Invasive Species: Plants

Invasive Species: Plants RC Godfree, Commonwealth Scientific and Industrial Research Organisation (CSIRO) Plant Industry, Canberra, ACT, Australia BR Murray, University of Technology Sydney, Broadway, NSW, Australia r 2014 Elsevier Inc. All rights reserved.

Glossary Biotic resistance hypothesis The hypothesis that competition and other biotic interactions with native plant species explain the tendency for less disturbed plant communities to suffer relatively low rates of invasion. EICA hypothesis The hypothesis that the lack of native pathogens, predators, and herbivores experienced by nonindigenous plant species in new habitats allows re-allocation of resources away from defense and toward development and growth, in turn resulting in increased invasiveness. Exposure studies Studies that are conducted to determine the level of risk posed by an organism to the environment, often involving tiered assessment of the likelihood of each step of a given hazard scenario occurring. Invasion barriers Biotic, abiotic, and geographical factors that prevent nonindigenous plant species from moving between invasion stages.

Introduction As human societies have developed they have increasingly broken down the spatial barriers that historically separated species, communities, and ecosystems. The result has been an ever-increasing rate of long-distance dispersal of both animal and plant species to new habitats and the concomitant taxonomic homogenization of previously distinct floral and faunal assemblages. Today, this process continues virtually unabated, and invasive species are considered to pose one of the primary contemporary threats to biodiversity conservation and agricultural productivity in both native and agricultural systems on a global scale. Indeed, it is no exaggeration to state that synergistic interactions between invasive plants and other agents of global change, including anthropogenic global warming (AGW), rising atmospheric carbon dioxide concentrations, nitrogen deposition, habitat fragmentation, and disturbance, are likely to leave few if any natural or agroecosystems unchanged over the next century. The economic costs of plant invasions alone are staggering. It has been estimated that in 2005 the direct annual economic impact of weedy plants on US agricultural production exceeded US$ 25 billion, with losses mainly caused by reduced crop and pasture yields and the cost of weed control (Pimentel et al., 2005). Similarly, a thorough assessment of the economic impact of weeds in Australia in 2004 revealed annual losses associated with agriculture of AU$ 4 billion, and costs of invasive species management in natural systems of at least AU$ 19 million. A further AU$ 80–90 million was spent on weed management and research by public authorities and the

66

Naturalization The process during which a plant species establishes viable populations in a new location. Nontarget environments Environments that may be placed at risk by the unwanted effects of a novel organism that lies outside the area originally targeted for its release. Phenotypic plasticity The ability of an organism to alter its phenotype in response to changing environmental conditions. Polyploid An organism containing more than two sets of homologous chromosomes. Preadaptation In the context of invasive species, referring to the ability of nonindigenous species to survive in newly colonized habitats by expressing preexisting adaptive traits. Propagule pressure A measure of the number of individuals of a nonindigenous species released into a given region, incorporating both the number of release events and the number of individuals per event.

Australian government (Sinden et al., 2004). The direct impacts of nonindigenous species on natural and seminatural ecosystems are also severe: In addition to direct negative impacts on biodiversity such as species extinction, many of the most aggressive nonindigenous species alter the structure, composition, and function of recipient plant communities, processes that are driving change in ecosystems on a global scale. For these and other reasons, plant invasions have long been of interest to ecologists. In the five decades since Charles Elton published his classic 1958 book ‘The Ecology of Invasions by Animals and Plants’ (Elton, 1958), ecologists and agronomists have sought to develop a unified understanding of how, why, and under what circumstances plant invasions occur, and the strategies that may be most successfully adopted for their control. Although much of this research has proved extremely fruitful, leading to successful systems for both identifying potentially invasive species and managing their impacts once established, some areas remain difficult or controversial or are compromised by a lack of data. Indeed, at the most basic level our understanding of the biological mechanisms that allow some species to dominate plant assemblages following introduction to new habitat remains limited, and integrating new data and theory from across ecology, evolutionary biology, and ecophysiology will be a critical ongoing challenge for invasion biologists in coming years (Richardson and Pyšek, 2008). This article is intended to provide a basic overview of current knowledge in invasion biology and to highlight active areas of research. The authors first describe the basic invasion

Encyclopedia of Agriculture and Food Systems, Volume 4

doi:10.1016/B978-0-444-52512-3.00202-3

Invasive Species: Plants

pathway, and consider the barriers that nonindigenous species must overcome before becoming invasive in a new geographic range. They illustrate key principles associated with the invasion pathway using data from Australia, which provide an excellent case study of invasion of a geographically isolated land mass by introduced plants. Then they consider evolution and other genetic processes associated with the establishment and spread of invasive species and turn our attention to the mechanisms of their impact on natural ecosystems and their roles as ecosystem engineers. They then discuss two current questions of great importance in contemporary invasion biology. The first is whether climate change is likely to confer an overall fitness advantage on nonindigenous species relative to cooccurring native plant species, and why. The second is whether genetically modified crops or their adoption is likely to cause increased weed pressure on target and nontarget natural and agricultural ecosystems – a matter of great controversy over the past decade. They conclude by briefly discussing opportunities for future research. Unfortunately, invasion biology has long been plagued by the use of vague or contradictory terminology to describe processes or stages of the invasion pathway and the plants involved (Richardson et al., 2000). The authors begin, therefore, by defining some important terms that will be used throughout the article. The term nonindigenous plant species (NIPS) refers to species occurring outside of their native range, and is used interchangeably with terms such as ‘alien’ and ‘exotic’ species. Invasion refers to the process in which introduced plants reproduce outside of the original immediate area of introduction and does not include reference to any damage cause by the plant. Weeds are invasive species, especially in agricultural systems, that cause harm or are otherwise undesirable.

The Biology of Plant Invasions The Invasion Pathway and the Australian Experience The end result of the successful introduction of NIPS to new regions can be a situation where introduced species become geographically widespread, overabundant, and cause an enormous amount of ecological and economic damage. To get to this situation, however, plant species require the ability to traverse several invasion stages (Theoharides and Dukes, 2007) and to surpass several important ‘invasion barriers’ (see Figure 1; Satai et al., 2001). The first stage of the invasion pathway, ‘Transport,’ is associated with NIPS overcoming a significant natural barrier of geography, one that NIPS could not disperse across naturally. Somehow, NIPS must cross vast distances over land and/or sea. The most striking characteristic of modern-day plant invasions is that this transport stage is facilitated by a diverse range of human activities. For instance, many plant species are deliberately brought into new regions for ornamental purposes (Groves et al., 2005). A good example of global translocation of NIPS is the Singapore Daisy (Sphagneticola trilobata), which was originally brought into Australia as a garden ornamental and deliberately planted as a roadside and railway embankment stabilizer in the north-east of the

67

Species in native range

Barrier: Geography Transport stage

Transport to new range

Barrier: Abiotic conditions Colonization stage

Survival in new range

Barrier: Biotic interactions Establishment stage

Population growth

Barrier: Landscape factors Landscape spread stage

Landscape spread

Figure 1 Typical invasion pathway showing the four main invasion stages (transport, colonization, establishment, landscape spread) and barriers faced by plants associated with each stage. Adapted from Hellmann, J.J., Byres, J.E., Bierwagen, B.G., Dukes, J.S., 2008. Five potential consequences of climate change for invasive species. Conservation Biology 22, 534–543.

continent. Some NIPS have been introduced to new regions to be used for a variety of purposes. For example, Mesquite species (Prosopis spp.) were introduced to Australia as fodder for livestock and as ornamental species in station homestead or town gardens, and were also used in mine dumps and other soil stabilization programs. Importantly, a large number of NIPS can be introduced to new regions outside their natural geographic ranges accidentally, and thus can be considered unintentional introductions. For instance, Whiskey Grass (Andropogon virginicus) was introduced accidentally to Australia when it was used as protective packaging for American bottles of whiskey. It would appear that irrespective of the reason for introduction, there is the potential for any NIPS to become a serious agricultural weed, whether introduced purposefully or accidentally. Indeed, all three plant examples provided above are now considered serious weeds of Australian agriculture. The second stage of the invasion pathway, ‘Colonization,’ or ‘Introduction,’ can be successfully traversed by NIPS in one of two ways. First, some species need to overcome a barrier in the form of cultivation. For example, many NIPS exist in cultivation beyond the limits of their native ranges, but they do not form self-sustaining populations outside of cultivation. The key concern here is the fact that some species can escape cultivation to become established in the wild. Using Australian examples to illustrate this idea, two extreme examples of escapees from cultivation, which are considered to have ‘jumped the garden fence,’ are Paterson's curse (Echium plantagineum), which costs agriculture AU$30 million per year, and Lippia (Phyla canescens), which costs the grazing industry an estimated AU$38 million per year. The second way NIPS can traverse the colonization stage is by avoiding the cultivation

68

Invasive Species: Plants

barrier; this happens when a species has been introduced accidentally by humans directly into the new environment. More than 27 000 NIPS are known to have been introduced to Australia; however, only approximately 10% of these species have surpassed the third stage of the invasion pathway, the ‘Establishment’ stage. For plants, this is also commonly referred to as the ‘Naturalization’ stage. An invasion ecology generality that appears to be fairly consistent globally is the Tens rule (Williaimson and Fitter, 1996), whereby only approximately 10% of the introduced flora of a region will ever become naturalized (although it is worth noting that this is an average figure and recent studies have documented much higher percentages in some regions of the world). Furthermore, approximately 10% of naturalized NIPS will transition to become invasive weeds (see below, the ‘Landscape spread’ stage). Interestingly, NIPS introduced as garden ornamentals are the dominant source of new naturalized plants and weeds in Australia. Of the roughly 2700 NIPS now known to be established in the Australian environment, approximately 1800 species are escaped garden plant species. The rate of increase in the number of naturalized plant species in Australia has tended to be linear until recent times. Recent work has reported an increase in the number of species naturalizing in Australia, leading to the conclusion that the rate of naturalization is increasing. What does it take for NIPS to become naturalized? A species is generally considered naturalized when there is evidence for the existence of a self-sustaining population over a period of time corresponding to multiple generations. It is rare to find published work that compares the attributes of species that have become naturalized with other species introduced to the same region that have not become naturalized. Recent comparative work, however, for North American plant species that have become naturalized in Europe has shown that introduction history, in particular increased planting frequency, is an important determinant of current naturalization success among these exotic invaders in Europe (Bucharova and van Kleunen, 2009). Life-history traits of species can also be important, but this depends on plant growth form. For example, the naturalization success of trees was found to increase positively with maximum plant height and seed spread rate. This was not the case for other life forms. The fourth and final stage of the invasion pathway is ‘Landscape spread.’ Arguably, this is the most important part of the invasion pathway. If NIPS fail to get past the barriers of dispersal and subsequent survival during population spread in the introduced range, then they do not have the potential to become serious weeds of the environment and/or agriculture. Few species pass through this invasion stage successfully. Indeed, current estimates of the number of invasive NIPS in Australia put the number at approximately 130 species. This figure is currently being reviewed, and it is likely to nearly double in the future to nearly 250 species. Interestingly, this puts the number of invasive NIPS in Australia at 10% of the total number of naturalized species, a percentage that follows the aforementioned Tens rule. A fundamental question of invasion ecology is just how long it takes NIPS to shift from the state of naturalization to become invasive. The length of time between these two states is referred to as the lag time or lag phase (although it can

sometimes refer to the period of time between introduction and invasion). Lag times are notoriously difficult to determine, as it is not often known when a species has become naturalized. Nevertheless, some studies have described lag times and it is clear that these can vary considerably. The majority of species do go through a lag phase, which appears to be approximately 50 years as a minimum for NIPS in Australia. Recent work examining 23 NIPS introduced to Hawai'i that eventually became invasive pests has revealed that the average lag time between introduction and first evidence of spread is 14 years for woody plants and 5 years for herbaceous plants. Such short lag times provide equally worrying short times for managers attempting to manage the spread of new invaders. Perhaps one of the most important questions for invasion ecology and for the quest for an understanding of just what it is that makes species invasive is what traits invasive NIPS possess that allow them to become successful invaders (Drenovsky et al., 2012). One approach that has been adopted over the years has been to compare the traits of invasive NIPS with resident native species. Unfortunately, this approach has yet to reveal any globally consistent biological differences. A more relevant approach has been the exploration of traits that differ between naturalized noninvasive NIPS and invasive NIPS. The idea here is that by using a ‘target-area’ approach, one can compare NIPS that cooccur within their introduced range for fundamental differences in their attributes. Given its long-term geographic isolation, but relatively recent settlement by nonindigenous peoples, Australia provides a unique opportunity to study the invasion process and the factors that contribute to invasion success in NIPS (Murray and Phillips, 2010; Phillips and Murray, 2012). A substantial body of evidence is emerging to show that invasive NIPS differ from noninvasive NIPS in both their introduction history and life history. Introduction history refers to traits of species related to where they have been introduced from, why they were introduced, and how long they have been resident in the novel range. Indeed, residence time has emerged as a highly important trait for successful invasion. NIPS that have been present in the new range for longer periods of time are more likely to have become invasive. Life-history traits of species refer to those characteristics of species related to growth, survival, and reproduction. Interestingly, recent work for Australia has shown that species with small seeds (correlated with increased reproductive output) are more likely to become invasive than larger-seeded species. Furthermore, the seeds of invasive NIPS are more likely to have had more resources invested by the parent plant into their dispersal adaptations (e.g., wings for wind dispersal, fleshy fruit for dispersal by vertebrate foragers).

Evolution in Invasive Species One of the most interesting paradoxes of invasion biology is that many invasive species, despite originating from numerically small founder populations, adapt extremely well to new environments and undergo rapid population growth and range expansion. Because bottlenecked populations typically have low genetic diversity and hence should have reduced evolutionary potential, it is argued that this phenomenon could be simply explained by high levels of ‘preadaptation’

Invasive Species: Plants

among the most successful exotic species with traits that lead to high fitness in new environments (e.g., high dispersal rates and competitive ability). Such species may remain highly invasive despite suffering some diminishment of genetic diversity in response to inbreeding. Although this may be broadly true, the fact remains that many of the best examples of rapid evolution come from invasive NIPS, and the critical role that evolutionary adaptation can play in the invasion process is now becoming widely recognized among biologists (Whitney and Gabler, 2008). Indeed, mounting evidence suggests that a range of genetic processes, even among small founder populations, can facilitate plant invasion of new habitats (Lee, 2002). Most of these involve increases in additive genetic variance, which must be present for evolution to occur in response to selection pressure. Two of the most important mechanisms include hybridization of invasive species with close relatives in the new range (interspecific) and admixture of previously separated genotypes in the introduced range (intraspecific). One of the most widely cited examples of interspecific hybridization leading to increased hybrid invasiveness is that of Spartina anglica in the British Isles (Raybould et al., 2008), an allopolyploid derived from Spartina maritima and Spartina alterniflora. Spartina anglica is now a noxious weed, whereas the parental species have relatively limited distributions. ‘Polyploid’ hybrids usually have greater fitness than diploid hybrids, and allopolyploidy has been linked to invasiveness in a number of taxa. Intraspecific genetic admixture is most likely in species with a history of multiple introductions, and can result in genetically diverse populations of invasive NIPS that have an increased evolutionary capacity (Frankham, 2005). Several other genetic processes can also apparently increase the fitness of invasive species. Additive genetic variance can be increased by conversion of epistatic or dominance variance by genetic drift following population bottlenecks; epistasis itself (interaction between different gene) provides a source of variance that could respond to selection pressure. In some species a small number of genes instead of total additive variance may play a critical role in determining plant fitness under new selection regimes. Of course, not all introduced species benefit significantly from these genetic processes; for instance, some invasive NIPS are known to have limited genetic diversity and evolutionary capacity as a result of small founder population size. All are also influenced to a greater or lesser extent by a range of factors such as genetic drift, small population size, trade-offs among fitness-related traits, and swamping of adapted genotypes by migration (i.e., migration-selection balance) that limit population adaptation and expansion in range-edge or marginal habitats. Evolutionary theory suggests that adaptation should occur following selection for fitness-related traits that have a heritable source of genetic variation. Sources of selection in new habitats are myriad, including climatic factors, competition with resident species, or new grazing or predation regimes. There is compelling evidence (Colautti et al., 2010) that many invasive species have undergone population differentiation or evolution in response to broad-scale climatic gradients within a matter of decades (e.g., Echium plantagineum, Hypericum perforatum, Solidago altissima, Senecio inaequidens, and

69

Lythrum salicaria; Konarzewski et al., 2012; Monty and Mahy, 2009), although the establishment of such clines is not ubiquitous. Mimicry is a fascinating evolutionary adaptation exhibited by some crop weeds in response to the strong directional selection typical of agricultural cultivation (Barrett, 1983). Under these predictable and repetitive selection regimes mimicry can develop in seed, which increasingly resembles that of the associated crop species in terms of weight, size, and overall appearance, or vegetative growth traits, which also can converge. Studies have shown that Echinochloa crus-galli var. oryzicola, for instance, is closer to rice in phenology, growth, and morphology than to its close relative Echinochloa crus-galli var. crus-galli. Rapid evolutionary responses to agricultural herbicide application also occur frequently, with crop weeds developing herbicide resistance. The implications of this process associated with the cultivation of genetically modified crops are discussed in another section of this article. Over the past decade there has been much focus on the potential for plants to experience release from specialist natural enemies (herbivores and diseases) when moved outside of their native range, resulting in increased fitness (the enemy release hypothesis; Keane and Crawley, 2002). Theory suggests that plants experiencing reduced rates of herbivory will reallocate resources away from resistance or tolerance strategies to increased growth, provided that the production of defense chemicals or tolerance mechanisms comes at a fitness cost. Evidence for evolution of increased competitive ability involving this mechanism (known as the ‘EICA hypothesis’; Blossey and Nötzold, 1995), along with other evolutionary processes, has been postulated as explaining the tendency for some NIPS to exhibit a lag phase before becoming invasive. Evidence for the EICA hypothesis was first observed in Lythrum salicaria but since then has only found partial experimental support; recently it has been suggested that resource allocation under enemy release may affect many fitness-enhancing traits, not just those relating to growth and competition. Traditionally, much of the focus of research of invasion biology has been on developing an understanding of the demographic and life-history predictors of invasiveness, on their impacts on native and agricultural ecosystems, and on means for more efficient weed control in different settings. However, given the strong emerging links between genetic background, evolutionary capacity, and invasion dynamics, it is clear that treating invasive species as uniform entities with fixed genotypes will fail to adequately capture their likely behavior when faced with changing environmental conditions and selection regimes. One area where this clearly applies, for example, is in the modeling of weed spread under future climate change scenarios. Successful integration of evolutionary and ecological concepts into a more detailed and predictive understanding of the process of plant invasions will remain one of the most interesting but challenging aspects of invasion research for years to come.

Impacts of Invasive Species on Natural and Agricultural Systems Invasive NIPS can have an impact on individual species to entire landscape scales through a wide range of genetic,

70

Invasive Species: Plants

demographic, and evolutionary mechanisms (Levine et al., 2003). These impacts can be direct, such as competition for light, water, and nutrients with native community associates, or indirect, involving complex changes to ecosystem processes. As with other components of the invasion process (Ehrenfeld, 2010), however, invasive NIPS differ significantly in their capacity to directly and indirectly drive change in recipient ecosystems. Determining when and how impacts arise, which species are likely to significantly transform ecosystems, and which communities are most susceptible to change has direct application to more effective management of extant invasive species and efficient restoration of degraded ecosystems once they have been removed. Arguably the most obvious and socioeconomically damaging direct impacts of invasive NIPS occur in agricultural systems where competition for light, water, and nutrients by weedy species reduces crop yields to a point where intensive and costly control is required. These systems, which are characterized by both impoverished faunal and floral assemblages and abundant supply of resources, are especially prone to invasion, and many crop weeds are ruderal, r-selected (fast growing) species that are generally intolerant of competition and abiotic stress. It is estimated that globally weeds reduce crop yields by approximately 9%, although figures vary across crops and depend strongly on the crop protection measures used to reduce potential losses: less developed regions suffer considerably higher losses. Along with other crop pests, weeds are expected to pose a significant challenge to global food security in coming decades, and thus a range of strategies including the development of genetically modified herbicideand pest-resistant crops (see Section Transgenic Plants as Invasive Species) are being adopted to enable more efficient weed control. A great many ecological studies have quantified changes in the composition and diversity of terrestrial plant assemblages following invasion by NIPS, usually by comparing invaded with uninvaded communities. In many cases, competition is cited as the reason for loss of native species. This a plausible explanation because virtually all plants compete for a limited set of common resources (e.g., light, water, nutrients), but relatively few studies have investigated the implied competitive relationships among native and exotic species using formal removal or addition experiments, and the link between specific resources and competitive exclusion is often unclear. However, water and light competition associated with invasive species is a frequent driver of vegetation decline or change. Carpobrotus edulis, for example, is an invasive succulent plant that, by competing for water, affects the morphology and growth of cooccurring native shrubs in coastal California, while competition with annual exotic species for light during spring has been proposed as a key mechanism that has led to the decline of the native bunchgrass Nassella pulchra in Californian grasslands. Interestingly, the competitive advantage experienced by invasive species is highly dependent on resource availability; many invasive species are fast-growing species with high biomass production when resource supply is adequate but perform less well than native species in lowresource environments. Biomass production, in fact, is perhaps the most widespread process by which invasive species both directly and indirectly impact on native associates.

Invasive NIPS that are taller or denser than native associates or have novel traits can alter ecosystem attributes such as light regimes, soil moisture patterns, litter composition, and sediment deposition to such an extent that they become ecosystem engineers (also called transformers or keystone species). NIPS capable of fixing atmospheric nitrogen (nitrogen fixation) are one widely studied group that transform ecosystems by increasing soil nitrogen availability, which can significantly alter plant community structure and composition, especially in environments where soil nitrogen is low and there are few or no native nitrogen fixers. An excellent example is Myrica faya, an invader of volcanic soils in Hawai'i (Vitousek and Walker, 1989). The rate of nitrogen fixation of M. faya is two orders of magnitude higher than that of associated native species, leading to overall increases in nitrogen availability at the ecosystem level. Such processes can also facilitate invasion by other nonnitrogen-fixing species, such as the stimulation of Bromus diandris growth in soil obtained from patches of the invasive nitrogen-fixing shrub Lupinus arboreus. Other NIPS alter soil chemistry through salt concentration or produce allelopathic chemicals that negatively impact on native species. Morphological and physiological differences between NIPS and associated native species can also alter the spatial and temporal distribution of water through the soil profile, in turn affecting ecosystem composition and structure. Grasses such as Pennisetum setaceum, an invader of tropical dry forests in Hawaii, can reduce water availability in the upper rhizosphere, in turn reducing the health of coexisting native tree species (Cordell and Sandquist, 2008). In contrast, deeprooted woody or taprooted species can alter soil water balances by drawing water from deeper in the soil profile, changing subsoil drainage patterns and patterns of primary production. Differences in phenology can also be critical; for example, the displacement of the summer-active C4 species Themeda australis by the invasive spring-active C3 species Nassella neesiana is believed to have altered the timing of peak soil water availability in temperate grasslands and pastures of south-eastern Australia. Similarly, exotic species including the winter-active grass Bromus tectorum shift soil water extraction in semiarid shrub-steppe ecosystems in western North America to earlier in the growing season, reducing later availability for native species. Similar interference patterns occur in crops and pastures invaded by agricultural weeds. At broader scales, NIPS can significantly alter ecosystem processes by changing nutrient pools, the flow of energy though plant communities, or the structure and behavior of entire food webs and trophic levels. One frequently observed effect of increased biomass production by invasive species is an increase in the size of above- and below-ground carbon pools, and associated fluxes. Carbon inputs are generated by plant net primary productivity, which is often, although not always, greater in invasive species, and litter produced by invasive species typically decomposes faster due to lower C:N ratios and lower lignin:N ratios in plant tissue. As noted previously, soil N pools and fluxes can be radically altered by invasive species, especially N-fixers, although significant changes have also been observed in response to invasion by grasses. Soil and plant respiration are the major pathways of

Invasive Species: Plants

carbon loss in terrestrial systems, and these can also be altered by plant invasion, as can N loss via nitrification and denitrification. However, the magnitude and direction of these changes often vary significantly across different habitats and different invasive species, and more work is needed to determine whether any broad patterns can be discerned across invaded systems and species more generally. One dramatic way in which invasive species can significantly alter both the structure and processes of entire ecosystems is by inducing change in fire regimes. Because plant species have multiple strategies for maximizing fitness in fireprone habitats (e.g., resprouters vs. obligate seeders), changes in the timing, intensity, frequency, and extent of fires in response to elevated fuel loads have the potential to drastically change vegetation composition and structure. These changes to vegetation can in turn alter fire behavior in self-perpetuating positive feedback loops. An excellent example of such a process is the ‘grass-fire cycle’ in which high biomass production by invasive grasses results in hotter fires, which increases mortality in overstorey vegetation. This in turn allows further invasion, leading to the eventual replacement of savanna, woodland, or forest plant communities by grassland – a process that is occurring in northern Australia where savanna vegetation is being invaded by the highly productive Gamba grass (Andropogon gayanus). Interestingly, the high-severity fires generated by Gamba grass and associated loss of tree species significantly reduce above-ground carbon storage and may increase greenhouse emissions (Setterfield et al., 2010), demonstrating the impact that invasive species can have on ecosystem processes that are critically important at the global scale. The authors have discussed some of the direct and indirect mechanisms by which invasive plants can alter the composition and diversity of plant communities and the processes that structure entire ecosystems. NIPS, however, can also have profound genetic effects on native or agricultural species. One important mechanism involves the mixing of gene pools of formerly distinct taxa; in this context the authors are mainly concerned with hybridization and introgression between previously isolated native species and NIPS, known as genetic assimilation or genetic swamping (Antilla et al., 1998). This process can potentially lead to extinction, especially if the native species is rare, and is of increasing concern to conservation biologists due to the high levels of sexual compatibility observed among plant species in general. Perhaps the best documented example is of sunflowers in North America, where introgression between weedy Helianthus annuus and rarer endemic species (e.g., Helianthus bolanderi) threatens the endemic species with local or even range-wide extinction. In the case of H. annuus and H. bolanderi, more recent introgression has led to the development of populations that are morphologically more similar to H. annuus (Carney et al., 2000). This process can occur even if the invasive species is rarer relative to the native species, as is the case in Spartina alterniflora in California, which, due to high male fitness, threatens to assimilate the more widespread native Spartina foliosa. In this example, competition between S. alterniflora and S. foliosa pollen tubes is thought to favor S. alterniflora on the stigmas of both species, which is an interesting form of reproductive interference. Introgression can

71

range from primarily unidirectional, as with gene flow from introduced Siberian elm (Ulmus pumila) to red elm (Ulmus rubra) in the eastern US, to bidirectional. Interestingly, recent studies indicate that biparental introgression can lead to the replacement of parental lineages by hybrid genotypes (e.g., California radish; Raphanus sp.), provided hybrids and hybrid descendents do not suffer from outbreeding depression. It should also be mentioned that hybridization can pose a threat to small plant populations even in the absence of introgression. Genetic effects of NIPS on recipient plant communities and ecosystems are not restricted to those associated with gene flow and hybridization. For example, evolution can occur in other trophic groups, especially insects, in response to the establishment of novel plant populations, and the mixing of biota may lead to the generation of new kinds of mutualistic interactions that stimulate further coevolution (Mooney and Cleland, 2001). The dramatic decline in populations of native species in response to incursion by invasive species is likely to increase the probability of genetic drift, inbreeding, and other genetic consequences associated with population bottlenecks. Finally, given that many invasive plants differ in phenology and morphology from associated native species, shifts in the timing, severity, and resource specificity of competitive relations among native and exotic plants are likely to significantly alter the intensity and direction of selection, and hence evolution, in both groups.

Ecosystem Invasibility The concept of ‘invasibility,’ which refers to the vulnerability of an ecosystem to invasion, arose in response to the basic observation that plant communities and biomes differ significantly in their tendency to be invaded by NIPS (Lonsdale, 1999). Explaining why this is so has been one of the most heavily researched areas in invasion biology. One commonly observed pattern is that invasions tend to be promoted by disturbance, fragmentation, high resource availability, and slow recovery of native vegetation, such that late-successional plant communities tend to be invaded by few NIPS compared with those that have been highly modified. As noted previously, a large proportion of invasive species worldwide can be classified as r-strategists, characterized by rapid growth and biomass accumulation and high reproductive potential, and these species tend to be favored by disturbance and high nutrient availability. It is therefore not surprising that disturbed or modified systems tend to be heavily invaded, and that so many agricultural systems suffer invasion by weedy species – both cropping and pasturebased agricultural systems create homogenous, highly modified environments characterized by high resource availability, low competition, and recurring disturbance. Indeed, disturbance-driven intermittent resource enrichment or release leading to differences between resource supply and resource uptake has been proposed as a general process that increases the susceptibility of plant communities to invasion by alien species. The ‘biotic resistance hypothesis’ has long been advanced to explain why less disturbed systems tend to be less invaded

72

Invasive Species: Plants

by exotic species. Competition with native species is well known to represent a major barrier to invasion, especially when resources are limited, because nonindigenous competitive advantage appears to be strongly linked to overall resource availability. Functional similarity between native and nonindigenous species can increase competitive interference, and based on associated niche-packing models it has therefore been argued that more diverse or species-rich native assemblages should be less invasible than impoverished ones. Indeed, depauperate communities may even contain vacant niches that allow preadapted species to readily establish invasive populations – Ammophila arenaria in Californian coastal dunes ecosystems is a frequently cited example. This has proved, however, to have been a contentious idea in invasion ecology, because the empirical data inconsistently show a negative relationship between native and exotic species diversity. Indeed, a frequently observed pattern is that native and exotic plant diversity tends to be negatively correlated at small spatial scales, due to competitive exclusion, but positively correlated at large scales, where differences in the diversity and availability of resources affect both native and exotic species in similar ways. The diversity of species and functional groups occurring within different trophic levels may also impact on community invasibility in some ecosystems. Abiotic resistance is also seen as an important driver of invasibility. At the biome scale, for instance, deserts, savannas, and other harsh environments tend to be less invaded by exotic plants than more mesic habitats, the latter being more generally favorable for seedbank germination and seedling survival. However, these patterns also to some extent reflect the specific nature of the pool of nonindigenous species that are potential sources of invasive species – for example, some extreme environments face little ‘propagule pressure’ from species that are adapted to those conditions. This can arise from the history and composition of introductions in a given region, which reflect patterns of agricultural development, seed movement, ornamental plantings, and many other social and economic factors. An excellent example of this occurs in temperate south-eastern Australia, where numerous introductions have occurred through agricultural practices including the long-distance transport of seed and hay from other temperate areas. Given their provenance, these species are typically unable to survive in intact Australian alpine vegetation, which appears to be less invasible than lowland temperate ecosystems. However, pressure from species that are better adapted to such conditions is now also increasing, and it is likely that such systems will turn out to be more invasible than was initially suspected. Collectively, these lines of evidence indicate that variation in the diversity and abundance of invasive plants in different habitats, which may at first glance be causally linked solely to invasibility, in fact reflects differences in species traits, characteristics of the recipient community, and propagule pressure. Untangling the interactions between these drivers will continue to be a focus of invasion biology research with capacity to improve our management of agricultural and natural landscapes in the face of an increasingly large pool of potentially invasive species.

Contemporary Issues in Invasion Ecology Climate Change and Plant Invasions There is now overwhelming evidence that anthropogenic emission of CO2 is altering the climate of the earth. Over the past century global temperatures have risen by approximately 0.75 °C and it is estimated that 1.8–4.0 °C of warming (relative to 1980–99) is likely to occur by the end of the twenty-first century, depending on the emissions scenario. A major emerging concern is that future climate change linked to anthropogenic global warming (AGW) could increase the fitness of NIPS relative to native species, thus magnifying their impacts on natural and agricultural ecosystems. There are very good reasons to suggest that many extant invasive NIPS will have an enhanced capacity relative to native species to respond to changes in abiotic and biotic stressors (Dukes and Mooney, 1999). To become successful invaders, all exotic species must pass through a series of environmental filters (Figure 1); those that fail to do so ultimately lack the ability to establish populations in new environments and undergo landscape-level spread. Barriers to success include geographical limits to colonization, abiotic limits to establishment and survival of founder populations, biotic limits to population expansion, and landcape-level limits to population spread. Only exotic plants that possess certain combinations of traits associated with dispersal, growth, competition, and adaptation can pass through these filters, and those that do tend to become widespread and abundant. These filters select for successful invaders that contrast with native species in a number of key ways. First, most native species are rare or have restricted geographical ranges, whereas invasive species tend to have broad physiological tolerances to abiotic stress and occupy large ranges (Hellmann et al., 2008). Rare species usually have slower population growth rates, lower competitive ability, and a reduced capacity to cope with abiotic and biotic change. Exotic species also often exhibit high levels of genetic diversity in their native range, which increases the chance of matching specific, locally adapted genotypes to climatic conditions of the invaded range (preadaptation). Opportunities for preadaptation are enhanced in exotic species with a history of multiple introductions, which has occurred in many of the worst ornamental and pasture weeds. A prime example is St. John's wort (Hypericum perforatum), which has been introduced into North American from Europe on multiple occasions, and which displays climate matching of source genotypes in the invasive range. Populations of invasive species also frequently undergo rapid evolution, including for climate-related traits that increase fitness in new environments. Adaptive differentiation of populations along broad climatic gradients has been observed in many species, including purple loosestrife (Lythrum salicaria), invasive North American populations of which have developed strong latitudinally based clinal variation for phenological traits (flowering time and size at flowering time) that are similar to those that are found in native European populations. Other examples include ragweed (Ambrosia artemisiifolia), African ragwort (Senecio inaequidens), and sunflower (Helianthus maximiliani; Kawakami et al., 2011). Adaptive evolution can also arise from the admixture of previously

Invasive Species: Plants

separated and distinct genetic lineages and subsequent generation of novel genotypes, providing invasive species with added capacity to invade new habitats. ‘Phenotypic plasticity’ has also been suggested as a key driver of plant invasiveness, because having the ability to maintain positive population growth rates in a wide range of environments is likely to confer an advantage over more specialized native or crop species (Hulme, 2008). Enhanced niche breadth linked to phenotypic plasticity can be striking: Brazilian pepper (Schinus terebinthifolius), for example, a highly invasive species in the southern US, can grow as a tree, vine, or scrambling shrub, depending on growing conditions. Plasticity in reproductive allocation, dry matter partitioning, and growth have been observed in numerous crop and pasture weeds (e.g., Echium plantagineum; Ammannia spp.), and can alter competitive relationships among plants when grown under conditions of limited resource availability. By allowing plants to more efficiently exploit a wide range of environments and by compensating for the low genetic diversity that typifies many founder weed populations, phenotypic plasticity is also likely to place at an advantage plant populations subjected to increasing climatic variability or enhanced abiotic stress (Nicotra et al., 2010). Seeds produced by invasive NIPS often have specific morphological adaptations for wind, water, or animal dispersal that increase the rate and distance of propagule movement into new habitats (Murray and Phillips, 2010). For example, low seed mass often facilitates wind dispersal, while many ornamental species produce fruit that are spread long distances by birds and animals (zoochory). Some species have multiple modes of seed production: An excellent example is the invasive South American tussock grass Nassella neesiana, which produces seeds in panicles, stems, and at the base of the tussock. This gives N. neesiana the capacity to spread rapidly via multiple vectors (e.g., grazing animals, roadside mowing), and as a result it has become one of Australia's worst invaders of native grassland and grazed pastures. Rapid generation time, typical of many r-selected species, is another life-history adaptation that facilitates the growth and spatial expansion of plant populations. Shifting climatic envelopes are expected to benefit species that can rapidly extend populations into new climatically suitable ranges, and it is likely that many invasive species, especially those that are already efficient dispersers, will benefit. One of the most often-cited differences between native and invasive NIPS is in competitive ability and overall growth and biomass production (Vilà and Weiner, 2004). Typically, invasive species have rapid growth rates, and numerous studies suggest that most are highly competitive in both natural and agricultural settings. However, in most circumstances the magnitude of any growth or competitive advantage enjoyed by invasive species depends on resource availability: in resourcepoor environments native species often have superior or equal fitness (Daehler, 2003). In typical high-fertility cropping situations such conditions are unlikely to arise, but nutrient limitation often negatively impacts on the performance of exotic species in seminatural (e.g., unimproved pasture) and natural vegetation. In these environments, disturbance favors invasive species because it frees up resources that would otherwise be exploited by existing native vegetation. It is

73

therefore clear that interactions between climate change and other drivers of global change (Marini et al., 2012), including anthropogenic disturbance and habitat fragmentation, will determine the relative fitness of invasive species in the coming century. Rising atmospheric carbon dioxide concentrations [CO2] associated with the burning of fossil fuels is another driver of global change that will impact on all plant populations. Because trade-offs exist between acquisition of atmospheric CO2 and water loss, altering [CO2] changes physiological processes taking place in plant tissue, in turn altering plant growth and water use efficiency. Many invasive and native plants increase biomass production and fecundity under elevated [CO2], a response known as the CO2 fertilization effect (Godfree et al., 2013). This process has been implicated in the success of a number of invasive species, including cheatgrass (Bromus tectorum) in semiarid rangelands of the US. However, although the high phenotypic plasticity of invasive species may increase their ability to shift their physiological optimum to take advantage of CO2 fertilization, at present it appears questionable whether invasive species should in general disproportionately benefit more from rising [CO2]. Indeed, a large proportion of the world's worst weeds are C4 species, and many compete with C3 crops, which in general tend to be more responsive to CO2 fertilization. Overall, developing a level of knowledge sufficient to answer the question of nonindigenous advantage under AGW with any certainty is likely to be a challenging task, especially given the somewhat chequered history of predicting the success of invasive species in general. Nonetheless, understanding the links between the traits and attributes of NIPS and key climatic drivers will underpin our ability to manage or avoid the detrimental impacts of NIPS on a global scale.

Transgenic Plants as Invasive Species The impact of genetic modification (GM) or engineering on the biosafety of recipient organisms has been one of the most contentious issues in ecology of the past two decades (Hails, 2000). Genetic modification involves the transfer of genetic material (a transgene), usually coding for a beneficial trait, from one organism into a target organism (the genetically modified organism or GMO), which then expresses the trait. Since 1996 GM technology has been rapidly adopted across the globe, and in 2011 over 160 million hectares of GM crops were grown across 29 countries, a figure that is currently increasing at a rate of approximately 8% per year. A small number of widely grown species (cotton, corn, soybean, and canola) expressing herbicide tolerance (HT) and virus and insect resistance (VR and IR) still dominate global plantings of transgenic crops. In the late 1970s to mid-1980s arguments were widely put forward that no GMOs should pose any risks to the environment. These were based on a range of premises, such as that genetic modification does not produce any genetic novelty that could not otherwise be generated by traditional selective breeding, or that all GMOs will suffer metabolic costs associated with transgene expression leading to reduced environmental fitness. This view, however, was challenged shortly

74

Invasive Species: Plants

after when it became clear that genetic engineering had the potential to create truly novel combinations of potentially adaptive traits, especially where phylogenetic leapfrogging was involved. Pleiotropic or other unintended effects associated with the transformation process could also impact strongly on plant phenotype in unpredictable ways. Since then, the development and release of GMOs has been placed under strong regulatory oversight in most countries, and methods for quantifying risk associated with GMO deployment have been the focus of sustained scientific endeavor (Hilbeck et al., 2011). One of the most widely discussed potential risks associated with GMO release involves transgenic organisms or wild relatives becoming weedy either in crop or nontarget environments. This can occur via two main pathways (Figure 2): either the GMO itself can spread into nontarget areas outside of the cropping area and become invasive, or gene flow can occur into wild relatives growing in nontarget habitats, which then become invasive. For the majority of transgenic crops, the highest-risk situations occur when the transgene confers a clear fitness advantage under natural selection regimes and where hybridization with wild relatives is possible. Many transgenes are unlikely to increase the fitness of either crop plants or wild relatives in noncrop situations, for example, herbicide (glufosinate or glyphosate) tolerance in natural ecosystems. However, other traits such as pathogen or insect resistance could increase the fitness of plants containing the transgene significantly. Probably the most widely studied transgenic plants that fall into this category contain cry genes Bacillus thuringiensis (Bt) that express endotoxin proteins that provide resistance to insect pests. Many crops such as maize and cotton that express cry genes are unlikely to become themselves invasive because they are highly domesticated and lack critical demographic traits that are essential for establishment of feral populations in nontarget environments, such as seed dormancy or the ability to compete with native vegetation for light, water, and nutrients. Of much greater concern in these crops is the potential for gene flow to wild relatives, a process involving pollen flow, the formation of viable wild-crop hybrids, and (usually), if the transgene present confers a selection advantage, gene introgression into wild populations (Figure 2). Given that virtually all the world's most important crops hybridize with wild relatives somewhere in the world (Ellstrand et al., 1999), it seems virtually inevitable that gene flow from transgenic crops to wild relatives will occur, although this does not mean that a significant risk will be posed to the environment in all cases. Examples of GM crops in which the potential for hybridization with sexually compatible wild species or wild genotypes have been investigated include sorghum (with Sorghum halapense), maize (with teosinte), sunflowers, canola (with Brassica rapa and other species), sugar beet, and cotton (with other Gossypium species). Transgenic pasture plants probably pose a greater invasive risk to the environment than most crop species because they have a wide range of traits that allow them to persist in mixed plant communities in which competition for limiting resources and abiotic stresses is more intense. Pasture species have a history of invasiveness, and many have shown the capacity to rapidly undergo range expansion and local

adaptation following introduction. An increase in fitness arising from the presence of disease or pathogen- or stressrelated transgenes in such species could therefore have serious consequences for nontarget plant communities. One of the best studied cases is that of virus-resistant white clover (Trifolium repens), which in south-eastern Australia, in addition to being an important pasture plant, is also an invader of endangered native grassland and woodland communities. Studies have shown that a range of viruses including Clover yellow vein virus limit the growth rate of wild T. repens populations, and thus the introgression of transgenes for virus resistance from sown pastures could increase the invasiveness of wild populations in these habitats (Godfree et al., 2007). Another example is creeping bentgrass (Agrostis stolonifera), a highly competitive turf grass species. Flow of transgenes from HR genotypes to wild A. stolonifera populations and another related species, Agrostis gigantea, has been observed under field trial conditions, potentially making their control in other agronomic settings more difficult. Ecological and evolutionary processes associated with the use of GM technology can also lead to weed problems within agricultural settings. GM plant lines can themselves volunteer in subsequent crops, especially if they produce a persistent seedbank. This is of particular concern in farming systems that rely on sequential rotations of crops that all express the same type of HT (e.g., glyphosate tolerance). Gene stacking can also lead to the development of plant lines resistant to multiple herbicides, as has been demonstrated in canola in Canada. Evidence suggests that the repeated use of a single herbicide group for control of weeds in GM crops can also lead to the development of herbicide-resistant crop weed biotypes, for example, Roundup resistance in Conyza canadensis. Both situations have significant implications for farm management and economics. Finally, the use of GM technology allows for greater control of weeds, but this may adversely affect other faunal associates that depend on plant biodiversity as a resource. Over the past two decades, increased regulation of GMOs around the world has necessitated the development of new strategies for quantifying the potential direct and indirect risks posed to human health and the environment. Risk is defined by the equation risk ¼ f(hazard, exposure) (Wilkinson et al., 2003), where the hazard is the severity of the unwanted change and exposure is the probability of the hazard occurring. In the scenarios described above, hazard refers to the severity of the impact of increased invasiveness of GM crops or associated wild relatives on natural or agricultural systems. Hazards are normally identified using structured hazard identification studies in which potential scenarios leading to hazard realization are identified; several such scenarios are shown in Figure 2. In most cases, the efficacy of the genetic transformation and the potential impact of GM plants on target and nontarget organisms are initially evaluated by direct comparison of transgenic lines with genetically similar or identical isolines so that the impact of the introduced trait on plant phenotype and fitness can be directly quantified. Detailed ‘exposure studies’ are then conducted to determine the probability of each step in a given hazard scenario occurring following release of a transgenic plant into the environment. Risk assessments are often conducted in a carefully structured tiered procedure in which hazards are first indentified under

Invasive Species: Plants

Resistance development in nontarget species

Repeated use of herbicides to control weeds in GM crops

Invasiveness or weediness of GM crops

Survival of GMO within cultivated areas

Spread and survival of GMO outside cultivated areas

75

Gene flow into wild relatives

Pollen flow from GMO to wild relative

Hybrid formation

Selection for herbicide resistant weed species

Establishment of volunteer populations in subsequent crops

Reproduction Hybrid survival and reproduction, gene introgression

Crop/relative population containing transgene with enhanced fitness

Altered pest control strategies

Spread and invasion of nontarget systems

Negative impact on farm agricultural economics

Negative impacts on nontarget ecosystems

Figure 2 Framework for identifying some possible invasion hazards (bold) to nontarget and agricultural systems associated with the release of transgenic plants into the environment. Potential direct and indirect pathways leading to realization of hazard are shown; the risk assessment processes seek to quantify the probability of each stage in the hazard pathway occurring. Adapted from Sanvido, O., Romeis, J., Bigler, F., 2007. Ecological impacts of genetically modified crops: Ten years of field research and commercial cultivation. Advances in Biochemical EngineeringBiotechnology 107, 235–278.

‘worst-case scenario’ conditions using first-tier laboratory experiments and then more detailed second- and third-tier semifield and field experiments aimed at refining probabilities associated with specific steps within a given hazard scenario. The combined experience from the past two decades suggests that the relatively small number of mainly herbicidetolerant and insect- or pathogen-resistant GM crop species currently grown on a wide scale pose only a limited direct threat as invasive species. Although gene flow to weedy relatives will occur in some species, there are few examples to date of where this has led to realization of hazard in nontarget systems. However, a wide variety of ‘new generation’ GM crops engineered for oil and pharmaceutical production, drought tolerance, fungal resistance, altered phenology, bioremediation, and many other traits are also now in development or commercial adoption worldwide. Interactions between these plants and nontarget species and ecosystems are certain to be complex, and risk assessments that explicity test for increased

invasiveness under field conditions will remain central to their successful, hazard-free release into the environment.

Using the Past to Inform the Future Directions for Study and Management of Invasive Species in a Changing World As this article has shown, invasion ecology is a dynamic, rapidly evolving field of research that continues to provide novel insights into the ways in which plants establish and maintain viable populations in new habitats and how they impact recipient agricultural and natural systems. Yet, many questions remain, and new approaches are being developed to better understand the role of plant invasions of agents of global change. For example, very recently, a small number of studies in invasion ecology have examined historical patterns

76

Invasive Species: Plants

of introduction of invasive NIPS to particular regions of the world. Such comparative analyses of NIPS introductions can pinpoint periods in history when different regions of the world have received particularly high numbers of species. Such knowledge can provide an important historic baseline for determining the factors driving the successful spread of NIPS. An understanding of temporal introduction patterns can inform management aimed at limiting the ecological, economical, and social impacts of invasive NIPS on native biodiversity. Such a historical approach is somewhat analogous to using historical information on climate variation to predict future climate scenarios. Australian work in this context has revealed distinct peaks since permanent European settlement of the continent began in the late 1700s. Peaks during early European settlement (1810–20) and human range expansion across the continent (1840–60) both coincided with considerable growth in Australia's human population. It is likely that human population growth during these times increased the likelihood of NIPS becoming invasive as a result of increased colonization and propagule pressure. Although this sort of research is in its infancy, there are many promising avenues to pursue, and valuable information to be gleaned from better understanding previous temporal patterns of NIPS introductions. Hopefully, such approaches will further increase our ability to predict and manage the impacts of NIPS globally in years to come.

See also: Climate Change and Plant Disease. Invasive Aquatic Animals. Plant Biotic Stress: Weeds. Regulatory Challenges to Commercializing the Products of Ag Biotech

References Antilla, C.K., Daehler, C.C., Rank, N.E., Strong, D.R., 1998. Greater male fitness of a rare invader (Spartina alterniflora, Poaceae) threatens a common native (Spartina foliosa) with hybridization. American Journal of Botany 85, 1597–1601. Barrett, S.C.H., 1983. Crop mimicry in weeds. Economic Botany 37, 255–282. Blossey, B., Nötzold, R., 1995. Evolution of increased competitive ability in invasive nonindigenous plants: A hypothesis. Journal of Ecology 83, 887–889. Bucharova, A., van Kleunen, M., 2009. Introduction history and species characteristics partly explain naturalization success of North American woody species in Europe. Journal of Ecology 97, 230–238. Carney, S.E., Gardner, K.A., Rieseberg, L.H., 2000. Evolutionary changes over the fifty-year history of a hybrid population of sunflowers (Helianthus). Evolution 54, 462–474. Colautti, R.I., Eckert, C.G., Barrett, S.C., 2010. Evolutionary constraints on adaptive evolution during range expansion in an invasive plant. Proceedings of the Royal Society of London B 277, 1799–1806. Cordell, S., Sandquist, D.R., 2008. The impact of invasive African bunchgrass (Pennisetum setaceum) on water availability and productivity of canopy trees within a tropical dry forest in Hawaii. Functional Ecology 22, 1008–1017. Daehler, C.C., 2003. Performance comparisons of co-occurring native and alien invasive plants: Implications for conservation and restoration. Annual Review of Ecology, Evolution and Systematics 34, 183–211. Drenovsky, R.E., Grewell, B.J., D'Antonio, C.M., et al., 2012. A functional trait perspective on plant invasion. Annals of Botany 110, 141–153. Dukes, J.S., Mooney, H.A., 1999. Does global change increase the success of biological invaders? Trends in Ecology and Evolution 14, 135–139. Ehrenfeld, J.G., 2010. Ecosystem consequences of biological invasions. Annual Review of Ecology, Evolution and Systematics 41, 59–80. Ellstrand, N.C., Prentice, H.C., Hancock, J.F., 1999. Gene flow and introgression from domesticated plants into their wild relatives. Annual Review of Ecology and Systematics 30, 539–563.

Elton, C.S., 1958. The Ecology of Invasions by Animals and Plants. London: Methuen Frankham, R., 2005. Resolving the genetic paradox in invasive species. Heredity 94, 385. Godfree, R.C., Robertson, B.C., Gapare, W.J., et al., 2013. Nonindigenous plant advantage in native and exotic Australian grasses under experimental drought, warming and atmospheric CO2 enrichment. Biology 2, 481–513. Godfree, R.C., Thrall, P.H., Young, A.G., 2007. Enemy release after introduction of disease-resistant genotypes into plant-pathogen systems. Proceedings of the National Academy of Sciences of the USA 104, 2756–2760. Groves, R.H., Boden, R., Lonsdale, W.M., 2005. Jumping the garden fence: Invasive garden plants in Australia and their environmental and agricultural impacts. A CSIRO Report for WWF-Australia. Ultimo: CSIRO. Hails, R.S., 2000. Genetically modified plants − The debate continues. Trends in Ecology and Evolution 15, 14–18. Hellmann, J.J., Byers, J.E., Beirwagen, B.G., Dukes, J.S., 2008. Five potential consequences of climate change for invasive species. Conservation Biology 22, 534–543. Hilbeck, A., Meier, M., Römbke, J., et al., 2011. Environmental risk assessment of genetically modified plants − Concepts and controversies. Environmental Sciences Europe 23, 13. Hulme, P.E., 2008. Phenotypic plasticity and plant invasions: Is it all Jack? Functional Ecology 22, 3–7. Kawakami, T., Morgan, T.J., Nippert, J.B., et al., 2011. Natural selection drives clinal life history patterns in the perennial sunflower species, Helianthus maximiliani. Molecular Ecology 20, 2318–2328. Keane, R.M., Crawley, M.J., 2002. Exotic plant invasions and the enemy release hypothesis. Trends in Ecology and Evolution 17, 164–170. Konarzewski, T.K., Murray, B.R., Godfree, R.C., 2012. Rapid development of adaptive, climate-driven clinal variation in seed mass in the invasive annual forb Echium plantagineum L. PLoS One 7, e49000. Lee, C.E., 2002. Evolutionary genetics of invasive species. Trends in Ecology and Evolution 17, 386–391. Levine, J.M., Vilà, M., D'Antonio, C.M.D., et al., 2003. Mechanisms underlying the impacts of exotic plant invasions. Proceedings of the Royal Society of London B 270, 775–781. Lonsdale, W.M., 1999. Global patterns of plant invasions and the concept of invasibility. Ecology 80, 1522–1536. Marini, L., Battisti, A., Bona, E., et al., 2012. Alien and native plant life-forms respond differently to human and climate pressures. Global Ecology and Biogeography 21, 534–544. Monty, A., Mahy, G., 2009. Clinal differentiation during invasion: Senecio inaequidens (Asteraceae) along altitudinal gradients in Europe. Oecologia 159, 305–315. Mooney, H.A., Cleland, E.E., 2001. The evolutionary impact of invasive species. Proceedings of the National Academy of Sciences of the USA 98, 5446–5451. Murray, B.R., Phillips, M., 2010. Investment in seed dispersal structures is linked to invasiveness in exotic plant species of south-eastern Australia. Biological Invasions 12, 2265–2275. Nicotra, A.B., Atkin, O.K., Bonser, S.P., et al., 2010. Plant phenotypic plasticity in a changing climate. Trends in Plant Science 15, 684–692. Phillips, M., Murray, B.R., 2012. Invasiveness in exotic plant species is linked to high seed survival in the soil. Evolutionary Ecology Research 14, 83–94. Pimental, 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, 273–288. Raybould, A.F., Gray, A.J., Lawrence, M.J., Marshall, D.F., 2008. The evolution of Spartina anglica C.E. Hubbard (Gramineae): Origin and genetic variability. Biological Journal of the Linnean Society 43, 111–126. Richardson, D.M., Pyšek, P., 2008. Fifty years of invasion ecology − The legacy of Charles Elton. Diversity and Distributions 14, 161–168. Richardson, D.M., Pyšek, P., Rejmánek, M., et al., 2000. Naturalisation and invasion of alien plants: Concepts and definitions. Diversity and Distributions 6, 93–107. Satai, A.K., Allendorf, F.W., Holt, J.S., et al., 2001. The population biology of invasive species. Annual Review of Ecology and Systematics 32, 305–332. Setterfield, S.A., Rossiter-Rachor, N.A., Hutley, L.B., Douglas, M.M., Williams, R.J., 2010. Turning up the heat: The impacts of Andropogon gayanus (gamba grass) invasion on fire behaviour in northern Australian savannas. Diversity and Distributions 16, 854–861. Sinden, J., Jones, R., Hester, S., et al., 2004. The economic impact of weeds in Australia. CRC for Australian Weed Management Technical Series No. 8. SA, Australia: CRC for Australian Weed Management, Waite Campus, University of Adelaide, PMB 1, Glen Osmond.

Invasive Species: Plants

Theoharides, K.A., Dukes, J.S., 2007. Plant invasion across space and time: Factors affecting nonindigenous species success during four stages of invasion. New Phytologist 176, 256–273. Vilà, M., Weiner, J., 2004. Are invasive plant species better competitors that native plants species? − Evidence from pair-wise experiments. OIKOS 105, 229–238. Vitousek, P.M., Walker, L.R., 1989. Biological invasion by Myrica faya in Hawai'i: Plant demography, nitrogen fixation, ecosystem effects. Ecological Monographs 59, 247–265. Whitney, K.D., Gabler, C.A., 2008. Rapid evolution in introduced species, ‘invasive traits’ and recipient communities: Challenges for predicting invasive potential. Diversity and Distributions 14, 569–580. Williamson, M.H., Fitter, A., 1996. The varying success of invaders. Ecology 77, 1661–1666. Wilkinson, M.J., Sweet, J., Poppy, G.M., 2003. Risk assessment of GM plants: Avoiding gridlock? Trends in Plant Science 8, 208–212.

Relevant Websites http://academic.sun.ac.za/cib/index.asp Centre of Excellence for Invasion Biology. http://www.ibot.cas.cz/invasions/index.htm Institute of Botany, Academy of Sciences of the Czech Republic. http://www.cassey-invasion-ecology.org/index.php The University of Adelaide. http://sydney.edu.au/science/biology/shine/ The University of Sydney. http://www.uts.edu.au/staff/brad.murray University of Technology Sydney.

77