floras play important roles in invasions. Such traits include seed bank persistence (Gioria et al. 2019), germination characteristics (Brändle et al. 2003; Gioria and Pyšek 2017), reproductive traits such as fecundity (Moravcová et al. 2010, 2015), and karyological characteristics such as genome size and ploidy levels (Kubešová et al. 2010; Pandit et al. 2014). The results of our models are only as good as the information available, and not considering a key trait can result in the influence of another trait being spuriously over-emphasised. Similarly, it has been shown in birds that missing important factors in the analyses might identify spurious effects determining invasion success. For example, propagule pressure is a major driver of establishment success and has been shown to be correlated to many species’ traits in alien birds, like native range or body size (Cassey et al. 2004c). Analyses ignoring propagule pressure misidentified such species’ traits as drivers of invasion success (Blackburn and Duncan 2001b).

In a study of European plants naturalised in North America, the effects of species traits on invasion were indirect, via their effect on the number of native-range habitats occupied and frequency of cultivation in the native range, and the importance of the biological traits was nearly an order of magnitude less than that of the breadth of the habitat niche, propagule pressure, and residence time (Fig. 5; Pyšek et al. 2015). This agrees with a previous study that reported direct effects of biological traits on the global invasion of Central-European species only during the most advanced stage of invasive spread, while the effects of traits on the probability of a species becoming naturalised were indirect (Pyšek et al. 2009a). Both these plant studies used the source-area approach (Pyšek et al. 2004b), looking at the pool of native European species invading elsewhere, therefore ignoring potential selection effects and post-invasive evolution in traits (Guo et al. 2018), but this approach is justified by the fact that a large fraction of species do not need to undergo evolutionary change for invasion (Parker et al. 2013; Colautti et al. 2014) and behave the same way abroad as at home (Firn et al. 2011; Petitpierre et al. 2012).

Moreover, the traits that confer an advantage at one stage of the process and in a particular habitat may be neutral or even detrimental at another phase and/or in a different habitat. For example, while small genome size played a role in the naturalisation of alien species in the Czech Republic, it did not separate invasive species from those that are not invasive (Kubešová et al. 2010; see also Küster et al. 2008).

Figure 5.

The number of North American regions in which Central-European species have become naturalised is driven by the combination of factors related to geographic attributes (the species’ performance in its native range, i.e. habitat niche and distribution); propagule pressure (measured by using proxies related to human use of the species both in its native and invaded range) and residence time (the time since introduction to North America) that represent the event-related factors; and a suite of alien species traits that affect the species’ invasion success indirectly, via their effect on the habitat niche in the native range (see Fig. 1 and 2 for explanation of colour codings); significant traits are shown in bold. The width and magnitude of numbers on arrows showing relationships between drivers is proportional to the value of the coefficient. Significance is indicated as: *** p < 0.001. Adapted from Pyšek et al. 2015.

To know whether a region, community or habitat is more invasible we need to ask not only whether it has more alien species, but whether it is intrinsically more susceptible to invasions. Intrinsic invasibility can only be determined if processes of immigration and extinction are taken into account (including colonisation pressure), as pointed out by Lonsdale (1999), and if the relative invasiveness of the pool of invading species is also considered (Catford et al. 2012). Lonsdale’s concept of invasibility has proved extremely useful in emphasising the role of colonisation pressure (although he used the term ‘propagule pressure’) and pointing out the difference between invasibility (or vulnerability to invasion) of a region, community or habitat and a simple number of invasive species it harbours; for the latter the term ‘level of invasion’ has become broadly used (Chytrý et al. 2005; Hierro et al. 2005; Catford et al. 2012).

There is a consensus in the research community that in biological invasions, the invaded habitats and invading species are ‘a key-lock principle’, and need to be studied in concert for a complete picture (Shea and Chesson 2002). The majority of hypotheses in invasion ecology have received support in some circumstances (and failed in others), but those hypotheses that merge the habitat- and species-perspective perform best (Richardson and Pyšek 2006; Jeschke et al. 2012). At the regional scale of temperate Europe, the type of habitat that is invaded by alien plants has been shown to play an even greater role than climate and propagule pressure (Chytrý et al. 2008b). Yet, studies exploring factors underlying the outcome of species introductions at the regional and global scale, even those that do include a number of different factors, usually do not consider the identity and characteristics of habitats (e.g. structure, disturbances regimes, nutrient or water supply, etc.), in either native nor alien distribution range (Appendix I). This is of key importance because these habitat characteristics determine the mechanisms of invasion acting in a particular site; yet, papers that to some extent combine the effect of habitats with other factors are exceptions rather than the rule (Pyšek et al. 2015).

Available analyses comparing the range of habitats occupied by species in their native and invaded range suggest that for some species there is a shift in habitat use attributable to the invasion process. While naturalised plant species inhabit a comparable spectrum of habitats in both ranges, invasive species tend to occupy a wider range of habitats in their invaded than in their native range (Hejda et al. 2009). This supports the idea that the invasion phase of the process is associated with extension of the spectrum of occupied habitats, hence broadening species’ habitat niches (Pyšek et al. 2009a). Another research direction in habitat-oriented invasion ecology is looking at habitat affinities that alien species exhibit in their native range and analysing how this preadaptation affects their success as invaders (Hejda et al. 2015; Kalusová et al. 2017). In a study of European plants introduced to North America, the direct effect of native-range habitat legacy and residence time were the main factors associated with the likelihood that a species would naturalise – more important than propagule pressure measured by a proxy related to species’ human use (Fig. 5; Pyšek et al. 2015). This key role of habitat legacy in shaping invasion dynamics accords with studies showing the strong effect of the breadth of habitat niche on invasion success (Hejda et al. 2009; Kalusová et al. 2013) and supports the notion that abundant, widely distributed species are superior competitors due to their ability, acquired over evolutionary history, to tolerate a wide range of abiotic conditions, use a broad spectrum of resources, and resist a large number of potential enemies (Sax and Brown 2000). Macroecological studies that explore how species with different traits interact with habitat characteristics are rare (but see Divíšek et al. 2018); more work on this topic is needed to improve our understanding of this kind of context dependence in invasion macroecology.

One of the main reasons why, in the majority of models of plant naturalisation and invasion, habitats are not considered is the lack of data on habitat affinities of alien species for most continents other than Europe (see Chytrý et al. 2016), and on the variation in this characteristic by regions. Since habitats have a strong effect on the outcome of invasion (Chytrý et al. 2008a, b) and on the way alien species integrate into local communities (Divíšek et al. 2018), such models may provide biased results or yield a low predictive ability due to exclusion of this important determinant. Similarly, testing of hypotheses in invasion ecology without taking habitats into account may mask the validity of concepts that do not hold across all environments, but may still be true under specific circumstances.

Another aspect of the interaction of habitat with pathway is that alien species intentionally brought into new regions (e.g. pets, aquarium related introductions, and horticulture) often escape or are released in places with suitable local conditions (e.g. similar habitats as in their native range) or close to human settlements and other sites favourable for alien species spread such as harbours, roads, etc. Given that the majority of successful alien plants are introduced through horticulture (Hanspach et al. 2008; Lambdon et al. 2008; Pyšek et al. 2012; van Kleunen et al. 2018), this phenomenon may have important consequences for macroecological patterns.