Theoretical support for the regulation of host populations by parasites (including pathogens, parasitoids and other parasites) accumulated in the late 1970s following the publication of two landmark papers (Anderson & May 1978; May & Anderson 1978), but empirical evidence for population-level effects of parasites on their hosts has been limited (Tompkins et al. 2010). However, the effects of parasitism extend beyond populations, to communities and even ecosystems, because of direct and indirect interactions at multiple scales (Tompkins et al. 2010). Limiting studies to the population level could therefore result in misleading conclusions about parasite interactions. Here, we explore the contribution of parasite interactions, in the context of community ecology, with specific reference to the success of an invasive alien predator, the harlequin ladybird, Harmonia axyridis. We begin by defining the main theories and processes relating to parasite–host interactions in the context of invasion biology, before outlining the advantages of a community ecology approach, and particularly ecological network analysis (ENA), for revealing dominant mechanisms in determining the success of invasion.
- There are a number of theories and processes relating to invasive alien species (IAS) and their interactions with natural enemies (predators, parasites and pathogens), including the enemy release hypothesis, spillover and spillback.
- Most empirical studies focus on pairwise interactions and are limited to the population-level avoiding the complexities of a community approach. Oversimplified studies could result in misleading conclusions because of the importance of positive and negative feedback mechanisms, mediated by parasites, throughout communities and beyond.
- Recent advances in ecological networks analysis (ENA) provide a powerful framework to investigate the complexity of interactions within ecological communities and response to disturbance, such as the arrival of an IAS. Molecular data, from either standard PCR or next-generation approaches, can be incorporated into ecological networks to enable the number and strength of interactions to be quantified; however, this has rarely been performed so far.
- In this paper, we explore the contribution of parasite interactions, in the context of community ecology, to the success of invasive alien predatory insects with specific reference to the harlequin ladybird, Harmonia axyridis. We begin by defining the main theories and processes relating to parasite–host interactions in the context of invasion biology, before exploring the advantages of a community ecology approach, and particularly ENA, for revealing dominant mechanisms in determining the success of invasion.
Invaders and their parasites: theoretical concepts and processes
The population growth rate of an invasive alien species (IAS) is determined by three factors that vary at spatial and temporal scales: available resources, interactions with natural enemies (predators and parasites including parasitoids and pathogens) and the abiotic environment (Shea & Chesson 2002). The relative importance of these factors varies between species and determines the ability of the species to invade. As an IAS increases in density, it will influence the invaded locality through interactions with other species within the community, for example by recruitment of natural enemies. For some invaders, there is a distinct time lag in this recruitment process, while native natural enemies adapt to the specific life-history traits (morphological, physiological, behavioural and ecological) of the alien species (Cornell & Hawkins 1993). Therefore, the process is dynamic and specific outcomes of interactions will undoubtedly vary over time. However, there are many facets of community ecology theory which provide opportunities for unravelling some of the complex interactions between invaders and their natural enemies including parasites.
Enemy release and evolution of increased competitive ability
There are a number of theories relating to IAS and their interactions with natural enemies. Perhaps, the most evocative is the Enemy Release Hypothesis (ERH), also referred to as enemy escape or escape-from-enemy hypothesis (Elton 1958; Jeffries & Lawton 1984; Roy et al. 2011a). The ERH predicts that an alien species introduced to a new region will increase in distribution and abundance because of reduced impacts from natural enemies (Roy et al. 2011a). This theory relies on the notion that natural enemies are important in regulating populations and, furthermore, that such natural enemies have a stronger regulatory effect on native species than they do on alien species in the introduced range. It is this disparity in enemy regulation that enables the alien species to exhibit increased population growth. Two mechanisms have been proposed, depending on the evolved host defences of the alien species, which could lead to increased population growth: regulatory or compensatory release (Colautti et al. 2004). A reduction in enemies in the introduced range, for hosts that are strongly regulated by enemies in their native range, may lead to direct changes in survivorship, fecundity, biomass or other demographic parameters (regulatory release). Conversely, a reduction in enemies may be of little immediate consequence for hosts that are well defended and, consequently, lack natural enemies within their native range. In this case, fewer enemies may lead to a reallocation of resources from defence to population growth over ecological time (Blossey & Notzold 1995), sometimes referred to as the Evolution of Increased Competitive Ability (EICA), or selection of genotypes with reduced defences over evolutionary time (compensatory release). Empirical evidence for the role of the ERH or EICA in invasion success is lacking (Roy et al. 2011a), particularly for invertebrates, although there are a few landmark studies supporting the ERH as an important mechanism in invasion success (Torchin et al. 2003; Torchin & Mitchell 2004).
There have been two main approaches to studying the ERH: ‘biogeographical studies’ that compare richness and impacts of enemies in native and introduced populations of an alien host, and ‘community studies’ that compare native and alien species occurring within the same community (Colautti et al. 2004). Biogeographical richness-based assessments of the ERH are rare for animal invasions, but the few studies available have generally found more species of enemy associated with an alien species in its native rather than introduced range. A major criticism of such studies is the bias of higher sampling effort in the native relative to the introduced range, particularly for recent invasions (Colautti et al. 2004). In addition, invasive populations should be directly compared with their probable source population(s) in the native range to avoid exaggerated estimates of enemy release (Colautti et al. 2004; Slothouber Galbreath et al. 2010). This second criticism was recently addressed in a test of the ERH in the invasive amphipod, Crangonyx pseudogracilis (Slothouber Galbreath et al. 2010). Putative source population and host genetic diversity were first determined by sequencing nuclear and mitochondrial genes. Microsporidian diversity was then compared between invasive and source populations. No difference in microsporidian diversity, despite reduced genetic diversity in the invasive population, suggested enemy release has not occurred in this system. An additional important consideration, with both ‘biogeographical studies’ and ‘community studies’, is how the loss of enemy diversity translates into population regulation; a small number of enemies may have large effects. Indeed, functional diversity of enemies may be a better predictor of impacts upon hosts than overall diversity (Colautti et al. 2004).
Parasite spillover and spillback
Two processes, which have received considerably less attention than ERH and EICA, relate to the potential role of the IAS as a reservoir for parasites (Fig. 1). An invader could facilitate the dispersal of new parasites to native species through the process of ‘spillover’ or act as a new host for parasites that they transmit or ‘spillback’ to native species (Kelly et al. 2009). The term ‘parasite spillover’ has been used to describe the transmission of disease from wild animals to domesticated animals or humans, for example H5N1 or avian influenza which can spread from wild birds to humans (Wolfe, Dunavan & Diamond 2007).
There are very few examples of spillover in wild animals because parasite outbreaks in wildlife go largely unnoticed unless humans are in some way affected (Otterstatter & Thomson 2008). This is particularly the case for insect diseases (Goulson 2003). The most thoroughly documented, although still circumstantial, evidence for parasite spillover in insects concerns the transfer of parasites from introduced (commercially reared) bumble bees to wild bumble bees, Bombus spp. (Otterstatter & Thomson 2008). There is no empirical evidence of spillover, from invasive alien insects to native insect species within an introduced range, but the potential exists and should be examined.
‘Parasite spillback’ has been described as ‘a neglected concept in invasion ecology’ (Kelly et al. 2009). Parasite spillback could occur when an alien species is a competent host for a native parasite or pathogen. The additional alien host provides a reservoir for a parasite that then increases in prevalence and can spill back into native species. Parasite spillback is an example of apparent competition in which two or more species that do not directly compete for the same resource share a natural enemy (Kelly et al. 2009). Empirical evidence is lacking for spillback in invaded systems, but there are many examples of alien species acquiring generalist pathogens or parasites after invasion (Cornell & Hawkins 1993; Torchin et al. 2003). Indeed, Cornell & Hawkins (1993) demonstrated that 25% of alien insect herbivores acquired more than 10 parasitoid species. However, the acquisition of native parasites and pathogens by an alien species will not necessarily lead to spillback into native fauna; the process depends on the alien species disseminating the parasite or pathogen and acting as a reservoir. It is possible that alien species could be sinks for the pathogen or parasite and so reduce infection prevalence in the native fauna (Heimpel, Neuhauser & Hoogendoorn 2003; Keesing, Holt & Ostfeld 2006).
The influence of parasite life history during invasions
Parasites and pathogens are taxonomically diverse, and this is reflected in their life histories. The impact of a parasite on invasion is likely to be influenced by specific life-history traits. For example, the survival of a parasite during translocation from the native to invaded range could depend on the mode of transmission. Vertically transmitted parasites are more likely to be introduced with their invading host than horizontally transmitted parasites (Mitchell & Power 2003; Prenter et al. 2004; Dunn 2009), and by definition, less important in spillover or spillback. Vertically transmitted parasites are not dependent on host density and are associated with low virulence. Both factors will increase the probability of host and parasite survival through the invasion process. Sex-ratio-distorting bacteria, such as Wolbachia, are maintained in host populations by a combination of predominantly vertical transmission and conferring indirect fitness benefits on female hosts. For example, female neonate ladybird larvae from mothers infected with male-killing bacteria obtain an indirect fitness advantage by consuming undeveloped male eggs (Roy et al. 2007; Elnagdy, Majerus & Lawson Handley 2011). This is important in the case of invasions because of the implications for host fitness and adaptation. Theoretically, female-biased sex-ratio distortion, combined with enhanced female fitness, could facilitate host invasion, as implicated in the sweet potato whitefly, Bemisia tabaci (Himler et al. 2011), and amphipod, C. pseudogracilis (Slothouber Galbreath et al. 2004). Male-killers should also promote dispersal in their hosts to facilitate their own spread (Bonte, Hovestadt & Poethke 2008). However, few other studies have investigated male-killers or other vertically transmitted endosymbionts in the context of biological invasions (Shoemaker et al. 2000; Zindel, Gottlieb & Aebi 2011), and further investigation is needed to fully understand whether endosymbionts are likely to generally facilitate or inhibit host invasion (Zindel, Gottlieb & Aebi 2011).
A second life-history trait of parasites relevant to invasion is their host breadth, and it is essential to recognise the distinction between generalist and specialist parasites when exploring invasion processes. This distinction is of intuitive and obvious importance when considering spillover and spillback, as both concepts are more relevant to generalists than specialists, but it is also highly relevant to the role of ERH in invasion success (Müller-Schärer, Schaffner & Steinger 2004). The number of hosts that different species of parasites and pathogens attack is extremely variable. Generalist parasites and pathogens attack many diverse host species and are not constrained by the generation time of a particular target species; therefore, they have the potential to be widely distributed. Conversely, specialists may have only a single or a few closely related host species and an intimate relationship with their hosts (developing within and having a similar generation time); therefore, it is unlikely that they will occur outside the native range of their host (Snyder & Ives 2003). Consistent with this, and the prediction that release from vertically transmitted parasites is more likely than release from horizontally transmitted parasites, as discussed above, Mitchell & Power (2003) confirmed that invasive plants are less likely to escape seed-transmitted viruses with typically broad host ranges than their more specialist fungal pathogens.
Of course, there is a spectrum from specialist to generalist, and a single parasite or pathogen species can vary across this spectrum depending on a number of factors including environmental conditions, which may vary considerable in the invaded vs. native range. There is a further complication when discriminating between specialists and generalists and that is the concept of ‘faux generalists and specialists’ (Radtke, Mclennan & Brooks 2002). ‘Faux generalists’ are parasites that specialise on a particular resource that is phylogenetically widespread as opposed to a particular host, while ‘faux specialists’ are generalists restricted to a single or few target species by ecological factors (such as competition, allopatric distributions, local climatic conditions or habitat substrate). Host switching can be initiated for both faux specialists and generalists by a change in ecological conditions (appearance of an alien host or the decline in a native host). Therefore, faux specialists and generalists can complicate the predictions of the ERH by diffusing the gradient between the relative proportions of specialist to generalist enemies in the native and introduced ranges (Roy et al. 2011a).
Keane & Crawley (2002) outlined three key predictions of the ERH based on the specialist or generalist nature of natural enemies. When a species is introduced into a new region, (i) its specialist enemies will be absent from the new region; (ii) host switching by specialist enemies of native congeners will be rare (therefore, the introduced species is unlikely to be attacked by native specialists); and 3) generalists will have greater impact on native species (with which they might they be better adapted) than on introduced competitors, which could be expected to reduce interspecific competition for the introduced species although there is so far little empirical evidence for this (Keane & Crawley 2002). Roy et al. (2011a) explored these predictions in relation to alien arthropod predators and parasitoids and expanded the predictions, acknowledging that specialists and generalists will co-occur, through a consideration of both potential competitive and trophic interactions between specialists and generalists. The most important interaction in this context is likely to be intra-guild predation (IGP), which is common among generalists (Roy & Pell 2000), and can exacerbate prey outbreaks by removing key specialists (Rosenheim et al. 2004). Detailed assessment of life-history traits associated with invasion and acquisition of parasites by invaders could reveal intriguing insights and elucidation of broad patterns with respect to parasites and invasions (Prenter et al. 2004).
While the ERH and EICA emphasise species characteristics (or autecology) in the context of the introduced habitat, other hypotheses dealing exclusively with the habitat or community into which the alien species is introduced have also been evoked to explain invasion success. These include the Biotic Resistance Hypothesis, aka Species Richness (Elton 1958), the Fluctuating Resource Hypothesis (Davis, Grime & Thompson 2000) and the Disturbance Hypothesis (Hobbs & Huenneke 1992). Frameworks are required in which the interactive effects of resources, natural enemies, native competitors and the physical environment on invasion success of alien species can be evaluated (Shea & Chesson 2002).
A community approach to studying invaders and their parasites: integrating ecological network analysis with modelling and molecular tools
Community ecology, and particularly niche concepts, has been proposed as a theoretical framework for understanding invasions (Shea & Chesson 2002). Elucidating the major mechanisms that contribute to an alien species becoming invasive is seen as essential for limiting the impacts of alien species on biodiversity (Keane & Crawley 2002; Allendorf & Lundquist 2003). It is widely accepted that characteristics of both invaders (such as reproductive strategies and resource acquisition) and invaded communities (such as community maturity and niche opportunities) influence the outcome of invasion (Shea & Chesson 2002). Furthermore, organisms within ecological communities are both directly and indirectly linked to one another as predators, prey, hosts, parasites or competitors, leading to complex networks of interactions that underpin all ecological processes. Understanding how and to what extent interspecific interactions influence community structure, species coexistence and biodiversity, and how these interactions respond to global environmental change, is one of the biggest challenges facing ecologists (Hatcher, Dick & Dunn 2006). However, most studies still focus on pairwise interactions and avoid the complexities of a community approach. This is particularly the case for studies considering the role of parasites within an invasion process. Niche concepts provide a tangible framework, but a broad community ecology approach, encompassing ecological network analysis (ENA), could prove even more informative.
Ecological Network Analysis
Ecological network analysis arguably provides the most exhaustive method for attempting to holistically assess community interactions (Fath et al. 2007). Most studies on ecological communities include only a small subset of the interactions occurring within an ecosystem, thereby excluding the majority of species in the community and abiotic processes. Consequently, many studies conclude that only a few species or processes are influential and neglect to consider the embedded nature of these within a rich web comprising many interactions (Fath et al. 2007). Quantitative ecological networks (or ‘interaction webs’), as depicted in Fig. 1, provide a graphical representation of community structure and a scaffold for investigating the strength of host-parasite, predator-prey, plant-pollinator or other species interactions (represented by links between species) and robustness of communities to disturbance, such as species declines and extinctions (Pocock, Evans & Memmott, 2012). ENA provides an opportunity to assess the structural and functional properties of entire webs of interactions. The ambition of such an approach is appealing but challenging.
Fath et al. (2007) outlined a step-by-step approach to ENA. Initially, the boundary of the network or ecosystem of interest is identified, and subsequently, a list of all the major species and functional groups within the ecosystem is compiled. The model can be aggregated, with few groupings (producers, decomposers, consumers), or disaggregated, with many groupings (producers, herbivores, carnivores, omnivores, decomposers and detritivores) or indeed fully disaggregated to consider the network surrounding each individual species (Fath 2004). The next step involves determining the relevant energy-matter flow currency such as biomass or energy. It is then necessary to elucidate all potential interactions through which the currency is exchanged and test these empirically or infer on the basis of available and defined evidence such as literature sources. ENA can then be applied to the network.
Invasive alien species (or other range-shifting species) can potentially influence ecological networks by altering diet breadth and strength of predator–prey interactions and triggering trophic cascades and extinctions (Montoya & Raffaelli 2010). A key question is whether ecosystem services are maintained when ecological networks are infiltrated by IAS (Aizen, Morales & Morales 2008; Tylianakis 2008). Although ecological networks analysis is arguably the most powerful tool currently at our fingertips for studying community interactions, it has yet to reach its potential in the context of invasions. A few pioneering studies have however used ENA to investigate the impact of IAS on the structure of pollination networks (Henneman & Memmott 2001; Memmott & Waser 2002; Sheppard et al. 2004; Tylianakis 2008; Vila et al. 2009). For example, Aizen, Morales & Morales (2008) compared the structure (overall connectivity) and interaction strength of plant-pollinator networks in several geographical replicates of highly invaded and less invaded communities. Key findings were that IAS were highly generalist compared to native pollinators and replaced native pollinators in the networks, although overall network connectivity remained the same. Moreover, IAS affected the strength of mutualistic interactions, with highly invaded communities exhibiting weaker mutualism than less invaded communities (Aizen, Morales & Morales 2008).
An ecological networks approach could be enlightening for investigating enemy release, parasite spillover and spillback. A simple prediction of the ERH is that parasite diversity will be lower in the invasive compared to native range. However, the strength of host–parasite interactions is also crucially important. Quantitative interaction webs could be used to test the prediction that the breadth and strength of host–parasite interactions are lower in the invasive than the native range of IAS (Fig. 2). Spillover and spillback can be investigated by characterising the community of parasites shared between the IAS and other potential hosts. Under spillover, parasites found in the native range of an IAS will be shared with the IAS and closely related suitable hosts in the invasive range. These parasites should not however be present in naïve, but geographically similar, communities in the invasive range. Under spillback, native parasites will be shared between the IAS and closely related species in the native range. Again, the benefit of addressing these hypotheses in a network framework is that the strength of interactions can be investigated, and the response of entire communities to the presence of IAS and their parasites can be tested through simulation.
Modelling ERH within an ecological network framework
Many studies assessing the ERH assess the importance of parasites in invasion success without attempting to reject the importance of other factors (Colautti et al. 2004), largely ignoring for example the characteristics of the habitat being invaded or the time since introduction (Barney & Whitlow 2008). Roy et al. (2011a) reviewed empirical modelling approaches for testing the ERH (against alternative hypotheses for invasion success), summarising conclusions of recent comprehensive meta-analyses of evidence for and against ERH (Keane & Crawley 2002; Colautti et al. 2004; Liu & Stiling 2006) and discussing their potential application to invasive alien arthropod predators and parasites.
Population matrix models could be very useful for understanding differential impacts of enemies on pairs of native and alien species. However, ENA and food web analysis may provide a more tractable alternative when dealing with multiple interacting enemy or alien species (Hatcher, Dick & Dunn 2006; Lawson Handley et al. 2011). Theoretical population models have also been developed to test particular assumptions of the ERH. Drake (2003) suggested that the role of enemy release may differ among the introduction, establishment and spread phases of an invasion and evidence is building in support of this prediction. For example, even if a parasite is introduced with an alien host species, it would have to exhibit a very high virulence (killing more than 90% of the population per generation) to have an impact on host establishment. Modelling approaches have also led to the conclusion that parasites may help to stabilise host population dynamics, therefore favouring the successful establishment of an IAS (Lively 2006).
In proposing community ecology as a framework for biological invasions, Shea & Chesson (2002) highlighted the importance of considering the interactions between resource opportunity and escape opportunity in governing invasion and of controlling for community maturity and covarying extrinsic abiotic factors in studies. Barney & Whitlow (2008) proposed instead a state factor model that integrates all identified aspects of the invasion process encompassing those implicit in both species-based and habitat-based hypotheses for invasion success. The model equation relates any quantifiable properties of an invasion (e.g. impact or extent of spread) to a number of factors, including propagule pressure, properties of the introduced habitats, species autecology, native habitat and the time since introduction. The value of this approach lies in the explicit experimental recognition of all contributing variables characterising invasions, the identification of critical knowledge gaps and the acknowledgement of the involvement of some key variables such as time since introduction that we may or may not be able to control.
Incorporating molecular data into ecological networks
The last decade has seen considerable interest in using molecular data for characterising trophic interactions (Zaidi et al. 1999; Symondson, Sunderland & Greenstone 2002; Aebi et al. 2011; Gagnon et al. 2011; King et al. 2011; Pompanon et al. 2011), but while this data holds great promise for unravelling food webs (Sheppard & Harwood 2005), to our knowledge, direct integration of molecular data into ecological networks has so far been overlooked. Instead, constructing ecological networks has relied on comprehensive field surveys, traditional taxonomy and rearing experiments or feeding trials, which are laborious and may be underestimating the diversity of interactions. This is particularly the case for host–parasite interactions, as distinguishing parasite species can be very difficult using morphology alone, interactions are rarely witnessed in the wild, and parasites are often cryptic or covert (Roy et al. 2009; Hesketh et al. 2010; Lawson Handley et al. 2011). Here, we propose that molecular data, generated from standard DNA barcoding or second-/next-generation sequencing technologies (outlined below), offer a potentially powerful counterpart to traditional taxonomic methods and could offer unique insights into the structure of ecological networks.
Standard DNA barcoding methods involve determining a species' ID from a short region of DNA (typically the mitochondrial cytochrome oxidase subunit I gene, COI), or polymerase chain reaction (PCR) assays that target a particular species with specifically designed PCR primers (Symondson 2002; King et al. 2008; Rougerie et al. 2010). Standard DNA barcoding requires DNA isolation of single parasite (or prey) species from the host (or predator), so that contamination is avoided and target DNA sequences can be unambiguously determined. The main constraints of standard approaches are therefore that assays target only a small, known fraction of the parasite or prey community, and development of species-specific PCR primers can be laborious and are often unsuccessful.
The drawbacks associated with standard barcoding approaches can be overcome by using second or next-generation sequencing technologies such as ‘massively parallel sequencing’ or ‘ultra-deep sequencing’ (Lawson Handley et al. 2011; Pompanon et al. 2011). In massively parallel sequencing, which is essentially DNA barcoding on a large scale, DNA from all the organisms within an environmental sample or the entire community of parasites within a single host (or prey within a predator's gut) is extracted simultaneously. A standard DNA barcoding gene (again typically COI or nuclear small subunit (SSU) ribosomal genes, for example 18S) is PCR amplified using conserved primers, so that PCR product is obtained for each individual organism within the pooled sample. Single DNA sequences from the pool are then separated (for example on individual tagged beads), and then PCR amplified again and sequenced (for example) on a Roche 454 FLX-Titanium platform (Mardis 2008; Creer et al. 2010). While massively parallel sequencing is an established method for characterising biodiversity in environmental samples (Fonseca et al. 2010; Andersen et al. 2011) and has also occasionally been used to study diet (Soininen et al. 2009; Valentini et al. 2009). Cox-Foster et al. (2007) used massively parallel sequencing to survey the microbial community in honeybees, Apis mellifera, from normal hives, compared to those affected with colony collapse disorder (CCD). Their results suggested a correlation between the presence of Israeli acute paralysis virus and CCD (Cox-Foster et al. 2007).
More recently, the honeybee microbiome was investigated further, by characterising new pathogens using ultra-deep sequencing and screening for known arthropod pathogens using a custom built microarray (Runckel et al. 2011). In the ultra-deep sequencing approach, whole target genomes (in this case honeybee), along with their parasites, pathogens and symbionts, are sequenced (as opposed to a single homologous gene in the massively parallel approach). In the majority of whole-genome sequencing projects, parasite and other ‘contaminating’, non-target sequences are typically removed during the initial bioinformatics stages. However, in this case, non-honeybee sequences were specifically targeted, and analyses revealed four novel RNA viruses; one of which (Lake Sinai Virus 2, LSV2) was the most abundant component of the honeybee microbiome (Runckel et al. 2011). This virus was previously undetected, probably because of the extreme divergence from other insect viruses (Runckel et al. 2011). This highlights a potential major advantage of a ‘shotgun’ ultra-deep sequencing approach compared to massively parallel sequencing, as the former does not rely on ability of primers to bind to and amplify all taxonomic groups. As ‘contaminating non-target’ sequences are a natural bi-product of whole-genome sequencing, it seems wise to make the most of this resource in the future.
The studies discussed here have provided unique insights into trophic interactions and community diversity, but standard or next-generation barcoding data have yet to be employed to test simple predictions of enemy release, spillover and spillback, let alone be integrated into ENA. We believe that second and next-generation sequencing technologies will revolutionise the study of ecological networks, because methods are extremely sensitive and (at least semi-) quantitative, and the gut contents and parasite/pathogen community of a target species can be analysed simultaneously.
Case study: harlequin ladybird, H. axyridis – a model species
The predatory ladybird, H. axyridis (Pallas) (Coleoptera: Coccinellidae), is documented as one of the fastest spreading IAS (Brown et al. 2008, 2011; Roy et al. 2011c). It is native to Asia but has spread at variable rates across the globe ranging from 100 to 500 km year−1. Harmonia axyridis is now considered as established and resident in at least 38 countries over four continents across its introduced range (Brown et al. 2011). This IAS is a voracious predator and was introduced as a biological control agent of aphids and coccids in a range of crops including alfalfa, cotton, maize, soybean, tobacco and winter wheat (Koch 2003; Majerus, Strawson & Roy 2006). There is empirical evidence that H. axyridis has succeeded in controlling insect pests on crops but, as a generalist predator, the negative ecological effects of this IAS are a major concern (Roy & Wajnberg 2008). A large-scale and long-term study demonstrated declines in historically common and widespread ladybirds in three European countries in response to the arrival of H. axyridis (Roy et al. 2012).
The parasites and pathogens of ladybirds are well documented (Ceryngier, Roy & Poland 2012) and their association with H. axyridis has received some preliminary attention (Roy et al. 2011b). Therefore, H. axyridis provides an opportunity for studying the role of parasites in the invasion process at a community scale (Fig. 3). Here, we provide a brief overview of the parasites and pathogens of H. axyridis and highlight possibilities for further study. We consider the parasites for which H. axyridis already acts as a host and discuss the potential for novel host–parasite interactions within the context of community ecology specifically assessing the potential for ERH, spillover and spillback. It seems that H. axyridis may have ‘left behind’ parasites within the native range, and evidence from both North America and Europe suggests that parasites of native coccinellids in these regions are beginning to utilise H. axyridis as a host (Roy et al. 2011b).
The Coccinellidae are host to a number of different parasites including fungal pathogens, bacterial endosymbionts, nematodes, hymenopteran and dipteran parasitoids [Table S1, Supporting information (Ceryngier, Roy & Poland 2012; Roy & Cottrell 2008c)]. Roy et al. (2011) provide a detailed account of these parasites. Some of these natural enemies occur globally, attacking H. axyridis in both its native (Asia) and introduced range, while others are more limited in their distribution, attacking H. axyridis in its native but not invaded regions, or vice versa. Figure 3 provides a schematic of possible interactions between these parasites and the IAS, H. axyridis.
Parasites of H. axyridis with emphasis on the cosmopolitan braconid Dinocampus coccinellae
Within Asia, H. axyridis is attacked by several parasitoid wasps, including D. coccinellae (Schrank) (Hymenoptera: Braconidae) (Liu 1950; Kuznetsov 1997), Homalotylus flaminius (Dalman) (Encyrtidae) (Kuznetsov 1997) and Oomyzus scaposus (Thomson) (Eulophidae) (Kuznetsov 1997). It is also host to several parasitoid flies, including Phalacrotophora philaxyridis Disney, P. berolinensis and P. fasciata (Phoridae) (Disney & Ve Beuk 1997), Strongygaster triangulifera (Loew), Medina luctuosa and M. separata (Tachinidae). There have been a number of reports of Phalacrotophora sp. emerging from H. axyridis within the native and invaded range of the host (Osawa 1992).
Most research on the natural enemies of H. axyridis has focussed on the braconid parasitoid D. coccinellae. Dinocampus coccinellae is a solitary, koinobiont endoparasitoid, that has been recorded from adults of 22 species of coccinellid in Europe, all from the subfamily Coccinellinae (Ceryngier, Roy & Poland 2012). This parasitoid has a worldwide distribution and is the most prevalent of all coccinellid parasitoids (Ceryngier, Roy & Poland 2012). It attacks all life stages, but prefers, and is most successful in adults (Hoogendoorn & Heimpel 2002).
Harmonia axyridis is attacked by D. coccinellae within its native range (Liu 1950; Maeta 1969; Kuznetsov 1997), but there is very little information on the parasitoid-host dynamics. There is a similar lack of information within the invaded range of H. axyridis although there have been a number of reports of parasitisation. Indeed, D. coccinellae parasitism of H. axyridis was first reported in North America (Hoogendoorn & Heimpel 2002) and Canada (Firlej et al. 2005). Successful parasitisation, although with low prevalence, of H. axyridis by D. coccinellae has also been documented in Denmark (Steenberg & Harding 2009b). There have been occasional sightings of D. coccinellae parasitising H. axyridis in the field in Britain but at considerably lower prevalence in comparison with the number of C. septempunctata parasitised (UK Ladybird Survey).
Understanding the population dynamics of this parasitoid in association with H. axyridis and co-occurring native ladybirds (particularly C. septempunctata in Europe) could yield fascinating insights from ecological and evolutionary perspectives. Harmonia axyridis is a marginal host and could be acting as a sink for D. coccinellae. However, D. coccinellae is likely to adapt to this super abundant invading host population and as such be implicated in spillback to native species, further exacerbating negative effects. There is a unique opportunity to assess this relatively new invader (first record of H. axyridis in Britain was 2004) through large-scale and long-term studies to assess the dynamics of this parasitoid with both H. axyridis and native hosts. A community ecology framework will be pivotal to unravelling the role of this parasite in the invasion process because many interacting factors will be important including population density of the various host species and interactions with other parasites such as microsporidia.
Conclusions and further work
The role of parasites in invasion has mostly been studied within the context of the ERH and other associated theories concerning direct interactions; however, parasites have also been implicated in mediating indirect interspecies interactions (competition and predation). Apparent competition, although difficult to demonstrate conclusively, is likely to have a role in invasion success (Prenter et al. 2004). Spillback can be considered a form of apparent competition in which two species interact through a shared parasite. There is little empirical evidence for apparent competition (Bonsall & Hassell 1998; Prenter et al. 2004), but it could be an important process driving declines of native species after the arrival of an invader and structuring post-invasion communities. As such, studies on this subtle interspecific interaction should be a research priority.
Cannibalism and IGP commonly occur between closely related species, even though enhanced transmission of parasites through ingested material should select against such interactions (Polis, Myers & Holt 1989). Parasites have been shown to mediate predation between natives and invaders at the same trophic level. For example, the microsporidian parasite Pleistophora mulleri increases the vulnerability of a native amphipod to predation by a large invasive alien amphipod which is not susceptible to the parasite (MacNeil, Dick, Hatcher et al., 2003). It is possible that D. coccinellae could transmit microsporidia from infected to uninfected coccinellid hosts, but there is no information on such interactions. Indeed, there is very little known about microsporidia infection in coccinellids even though microsporidia are known to be relatively common pathogens of predacious coccinellids. Again studies on such interactions are lacking but are worthy of additional attention.
Parasites can manipulate and mediate interspecific and intraspecific interactions both directly and indirectly through competition and predation. The subtle forces that parasites exert in structuring communities are undoubtedly challenging to study but could be pivotal in facilitating invasion processes. Very few studies extend beyond one parasite – two host systems (Prenter et al. 2004) and most are restricted to relative abundance and diversity of parasites in native compared to invading populations (Colautti et al. 2004) with a few assessing the catastrophic outbreak of disease in native populations through spillover from an invading population. In all cases, studies on insect communities are lacking. Insects could provide model systems in which to study the subtle (sublethal and indirect interactions) effects of parasites in driving community change after invasion. Further research should tackle parasite–host interactions at large spatial and temporal scales acknowledging the complex and dynamic nature of the relationships. There are clearly many advantages of adopting a community ecology approach, and particularly ENA, for revealing dominant mechanisms in determining the success of invasion. New molecular techniques coupled with modelling tools will undoubtedly support the process of moving towards a holistic framework for unravelling interactions between invaders and their parasites.
We thank Alison Dunn (University of Leeds) and Sarah Perkins (University of Cardiff) for kindly inviting us to contribute to this special issue of Functional Ecology. HER is funded through the NERC Centre for Ecology & Hydrology and the Joint Nature Conservation Committee. LLH's contribution to this work was supported by a Royal Society Research grant. We are grateful to Darren Evans for providing and allowing reproduction of Fig. 2.