The following is the introduction chapter (Chapter 1) from my Ph.D thesis Invader-invader mutualism facilitates secondary invasions in rainforest on Christmas Island.
Biological invasions are a major and increasing element of global change (Mack et al. 2000, Wonham and Pachepsky 2006, Pyšek and Richardson 2010). The impact of species invasions poses a significant threat to biodiversity through species loss and altered ecosystem function (Mack et al. 2000, Crooks 2002, Ricciardi 2007) as well as having significant economic consequences (Luque et al. 2014). These costs are associated with, i) the loss of primary production (Wilby and Thomas 2002), ii) loss of biodiversity and other impacts on natural ecosystems (Vilà et al. 2011, Simberloff et al. 2013), iii) the effort and resources required to reduce the entry and spread of new invasive species and the eradication of those already established (Eiswerth and Johnson 2002, Pyšek and Richardson 2010). So extensive is the problem that the research discipline of invasion biology (Elton 1958, Simberloff and Vitule 2013) has developed over the past few decades to examine important questions related to the spread and impacts of invasive species. Along with climate change and habitat loss, biological invasions, and their impacts, are considered a main driver of global environmental change (Didham et al. 2005, Tylianakis et al. 2008). Understanding the determinants of invasion success continues to be intensely studied (Catford et al. 2009, Lockwood et al. 2009, Blackburn et al. 2011) in order to better predict which species will become invasive and what the impacts of these successful species are likely to be.
Invasive species, by definition, are species whose presence in an area is due to intentional or accidental introduction as a result of human activity. However, the term ‘invasive’ is often reserved only for those introduced species that reproduce in large numbers, have great potential for spread, and may cause significant impact (Richardson et al. 2000). This is because ‘invasion’ evokes anthropogenic concepts like aggression, assault, intrusion, and encroachment (Richardson et al. 2000). Introduced species are also referred to interchangeably as ‘alien’, ‘exotic’, ‘naturalized’, ‘non-native’, ‘nonindigenous’, or simply ‘weeds’ and ‘pests’. The term ‘naturalized species’ has the longest history, being widely used in the middle 19th century to describe the widespread phenomenon of introduced taxa behaving like natives (e.g. Darwin, 1859). Around this time, species introductions were widely celebrated as ‘enriching’ the flora and fauna in many regions worldwide. As a result of both intentional and unintentional introductions, introduced species now make up a substantial part of the biodiversity in many places (Vitousek et al. 1997).
Richardson et al. (2000) have related the multiple terms used to describe introduced species to various stages of the invasion pathway (Fig. 1). This scheme reserves ‘naturalized’ for species that have entered and established in a community, ‘invasive’ for species spreading and potentially causing impact, and allows the use of other terms to refer to species throughout the entire pathway. The pathway itself describes the series of steps, or filters, that need to be overcome for a species to enter, establish and spread at a new location – referred to as ‘invasion success’. Invasion success is determined by the interaction of three key attributes; propagule pressure, traits of the invader, and properties of the recipient community (Colautti et al. 2006, Catford et al. 2009, Lockwood et al. 2009, Blackburn et al. 2011). While species traits, such as high reproductive capacity or disturbance tolerance, are clearly important, their effectiveness depends strongly on the nature of the recipient community and propagule pressure (Simberloff 2009). The context-specific complexity of species introductions means that there exist many examples of invasion success being strongly driven by abiotic or biotic aspects of the recipient community (Maron and Vila 2001, Von Holle and Simberloff 2004), or propagule pressure (Johnston et al. 2009, Lockwood et al. 2009, Driscoll et al. 2014) or an interaction between the two (Britton-Simmons and Abbott 2008, Catford et al. 2011, Green et al. 2011).
Since the establishment of modern invasion biology (Elton 1958), numerous hypotheses have been proposed to explain the determinants of invasion success at a variety of temporal and spatial scales (Richardson and Pyšek 2006). Most hypotheses attribute invasion success to characteristics of the invader or characteristics of the invaded ecosystems, with few integrating the two (Davis et al. 2000, Colautti et al. 2004, Blumenthal 2006). Many hypotheses overlap or share similarity with pre-existing hypotheses, a phenomenon not uncommon in ecology (McGill et al. 2007). Until recently, hypotheses concluding invasion success was the outcome of negative interactions dominated the literature. Biotic resistance and diversity-invasibility hypotheses describe invasion success as being inhibited by diverse, native communities full of competitors, herbivores, predators and pathogens (Elton 1958, Lake and O’Dowd 1991, Levine and D’Antonio 1999, Levine et al. 2004). Similarly, passive or neutral mechanisms also featured prominently, for example, enemy-release and empty-niche hypotheses posit invasion success as a result of unavoidable traits in the recipient community (Berdegue et al. 1996, Colautti et al. 2004, Holzmueller and Jose 2011). Over time, the idea that invasion success can be facilitated by positive interactions has attracted greater attention. This has led to the development of hypotheses describing direct or indirect facilitative relationships between invaders and natives, or among other invaders (Jones et al. 1997, Simberloff and Von Holle 1999, Floerl et al. 2004, Rodriguez 2006). Although invasion success is likely to be context-dependent and due to a combination of factors and mechanisms, a more general framework explaining the determinants of invasion success will advance understanding through reducing theoretical redundancy (e.g. Catford et al., 2009).
The invasional meltdown hypothesis
The invasional meltdown hypothesis describes the process by which a group of exotic species facilitate each other, increasing the likelihood of survival and/or ecological impact (Simberloff and Von Holle 1999). It is distinguished from classic models of invasion success that focus on negative interactions (i.e. biotic resistance) as it predicts an accelerating accumulation of introduced species and effects within a community, rather than a deceleration (Simberloff and Von Holle 1999). Although the study of interactions between exotic species had emphasised negative associations, Simberloff and Von Holle (1999) concluded that introduced species frequently interact with one another and that facilitative interactions are at least as common as detrimental ones. The logic behind invasional meltdown was introduced via original examples from Elton (1958) who described a mutualistic interaction between a neotropical ant and an Asian scale insect in California, suggesting facilitation among exotic species need not be generated by shared evolutionary history. Highlighting the many ways that exotic species formed novel facilitative associations through either direct or indirect mechanisms provided researchers with an appropriate theoretical framework for testing ideas about invasion success within heavily invaded ecosystems (Ricciardi 2001, Adams et al. 2003, O’Dowd et al. 2003, Bourgeois et al. 2005, Grosholz 2005, Johnson et al. 2009).
The strongest evidence of invader-invader facilitation occurring through direct mechanisms came from animals pollinating and dispersing plants (Simberloff and Von Holle 1999). For example, the introduced honey bee more readily pollinates exotic plants in many locations (Butz Huryn 1997, Barthell et al. 2001). Similarly, a frugivorous avian invader significantly enhances the spread of alien plant species through effective seed dispersal (Macdonald et al. 1991). Invaders also facilitate each other indirectly by modifying the recipient community (Simberloff and Von Holle 1999). For example, large introduced herbivores can facilitate the invasion of grassland weeds by increasing disturbance (Hobbs and Huenneke 1996) and introduced N-fixing shrubs can facilitate the invasion of other weeds by altering soil nutrient properties (Vitousek and Walker 1989). These examples, and others, help form the invasional meltdown hypothesis (Simberloff and Von Holle 1999) that has since gone through rigorous testing as other researchers have looked for evidence either for or against it (Ricciardi 2001, O’Dowd et al. 2003, Bourgeois et al. 2005, Grosholz 2005, Green et al. 2011, DeVanna et al. 2011).
Few studies have been able to demonstrate a complete invasional meltdown. The majority of suggested instances of invasional meltdown have been described as plausible scenarios of long-term consequences based on short-term observations of facilitative interactions between individuals of two species (Simberloff 2006). The invasional meltdown hypothesis predicts facilitative interactions between two invaders will lead to i) positive population-level effects for both species, ii) enhancement of the impact on the native community, and iii) increased establishment of other introduced species. Although many examples of introduced species facilitating each other has provided strong evidence of population-level impacts (Ricciardi 2001, O’Dowd et al. 2003, Bourgeois et al. 2005), few have conclusively demonstrated interspecific facilitation leading to accelerated impact and establishment of other exotic species (Simberloff 2006). The lack of empirical support for that aspect of the hypothesis has lead some to question the validity of the concept and suggest alternatives (DeVanna et al. 2011). Despite some criticism (Gurevitch 2006), invasional meltdown continues to be better supported than other hypotheses that do not strongly consider interaction of invaders with their new environment (Jeschke et al. 2012).
The strongest evidence in support of the invasional meltdown hypothesis comes from an invasive ant-scale insect mutualism on Christmas Island, Indian Ocean (O’Dowd et al. 2003); a conclusion supported by Simberloff (2006) when considering the initial impact of the hypothesis. Mutualism between an exotic ant and introduced scale insects has been demonstrated to dramatically increase the populations of both (Abbott 2006, Abbott and Green 2007), cause significant impact on the native community (O’Dowd et al. 2003, Davis et al. 2010), and be responsible for the invasion success of a previously unsuccessful exotic species (Green et al. 2011). As a result, Christmas Island continues to be an excellent study system for demonstrating invasional meltdown and the direct and indirect impacts of facilitative interactions among invaders.
The oceanic island
Christmas Island (Fig. 2) is an isolated oceanic island in the north-eastern Indian Ocean (105°40’E, 10°30’S). Lying 360 km south of Java, the island’s climate is monsoonal with distinct wet (November – May) and dry (June – October) seasons and a mean annual precipitation of ~2000 mm (Falkland 1986). Daily temperatures range from 20 to 35 °C. The limestone island is 135 km2 and rises in a series of cliffs and terraces that largely define the rainforest vegetation patterns (Du Puy 1993). ‘Plateau’ rainforest is widespread above ~200 m elevation, and is dominated by a dozen or so canopy species, the frequencies of which vary locally (Du Puy 1993). Green (1997) recorded 13 tree species on a 0.3 ha plot on the island, several times fewer species than on similar-sized plots in diverse continental tropical forests. However, tree density (404 stems/ha) and basal area (48 m2/ha) are similar to mainland rainforests (Green 1997). The canopy of plateau rainforest reaches 40 m and is evergreen, although it thins noticeably during the dry season (Green et al. 1997). The majority of the remaining 70% of original forest that was not cleared for phosphate mining is designated National Park. The rainforest community of Christmas Island is characterised by high species endemism (Du Puy 1993, Adler 1994, Green 1997, Allen 2008).
Almost all natural processes occurring in rainforest on Christmas Island are due to the actions of the dominant omnivore-detritivore, the red land crab Gecarcoidea natalis (Brachyura: Gecarcinidae) (Green 1997, Green et al. 1997, 1999, 2008). Endemic to the island, these large (to > 120 mm carapace width and 500 g live weight), burrow-dwelling crabs are naturally highly-abundant wherever native rainforest is still intact, occurring at mean densities of 0.75 – 1.2 crabs m-2 and mean biomass of 1137 kg ha-1 (Green 1997). It is not uncommon for land crab species to dominate some islands owing to possibly few natural enemies and competitors, high productivity potential, and their ability to survive on a low-quality diet (Lindquist et al. 2009). Red crab activity varies seasonally, with high daily activity occurring during the wet season and early dry season, and relatively low activity during the peak of the dry season (Green 1997). Red crabs are most active during the day when relative humidity is high and ambient temperature is low (dawn and dusk); however, they can maintain constant activity during periods of almost daily rainfall (Green 1997). Bursts of surface activity will coincide with brief showers and sharp increases in relative humidity, but during the driest time of year, red crab surface activity may cease (Green 1997).
At the beginning of each wet season, around half of all red crabs migrate to the coast of the island to breed. These annual breeding migrations are a true spectacle of nature as millions of crabs converge at a number of points around the island. The annual event has been a focus of many nature documentaries, including the BBC’s Trials of Life, presented by Sir David Attenborough. The actual timing of this movement en masse to the coast is determined by the phase of the moon and adequate rainfall (Hicks 1985, Green 1997). A typical migration event can last up to three months, as follows (Hicks 1985). Male crabs generally arrive on the coast first, in order to excavate a large burrow for breeding, followed closely by the females. Fights between males occur as they compete for burrow space and access to females, and many lose limbs as a consequence. Following copulation, males leave and the females remain in the burrows brooding their eggs, a process that usually takes 12 – 13 days. Females release their eggs at night on the turn of the high tide between the last quarter of the moon and the new moon before returning en masse to the plateau. Zoeae hatch immediately on contact with salt water, and baby crabs emerge on the island at the first crab stage around 27 days after egg release. Although quantitative data on this return is lacking, anecdotal accounts suggest baby crabs return to the island in large numbers only every three or so years (M. Orchard pers. comm.). Once adults have returned to the forest, the majority of breeders seal themselves in burrows to moult, leading to another period of relatively low crab activity (Green 1997)
Due to their behaviour and naturally high-abundance, red crabs are recognised ecosystem engineers on Christmas Island, with essentially all key ecological processes in intact rainforest driven by their presence and influence (O’Dowd and Lake 1989, 1991, Green et al. 1997, 1999, 2008). Leaf litter cover and biomass is seasonally dynamic, from almost complete cover and high biomass at the end of the dry season to almost total absence of litter at the end of each wet season (Green et al. 1999). This pattern mirrors the activity pattern of the red crab (Green 1997), and experimental evidence has demonstrated red crabs monopolize litter processing on the island, removing between 39 and 87% of the annual leaf fall from the forest floor (Green et al. 1999). Red crabs occur at biomass densities far greater than those reported elsewhere for entire litter faunas (Fittkau and Klinge 1973, McGlynn et al. 2007), making them important physical ecosystem engineers (Jones et al. 1994) by translocating organic matter and nutrients into the soil and reducing available habitat for other animals (Green et al. 1999). Red crabs also readily consume fruit, seeds and seedlings, which strongly influences seedling recruitment dynamics and impacts plant species composition on the island (Green et al. 1997). Almost all plant species are palatable to red crabs, such that very few individuals persist past an initial germination phase (Green et al. 1997). As a result, the rainforest understory on Christmas Island is structurally simple and very open, with few seedlings of only the couple of species resistant to red crab herbivory (Green et al. 1997). Shifting mosaics in crab densities over space and time may offer the best explanation for the recruitment of species vulnerable to red crab predation that are nonetheless represented in the overhead canopy (Green et al. 1997). Red crabs also significantly reduce the abundance of litter invertebrates indirectly through their removal of leaf litter (Green et al. 1999) and can exclude species completely through direct predation (Lake and O’Dowd 1991, Green et al. 2011).
There are 37 species of terrestrial or intertidal crabs on Christmas Island, many of which are endemic, but none are as abundant as the red crab (Orchard 2012). Also common in intact rainforest, in low densities, are robber (or coconut) crabs Birgus latro. Christmas Island remains one of the last places in which an abundant population of robber crabs persists (Orchard 2012). Elsewhere across its Indo-Pacific range, robber crabs have been so prized as a human food resource that their numbers have been drastically reduced (Lavery et al. 1996). Similarly, the endangered seabird Abbott’s booby Papasula abbotti is now endemic to the island after it became extinct from other islands within its former range (Yorkston and Green 1997). Concern for the future of this species was instrumental in the creation of the Christmas Island National Park in 1980, and extensions to the park throughout the 1980’s were aimed primarily at protecting habitat and known breeding sites for the species (Yorkston and Green 1997). The avian fauna on Christmas Island consists of 23 resident species, of which almost half (11) are endemic (including the Christmas Island frigate bird Fregata andrewsi, goshawk Accipiter fasciatus ssp. natalis, thrush Turdus poliocephalus ssp. erythropleurus and hawk-owl Ninox natalis) and only three are exotic (feral chicken Gallus gallus, tree sparrow Passer montanus and Java sparrow Lonchura oryzivora). The island only supports a single native mammal (the endemic flying-fox Pteropus melanotus ssp. natalis) following the extinctions of Christmas Island rats (Rattus macleari and Rattus nativitatis), shrew Crocidura trichura, and most recently the pipistrelle Pipistrellus murrayi. Similarly, only a single native reptile is commonly observed in the wild (the giant gecko Cyrtodactylus sadleiri) with the endemic blue-tail skink Cryptoblepharus egeriae and Lister’s gecko Lepidodactylus listeri existing exclusively in captivity (Smith et al. 2012), and the forest skink Emoia nativitatis recently extinct. A single individual of the endemic Christmas Island blind snake Ramphotyphlops exocoeti was recently collected in the rainforest, a species rarely observed since initial faunal collections were conducted on the island (c. 1890’s) (Maple et al. 2012). The dire state of native biodiversity on Christmas Island is either known to be, or suspected to be, the result of significant invasional processes impacting the rainforest community (O’Dowd et al. 2003, Davis et al. 2008, 2010, Beeton et al. 2010, Boland et al. 2011, Green et al. 2011, Smith et al. 2012).
Island ecosystems are impacted highly by invasive species because of low functional redundancy in the natural ecosystem and relatively simple food webs (Vitousek et al. 1996, Denslow 2003). Beginning with the rapid extinction of the native rats soon after human settlement, due to disease introduction by black rats Rattus rattus (Green 2014), the impacts across Christmas Island of the many invasive species are diverse and extensive. Some introductions are considered benign, such as the presence of the tree sparrow in settlement areas. Other introductions, like the invasive Asian wolf snake Lycodon capucinus and giant centipede Scolopendra subpinipes, are thought to be significant drivers of biodiversity loss (Smith et al. 2012); yet to date, no detailed study of their impacts has been undertaken. One of the most devastating invasions in terms of impacts on the island, and the one most well-studied and understood, is that of the pan-tropical invader, the yellow crazy ant Anoplolepis gracilipes (O’Dowd et al. 2003, Abbott 2006). When in mutualism with exotic honeydew-secreting scale insects, yellow crazy ants form expansive multi-queened supercolonies that cause local extinctions of the red crab, significantly disrupting natural litter decomposition and seedling recruitment processes (O’Dowd et al. 2003).
Yellow crazy ants have been present on Christmas Island from the early 1900’s (Donisthorpe 1935), but the distribution of the species was unknown until populations exploded in the early to mid-1990’s. The distribution of yellow crazy ants was assessed with a formal island-wide survey in 2001 that found that supercolonies, which consist of workers at extremely high densities (thousands of ants m-2), occupied a total area of over 3000 ha (Abbott 2006). Considerable research effort determined that this expansion was due to the subsequent introduction and establishment of exotic honeydew-producing scale insects, Tachardina aurantiaca (Kerridiae), Coccus celatus, C. hesperidium and Saisettia coffeae (all Coccidae), among others (O’Dowd et al. 2003, Abbott 2006, Abbott and Green 2007, Green and O’Dowd 2009). This mutualism benefits the scale insects through protection from parasitoids, parasites and predators, and the ants benefit from access to a constant carbohydrate food resource. Identified supercolonies ranged in size by nearly three orders of magnitude from 0.9 to 787 ha (Abbott 2006). Without intervention, supercolony boundaries were highly dynamic, with most expanding at a rate of 0.03 – 0.50 m day-1 (Abbott 2006). However, these boundaries were rarely sharply defined, but rather ant density decreased gradually from an area of highest abundance to form transition zones 46 – 153 m wide (Abbott 2006). Within transition zones, yellow crazy ants can coexist with other ant species as well as red crabs (Abbott 2006).
The general operational definition of a supercolony for both managers and researchers on Christmas Island is where the density of yellow crazy ants is sufficient to kill red crabs. This has been determined as ≥ 37 ants counted by a standard method of the number of ants crossing 11, 25 x 25 cm white cards, evenly placed along a 50 m transect, for 30 s each (Green and O’Dowd 2009, Boland et al. 2011). However, when supercolonies are at their highest densities, the total number of ants counted using this method is in the 100s. The most significant direct impact of crazy ant supercolonies is the extirpation of red crab populations, which has long-term implications for forest structure and composition through indirect effects on seedling and leaf litter dynamics (O’Dowd et al. 2003). Yellow crazy ants kill red crabs through sheer force of numbers and constant activity, overwhelming the crabs by spraying formic acid over their eyes and mouthparts (O’Dowd et al. 2003, Abbott 2006). Dead crabs can be abundant in recently formed supercolonies (O’Dowd et al. 2003). Also, many red crabs are killed in transit during their annual migration when migratory pathways intercept supercolonies (O’Dowd et al. 2003, Green et al. 2011). It has been estimated that potentially one-third of the entire red crab population, or around 30 million crabs, has been killed by this invasive ant since the initial island-wide survey in 2001 (D. Maple pers. comm.)
The local extinction of red crabs has a dramatic effect on rainforest structure and processes on Christmas Island (O’Dowd et al. 2003). First, leaf litter is allowed to build-up and persist throughout much of the year. This, along with an absence of burrowing activity, significantly alters the nutrient cycling dynamics (O’Dowd et al. 2003), which may be having an impact on subterranean biotic and abiotic properties. This persistence of a habitat and food resource may facilitate increased densities of litter invertebrates that were previously under-represented in intact rainforest (Green et al. 1999). Second, seedling recruitment dynamics are significantly altered (O’Dowd et al. 2003). Seedlings are no longer consumed immediately after germination (Green et al. 1997) and are allowed to grow and persist, resulting in high seedling density and increased seedling species richness in ant-impacted areas (O’Dowd et al. 2003). This creates rainforest that has a dense and complex understory compared to the open and structurally simple understory of intact rainforest. Third, red crabs are known to provide direct ‘biotic resistance’ through their predatory behaviour (Lake and O’Dowd 1991). Their removal creates the enemy-free space necessary for previously unsuccessful species to now enter the rainforest, establish reproductive populations and become invasive (Green et al. 2011). The invasive ant-scale mutualism on Christmas Island has created a recipient community with significantly altered properties compared to its intact state.
The invasional meltdown hypothesis describes a community-level phenomenon in which the net effect of facilitations leads to an increasing rate of establishment of introduced species and/or accelerating impact (Simberloff and Von Holle 1999, Simberloff 2006). The invader-invader mutualism on Christmas Island provides the best case study, to date, of invasional meltdown because not only have community-level consequences been revealed (O’Dowd et al. 2003, Abbott and Green 2007), the establishment of another introduced species has been demonstrated to be completely contingent on these impacts (Green et al. 2011). This idea of species invasions being facilitated by the presence and influence of previous invaders is referred to most commonly as ‘secondary invasion’ (e.g. Floerl et al. 2004, Heneghan et al. 2007, Green et al. 2011) – although not explicitly by Simberloff and Von Holle (1999). As the number of studies considering invasional meltdown has increased, examples of the invasion success of one species being significantly influenced, or even contingent, on other exotic species have emerged (Adams et al. 2003, Grosholz 2005, Johnson et al. 2009, Green et al. 2011).
The invasive ant-scale mutualism has facilitated the secondary invasion of the giant African land snail Achatina (Lissachatina) fulica in impacted rainforest on Christmas Island (Green et al. 2011). Although present on the island since the mid-1940’s, A. fulica never penetrated beyond the disturbed margins of rainforest because the snails were rapidly discovered and consumed by abundant red crabs (Lake and O’Dowd 1991). Green et al. (2011) used both survey and experimental data to describe the spatiotemporal pattern of spread of A. fulica across the island over seven years, and explored whether changes in the recipient community and propagule pressure wrought by invader-invader mutualism influenced the probability of invasion by A. fulica. Modelling of A. fulica spread across the island showed that invasion was facilitated 253-fold in ant supercolonies but impeded in intact forest where predaceous native red crabs remained abundant (Green et al. 2011). Predation pressure on A. fulica was lower, survival higher, and abundance 20-fold greater in supercolonies than in intact rainforest (Green et al. 2011). This study revealed that ant supercolonies, by killing red crabs but not A. fulica, disrupted biotic resistance and provided enemy-free space, thereby facilitating the entry of a secondary invader, propagating their spread at the whole-ecosystem level (Green et al. 2011).On Christmas Island, trophic interactions between exotic ant-scale mutualists reconfigure interaction networks (Fig. 3).
Invasive species are known to alter both the abiotic and biotic components of a community via a number of mechanisms. The entry and establishment of certain plants can alter soil nutrient properties (Vitousek and Walker 1989) as well as species diversity patterns (Darrigran 2002, O’Loughlin et al. 2015) and significantly alter vegetation structure (Braithwaite et al. 1989). As invasion success can be determined strongly by the properties of recipient communities, understanding how these are altered by the presence and influence of successful invaders in paramount to understanding secondary invasion. When considering secondary invasion, we were interested in how the properties of a recipient community had initially provided a filter limiting invasion success, and how these properties were altered by other exotic species.
The invasive ant-scale mutualism on Christmas Island significantly alters multiple properties of the recipient community, and as such, may be responsible for facilitating secondary invasions of other exotic species. Indirect habitat augmentation that increases habitat and resources in the form of persistent leaf litter and seedling recruitment, along with the creation of enemy-free space, provide a rainforest community with less biotic resistance to invading propagules. Achatina fulica is just one of 22 exotic land snail species present on Christmas Island (Kessner 2006), the remainder of which are considerably smaller (< 20 mm). A further 14 species are native to the island (Kessner 2006). Due to their small size, these species of land snail may avoid direct detection and predation by red crabs and not be inhibited strongly from invading rainforest on the island. However, where red crabs have been eliminated, these species may experience ‘release’ from a habitat of low resources (leaf litter) and potential indirect predation imposed on them by these abundant land crabs.
Overview of thesis objectives
This thesis uses the invader-invader mutualism on Christmas Island to continue the investigation of invasional meltdown broadly, and the nature of secondary invasion more specifically. Four main objectives are investigated: i) defining the concept of secondary invasion, ii) documenting a pattern of secondary invasion, iii) determining the mechanism of secondary invasion, and iv) assessing the impact of a secondary invasion. Each objective is addressed in one of four chapters, and synthesised in a final chapter that discusses the implications and significance of this work, as well as future directions for research (Fig. 4).
Objective 1: defining the concept
Chapter 2 consists of a concept / review paper that aims to define and explore the idea of secondary invasion. Currently, invasion success being facilitated by other invaders is not formally defined in the ecological literature, and the term ‘secondary invasion’ is used inconsistently to refer to a number of ecological phenomena. The specific aims of this chapter are three-fold; i) to place secondary invasion within existing heuristic models of invasion success and propose theoretical pathways, ii) discuss the mechanisms by which primary invaders indirectly (or directly) facilitate secondary invaders via changes to the properties of recipient communities, and iii) examine the kinds of data required to determine if secondary invasion is occurring.
Objective 2: documenting a pattern
Chapter 3 explores whether the entire land snail community on Christmas Island responds favourably to rainforest impacted by the invasive ant-scale mutualism, and therefore, whether those exotic species represent secondary invaders. The specific aims of this chapter are three-fold; i) to document how land snail abundance, species richness and composition changes between different forest states that have arisen as a result of the ant-scale invasion, ii) to observe whether exotic land snail species were only present where the recipient community had been altered, and iii) to note whether native land snail species were also facilitated by these changes.
Objective 3: determining the mechanism
Chapter 4 directly follows the findings of Chapter 3 and aims to experimentally determine the relative importance of the direct and indirect changes to the recipient community on the invasion success of these land snails. This chapter specifically asks, i) what is more important to the invasion success of land snails on Christmas Island – the creation of enemy-free space (due to the direct removal of the red crab) or the release of habitat and resources in the form of leaf litter (an indirect change associated with the removal of the red crab), and ii) is the mechanism of invasion success dependent on land snail size?
Objective 4: assessing the impact
Chapter 5 details whether a successful secondary invader is causing further impact on the invaded community through their high abundance. Under the invasional meltdown hypothesis, the presence of secondary invaders would increase and amplify impacts on the invaded community. This chapter uses a number of approaches to assess the impacts of highly abundant A. fulica in rainforest on Christmas Island. The specific aims of this chapter are three-fold; i) to establish whether A. fulica are more herbivorous or detritivorous in their dietary preferences, ii) to assess whether A. fulica is having an impact on seedling recruitment and/or survival, and iii) to ascertain if A. fulica are having an impact on leaf litter dynamics.
This thesis has been written with the intention that chapters 2-5 will be submitted to peer-reviewed journals for publication, and as such, they are presented here as stand alone pieces of work. This means that there is some overlap between chapters in the material presented (particularly in the descriptions of the study system).
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