• biotic resistance;
  • climate change;
  • disturbance corridors;
  • exotic plants;
  • global change;
  • invasive spread;
  • landscape pattern;
  • nonindigenous plants


  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References


II.Stages of invasion260
III.A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change268

I. Introduction

  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References

The invasion of nonindigenous plant species (NIPS) is an important component of global environmental change. Invasive NIPS disrupt ecosystems, compete with native species and cause economic losses (Hobbs & Humphries, 1995; Vitousek et al., 1996; DiTomaso, 2000; Levine et al., 2003; Dukes & Mooney, 2004; D’Antonio & Hobbie, 2005). In the USA alone, the estimated economic impact of invasive NIPS is $34 billion per year (Pimentel et al., 2005). Invasive NIPS can act synergistically with other elements of global change, including land-use change (Vitousek et al., 1996; Hobbs, 2000), climate change (Dukes & Mooney, 1999; Simberloff, 2000; Kriticos et al., 2003), increased concentrations of atmospheric carbon dioxide and nitrogen deposition (Dukes & Mooney, 1999; Dukes, 2002; Weltzin et al., 2003).

Studies of traits that make NIPS invasive (Baker, 1965; Rejmanek & Richardson, 1996), characteristics of invasible communities (Elton, 1958; Lonsdale, 1999; Davis et al., 2000), and ecosystem or community responses to invasion (Levine et al., 2003; D’Antonio & Hobbie, 2005) have increased our understanding of the invasion process. However, patterns of invasion observed in the field at one site may be difficult to extrapolate to other sites because those observations are specific to a time, place and spatial scale (Table 1). Studies may employ different methods or measures of invasion success and there is often no standardized method for evaluating the generality of these conclusions (Cadotte et al., 2006). Recognition of this problem led to field studies and synthesis papers that examined invasion across spatio-temporal scales, often exploring the role of one or two factors (e.g. propagule pressure) in the invasion process (e.g. Kolar & Lodge, 2001; Hamilton et al., 2005; Colautti et al., 2006; Dietz & Edwards, 2006; Pauchard & Shea, 2006; Melbourne et al., 2007). These papers suggest that the failure to identify a general suite of factors underlying invasion success may have resulted from attempts to extract generalities from diverse studies that do not address invasion at the same spatio-temporal stage (e.g. Kolar & Lodge, 2001).

Table 1.  Different approaches to invasion research that may complicate comparisons among studies
Potential differenceExampleExplanation of critical differencesSolutions
  1. NIPS, nonindigenous plant species.

ScaleLandscape-scale observational study vsLandscape patterns are generally observational and correlativeConduct studies at more than one scale
experimental neighborhood- scale studyExperimental factors at a small scale are easy to manipulateRelate local-scale studies to large-scale spatial pattern studies
 It is easier to isolate invasion mechanisms in small-scale studies 
StageMature forest study vs microcosmNIPS may respond to different factors at different stagesConduct studies at multiple stages
assembled community studyInitial success may not predict landscape spread 
Definition of invaderNative species not in original plot vs NIPSDifferent factors may be responsible for success of this new speciesOnly compare studies that use all NIPS or all native invaders
 Cannot test the same questions on native and NIPS invadersCompare similar traits between natives and NIPS
Definition of invader success or invasibilityBiomass of one specific invader vs diversity of exotic species within the entire communityDiversity of NIPS does not necessarily correlate with NIPS success or potential impactUse biomass and other abundance or cover-related measures to get at invader success
Biomass may correlate more strongly with impact and success 
Ecosystem studiedForest vs grasslandForests have more complex canopy structure with more diffuse competitionForest succession is longer, and forest plants are longer livedDirect comparisons between studies in different ecosystemsshould focus on general concepts, not specific results

In this review, we examine the broad categorical filters that NIPS pass through at four well-established (e.g. Vermeij, 1996) spatio-temporal stages of invasion: transport, colonization, establishment, and landscape spread (Figs 1, 2). These stages are not discrete and filters will likely affect more than one stage. However, separating invasion into stages allows us to compare patterns of NIPS success from disparate studies and to discuss the relative importance of filters to invasion at each stage. Further, identifying the stage at which an invasion fails may allow us to understand the interaction of invasion filters with invasion character (e.g. number of introduction events), species traits, and ecosystem characteristics (Table 2, Fig. 3). Generalities arising from this type of synthesis can then be used to predict the outcome of invasion events, or to explore mechanisms responsible for deviations from these generalizations (Cadotte et al., 2006).


Figure 1. Four stages of invasion, and the factors affecting nonindigenous plant species (NIPS) success at each stage. Graphics at the top of the figure depict the dominant process occurring at a stage, and conceptual diagrams below identify dominant mechanisms determining the success of NIPS at the relevant stage. The three conceptual figures on the left share the same vertical axis. The conceptual figure for spread shows the dependence of spread rates on success of a species at passing through the previous stages, in addition to its typical propagule dispersal characteristics and the characteristics of the landscape. Conceptual figures are hypothetical, although many featured mechanisms are widely supported by experimental results (see text). Figures are meant to be illustrative, but not comprehensive.

Download figure to PowerPoint


Figure 2. The spatial and temporal scale of the four stages of invasion relative to other key biological processes. Adapted from Chapin et al. (2002).

Download figure to PowerPoint

Table 2.  The major filters, character of the invasion, nonindigenous plant species (NIPS) traits and physical factors in the recipient ecosystem that interact to affect NIPS success at each stage of invasion
Invasion stageMajor filtersCharacter of the invasion eventNIPS traitsPhysical factors in recipient ecosystem
  1. NA, not applicable.

TransportGeographic distance and barriers to long-distance dispersal (waterbodies, intervening deserts, etc.)Type of introduction (accidental vs deliberate, etc.)Wide native rangeNA
Propagule pressureSeed longevity 
Cause of species transportAssociation with humans 
 Human desirability 
ColonizationAbiotic barriers to colonizationPropagule pressurePhenotypic plasticityClimate
Genetic variation contained in introductionsWide abiotic tolerancesSoil
Reason a species is introducedFast growth/short juvenile periodResource availability
Spatial distribution of introductionsSelf-compatibilityDisturbance regime
 Germination without pretreatment 
EstablishmentBiotic filters to population growth and establishmentPropagule pressureCompetitive abilityClimate
Reason a species is introducedFast growthSoil
Genetic variationEfficient resource useResource availability
  Disturbance regime
SpreadLandscape filters to dispersal and establishment in new areasPropagule pressureEffective long- and short-distance dispersalDisturbance regime
Number of invasive fociHigh fecundityPatch attributes
Distribution of invasive fociPhenotypic plasticityPresence of suitable patches for
 Dispersal by seedcolonization and establishment
 GeneralistDispersal corridors
 Fast generation timeHeterogeneity of landscape

Figure 3. The assembly of local communities is influenced by filters at local, landscape and regional scales. The regional species pool is assembled through speciation, migration, biotic exchange and geological events. Transport into this regional pool occurs on a much faster time-scale than most natural movements of species. Here transport is shown outside the axis of spatial and temporal scales. Following transport, colonizing nonindigenous plant species (NIPS) move through local abiotic filters to colonization success, biotic filters to establishment success, and dispersal barriers to success in landscape spread. As the NIPS moves from colonization to landscape spread, the temporal and spatial scales of processes underlying the invasion increase.

Download figure to PowerPoint

Integration of processes occurring at different spatial and temporal scales (Figs 2, 3) will allow us to connect local and regional invasion filters with the factors affecting NIPS success (Table 2, Fig. 3). The importance of local and regional processes has been widely debated, but it is clear that communities are shaped by both (Ricklefs, 1987, 2004; Schluter & Ricklefs, 1993). Studies of local filters focus on species interactions and niche-based processes, as well as environmental constraints. Local communities assemble from the larger regional species pool (Fig. 3), which is shaped by history, biogeography, range expansions, evolution, and extinction (Ricklefs, 1987). The success of NIPS likewise reflects the interaction of local filters that reduce diversity and regional processes that enhance it (Fig. 3; Davis et al., 2005; Smith & Shurin, 2006). To enter a new regional species pool, a NIPS must first be transported over long distances. Upon arrival, local environmental conditions, biotic interactions, and demographic processes limit its entrance into the new community. Following local establishment, the NIPS may spread across the landscape, navigating across or around dispersal barriers (Fig. 3). Successful landscape spread can entrench the NIPS in the species pool of a new region. The invasion process therefore reflects a series of regional, local, and landscape filters that limit NIPS success during four stages of invasion.

II. Stages of invasion

  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References

1. Long-distance transport

Transport involves the intercontinental movement of a species into a new region. Although such species movements have always occurred, current species movements are happening faster than before and from more distant regions, primarily as a result of global commerce and travel (Huenneke, 1997; Mack et al., 2000; Reichard & White, 2001; Le Maitre et al., 2004). However, many species are unlikely to be purposefully or accidentally transported by humans or may not survive such transport (Perrings et al., 2005). The factors that allow NIPS to pass through geographic filters are sometimes elusive because most transport events are studied long after they have occurred. We can gain insight from studies examining past transport events or studies addressing general patterns of human-mediated transport at global scales.

i. Invasion character and species traits  The character of the invasion (Table 2) will affect NIPS transport success. In general, higher numbers of propagules increase the likelihood that species survive transport (Kolar & Lodge, 2001; Lockwood et al., 2005). NIPS originating from large native ranges that are introduced across a wide swath of the nonnative range may be particularly successful. The invasion of Phalaris arundinacea (reed canary grass) was facilitated by multiple introduction events from a variety of sources within the native European range (Lavergne & Molofsky, 2007). Multiple introductions resulted in the transport of continental-scale genetic variation from the native range of P. arundinacea and subsequent reshuffling within North American populations. High genetic variation alleviated bottlenecks at expanding invasion fronts and increased genetic diversity. New genotypes coding for advantageous traits such as vegetative reproduction have also evolved in North American populations through genetic recombination (Lavergne & Molofsky, 2007).

Species traits may also influence transport. It is important to distinguish between traits that allow NIPS to survive transport and traits preferentially selected by humans before transport. Humans may preferentially transport plants with qualities that make them strong horticultural or agricultural species such as cold hardiness, disease resistance and showy flowers. Human-selected traits may not confer invasiveness to NIPS, although many human-selected traits may be advantageous. Traits that correlate with successful transport and introduction relate mainly to geographic origin, native range extent and dispersal ability. Widespread species may have a higher overall chance of transport as they are more likely to come into human contact (Goodwin et al., 1999; Cadotte et al., 2006).

ii. Human activities and NIPS transport  Humans are the primary dispersers of NIPS during the transport stage (Vermeij, 2005; Pauchard & Shea, 2006), so understanding patterns of trade, travel, and human desires may allow predictions of the types of NIPS that will be transported, common origin and destination regions of NIPS, and the potential success of different types of introduction events.

Today there are higher numbers of NIPS in the Americas and Africa than in Eurasia (Vermeij, 2005). Pysek (1998) reports that Eurasia, with 4.4% of the world's total floral diversity, contributes 58.9% of nonnative species to other regions. This asymmetrical pattern closely parallels historical trends in human colonization, agriculture, horticulture, and trade (Delcourt, 1987; Lonsdale, 1999; Williamson, 1999). Large-scale movements of plant species began with the establishment of European colonies (1500 AD) and trade routes running to the New World from Europe (Mack & Lonsdale, 2001; Le Maitre et al., 2004).

Intercontinental plant transport has occurred accidentally, for utilitarian reasons, and for aesthetic purposes (Huenneke, 1997; Mack & Lonsdale, 2001). Accidental introductions occurred (more frequently in the past than today) in ships’ cargo, in seed stock, or with livestock and travelers from other regions (Gerlach, 1997; Mack & Lonsdale, 2001; Perrings et al., 2005). These NIPS tended to be ruderal species, capable of fast growth and high resource uptake (Mack & Lonsdale, 2001). NIPS were also deliberately introduced for food, fuel, forage, lumber and medicinal purposes in many European colonies. The introduction of European food crops in the Americas reflects an ingrained human avoidance of novel food sources (Mack, 1999), while forage and fuel crops were often introduced in areas with ‘insufficient’ native species. Utilitarian plant introductions continue today as developed and developing nations struggle to keep pace in a global economy (Huenneke, 1997; Le Maitre et al., 2004). Some countries have adopted high-yield NIPS to increase agricultural production despite knowledge that these plants are invasive elsewhere (Le Maitre et al., 2004).

The human desire for both familiar and exotic species resulted in the introduction of NIPS for aesthetic purposes during the 19th century and continues today (Mack & Lonsdale, 2001). Species introduced ornamentally may have a significant advantage over species introduced accidentally. Because species introduced for human use are intentionally cultivated, they may suffer less as a result of environmental stochasticity and low population size (Mack, 1995, 2000). Species growing under human care can form stable source populations that may eventually spread into natural areas. Additionally, NIPS introduced for horticulture typically pass through a climate-matching process to determine where they will best grow. Climate matching, combined with intentional cultivation, greatly increases the likelihood that the species will escape cultivation (Huenneke, 1997; Mack & Lonsdale, 2001). Invasive species introduced for ornamental purposes in the USA include Cortaderia jubata (pampas grass), Fallopia japonica (Japanese knotweed) and Lonicera japonica (Japanese honeysuckle).

2. Colonization

Species that pass through the transport phase do not necessarily colonize their destination area. Survival depends on environmental conditions (e.g. soil type and climate) and biotic processes at the neighborhood scale. Arriving populations must survive and achieve positive growth rates at low densities (Chesson, 2000; Sakai et al., 2001). Founder populations of NIPS are also strongly regulated by incoming propagule pressure. Because of small population sizes, colonizing NIPS must overcome environmental and demographic stochasticity, lack of genetic variability and allee effects (Mack, 1995; Sakai et al., 2001). As a very approximate rule, Williamson & Fitter (1996) suggest only 10% of imported species give rise to naturalized populations.

i. Propagule pressure  Propagule pressure strongly influences NIPS colonization success (Williamson, 1999; Lockwood et al., 2005; Colautti et al., 2006; Pauchard & Shea, 2006). Propagule pressure is the combined measure of the number of individuals reaching a new area in any one release event and the number of discrete release events. Propagule pressure may range widely. For example, forage species have sometimes been introduced in large numbers by airplane, while populations of Salix babylonica (weeping willow) in New Zealand may have invaded from a single cutting (Mack, 1995). NIPS introduced across a wide area of the new region may have a better chance of landing in suitable locations for colonization (Lockwood et al., 2005) Repeated introductions from the source region may save a population at the brink of extinction (Mack, 1995), and greater genetic variation may allow NIPS to adapt to novel conditions (Sakai et al., 2001; Lavergne & Molofsky, 2007). The importance of propagule pressure can vary based on local conditions (Foster et al., 2004; Lockwood et al., 2005). Very little propagule pressure may be necessary for colonization to occur in benign environments where disturbance has eliminated native competitors. However, in locations with intense competition or harsh abiotic conditions, high propagule pressure may be necessary (D’Antonio et al., 2001; Foster et al., 2004; Lockwood et al., 2005).

ii. Abiotic filters and species traits  Climate sets the broad limits to plant distribution and productivity and may cause NIPS to fail immediately during colonization (Sakai et al., 2001). While many NIPS become naturalized in new ranges with similar climates to their native range, there are examples of species moving to areas with very different climates (e.g. Conyza canadensis), as well as NIPS failing to establish under similar climates (e.g. Lantana trifoliata) (Mack, 1995). Plants with wide geographic ranges in their native region may be more likely to survive in a new region, as a result of broader climatic tolerances (Goodwin et al., 1999). Additionally, phenotypic plasticity and high levels of genetic variability may allow NIPS to adapt to less favorable conditions or environmental variability (Sakai et al., 2001; Lavergne & Molofsky, 2007). Fast growth, self-compatibility, a short juvenile period, and seeds that germinate without pretreatment may also be advantageous (Goodwin et al., 1999; Sakai et al., 2001).

Many successful NIPS invasions occur in areas of high resource availability or under fluctuating resource conditions where temporal heterogeneity in resource availability opens a window for colonizing NIPS (e.g. Burke & Grime, 1996; Davis et al., 2000; Tilman, 2004; Leishman & Thomson, 2005; but see Funk & Vitousek, 2007). Increased light, moisture, and soil nutrients have been shown to increase NIPS success and alter community dynamics (Huenneke et al., 1990; Burke & Grime, 1996; Parendes & Jones, 2000; Davis & Pelsor, 2001). In California's nutrient-poor serpentine grassland, Huenneke et al. (1990) found that macronutrient additions increased the overall productivity of the community, decreased species richness, and increased NIPS biomass with or without soil disturbance. Results from this and other studies, as well as theoretical findings, indicate that some NIPS respond more strongly to increased resource availability than native species. Others capitalize on resource opportunities following disturbance events that remove native vegetation or directly add resources to a community (Huenneke et al., 1990; Burke & Grime, 1996; Davis et al., 2000; Davis & Pelsor, 2001; Leishman & Thomson, 2005). However, in order for a plant to establish it must continue to increase from low density over the long term. While short windows of resource availability may allow colonization success, periods of low resource availability may not allow NIPS to establish. Alternately, NIPS could retain gains made during high resource availability or in high-resource locations through the storage effect (e.g. storing temporal gains in storage organs) (Melbourne et al., 2007). We will discuss resource availability and plant invasion further in the context of establishment.

3. Establishment

To establish, a NIPS must colonize a site and develop self-sustaining, expanding populations. Establishment may last longer than colonization and occurs on a slightly larger spatial scale (Fig. 2). At this stage, small subpopulations of individuals may be tightly linked through dispersal (Melbourne et al., 2007). During establishment, biotic filters that constrain the population size of NIPS may be most important, although they will interact with environmental conditions, species traits, and continued propagule pressure from source regions (Table 2). Biotic filters are barriers to invasion created by the actions or presence of living organisms. While biotic filters will not necessarily prevent the germination of seeds or the spread of NIPS, these filters can affect survival, growth, and reproduction.

Traits that enhance competitive performance, reduce niche overlap between NIPS and natives or increase enemy resistance may be most important during establishment (Lloret et al., 2005; Dietz & Edwards, 2006). Species that share similar resource acquisition traits are likely to compete strongly. Conversely, NIPS representing functional groups not present or in low abundance within a new community may encounter less competition with native species, especially in regions with a number of different resources or heterogeneous resource conditions (Lloret et al., 2005; Turnbull et al., 2005; Melbourne et al., 2007). Other advantageous traits include secondary chemical compounds that deter herbivores, ‘novel weapons’, such as root exudates that negatively impact other plants, fast growth, and high fecundity (Rejmanek, 1996; Callaway & Ridenour, 2004; Richardson & Rejmanek, 2004; Dietz & Edwards, 2006). Although specific traits conferring these abilities may vary among habitats, examples of traits that correlate with competitive ability include vegetative reproduction, leaf size, stem height and flowering phenology (Goodwin et al., 1999; Lloret et al., 2005).

i. Plant–plant interactions: competition (–), novel weapons (–), and facilitation (+)  Competition is likely the best studied of the biotic filters of invasion, although this filter alone appears unlikely to fully exclude invasive plant species (Levine et al., 2004). Competition or, more specifically, exploitation competition occurs at local scales when plants reduce the growth of their neighbors by consuming resources. Because invasive NIPS are generally most successful in areas with high resource availability (Dukes & Mooney, 1999; Davis et al., 2000; but see Funk & Vitousek, 2007), competition undoubtedly reduces the size, density, and impact of many NIPS. In some instances a single, strongly competitive species may slow the growth of an NIPS by reducing availability of a limiting resource. In other cases a suite of species may collectively reduce the availability of critical resources to levels that suppress growth of the NIPS. This latter scenario, along with growing recognition of the pace of global biodiversity loss, has given rise to dozens of studies examining the role of plant community diversity in determining invasibility (e.g. Knops et al., 1999; Levine, 2000; Naeem et al., 2000; Dukes, 2001; Hector et al., 2001; Kennedy et al., 2002; Fargione et al., 2003; van Ruijven et al., 2003; Fargione & Tilman, 2005).

Taken together, results of these neighborhood-scale diversity–invasibility studies suggest that diverse plant communities often (but not always) provide greater competitive resistance to NIPS (Hooper et al., 2005). So, does resistance result from niche complementarity or reduced resource overlap (i.e. many species with different resource requirements collectively reducing the perceived availability of resources for the invader)? Or are diverse communities resistant to invasion simply because they are more likely to include the species that most strongly compete with a suite of NIPS (i.e. the much-discussed ‘sampling effect’ of the biodiversity literature) (Hooper et al., 2005)? While many early studies were unable to address this question (Wardle, 2001), it now seems that the answer may be: both.

Recent studies suggest three nonexclusive patterns of competition. (1) In some systems, growth of invasive species can be suppressed by species that are morphologically, phenologically, and physiologically similar, that is, species of the same functional type (e.g. Dukes, 2001; Fargione et al., 2003; van Ruijven et al., 2003). (2) In other cases (and even in some of the same systems), a single dominant species or functional group can most strongly suppress all or most invaders (Symstad, 2000; Fargione et al., 2003). (3) Finally, in some systems, an assemblage of species with different traits can compete more strongly with an invader than any one species alone (Fargione & Tilman, 2005; Milbau et al., 2005; Losure et al., 2007). Thus, niche complementarity among residents can contribute to a community's biotic resistance to invasion in cases where a single resident species is unlikely to out-compete the invader. The degree to which complementarity (and thus species diversity) plays a role in determining invasibility may be influenced by resource availability of a site, with more fertile sites being more prone to the influence of dominant species. In some cases, losses of even the least abundant native species can markedly increase the invasibility of resident communities (Lyons & Schwartz, 2001; Zavaleta & Hulvey, 2004). The critical variable in the diversity–invasibility relationship is likely to be whether the species that are lost contribute to lowering the availability of a limiting resource below some threshold level at a sensitive time for the invasive species (Davis & Pelsor, 2001). For example, in systems with a strong temporal component to resource availability (e.g. water in Mediterranean-climate systems), there may be greater opportunity for rare species to affect resource availability at these sensitive times (e.g. Dukes, 2001; Zavaleta & Hulvey, 2004).

Negative interactions between NIPS and native plants may also result from NIPS with novel weapons. Some NIPS have biochemical root exudates that act as allelopathic agents or alter plant–soil microbial interactions in the introduced range (Callaway & Ridenour, 2004). One mechanism through which NIPS root exudates can negatively impact native plants is through the disruption of beneficial relationships between native plants and soil biota. In forests of the northeastern USA, Allaria petiolata, an herbaceous mustard species, contains a type of phytotoxic glucosinolate that appears to disrupt the mutualism between arbuscular mycorrhizal fungi and hardwood canopy trees. Because the success of these juvenile hardwoods depends on the association with arbuscular mycorrhizal fungi, the invasion of A. petiolata results in tree mortality that favors further success of this invader because of reduced competition with tree species (Stinson et al., 2006).

Resident species do not always suppress growth of NIPS, and sometimes contribute to their success. Facilitation is less studied in invasion biology and perhaps generally in ecology, although recent studies suggest that it may be an important local regulator of community assembly (but see Prieur-Richard et al., 2000; Bruno et al., 2003). Facilitative relationships are most commonly observed in harsher abiotic environments where neighboring plants ameliorate microclimatic stressors (Bruno et al., 2005; Brooker, 2006), but facilitation is not limited to these environments. Smith et al. (2004) found that native dominants increased seedling establishment of the invasive Melilotus officinalis in a relatively productive North American grassland. Additionally, certain invasive species may facilitate the success of other invaders, leading to invasional meltdown (Simberloff & Von Holle, 1999). For example, invasions of nitrogen fixers into communities without native nitrogen fixers can increase the pool of soil nitrogen (Vitousek & Walker, 1989; Hughes & Denslow, 2005), facilitating the invasion of other NIPS previously limited by nitrogen availability (Yelenik et al., 2004).

ii. Interactions with other trophic levels  Herbivores, parasites, pathogens, mutualistic soil biota, pollinators, and dispersal agents also influence NIPS establishment. Escape from herbivory or disease may increase growth rates, and the chance of establishment in a new region. The enemy release hypothesis (ERH) suggests that NIPS benefit from transport outside the range of their natural enemies (Elton, 1958; Maron & Vila, 2001; Keane & Crawley, 2002; Carpenter & Cappuccino, 2005). Building on the ERH, the evolution of increased competitive ability (EICA) hypothesis (Blossey & Notzold, 1995) may also explain disproportionate success of invasive plants in new ranges. The EICA hypothesis suggests that, under reduced enemy pressure, selection may shift the resource allocation of NIPS from enemy defense to faster growth (Blossey & Notzold, 1995). Greater enemy pressure on native species should shift the competitive balance to favor NIPS (Keane & Crawley, 2002; Blumenthal, 2006).

There are mixed results for both the ERH and the EICA hypothesis (Keane & Crawley, 2002; Daehler, 2003). Studies show that some NIPS have longer life-spans, grow larger, and achieve higher reproduction in invaded ranges than in native ranges (Daehler, 2003; Leger & Rice, 2003). However, these studies have not always found mechanistic explanations linking increased NIPS growth to herbivory (Keane & Crawley, 2002). Covarying factors such as competition (Leger & Rice, 2003) and resource availability (Blumenthal, 2006) may also complicate predictions of the relative importance of herbivory. The Resource–ERH (Blumenthal, 2006) suggests that enemy release in combination with areas of high resource availability increases the success of fast-growing, ‘high resource use’ NIPS in novel environments (Fig. 4).


Figure 4. The resource–enemy release hypothesis (Blumenthal, 2006) suggests that nonindigenous plant species (NIPS) that require the most resources for growth will benefit most from enemy release in their introduced range. (a) Enemy regulation may be highest for high-resource-use species in their native range. (b) In an introduced range, high-resource-use species may have greater potential to increase growth in response to enemy release (solid line) relative to native competition (dashed line). Low-resource-use NIPS will be less influenced overall. Figure redrawn from Blumenthal (2006).

Download figure to PowerPoint

Herbivores also influence interactions between NIPS and the native plant community. For instance, intense grazing by introduced ungulates can increase the invasibility of native plant communities (D’Antonio et al., 2000). In a meta-analysis of 63 studies, Parker et al. (2006) found that native generalist herbivores suppressed introduced plants more than they suppressed natives, while native specialist herbivores did not suppress NIPS. Introduced generalist herbivores facilitated NIPS through their negative impact on natives. These results suggest that novel pressure from generalist herbivores may be an important line of defense against NIPS, but, in ecosystems heavily invaded by nonnative herbivores, native plants may also suffer from novel herbivore damage. Specialist enemies that switch from native hosts to NIPS, or that accompany NIPS from other regions, can limit the degree of enemy release. Although rare, host-switching has been observed among native and NIPS congeners (Creed & Sheldon, 1995).

Plant–soil feedbacks can strongly regulate the diversity and productivity of plant communities and affect NIPS success. Plant–soil interactions may be positive or negative, although negative feedbacks are most common (Reinhart & Callaway, 2006). Negative feedback is driven by soil pathogens, herbivores and parasites. These organisms reduce plant growth, provide density regulation and maintain higher degrees of diversity within plant communities. Positive feedback results from the presence of mycorrhizal fungi, nitrogen-fixing bacteria and other beneficial soil biota. Positive feedback may disproportionately facilitate the success of some species over others (Reinhart & Callaway, 2006). In general, interactions between native plants and soil communities tend to be negative, while positive feedbacks often occur between NIPS and soil biota in their introduced range (Klironomos, 2002).

Altered relationships and feedback with soil biota in the introduced vs native range may partially explain why some NIPS are so successful. Several studies have demonstrated that soil communities favor NIPS over native species (Reinhart et al., 2003, 2005; Callaway et al., 2004; Wolfe & Klironomos, 2005). In a California grassland, Klironomos (2002) found that four out of five nonnative species experienced positive soil feedbacks, while all five rare native plants experienced negative feedback. Reinhart et al. (2003) found that invasion of Prunus serotina (black cherry) was facilitated by soil communities of north-western Europe, while soil communities in the native range of the species inhibited its survival and growth. Reinhart & Callaway (2006) recently reviewed available biogeographical comparison studies investigating the effect of soil biota on NIPS in native and nonnative ranges. In all six studies the direction of soil–plant feedback was strongly negative in the native ranges of the NIPS. In the introduced ranges, feedback was strongly negative in only one case.

NIPS can directly affect the structure and function of soil biota (Wolfe & Klironomos, 2005), with a variety of consequences. In some cases, NIPS form novel mutualisms, increasing establishment success and changing the availability of soil nutrients (Richardson et al., 2000a; Callaway et al., 2004). For example, many NIPS increase soil nitrogen by forming associations with native nitrogen-fixing bacteria (Richardson et al., 2000a; Callaway et al., 2004). Increases in soil nitrogen resulting from these mutualisms may change native community structure and increase the success of future NIPS invasions (Vitousek et al., 1987; Vitousek & Walker, 1989; Yelenik et al., 2004). In other cases, NIPS may alter the prevalence of disease in a community. In a model with field-estimated parameters, Borer et al. (2007) showed that invasive annual grasses in California may increase the presence of generalist viral pathogens in native perennial communities. Annual grasses are inferior competitors in this system, but they may be able to successfully invade in part because of the negative effect of increased viral pathogens on native perennial grasses. Finally, as already discussed, NIPS may have biochemical exudates that act as ‘novel weapons’ and may disrupt beneficial mutualisms between native plants and soil fungi (Stinson et al., 2006).

Mutualisms with pollinators and seed dispersal agents in the introduced region are also necessary to ensure establishment of some NIPS (Richardson et al., 2000a), although seed dispersal agents are most important during spread. It is unlikely that plants with very tightly coevolved pollinator or disperser mutualisms will find replacements in their introduced range. Plants that are pollinated by generalists, display vegetative reproduction or are self-compatible may have significant advantages (Richardson et al., 2000a). Competition for pollination, similar to competition for resources, may occur between natives and NIPS (Brown & Mitchell, 2001). Showy NIPS may draw pollinators away from native species, reducing pollen quantity and seed set. Alternatively, these NIPS may attract more pollinators to natives, facilitating increased pollination (Brown et al., 2002).

iii. Lag phase  A lag phase often takes place between establishment and spread, when small populations of established NIPS adapt to their new community. This phase may correspond to a lack of genetic variation, which prevents rapid adaptation to novel conditions, or the time necessary for the population to reach a threshold size that allows it to spread (Sakai et al., 2001; Barney, 2006). Lag time may also reflect a lack of suitable local habitat, inclement environmental conditions, or a statistical artifact (Pysek & Hulme, 2005). During this period, multiple introductions, range expansion and migration of NIPS, and gene flow between populations of establishing NIPS may decrease the time spent in the lag phase (Sakai et al., 2001; Lavergne & Molofsky, 2007). Rapid evolution can sometimes produce new genotypes capable of surviving in different climates, competing more successfully with native species, or deterring enemies (Lee, 2002). For example, Abultilon theophrasti (velvetleaf) was originally introduced before 1700 in the USA. This species has only recently become an aggressive invader as a result of the evolution of different life-history strategies based on the nature of competition in its new environment (Weinig, 2000; cited in Lee, 2002).

4. Landscape spread

NIPS spread occurs at the scale of the regional metacommunity: a region containing groups of populations connected through long-distance dispersal (Melbourne et al., 2007). While transport occurs at an interregional scale, ‘spread’ refers to dispersal within a region over significantly longer time periods (Fig. 2). At the metacommunity scale the landscape is heterogeneous, and NIPS populations exist as interacting groups of species at different stages of colonization and establishment. In this regard, spread incorporates all three of the previous stages: regional spread rates of NIPS are influenced by landscape heterogeneity, the size and distribution of suitable habitat patches for colonization and establishment, the distance between suitable patches, and the population characteristics, growth rates, invasion history and dispersal ability of NIPS (Fig. 1, Table 2). The mosaic of local conditions or heterogeneity across the region will determine the interaction of local-scale population dynamics with local and long-distance spread. Heterogeneity includes both environmental (geomorphology, resource availability and soil types) and biotic (often measured as beta diversity) heterogeneity. In general, larger landscapes contain a greater heterogeneity of habitat patches and thus may maintain higher degrees of diversity of both natives and NIPS (Huston, 1994; Davies et al., 2005; Melbourne et al., 2007).

i. Invasion character, species traits, and dispersal  The spread rates of NIPS are primarily determined by landscape pattern and barriers to dispersal. However, many other factors will influence spread. For instance, at this scale, range expansion is faster if it stems from many small foci with the same aggregate area as a single large focus (Pysek & Hulme, 2005). Regional spread results from slow and steady local spread and rare long-distance dispersal (LDD) events. While local dispersal may result in linear rates of expansion moving out radially from the initial invasion foci, LDD tends to make spread rates nonlinear (Lewis & Kareiva, 1993; Kot et al., 1996; Neubert & Caswell, 2000; Hastings et al., 2005). Rates of local spread vary among species depending on the dispersal mechanism. Pysek & Hulme (2005) reported average local dispersal rates ranging from 2 to 370 m yr−1. Intraspecific variation can also be significant, suggesting that population dynamics and rare LDD can strongly influence dispersal. In Australia, Opuntia stricta invasions spread up to 18.5 km from their origin with an average rate of 370 m yr−1. However, in the first 2 yr, outlying populations were established up to 14 km away as a result of LDD early in the invasion. LDD may lead to aerial expansion of 3–500 km2 yr−1, allowing plants to spread significantly more rapidly than average local dispersal rates suggest (Pysek & Hulme, 2005). LDD is also largely decoupled from landscape pattern (With, 2004).

Traits promoting dispersal are most important during the spread stage of invasion (Lloret et al., 2005). Timing of flowering, length of juvenile period, mode of dispersal, phenotypic plasticity, and seed size may also affect spread (Kolar & Lodge, 2001; Garcia-Ramos & Rodriguez, 2002; Hamilton et al., 2005; Lloret et al., 2005; Pysek & Hulme, 2005; Cadotte et al., 2006; Dietz & Edwards, 2006). Pysek & Hulme (2005) argued that the available literature does not support close correlations between invasive traits and spread rates at the landscape scale. Wind, water and animal-mediated dispersal may be equally effective, although nonclonal species may spread marginally more rapidly than clonal species. The lack of correlation between dispersal-related traits may be a result of variations in the local success of NIPS.

Dispersal vectors also influence spread. NIPS dispersed by animals depend on the presence of these vectors (Richardson et al., 2000a), which may also be affected by landscape pattern. Dispersal agents and pollinators are unlikely to respond to the same features of landscape pattern as plants. In a German study conducted at landscape scales, the spread of invasive P. serotina depended on the presence of roosting trees across the landscape – locations where birds perch and defecate seeds (Deckers et al., 2005). NIPS may also come into contact with dispersal vectors more frequently in disturbance corridors (see section 4. iv.; D’Antonio et al., 2000) or at the interface of suburban and natural landscapes (Williams & Ward, 2006). For instance, where suburbs abut forest in the eastern USA, long-distance dispersal of NIPS by white-tailed deer (Odocoileus virginianus) may promote NIPS success (Williams & Ward, 2006). Deer range throughout both habitats, often defecating in areas of heavily browsed native vegetation. NIPS benefit both from transport and from competitive advantages as a result of reduced densities of native species. Humans also play a large role in intraregional dispersal. For example, Macdonald et al. (1989) and Lonsdale (1999) showed that the number of visitors to national parks in North America and South Africa is positively correlated with the number of exotic species in the park. However, it is not clear whether this finding results from higher propagule pressure or increased disturbances caused by heavy foot traffic (Pysek & Hulme, 2005).

ii. Landscape pattern and the disturbance regime  Landscape pattern – or the spatial arrangement of different landscape elements – affects the spread rate of NIPS (Neubert & Caswell, 2000; Richardson et al., 2000b; With, 2002; Hastings et al., 2005). Landscape pattern arises from a variety of geological and biological phenomena, and the disturbance regime in a region. The disturbance regime describes the frequency, spatial extent, severity, and intensity of killing events over time. Natural disturbance regimes are often linked to physical site characteristics, extrinsic factors (e.g. weather), and the biotic community. Geomorphology, vegetation patterns, and edge effects can influence the spread of disturbances such as fire. Anthropogenic disturbance tends to differ from natural disturbance, and may alter the regional disturbance regime (D’Antonio et al., 2000). Variable spread and timing of disturbance events create a mosaic of patches in various stages of succession. This pattern strongly influences the presence and persistence of different species across the landscape (Mouquet et al., 2003). Changes to the natural disturbance regime may dramatically alter landscape pattern, facilitating invasive spread.

Both anthropogenic activities and plant invasions can disrupt, intensify, or suppress the natural disturbance regime (Hobbs & Huenneke, 1992; D’Antonio et al., 2000; Hobbs, 2000; D’Antonio & Hobbie, 2005). Alterations of the disturbance regime that increase resource availability or change landscape pattern can promote NIPS spread by creating favorable patches for colonization and establishment (e.g. Hobbs & Huenneke, 1992; Burke & Grime, 1996; D’Antonio et al., 2000; Davis et al., 2000; Hobbs, 2000). Disturbances alter resource availability in a local site by killing resident individuals or by directly increasing resource supply (D’Antonio et al., 2000; Davis et al., 2000). Disturbance can also interact with other factors that influence NIPS success during colonization and establishment. In a study comparing the response of Centaurea solstialis (starthistle) to uniform disturbance treatments in two invaded ranges and its home range, response to disturbance was found to be significantly higher in both invaded ranges than in its native range (Hierro et al., 2006). The authors suggested that soil microbes may suppress the response of C. solstialis to disturbance in its home range. In the invaded range, escape from these microbes may allow the weed to capitalize on disturbance events that eliminate competitors (Hierro et al., 2006).

Alterations of natural landscapes may favor weedy NIPS that have coevolved with human land use and disturbances (Delcourt, 1987; Pyle, 1995; D’Antonio et al., 2000; Parendes & Jones, 2000; Stohlgren et al., 2001; Teo et al., 2003; Kim, 2005; Vermeij, 2005). For example, native perennial grasses in Australia and North America may suffer more damage from introduced ungulate grazing than from introduced annual species, resulting in a shift to higher NIPS abundance (D’Antonio et al., 2000). Both Kim (2005) and Pyle (1995) found that human disturbance regimes promoted NIPS invasion success, while natural disturbance regimes either had no relationship with NIPS success (Kim, 2005) or actually prevented invasion (Pyle, 1995). Similarly, in riparian systems, natives tend to respond positively to the natural disturbance regime, while disruptions to natural cycles favor NIPS (D’Antonio et al., 2000). In some cases NIPS invasions may lead to further land transformation, altering the natural disturbance regime, landscape pattern, and ecosystem function (Hobbs, 2000).

iii. Patch attributes and edge effects  Increasingly, humans have fragmented landscapes into habitat patches within a matrix of human land use. Patch attributes, patch connectivity, and dispersal corridors influence NIPS spread (Huston, 1994; With, 2002; Davies et al., 2005; Knight & Reich, 2005; Ohlemuller et al., 2006). The size, shape, and edge-to-interior ratio of a patch may affect NIPS success. While large patches often favor natives, smaller patches may promote NIPS (Timmens & Williams, 1991; Harrison et al., 2001; Ohlemuller et al., 2006). Edge effects are more pronounced in small patches, and increased light, space, and soil moisture may favor NIPS (Timmens & Williams, 1991; Parendes & Jones, 2000). Small patches may also experience a greater influx of propagules from the surrounding landscape (Saunders et al., 1991; Brothers & Spingarn, 1992; Trombulak & Frissell, 2000; Bartuszevige et al., 2006; Ohlemuller et al., 2006). Similarly, the shape of patches can influence the rate of NIPS introduction. Nature reserves with high edge-to-interior ratios may experience a higher rate of NIPS invasions than those of similar size that are more circular in shape (Timmens & Williams, 1991).

Habitat patches near developed edges may contain more NIPS than patches in interior habitat. In some cases, patches near edge experience increased resource availability or altered microclimate conditions. For example, forest sites abutting agricultural fields may have more light and soil nutrients (as a consequence of nearby fertilization), and less soil moisture (as a consequence of higher evapotranspiration) (Brothers & Spingarn, 1992; Trombulak & Frissell, 2000). Edge areas may also experience higher propagule pressure. In human-dominated systems, NIPS may be cultivated in gardens, or weedy NIPS may grow in areas of frequent disturbance, providing a source of propagules to neighboring natural areas (Esler, 1987; Timmens & Williams, 1991; Brothers & Spingarn, 1992; Rose, 1997; Searcy et al., 2006). For this reason, sites closest to development are often most heavily invaded (Gelbard & Harrison, 2003; Deckers et al., 2005; Knight & Reich, 2005; Bartuszevige et al., 2006; Ohlemuller et al., 2006) or differ significantly in composition from interior sites (Brothers & Spingarn, 1992; Rose, 1997; McDonald & Urban, 2006).

iv. Corridors, connectivity and metapopulation dynamics   Connectivity of suitable patches influences dispersal of NIPS, movements of other species, and metapopulation dynamics of NIPS populations. Metapopulation theory suggests that the balance between local extinction and migration determines the regional persistence of a species. Therefore, connectivity between NIPS populations may promote spread and persistence across the landscape (Murphy et al., 2006). Corridors between suitable patches provide transport for natives and NIPS across unfavorable landscape matrix, encouraging spread and facilitating interactions between local populations (With, 2002), although natives and NIPS may require different types of corridors to disperse (Harrison et al., 2001; With, 2002; Damschen et al., 2006).

Native plants often require wide undisturbed corridors of intact habitat, while NIPS may disperse best through strips of human-disturbed habitat or ‘disturbance corridors’ (D’Antonio et al., 2000; Parendes & Jones, 2000; Rubino et al., 2002; Searcy et al., 2006). Disturbance corridors include roads, trails and power-line rights of way. These habitats can facilitate rapid NIPS dispersal for two reasons. First, removal of native vegetation from disturbance corridors leads to disturbed soil, high light, altered hydrology, and destruction of the native seed bank (D’Antonio et al., 2000; Trombulak & Frissell, 2000). Thus, disturbance corridors often provide favorable conditions for NIPS colonization and establishment. Secondly, disturbance corridors may increase physical transport of NIPS by providing pathways for dispersal vectors. Humans and horses have been blamed for carrying NIPS propagules along trails (MacDonald et al., 1988; Timmens & Williams, 1991; Campbell & Gibson, 2001), and vehicles transport weedy species along Australian roadsides (Lonsdale & Lane, 1994). Deer and other small mammals may transport large numbers of NIPS between disturbance corridors and from suburban landscapes into forest interiors (Vellend, 2002; Meyers et al., 2004; Williams & Ward, 2006). It is not clear that disturbance corridors always facilitate invasion into the adjacent habitat matrix; corridors may act solely as habitat refugia for NIPS not able to establish in intact natural habitat (Rubino et al., 2002). However, studies conducted at landscape scales often reveal correlations between the distance to disturbance corridors and NIPS presence or abundance (Timmens & Williams, 1991; D’Antonio et al., 2000; Parendes & Jones, 2000; Rubino et al., 2002; Gelbard & Harrison, 2003; Watkins et al., 2003; Searcy et al., 2006). D’Antonio et al. (2000) reviewed 14 studies of disturbance corridors and found that half reported NIPS movement into adjacent undisturbed habitat, while the other half found that NIPS remained only in corridors. NIPS spread from corridors into adjacent natural systems likely depends on the nature of the ecosystem, the traits of the invader and the time since invasion (D’Antonio et al., 2000; Rubino et al., 2002).

Landscape structure and connectivity also affect gene flow, influencing the ability of NIPS to adapt to novel conditions (With, 2004; Taylor & Hastings, 2005). Isolation may be particularly detrimental at the expanding edge of a population where allee effects are most common as a result of patchy dispersal and pollen limitation (Lewis & Kareiva, 1993; Kot et al., 1996; Keitt et al., 2001; With, 2002; Davis et al., 2004). Small population size and landscape boundaries that limit connectivity among satellite populations may ultimately prevent NIPS spread or increase the lag time between local establishment and further spread (Lewis & Kareiva, 1993). Reproductive isolation in spreading populations may also lead to speciation events as a result of the interaction of a NIPS genotype with the environment and subsequent adaptation, or as a result of genetic drift (Lee, 2002).

v. Coexistence at landscape scales  Positive correlations between native and NIPS diversity at landscape scales have sometimes been used to suggest that native diversity is not an important barrier to invasion (e.g. Stohlgren et al., 1999). While native diversity is only a small component of a complex ‘defense system’ limiting invasion (Fig. 1), the importance of native diversity can be underestimated at larger scales (Davies et al., 2005; Smith & Shurin, 2006). Theory predicts that increasing heterogeneity in resource availability and site conditions should allow native species and NIPS with different functional traits, competitive abilities and resource optima to coexist at the regional metacommunity scale, resulting in high diversity of both (Grime, 1974; Davies et al., 2005; Smith & Shurin, 2006; Melbourne et al., 2007). Because resource levels vary among local sites, one patch may have greater resistance to invasion while another provides a niche opportunity to the NIPS (Shea & Chesson, 2002). While native diversity provides ‘biotic resistance’ at neighborhood scales, at the landscape or regional scale the correlation between native and NIPS diversity is merely indicative of high heterogeneity which promotes diversity of both (Smith & Shurin, 2006).

For these reasons, the potential impact of NIPS on native species is more difficult to predict at regional scales (Stachowicz & Tilman, 2005; Smith & Shurin, 2006). Evidence from historic biotic exchange events, as well as ongoing NIPS invasions, suggests that diversity almost always increases following species introductions (Vermeij, 2005; Smith & Shurin, 2006), especially in island ecosystems (Sax et al., 2002). However, Smith & Shurin (2006) noted that patterns of species diversity at regional scales may not reflect the impact of local biotic interactions between natives and NIPS. Invasions initially result in reduced local abundance, reproduction or range size of natives and small changes may not be readily observed (Levine et al., 2003; Miller & Gorchov, 2004). Melbourne et al. (2007) suggested that heterogeneity at the scale of the metacommunity reduces the impact of invasive NIPS on natives by providing coexistence opportunities not present in homogeneous environments. However, while native species may maintain viable populations at regional scales, escape from extinction may only be temporary (Tilman, 1994; Harding et al., 2006). Studies documenting declines in beta diversity (distinctness of species composition between local sites) suggest that homogenization may be occurring regionally (Smith & Shurin, 2006).

III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change

  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References

Addressing invasion in four stages helps to identify the different processes that affect NIPS success at each stage of invasion, and provides a conceptual ‘map’ with which to predict and test the effects of environmental changes on these filters. Changes in climate and atmospheric CO2 are already affecting plant communities at local and regional scales (Dukes, 2000; Weltzin et al., 2003). NIPS have the potential to benefit if climate change affects the filters that limit invasion success (Table 3).

Table 3.  The framework can help categorize consequences of climate change that may affect success of invasive plant species (see text for details)
StagePossible consequences of climate change that affect nonindigenous plant species (NIPS)
TransportHorticultural species imported to new areas
Crops imported to new areas (with associated weeds?)
New NIPS cultivated as biofuel?
Shifting commercial activities and shipping pathways
ColonizationClimatic range restrictions will shift
Potential ranges may increase or decrease
Some ornamental species may become weedy
Some invasives may spread from disturbance corridors into natural areas
Many invasive species share advantageous traits?
Broad environmental tolerances
Rapid evolution
High phenotypic plasticity
EstablishmentShifting resource availability alters competition
Reduced competition from natives?
Range shifts for herbivores, pathogens, mutualists
SpreadInvasive NIPS propagules spread to newly suitable climatic zones more rapidly than many natives?

Climate change and increasing CO2 are unlikely to directly alter transport of most NIPS, but may affect patterns of trade and introduction success. Warming is likely to increase horticultural imports into regions with cold winters. Similarly, patterns of transport of agricultural species are likely to change, and agricultural weeds may ‘hitchhike’ into new environments. Concerns over climate change may also lead to the intentional introduction of nonnative crops for biofuel production. Many potential biofuel species possess similar traits to established NIPS, suggesting that they could become invasive (Raghu et al., 2006).

Colonization and survival of NIPS may also change with changing climate. For many species, decreasing frequencies of lethal cold temperatures will allow poleward range expansions (Simberloff, 2000; Kriticos et al., 2003). Conversely, warming may cause drying and dessication at warmer range margins, decreasing colonization success (Kriticos et al., 2003; Brooker, 2006). In other cases, increased precipitation and/or increased plant water-use efficiency (WUE) as a result of higher concentrations of atmospheric CO2 may expand the warmer range boundaries of some species. Kriticos et al. (2003) modeled the range of an invasive Acacia nilotica under climate change scenarios and found that warming temperature increased the range poleward, while higher precipitation and enhanced WUE expanded the range inland.

Changes in resource availability resulting from climate change and CO2 enrichment are likely to alter competitive interactions during NIPS establishment. Potential effects of elevated CO2 on NIPS establishment are discussed elsewhere (Dukes, 2000; Weltzin et al., 2003). Increased moisture resulting from precipitation changes or greater WUE may favor some NIPS (Dukes & Mooney, 1999), especially in arid communities and regions with strong seasonal patterns of precipitation. For example, in years with high rainfall, exotic annual grasses successfully invaded resource-limited California serpentine grasslands that had previously repelled NIPS (Hobbs & Mooney, 1991). Warming would reduce physiological stress on some introduced NIPS. These NIPS might then compete more effectively with native plants (Dukes & Mooney, 1999; Shea & Chesson, 2002). Phenological shifts in the timing of spring leaf-out may also allow certain NIPS to compete more strongly (Brooker, 2006). Pathogen, mutualist and herbivore ranges may also shift with unpredicted consequences for NIPS and native plants.

Landscape spread may be influenced by shifts in the ranges of species. NIPS that could once only survive in gardens or disturbance corridors may be able to spread into natural areas if the climate becomes more favorable for their survival and growth. Plants that cannot shift ranges quickly enough to maintain populations in suitable climates may decline, while species that can may expand (Dukes & Mooney, 1999; Higgins & Richardson, 1999; Simberloff, 2000). Thus, rapid warming may disproportionately benefit NIPS with traits such as rapid dispersal, short juvenile periods, high fecundity and small seed mass (Rejmanek, 1996; Dukes & Mooney, 1999; Simberloff, 2000). Invasive NIPS with fast reproduction, short life cycles, and high phenotypic plasticity may also respond to change with rapid genetic or phenotypic adaptation (Dukes & Mooney, 1999; Schweitzer & Larson, 1999).

IV. Conclusion

  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References

The four-stage framework acknowledges the multiscale nature of the NIPS invasion process and attempts to integrate invasion patterns and the mechanisms underlying these patterns at the different stages. Future studies that approach at least two different stages of invasion (e.g. Levine, 2000; Davies et al., 2005; Knight & Reich, 2005) can provide excellent insights into invasions. Where possible, we recommend adopting this approach. We also recommend that management approaches explicitly consider the targeted stage, and enhance natural filters in order to prevent invasion success (Table 4). It is likely that managing multiple stages of the invasion process simultaneously will be most effective.

Table 4.  Potential management strategies to reduce nonindigenous plant species (NIPS) success at each stage of the invasion process
StageManagement strategies
TransportRegulate nursery trade more strictly
Promote native species in landscape design
Screen seed stock more effectively
Educate the public about the consequences of NIPS and how to prevent their introduction (not buying nonnative ornamentals, not planting house plants outdoors, etc.)
ColonizationUse climate envelope techniques to predict range of potential NIPS
Reduce habitat fragmentation and edge effects
Start cutting, mowing, and herbicide treatments immediately
EstablishmentIncrease health and seed recruitment of native plants
Promote intact native communities and trophic structure
Study the effects of native generalist herbivores on NIPS for potential control (make sure these species do not prefer natives)
Investigate other (native) biocontrols
Reduce human disturbances, promote natural disturbance regime
Continue cutting, mowing, and herbicide application
SpreadMinimize disturbance corridors through natural landscapes
Promote native species that can compete with NIPS at the edges of disturbance corridors (e.g. early successional natives)
Isolate source populations of NIPS
Eliminate or reduce transport vectors (such as deer) in natural areas during reproductive stage of NIPS growth
Prioritize use of local techniques to manage colonization and establishment n sites that are most susceptible to invasion (e.g. sites with high numbers of reproducing individuals, and adjacent sites, sites on the edge of landscapes).


  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References

We thank Ophelia Wilkins and Alex Theoharides for help with figures, and Heather Charles, Mary Costello, Mike Rex, and Ron Etter for comments on previous drafts. We also acknowledge Marc Cadotte, Rich Norby, and two anonymous reviewers for suggestions that greatly improved the quality of the manuscript. We acknowledge funding from an National Science Foundation (NSF) CAREER award (DEB-0546670 to JSD) and an NSF GK-12 teaching fellowship (to KT) through the Watershed Integrated Science Partnership at UMass Boston. An NSF travel grant allowed KT to present and discuss ideas from an earlier draft at Ecological Society of America's (ESA's) Diversity in an Era of Globalization conference in Merida, Mexico in 2006.


  1. Top of page
  2. I. Introduction
  3. II. Stages of invasion
  4. III. A case for the four-stage framework: forecasting the response of nonindigenous plant species to climate change
  5. IV. Conclusion
  6. Acknowledgements
  7. References
  • Baker HG. 1965. Characteristics and modes of origin of weeds. In: BakerHG, StebbinsGL, eds. The genetics of colonizing species. New York, NY, USA: Academic Press, 147169.
  • Barney JN. 2006. North American history of two invasive plant species: phytogeographic distribution, dispersal vectors, and multiple introductions. Biological Invasions 8: 703717.
  • Bartuszevige AM, Gorchov DL, Raab L. 2006. The relative importance of landscape and community features in the invasion of an exotic shrub in a fragmented landscape. Ecography 29: 213222.
  • Blossey B, Notzold R. 1995. Evolution of increased competitive ability in invasive nonindigenous plants – a hypothesis. Journal of Ecology 83: 887889.
  • Blumenthal DM. 2006. Interactions between resource availability and enemy release in plant invasion. Ecology Letters 9: 887895.
  • Borer ET, Hosseini PR, Seabloom EW, Dobson AP. 2007. Pathogen-induced reversal of native dominance in a grassland community. Proceedings of the National Academy of Sciences, USA 104: 54735478.
  • Brooker RW. 2006. Plant–plant interactions and environmental change. New Phytologist 171: 271284.
  • Brothers TS, Spingarn A. 1992. Forest fragmentation and alien plant invasion of central Indiana old-growth forests. Conservation Biology 6: 91100.
  • Brown BJ, Mitchell RJ. 2001. Competition for pollination: effects of pollen of an invasive plant on seed set of a native congener. Oecologia 129: 4349.
  • Brown BJ, Mitchell RJ, Graham SA. 2002. Competition for pollination between an invasive species (purple loosestrife) and a native congener. Ecology 83: 23282336.
  • Bruno JF, Fridley JD, Bromberg KD, Bertness MD. 2005. Insights into biotic interactions from studies of species invasions. In: SaxDF, StachowiczJJ, GainesSD, eds. Species invasions: insights into ecology, evolution and biogeography. Sunderland, MA, USA: Sinauer Associates, 1340.
  • Bruno JF, Stachowicz JJ, Bertness MD. 2003. Inclusion of facilitation into ecological theory. Trends in Ecology and Evolution 18: 119125.
  • Burke MJW, Grime JP. 1996. An experimental study of plant community invasibility. Ecology 77: 776790.
  • Cadotte MW, Murray BR, Lovett-Doust J. 2006. Ecological patterns and biological invasions: using regional species inventories in macroecology. Biological Invasions 8: 809821.
  • Callaway RM, Ridenour WM. 2004. Novel weapons: invasive success and the evolution of increased competitive ability. Frontiers in Ecology and the Environment 2: 436443.
  • Callaway RM, Thelen GC, Rodriguez A, Holben WE. 2004. Soil biota and exotic plant invasion. Nature 427: 731733.
  • Campbell JE, Gibson DJ. 2001. The effect of seeds of exotic species transported via horse dung on vegetation along trail corridors. Plant Ecology 157: 2335.
  • Carpenter D, Cappuccino N. 2005. Herbivory, time since introduction and the invasiveness of exotic plants. Journal of Ecology 93: 315321.
  • Chapin III FS, Mooney HA, Chapin MC, Matson PA. 2002. Principles of terrestrial ecosystem ecology. New York, NY, USA: Springer, 11.
  • Chesson P. 2000. Mechanisms of maintenance of species diversity. Annual Review of Ecology and Systematics 31: 343366.
  • Colautti RI, Grigorovich IA, MacIsaac HJ. 2006. Propagule pressure: a null model for biological invasions. Biological Invasions 8: 10231037.
  • Creed RP, Sheldon SP. 1995. Weevils and watermilfoil – Did a North-American herbivore cause the decline of an exotic plant? Ecological Applications 5: 11131121.
  • D’Antonio CM, Dudley TL, Mack M. 2000. Disturbance and biological invasions: Direct effects and feedbacks. In: WalkerLR, ed. Ecosystems of disturbed ground, Vol. 16. New York, NY, USA: Elsevier Science, 429468.
  • D’Antonio CM, Hobbie SE. 2005. Plant species effects on ecosystem processes. In: SaxDF, StachowiczJJ, GainesSD, eds. Species invasions: insights from ecology, evolution and biogeography. Sunderland, MA, USA: Sinauer Associates, 6584.
  • D’Antonio CM, Levine J, Thomsen M. 2001. Propagule supply and resistance to invasion: a California botanical perspective. Journal of Mediterranean Ecology 2: 233245.
  • Daehler CC. 2003. Performance comparisons of co-occurring native and alien invasive plants: Implications for conservation and restoration. Annual Review of Ecology, Evolution and Systematics 34: 183211.
  • Damschen EI, Haddad NM, Orrock JL, Tewksbury JJ, Levey DJ. 2006. Corridors increase plant species richness at large scales. Science 313: 12841286.
  • Davies KE, Chesson P, Harrison S, Inouye BD, Melbourne BA, Rice KJ. 2005. Spatial heterogeneity explains the scale dependence of the native-exotic diversity relationship. Ecology 86: 16021610.
  • Davis MA, Grime JP, Thompson K. 2000. Fluctuating resources in plant communities: a general theory of invasibility. Journal of Ecology 88: 528534.
  • Davis MA, Pelsor M. 2001. Experimental support for a resource-based mechanistic model of invasibility. Ecology Letters 4: 421428.
  • Davis HG, Taylor CM, Civille JC, Strong DR. 2004. An Allee effect at the front of a plant invasion: Spartina in a Pacific estuary. Journal of Ecology 92: 321327.
  • Davis MA, Thompson K, Grime JP. 2005. Invasibility: the local mechanism driving community assembly and species diversity. Ecography 28: 696704.
  • Deckers B, Verheyen K, Hermy M, Muys B. 2005. Effects of landscape structure on the invasive spread of black cherry Prunus serotina in an agricultural landscape in Flanders. Belgium Ecography 28: 99109.
  • Delcourt HR. 1987. The impact of prehistoric agriculture and land occupation on natural vegetation. Trends in Ecology and Evolution 2: 3944.
  • Dietz H, Edwards PJ. 2006. Recognition that causal processes change during plant invasion helps explain conflicts in evidence. Ecology 87: 13591367.
  • DiTomaso JM. 2000. Invasive weeds in rangelands. Species, impacts, and management. Weed Science 48: 255265.
  • Dukes JS. 2000. Will increasing atmospheric CO2 affect the success of invasive species?. In: MooneyHA, HobbsRJ, eds. Invasive species in a changing world. Washington, DC, USA: Island Press, 95113.
  • Dukes JS. 2001. Biodiversity and invasibility in grassland microcosms. Oecologia 126: 563568.
  • Dukes JS. 2002. Comparison of the effect of elevated CO2 on an invasive species (Centaurea solstitialis) in monoculture and community settings. Plant Ecology 160: 225234.
  • Dukes JS, Mooney HA. 1999. Does global change increase the success of biological invaders? Trends in Ecology and Evolution 14: 135139.
  • Dukes JS, Mooney HA. 2004. Disruption of ecosystem processes in western North America by invasive species. Revista Chilena de Historia Natural 77: 411437.
  • Elton CS. 1958. The ecology of invasions by animals and plants. London, UK: Methuen.
  • Esler AE. 1987. The naturalization of plants in urban Auckland, New Zealand. 1. Introduction and spread of naturalized plants. New Zealand Journal of Botany 25: 511522.
  • Fargione J, Brown CS, Tilman D. 2003. Community assembly and invasion: An experimental test of neutral versus niche processes. Proceedings of the National Academy of Sciences, USA 100: 89168920.
  • Fargione JE, Tilman D. 2005. Diversity decreases invasion via both sampling and complementarity effects. Ecology Letters 8: 604611.
  • Foster BL, Dickson TL, Murphy CA, Karel IS, Smith VH. 2004. Propagule pools mediate community assembly and diversity-ecosystem regulation along a grassland productivity gradient. Journal of Ecology 92: 435449.
  • Funk JL, Vitousek PM. 2007. Resource-use efficiency and plant invasion in low-resource systems. Nature 446: 10791081.
  • Garcia-Ramos G, Rodriguez D. 2002. Evolutionary speed of species invasions. Evolution 56: 661668.
  • Gelbard JL, Harrison S. 2003. Roadless habitats as refuges for native grasslands: Interactions with soil, aspect, and grazing. Ecological Applications 13: 404415.
  • Gerlach JD. 1997. How the West was lost. Reconstructing the invasion dynamics of yellow starthistle and other plant invaders of Western rangelands and natural areas. In: KellyM, WagnerE, WarnerP, eds. California Exotic Pest Plant Council Symposium, Vol. 3. Concord, CA, USA, 6772.
  • Goodwin BJ, McAllister AJ, Fahrig L. 1999. Predicting invasiveness of plant species based on biological information. Conservation Biology 13: 422426.
  • Grime JP. 1974. Vegetation classification by reference to strategy. Nature 250: 2630.
  • Hamilton MA, Murray BR, Cadotte MW, Hose GC, Baker AC, Harris CJ, Licari D. 2005. Life-history correlates of plant invasiveness at regional and continental scales. Ecology Letters 8: 10661074.
  • Harding KC, McNamara JM, Holt JD. 2006. Understanding invasions in patchy habitats through metapopulation theory. In: CadotteMW, McMahonSM, FukamiT, eds. Conceptual ecology and invasion biology: reciprocal approaches to nature. Dordrecht, the Netherlands: Springer, 371403.
  • Harrison S, Rice K, Maron J. 2001. Habitat patchiness promotes invasion by alien grasses on serpentine soil. Biological Conservation 100: 4553.
  • Hastings A, Cuddington K, Davies KF, Dugaw CJ, Elmendorf S, Freestone A, Harrison S, Holland M, Lambrinos J, Malvadkar U, Melbourne BA, Moore K, Taylor C, Thomson D. 2005. The spatial spread of invasions: new developments in theory and evidence. Ecology Letters 8: 91101.
  • Hector A, Dobson K, Minns A, Bazeley-White E, Lawton JH. 2001. Community diversity and invasion resistance: An experimental test in a grassland ecosystem and a review of comparable studies. Ecological Research 16: 819831.
  • Hierro JL, Villarreal D, Eren O, Graham JM, Callaway RM. 2006. Disturbance facilitates invasion: the effects are stronger abroad than at home. American Naturalist 168: 144156.
  • Higgins SI, Richardson DM. 1999. Predicting plant migration rates in a changing world: The role of long-distance dispersal. American Naturalist 153: 464475.
  • Hobbs RJ. 2000. Land-use changes and invasions. In: MooneyHA, HobbsRJ, eds. Invasive species in a changing world. Washington, DC, USA: Island Press, 5564.
  • Hobbs RJ, Huenneke LF. 1992. Disturbance, diversity, and invasion – implications for conservations. Conservation Biology 6: 324337.
  • Hobbs RJ, Humphries SE. 1995. An integrated approach to the ecology and management of plant invasions. Conservation Biology 9: 761770.
  • Hobbs RJ, Mooney HA. 1991. Effects of rainfall variability and gopher disturbance on serpentine annual grassland dynamics. Ecology 72: 5968.
  • Hooper DU, Chapin FS, Ewel JJ, Hector A, Inchausti P, Lavorel S, Lawton JH, Lodge DM, Loreau M, Naeem S, Schmid B, Setala H, Symstad AJ, Vandermeer J, Wardle DA. 2005. Effects of biodiversity on ecosystem functioning: a consensus of current knowledge. Ecological Monographs 75: 335.
  • Huenneke LF. 1997. Outlook for plant invasions: Interactions with other agents of global change. In: LukenJO, ThieretJW, eds. Assessment and management of plant invasions. New York, NY, USA: Springer-Verlag, 95103.
  • Huenneke LF, Hamburg SP, Koide R, Mooney HA, Vitousek PM. 1990. Effects of soil resources on plant invasion and community structure in Californian serpentine grassland. Ecology 71: 478491.
  • Hughes RF, Denslow JS. 2005. Invasion by a N2-fixing tree alters function and structure in wet lowland forests of Hawaii. Ecological Applications 15: 16151628.
  • Huston MA. 1994. Biological diversity: the coexistence of species on changing landscapes. Cambridge, UK: Cambridge University Press.
  • Keane RM, Crawley MJ. 2002. Exotic plant invasions and the enemy release hypothesis. Trends in Ecology and Evolution 17: 164170.
  • Keitt TH, Lewis MA, Holt RD. 2001. Allee effects, invasion pinning, and species’ borders. American Naturalist 157: 203216.
  • Kennedy TA, Naeem S, Howe KM, Knops JMH, Tilman D, Reich P. 2002. Biodiversity as a barrier to ecological invasion. Nature 417: 636638.
  • Kim KD. 2005. Invasive plants on disturbed Korean sand dunes. Estuarine Coastal and Shelf Science 62: 353364.
  • Klironomos JN. 2002. Feedback with soil biota contributes to plant rarity and invasiveness in communities. Nature 417: 6770.
  • Knight KS, Reich PB. 2005. Opposite relationships between invasibility and native species richness at patch versus landscape scales. Oikos 109: 8188.
  • Knops JMH, Tilman D, Haddad NM, Naeem S, Mitchell CE, Haarstad J, Ritchie ME, Howe KM, Reich PB, Siemann E, Groth J. 1999. Effects of plant species richness on invasion dynamics, disease outbreaks, insect abundances and diversity. Ecology Letters 2: 286293.
  • Kolar CS, Lodge DM. 2001. Progress in invasion biology: predicting invaders. Trends in Ecology and Evolution 16: 199204.
  • Kot M, Lewis MA, Van Den Driessche P. 1996. Dispersal data and the spread of invading organisms. Ecology 77: 20272042.
  • Kriticos DJ, Sutherst RW, Brown JR, Adkins SW, Maywald GF. 2003. Climate change and the potential distribution of an invasive alien plant: Acacia nilotica ssp. indica in Australia. Journal of Applied Ecology 40: 111124.
  • Lavergne S, Molofsky J. 2007. Increased genetic variation and evolutionary potential drive the success of an invasive grass. Proceedings of the National Academy of Sciences, USA 104: 38833888.
  • Le Maitre DC, Richardson DM, Chapman RA. 2004. Alien plant invasions in South Africa: driving forces and the human dimension. South African Journal of Science 100: 103112.
  • Lee CE. 2002. Evolutionary genetics of invasive species. Trends in Ecology and Evolution 17: 386391.
  • Leger EA, Rice KJ. 2003. Invasive California poppies (Eschscholzia californica Cham.) grow larger than native individuals under reduced competition. Ecology Letters 6: 257264.
  • Leishman MR, Thomson VP. 2005. Experimental evidence for the effects of additional water, nutrients and physical disturbance on invasive plants in low fertility Hawkesbury Sandstone soils, Sydney, Australia. Journal of Ecology 93: 3849.
  • Levine JM. 2000. Species diversity and biological invasions: Relating local process to community pattern. Science 288: 852854.
  • Levine JM, Adler PB, Yelenik SG. 2004. A meta-analysis of biotic resistance to exotic plant invasions. Ecology Letters 7: 975989.
  • Levine JM, Vila M, D’Antonio CM, Dukes JS, Grigulis K, Lavorel S. 2003. Mechanisms underlying the impacts of exotic plant invasions. Proceedings of the Royal Society of London Series B – Biological Sciences 270: 775781.
  • Lewis MA, Kareiva P. 1993. Allee dynamics and the spread of invading organisms. Theoretical Population Biology 43: 141158.
  • Lloret F, Medail F, Brundu G, Camarda I, Moragues E, Rita J, Lambdon P, Hulme PE. 2005. Species attributes and invasion success by alien plants on Mediterranean islands. Journal of Ecology 93: 512520.
  • Lockwood JL, Cassey P, Blackburn T. 2005. The role of propagule pressure in explaining species invasions. Trends in Ecology and Evolution 20: 223228.
  • Lonsdale WM. 1999. Global patterns of plant invasions and the concept of invasibility. Ecology 80: 15221536.
  • Lonsdale WM, Lane AM. 1994. Tourist vehicles as vectors of weed seeds in Kakadu National Park, Northern Australia. Biological Conservation 69: 277283.
  • Losure DA, Wilsey BJ, Moloney KA. 2007. Evenness–invasibility relationships differ between two extinction scenarios in tallgrass prairie. Oikos 116: 8798.
  • Lyons KG, Schwartz MW. 2001. Rare species loss alters ecosystem function – invasion resistance. Ecology Letters 4: 358365.
  • Macdonald IAW, Clark DL, Taylor HC. 1989. The history and effects of alien plant control in the Cape of Good Hope Nature Reserve, 1941–87. South African Journal of Botany 55: 5675.
  • MacDonald IAW, Graber DM, De Benedettii S, Groves RH, Fuentes ER. 1988. Introduced species in nature reserves in Mediterranean-type climatic regions of the world. Biological Conservation 44: 3766.
  • Mack RN. 1995. Understanding the processes of weed invasions: the influence of environmental stochasticity. In: StirtonCH, ed. Weeds in a changing world. Proceedings No. 64. Brighton, UK: British Crop Protection Council, 6574.
  • Mack RN. 1999. The motivation for importing potentially invasive plant species. A primal urge? In: EldridgeD, FreudenbergerD, eds. Proceedings of the VIth International Rangeland Congress, Vol. 2. Aitkenvale, Australia: International Rangeland Congress, 557562.
  • Mack RN. 2000. Cultivation fosters plant naturalization by reducing environmental stocasticity. Biological Invasions 2: 111122.
  • Mack RN, Lonsdale WM. 2001. Humans as global plant dispersers: Getting more than we bargained for. Bioscience 51: 95102.
  • Mack RN, Simberloff D, Lonsdale WM, Evans H, Clout M, Bazzaz FA. 2000. Biotic invasions: Causes, epidemiology, global consequences, and control. Ecological Applications 10: 689710.
  • Maron JL, Vila M. 2001. When do herbivores affect plant invasion? Evidence for the natural enemies and biotic resistance hypotheses. Oikos 95: 361373.
  • McDonald RI, Urban DL. 2006. Edge effects on species composition and exotic species abundance in the North Carolina Piedmont. Biological Invasions 8: 10491060.
  • Melbourne BA, Cornell HA, Davies KF, Dugaw CJ, Elmendorf S, Freestone AL, Hall RJ, Harrison S, Hastings A, Holland M, Holyoak M, Lambrinos J, Moore K, Yokomizo H. 2007. Invasion in a heterogeneous world: resistance, coexistence or hostile takeover? Ecology Letters 10: 7794.
  • Meyers JA, Vellend M, Gardescu S, Marks PL. 2004. Seed dispersal by white-tailed deer: implications for long-distance dispersal, invasion, and migration of plants in eastern North America. Oecologia 139: 3544.
  • Milbau A, Nijs I, De Raedemaecker F, Reheul D, De Cauwer B. 2005. Invasion in grassland gaps: the role of neighbourhood richness, light availability and species complementarity during two successive years. Functional Ecology 19: 2737.
  • Miller KE, Gorchov DL. 2004. The invasive shrub, Lonicera maackii, reduces growth and fecundity of perennial forest herbs. Oecologia 139: 359375.
  • Mouquet N, Munguia P, Kneitel JM, Miller TE. 2003. Community assembly time and the relationship between local and regional species richness. Oikos 103: 618626.
  • Murphy HT, VanDerWal J, Lovett-Doust L, Lovett-Doust J. 2006. Invasiveness in exotic plants: immigration in an ecological continuum. In: CadotteMW, McMahonSM, FukamiT, eds. Conceptual ecology and invasion biology: reciprocal approaches to nature. Knoxville, TN, USA: Springer, 65105.
  • Naeem S, Knops JMH, Tilman D, Howe KM, Kennedy T, Gale S. 2000. Plant diversity increases resistance to invasion in the absence of covarying extrinsic factors. Oikos 91: 97108.
  • Neubert MG, Caswell H. 2000. Demography and dispersal: Calculation and sensitivity analysis of invasion speed for structured populations. Ecology 81: 16131628.
  • Ohlemuller R, Walker S, Wilson JB. 2006. Local vs regional factors as determinants of the invasibility of indigenous forest fragments by alien plant species. Oikos 112: 493501.
  • Parendes LA, Jones JA. 2000. Role of light availability and dispersal in exotic plant invasion along roads and streams in the H. J. Andrews Experimental Forest, Oregon. Conservation Biology 14: 6475.
  • Parker JD, Burkepile DE, Hay ME. 2006. Opposing effects of native and exotic herbivores on plant invasions. Science 311: 14591461.
  • Pauchard A, Shea K. 2006. Integrating the study of non-native plant invasions across spatial scales. Biological Invasions 8: 399413.
  • Perrings C, Dehnen-Schmutz K, Touza J, Williamson M. 2005. How to manage biological invasions under globalization. Trends in Ecology and Evolution 20: 212215.
  • Pimentel D, Zuniga R, Morrison D. 2005. Update on the environmental and economic costs associated with alien-invasive species in the United States. Ecological Economics 52: 273288.
  • Prieur-Richard AH, Lavorel S, Grigulis K, Dos Santos A. 2000. Plant community diversity and invasibility by exotics: invasion of Mediterranean old fields by Conyza bonariensis and Conyza canadensis. Ecology Letters 3: 412422.
  • Pyle LL. 1995. Effects of disturbance on herbaceous exotic plant species on the floodplain of the Potomac River. The American Midland Naturalist 134: 244253.
  • Pysek P. 1998. Is there a taxonomic pattern to plant invasions? Oikos 82: 282294.
  • Pysek P, Hulme PE. 2005. Spatio-temporal dynamics of plant invasions: Linking pattern to process. Ecoscience 12: 302315.
  • Raghu S, Anderson RC, Daehler CC, Davis AS, Wiedenmann RN, Simberloff D, Mack RN. 2006. Adding biofuels to the invasive species fire? Science 313: 1742.
  • Reichard SH, White P. 2001. Horticulture as a pathway of invasive plant introductions in the United States. Bioscience 51: 103113.
  • Reinhart KO, Callaway RM. 2006. Soil biota and invasive plants. New Phytologist 170: 445457.
  • Reinhart KO, Packer A, Van der Putten WH, Clay K. 2003. Plant–soil biota interactions and spatial distribution of black cherry in its native and invasive ranges. Ecology Letters 6: 10461050.
  • Reinhart KO, Royo AA, Van Der Pattern WH, Clay K. 2005. Soil feedback and pathogen activity in Prunus serotina throughout its native range. Journal of Ecology 93: 890898.
  • Rejmanek M. 1996. A theory of seed plant invasiveness: The first sketch. Biological Conservation 78: 171181.
  • Rejmanek M, Richardson DM. 1996. What attributes make some plant species more invasive? Ecology 77: 16551661.
  • Richardson DM, Allsopp N, D’Antonio CM, Milton SJ, Rejmanek M. 2000a. Plant invasions – the role of mutualisms. Biological Reviews 75: 6593.
  • Richardson DM, Pyček P, Rejmánek M, Barbour MG, Penetta FD, West. CJ. 2000b. Naturalization and invasion of alien plants: concepts and definitions. Diversity and Distributions 6: 93107.
    Direct Link:
  • Richardson DM, Rejmanek M. 2004. Conifers as invasive aliens: a global survey and predictive framework. Diversity and Distributions 10: 321331.
  • Ricklefs RE. 1987. Community diversity – relative roles of local and regional processes. Science 235: 167171.
  • Ricklefs RE. 2004. A comprehensive framework for global patterns in biodiversity. Ecology Letters 7: 115.
  • Rose S. 1997. Influence of suburban edges on invasion of Pittosporum undulatum into the bushland of northern Sydney, Australia. Australian Journal of Ecology 22: 8999.
  • Rubino DL, Williams CE, Moriarity WJ. 2002. Herbaceous layer contrast and alien plant occurrence in utility corridors and riparian forests of the Allegheny High Plateau. Journal of the Torrey Botanical Society 129: 125135.
  • Van Ruijven J, De Deyn GB, Berendse F. 2003. Diversity reduces invasibility in experimental plant communities: the role of plant species. Ecology Letters 6: 910918.
  • Sakai AK, Allendorf FW, Holt JS, Lodge DM, Molofsky J, With KA, Baughman S, Cabin RJ, Cohen JE, Ellstrand NC, McCauley DE, O’Neil P, Parker IM, Thompson JN, Weller SG. 2001. The population biology of invasive species. Annual Review of Ecology and Systematics 32: 305332.
  • Saunders DA, Hobbs RJ, Margules CR. 1991. Biological consequences of ecosystem fragmentation – a review. Conservation Biology 5: 1832.
  • Sax DF, Gaines SD, Brown JH. 2002. Species invasions exceed extinctions on islands worldwide: a comparative study of plants and birds. American Naturalist 160: 766783.
  • Schluter D, Ricklefs RE. 1993. Species diversity: An introduction to the problem. In: RicklefsRE, SchluterD, eds. Species diversity in ecological communities. Chicago, IL, USA: University of Chicago Press, 110.
  • Schweitzer JA, Larson KC. 1999. Greater morphological plasticity of exotic honeysuckle species may make them better invaders than native species. Journal of the Torrey Botanical Society 126: 1523.
  • Searcy KB, Pucko C, McClelland D. 2006. The distribution and habitat preferences of introduced species in the Mount Holyoke Range, Hampshire Co., Massachusetts. Rhodora 108: 4361.
  • Shea K, Chesson P. 2002. Community ecology theory as a framework for biological invasions. Trends in Ecology and Evolution 17: 170176.
  • Simberloff D. 2000. Global climate change and introduced species in United States forests. Science of the Total Environment 262: 253261.
  • Simberloff D, Von Holle B. 1999. Positive interactions of nonindigenous species: invasional meltdown? Biological Invasions 1: 2132.
  • Smith SA, Shurin JB. 2006. Room for one more? Evidence for invasibility and saturation in ecological communities. In: CadotteMW, McMahonSM, FukamiT, eds. Conceptual ecology and invasion biology: reciprocal approaches to nature. Dordrecht, the Netherlands: Springer, 423447.
  • Smith MD, Wilcox JC, Kelly T, Knapp AK. 2004. Dominance not richness determines invasibility of tallgrass prairie. Oikos 106: 253262.
  • Stachowicz JJ, Tilman D. 2005. Species invasions and the relationships between species diversity, community saturation, and ecosystem functioning. In: SaxDF, StachowiczJJ, GainesSD, eds. Species invasions: insights into ecology, evolution and biogeography. Sunderland, MA, USA: Sinauer, 4164.
  • Stinson KA, Campbell SA, Powell JR, Wolfe BE, Callaway RM, Thelen GC, Hallett SG, Prati D, Klironomos JN. 2006. Invasive plant suppresses the growth of native tree seedlings by disrupting belowground mutualisms. PLoS Biology 4:140–145.
  • Stohlgren TJ, Binkley D, Chong GW, Kalkhan MA, Schell LD, Bull KA, Otsuki Y, Newman G, Bashkin M, Son Y. 1999. Exotic plant species invade hot spots of native plant diversity. Ecological Monographs 69: 2546.
  • Stohlgren TJ, Otsuki Y, Villa CA, Lee M, Belnap J. 2001. Patterns of plant invasions: a case example in native species hotspots and rare habitats. Biological Invasions 3: 3750.
  • Symstad AJ. 2000. A test of the effects of functional group richness and composition on grassland invasibility. Ecology 81: 99109.
  • Taylor CM, Hastings A. 2005. Allee effects in biological invasions. Ecology Letters 8: 895908.
  • Teo DHL, Tan HTW, Corlett RT, Wong CM, Lum SKY. 2003. Continental rain forest fragments in Singapore resist invasion by exotic plants. Journal of Biogeography 30: 305310.
  • Tilman D. 1994. Competition and biodiversity in spatially structured habitats. Ecology 75: 216.
  • Tilman D. 2004. Niche tradeoffs, neutrality, and community structure: a stochastic theory of resource competition, invasion, and community assembly. Proceedings of the National Academy of Sciences, USA 101: 1085410861.
  • Timmens SM, Williams PA. 1991. Weed numbers in New Zealand's forest and scrub reserves. New Zealand Journal of Ecology 15: 153162.
  • Trombulak SC, Frissell CA. 2000. Review of ecological effects of roads on terrestrial and aquatic communities. Conservation Biology 14: 1830.
  • Turnbull LA, Rahm S, Baudois O, Eichenberger-Glinz S, Wacker L, Schmid B. 2005. Experimental invasion by legumes reveals non-random assembly rules in grassland communities. Journal of Ecology 93: 10621070.
  • Vellend M. 2002. A pest and an invader: white-tailed deer (Odocoileus virginianus Zimm.) as a seed dispersal agent for honeysuckle shrubs (Lonicera L.). Natural Areas Journal 22: 230234.
  • Vermeij GJ. 1996. An agenda for invasion biology. Biological Conservation 78: 39.
  • Vermeij GJ. 2005. Invasion as expectation. In: SaxDF, StachowiczJJ, GainesSD, eds. Species invasions: insights into ecology, evolution and biogeography. Sunderland, MA, USA: Sinauer Associates, 315339.
  • Vitousek PM, Dantonio CM, Loope LL, Westbrooks R. 1996. Biological invasions as global environmental change. American Scientist 84: 468478.
  • Vitousek PM, Walker LR. 1989. Biological invasion by Myrica faya in Hawai’i: plant demography, nitrogen fixation, and ecosystem effects. Ecological Monographs 59: 247265.
  • Vitousek PM, Walker LR, Whiteaker LD, Mueller-Dombois D, Matson PA. 1987. Biological invasion by Myrica faya alters ecosystem development in Hawaii. Science 238: 802804.
  • Wardle DA. 2001. Experimental demonstration that plant diversity reduces invasibility – evidence of a biological mechanism or a consequence of sampling effect? Oikos 95: 161170.
  • Watkins RZ, Chen JQ, Pickens J, Brosofske KD. 2003. Effects of forest roads on understory plants in a managed hardwood landscape. Conservation Biology 17: 411419.
  • Weinig C. 2000. Plasticity versus canalization: population differences in the timing of shade-avoidance responses. Evolution 54: 441451.
  • Weltzin JF, Belote RT, Sanders NJ. 2003. Biological invaders in a greenhouse world: will elevated CO2 fuel plant invasions? Frontiers in Ecology and the Environment 1: 146153.
  • Williams SC, Ward JS. 2006. Exotic seed dispersal by white-tailed deer in southern Connecticut. Natural Areas Journal 26: 383390.
  • Williamson M. 1999. Biological invasions. London, UK: Chapman & Hall.
  • Williamson M, Fitter A. 1996. The varying success of invaders. Ecology 77: 16611666.
  • With KA. 2002. The landscape ecology of invasive spread. Conservation Biology 16: 11921203.
  • With KA. 2004. Assessing the risk of invasive spread in fragmented landscapes. Risk Analysis 24: 803815.
  • Wolfe BE, Klironomos JN. 2005. Breaking new ground: Soil communities and exotic plant invasion. Bioscience 55: 477487.
  • Yelenik SG, Stock WD, Richardson DM. 2004. Ecosystem level impacts of invasive Acacia saligna in the South African fynbos. Restoration Ecology 12: 4451.
  • Zavaleta ES, Hulvey KB. 2004. Realistic species losses disproportionately reduce grassland resistance to biological invaders. Science 306: 11751177.