Strongly interactive carnivore species: maintaining and restoring ecosystem function.

OF AUSTRALIACARNIVORESP A S T , P R E S E N T A N D F U T U R E

E D I T O R S : A . S . G L E N A N D C . R . D I C K M A N

041403 Carnivores of Australia 2pp.indd 1 4/07/2014 12:14 am

© A. S. Glen and C. R. Dickman 2014

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Carnivores of Australia: past, present and future/edited by A.S. Glen and C.R. Dickman.

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Carnivora – Australia. Carnivorous animals – Australia. Predation (Biology) – Australia.

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13Strongly interactive carnivore species:

maintaining and restoring ecosystem function

Chris R. Dickman, Alistair S. Glen, Menna E. Jones, Michael E. Soulé, Euan G. Ritchie and Arian D. Wallach

SummaryPredators can have dramatic and lethal effects on individual prey, but they can also have subtle yet powerful effects on non-prey species via webs of indirect interactions. Top predators may, for exam-ple, suppress the activity of smaller predators and in turn provide a net benefit for the prey of the smaller predators; they can also reduce the impacts of herbivores and thus indirectly alter vegetation dynamics. Species that have such pervasive effects on their communities, and the broader ecosystems to which they belong, are termed ‘strongly interac-tive’. Here, we begin by reviewing the kinds of effects that theoretically can be engendered by the presence of strongly interactive carnivores, and then present examples of such species among the native reptiles, birds and mammals of Australia. The examples include elapid snakes, varanid (mon-itor) lizards, day-active raptors and owls, dasyurid marsupials and the dingo. The dingo, in particular, has been shown in many studies to suppress the activity of smaller mesopredators and herbivores and to have broadly beneficial effects on biodiver-sity and ecosystem function. Using the dingo as a case study, we propose that this important but

persecuted species should be maintained in areas where it still occurs and that immediate considera-tion should be given to reintroducing it to areas from which it has been banished. We conclude that strongly interactive carnivores are key components of many ecosystems and should be retained where they still occur and reintroduced, where possible, elsewhere.

IntroductionMost carnivores kill for a living and thus have the potential to affect populations of their prey. These effects vary with the type of carnivore, the compo-nents of the prey population that are killed, and the relative densities of the interactants (Banks 1999; Cortez 2011). In Australia, for example, native mar-supial predators often have little evident effect on populations of prey with which they have co-evolved (Glen and Dickman 2005), whereas the impacts of their introduced counterparts can be catastrophic (Salo et al. 2007; see Chapter 5). The impacts of red foxes (Vulpes vulpes) and feral cats (Felis catus) are particularly damaging in Austral-ian ecosystems, and perhaps reflect the abundance


CARNIVORES OF AUSTRALIA302

of these carnivores in many habitats as well as the evolutionary naiveté of native prey to the hunting tactics of the newcomers (Banks and Dickman 2007; see Chapter 17).

The direct effects of carnivores are often con-spicuous, but a large body of recent research sug-gests that predators exert other effects that are just as powerful as those of direct killing, but operate more subtly and pervasively on both prey and the communities to which they belong. This shift in our understanding of carnivore ecology has two main components.

First, carnivores often drive shifts in prey activ-ity, habitat and resource use simply by their pres-ence (Lima and Steury 2005; Müller-Schwarze 2006). They can do so by increasing apprehension in prey individuals and reducing their fitness by suppressing both growth and reproduction (Hik 1995). This ‘ecology of fear’ can have a dramatic impact on the local distributions and abundances of both predators and prey, and can affect the long-term stability of their dynamics (Brown et al. 1999; Berger 2010). Second, carnivores frequently have wide-ranging but subtle effects on non-prey spe-cies via indirect interactions. For example, large carnivores may suppress populations of smaller predators and thereby benefit the prey species of those smaller predators (Crooks and Soulé 1999; Gittleman and Gompper 2005); they may benefit plant species if they suppress populations of herbi-vores (Beyer et al. 2007; Ripple and Beschta 2007). Large carnivores may also entrain community-level trophic and ecological cascades that exten-sively redistribute plant biomass and alter the composition of plant species (Polis et al. 2000; Schmitz et al. 2000; Vanak et al. 2014), thus having profound effects on the broader community. Carni-vores sometimes also exert their effects via trans-mission of disease to both prey and non-prey species; the spread of toxoplasmosis by feral cats into Australian native fauna is one example (Dick-man 1996a).

Systems with two or more species of carnivores provide opportunities for intraguild predation (Holt and Polis 1997), extreme interference compe-

tition (Moseby et al. 2012) and other, more complex effects. For example, different predators may domi-nate at different times or in different habitats and enhance or depress community diversity depend-ing on how they interact (Schmitz 2007). Alterna-tively, a dominant predator may alter the balance of advantages within communities of smaller species of predators via indirect amensal and commensal pathways (Glen and Dickman 2005, 2008), resulting in differential effects on smaller prey species (Elm-hagen et al. 2010). The removal or disappearance of the dominant or largest predator may lead, con-versely, to an increase in the abundance of smaller ones – mesopredator release (Soulé et al. 1988; Brashares et al. 2010).

Taken together, these observations suggest that the ecological effects of carnivores can ripple through the communities, and the broader ecosys-tems, to which they belong. Using as examples the sea otter (Enhydra lutris) and the wolf (Canis lupus lupus) in North America, Soulé et al. (2003) pro-posed that such species could be considered strongly interactive. The absence or removal of such species will lead to significant changes in some features of the ecosystems to which they belong, potentially including structural or compo-sitional modifications in habitat, loss of resilience, or declines in native species diversity. Although not stated explicitly by Soulé et al. (2003), the con-cept of strongly interactive species was intended to apply primarily to native or ecologically beneficial species that might be of conservation concern, rather than to introduced organisms. Thus, while red foxes could be considered strongly interactive in the sense that they affect populations of many prey species, in Australia the mostly negative nature of these effects makes the fox a target for control rather than for conservation. Soulé et al. (2005) proposed further that strongly interactive species be prioritised by managers to ensure that their populations remain at ecologically effective levels. Recent reviews by Soulé (2010), Estes et al. (2011), Ritchie et al. (2012) and Ripple et al. (2014) have emphasised the importance of top predators as strongly interactive species, and lament that the


Chris

Chris

et al. should be in italic.

30313 – STRONGLY INTERACTIVE CARNIVORE SPECIES: MAINTAINING AND RESTORING ECOSYSTEM FUNCTION

demise of these animals in many systems is lead-ing to hitherto unanticipated effects on ecological communities. These effects range from changes in species composition to shifts in disease dynamics, carbon sequestration, occurrence of wildfires, and even biogeochemical cycles.

In this chapter we first outline some ways by which strongly interactive carnivore species exert their effects on other community members, and then identify and describe the community-level effects of exemplar species of reptiles, birds and mammals. Finally, using the dingo (Canis dingo) as a case study, we ask whether the reintroduction of a strongly interactive carnivore can help restore ecological functioning in systems that have experi-enced trophic downgrading (sensu Estes et al. 2011) following the loss of the species at an earlier time.

Types of interactionsDirect interactions among carnivores, such as intraguild killing, intraguild predation and inter-ference competition have been described in detail elsewhere (see Chapters 1 and 12) as have direct interactions between predators and prey (see Chapters 5 and 8). Here, we focus on interactions that involve strongly interactive carnivores and two or more species – in the same trophic level or lower – that co-occur with them. In doing so, we are less concerned with carnivores that hunt and eat one or more different prey species in multi-species prey communities (and thus exhibit multi-ple pairwise predator–prey interactions) than with carnivores that induce suites of less obvious inter-actions via their predatory behaviour.

In general, if a species affects the interaction between two others, its effect can be considered indirect (Dickman 2006). If that species is a carni-vore, the indirect effects that it engenders can take many forms; we depict some of the simpler interac-tions involving three or four species in Fig. 13.1.

In a situation involving two carnivores and a shared prey species, depletion of prey numbers by one carnivore inevitably reduces the prey base for the other, and vice versa (Fig. 13.1a). The two

carnivores may seldom encounter each other, but both may nonetheless suffer reduced growth, sur-vival or reproduction because of the deficiency in the shared resource they have caused. This situa-tion – exploitation competition – is probably not common among carnivores, but may occur if the two species’ modes of hunting are so different that they seldom encounter each other directly. For example, during an outbreak of native rodents in central Australia, Pavey et al. (2008) showed that an avian predator, the letter-winged kite (Elanus scrip-tus), and three species of mammalian predators hunted the rodents almost exclusively. Direct encounters between the kites and the mammalian predators were not observed, but the mutual deple-tion in abundance of the shared rodent prey most likely reduced the population growth rates of all species, especially as rodent numbers declined from their peak.

In general, predatory species exhibit aggressive behaviours and may fight over shared resources even if those resources are not in short supply (see Chapter 12). This interaction – interference compe-tition – may result in the dominant carnivore restricting the activity time or the habitats used by the subordinate species (Creel et al. 2001; Brook et al. 2012), or even in death if the subordinate carni-vore cannot get away (Moseby et al. 2012).

If a carnivore hunts two or more prey species that do not interact, it may lead to a situation termed ‘apparent competition’ (Holt 1977) where the prey species appear to compete with each other (Fig. 13.1b). Here, the presence of several prey spe-cies elevates predator abundance to levels above those that could be sustained by a single prey spe-cies. The increased predation pressure keeps the prey populations at low levels, providing the erro-neous impression that they are competing. Other forms of apparent competition have been recog-nised that occur within a single trophic level (Fig. 13.1c), but these probably occur rarely in nature (Connell 1990).

More common types of interaction involving carnivores are depicted in Figs 13.1d and 13.1e. In Fig. 13.1d, a dominant carnivore suppresses the


Chris

Add a comma after the brackets: "... (see Chapters 1 and 12), .."


c) Apparent compe!!on d) Indirect commensalism

g) Keystone preda!on

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P2P1

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Fig. 13.1. Indirect interactions involving at least one strongly interactive carnivore species (C) and one or more prey species (P). Direct effects between species are shown by solid arrows and indirect effects by broken arrows. Arrow heads show the species affected, and the +, – and 0 symbols show the direction of the effect. Panel (a) depicts exploitation competition, (b) apparent competition involving one carnivore and two prey species, (c) apparent competition involving three species in the same trophic level (carnivores are shown, but other consumer species could be shown in the same way), (d) indirect commensalism between a dominant carnivore and prey (C represents the dominant carnivore in this interaction, and c depicts the subordinate), (e) a trophic cascade (H represents a herbivore and V represents vegetation), (f) indirect mutualism between carnivores, and (g) keystone predation (C again represents the dominant carnivore in this interaction, and c1 and c2 the two subordinate carnivores). Mesopredator release is not shown specifically, but can be thought of as a special kind of trophic cascade where c would replace H and P would replace V. The negative effect of C on c in this situation could arise from direct killing or fear effects, when the term ‘ecological cascade’ should be used. Adapted and redrawn from Glen and Dickman (2005) and Dickman (2006).



abundance of a subordinate carnivore via intragu-ild killing or reduces its activity by instilling fear; the reciprocal effect is small or negligible. If the dominant carnivore has little or no negative effect on a third species in the system which would oth-erwise be suppressed by the activity of the subor-dinate carnivore, the third species derives an indirect commensal benefit from the interaction (Glen and Dickman 2005). In Fig. 13.1e, which depicts a trophic cascade (Hairston et al. 1960), the carnivore depresses populations of herbivores, in turn indirectly benefiting primary producers in the system. More complex effects can arise when four

or more species interact (e.g. Figs 13.1f and 13.1g). It is often possible to identify key relationships between pairs of species, but predicting all the direct and indirect effects that occur becomes increasingly difficult in diverse multi-species assemblages. Webs of potential interactions can be mapped (e.g. Fig. 13.2), but confirming effects and quantifying interaction strengths have seldom been attempted.

Before discussing examples of strongly interac-tive species, we need to add a caveat: interactions occur between individuals, but the effects of these interactions are of most interest when they are

Fig. 13.2. Interaction webs involving dingoes, red foxes, feral cats and other species in different bioclimatic environments throughout Australia. The web at top left depicts potential interactions in arid environments (low rainfall, high potential evapo-transpiration, low primary productivity; dominated by cattle grazing), that at top right depicts interactions in tropical northern environments (> 600 mm rainfall per annum), that at bottom left depicts interactions in semi-arid rangelands (low rainfall, high potential evapo-transpiration, moderate primary productivity; dominated by sheep grazing), and the web at bottom right depicts interactions in temperate forest. Only some of the possible interactions are shown, based on previous research (dashed arrows), current research (solid arrows) and work that might be undertaken in future (dotted arrows). For simplicity the major interactants are shown to be mammals; more realistic but complex webs would include goannas, snakes and birds of prey. Figure from Visser et al. (2009).



manifest at the population level. For example, a predatory interaction occurs when one animal kills and eats another, but this event may have no numerical influence on the population of the prey species if non-reproductive individuals are hunted, if the harvesting rate is low, or if compensatory survival, immigration or reproduction occurs. The interaction becomes of interest and, for threatened prey species, of conservation relevance if the prey population is depressed by the effect of the preda-tor. This caveat has important practical conse-quences. Whereas killing or predation is easy to confirm from direct observation or from dietary analysis, and competition may be suspected if dif-ferent species are known to use the same resources, the effects of these interactions can be quantified at the population level only by careful observations or – ideally – by planned manipulations of the spe-cies that appear to drive the interactions. In most of the examples below, therefore, we focus on studies that document changes in the abundance of a strongly interactive species and resultant shifts that are triggered in the population performance of two or more species in the community.

Strong interactorsSome of the best-known examples of strongly inter-active species come from marine systems (e.g. Paine 1966; Estes et al. 1998). This may reflect in part the presence of many species of large preda-tors in marine waters, but also the fact that long-term records that document both predator population declines and the trophic consequences are available from commercial fisheries. For exam-ple, numbers of the blacktip shark (Carcharhinus limbatus) and other predators have declined sharply off the east coast of the United States since 1970, and this has released populations of smaller preda-tors such as the cownose ray (Rhinoptera bonasus) (Myers et al. 2007). As populations of these smaller predators have increased, their own prey species have been subjected to elevated levels of predation. Some of these species, such as the economically and socially significant Atlantic bay scallop

(Agropecten irradians), have suffered population declines or even local extinction as a result (Myers et al. 2007; Heithaus et al. 2008). Data sets of compa-rable duration and quality are scant or non-existent in terrestrial environments; nonetheless, a compel-ling case can be made that several land-dwelling carnivores in Australia are strongly interactive in their respective systems.

ReptilesSnakes and large varanid (monitor) lizards are highly effective and ubiquitous predators in Aus-tralia and provide several examples of species that can be considered strongly interactive. Several studies have monitored events that have caused changes in their populations and, consequently, in populations of their prey.

In one example, Read et al. (2011) reported an attempt to reintroduce the threatened woma python (Aspidites ramsayi) to a reserve in an arid region of South Australia. The reserve had been cleared of large mammalian predators that might otherwise have preyed upon the woma, but this intervention appeared to allow populations of varanids and the venomous mulga snake (Pseudechis australis) to increase. The woma reintroduction failed, probably because the released pythons were killed by mulga snakes. Read et al. (2011) speculated that mulga snakes might suppress populations of other snake species, and in doing so potentially release multiple prey species from snake predation. It is possible that other species of abundant elapid snakes, such as the eastern brown snake (Pseudonaja textilis), exert strong effects on their communities via the consumption of common prey species (Whitaker and Shine 2003; Hughes et al. 2010), but experimen-tal confirmation is wanting.

Varanid lizards provide several compelling examples of species that have strong and pervasive effects on their communities. Across tropical north-ern Australia, the arrival of the cane toad (Rhinella marina) in recent years has triggered wide-ranging ecological changes (Shine 2010), among them a dra-matic decline in the yellow-spotted monitor (Vara-nus panoptes). Like several other native carnivores,



this large (2 m) varanid is highly susceptible to bufonid poison and can be killed if it eats, or even ‘mouths’, a single cane toad. The yellow-spotted monitor is an active and wide-ranging hunter that tracks and eats other vertebrates, especially reptiles and their eggs. Following the arrival of cane toads at sites along the Daly River, Northern Territory, in 2004, numbers of this varanid fell by 77–92% and depredation of its main prey species also declined (Doody et al. 2006). The reduction in predation allowed a 20% increase in the survival of eggs of the pig-nosed turtle (Carettochelys insculpta), and a probable but unquantified increase in egg survival of the flatback sea turtle (Natator depressus) at other sites (Blamires 2004; Doody et al. 2006). Increased turtle survival could lead in turn to increased impacts on the plant and animal prey that these species consume, hence resulting in trophic cascades.

Similar collapses in yellow-spotted monitor populations on the Adelaide River floodplain, Northern Territory, most likely facilitated increases in populations of several snake species, especially the frog-eating keelback (Tropidonophis mairii) (Brown et al. 2011). Annual survival rates of both wild-hatched and laboratory-reared keelbacks released into the wild increased dramatically fol-lowing the demise of the varanid (Brown et al. 2013), potentially allowing concomitantly large increases in predation by the snake on native frogs. Although cane toads could have direct effects on snakes, both the Adelaide River studies were care-ful to emphasise the primacy of indirect effects in driving the snake responses.

In more-arid regions of Australia, the sand goanna (V. gouldii) is the most conspicuous large monitor. Following poison baiting of the red fox in semi-arid shrubland habitats in western New South Wales, Olsson et al. (2005) reported fivefold increases in goanna density compared to non-baited sites, and also recorded concomitant changes in the activity of smaller lizards. In the baited areas, capture rates of day-active skinks increased whereas those of nocturnal geckoes declined. Olsson et al. (2005) interpreted the increased skink

activity to be a consequence of reduced predation by foxes, but suggested that decreases in gecko activity arose from increased goanna predation. Intriguingly, the species richness of geckoes was 80% higher in areas where goannas were abundant compared with where they were suppressed by foxes, despite the control and treatment areas being matched in terms of habitat, fire and grazing his-tory (Olsson et al. 2005). This suggests a keystone effect driven by predation from the goanna. Pianka (1994) also proposed that the sand goanna acts as a keystone predator in the Great Victoria Desert, Western Australia, while Sutherland et al. (2011) discussed cascading effects that could be expected in faunal communities where sand goannas and other monitors were allowed to achieve ecologi-cally effective densities (see also Chapter 11).

Another large reptilian carnivore, the saltwater crocodile (Crocodylus porosus), can also have a strong influence on its prey. For example, agile wal-labies (Macropus agilis) alter their behaviour and habitat use to avoid riverbanks where they are at risk of predation by crocodiles (Doody et al. 2007), potentially reducing grazing and trampling effects on riparian vegetation.

BirdsAustralia has some formidable avian predators capable of exerting strong impacts on their prey and on other carnivores. For example, powerful owls (Ninox strenua) can drive populations of favoured prey such as greater gliders (Petauroides volans) to very low abundance (Kavanagh 1988), and may therefore compete strongly with arboreal predators such as quolls (Glen and Dickman 2011; see Chapter 12). Large raptors can also have direct effects on other carnivores through intraguild pre-dation, with potential for flow-on effects on prey. For example, raptors in Australia have been recorded to eat quolls, foxes, cats and varanid liz-ards, sometimes in large numbers (Aumann 2001; Bilney et al. 2006; Olsen et al. 2010). In Tasmania, masked owls (Tyto novaehollandiae castanops) are effective predators of eastern quolls (Dasyurus viverrinus) (Mooney 1993); both adult males and



recently weaned subadults show anti-predator responses when vocalisations of the masked owl are played (Jones et al. 2004).

MammalsAmong native Australian mammals, dasyurid marsupials and the dingo can be considered strongly interactive. The recently extinct thylacine (Thylacinus cynocephalus) probably also had wide-ranging effects on smaller mammalian predators and prey, but reconstructing these a century or so after the species became functionally extinct is challenging. We do not consider this species fur-ther here, but Chapter 4 discusses some aspects of predatory behaviour of the thylacine and its inter-actions with the dingo in the mid to late Holocene. It is possible that many of the smaller insectivorous dasyurids (< 100 g) also have marked effects on invertebrate assemblages. However, apart from dietary studies that document what these species eat, it remains unclear how strong or pervasive these effects may be (see Chapter 10). Larger dasy-urids have more obvious effects on the communi-ties to which they belong. The brush-tailed mulgara (Dasycercus blythi) can be considered a keystone predator in arid regions (see Chapter 10; Dickman 2006), and in more temperate regions both the quolls (Dasyurus spp.) and the Tasmanian devil (Sarcophilus harrisii) may be strongly interactive.

Devils are now apex predators in Tasmania, fol-lowing the extinction of the thylacine, and perform several important ecological functions. As special-ised bone-cracking scavengers, they remove car-rion, limiting scavenging opportunities for other carnivores. Devils are also dominant in aggressive encounters with smaller predators (Jones and Bar-muta 1998). Both devils and eastern quolls compete with spotted-tailed quolls (D. maculatus), which use different habitats and hunt more arboreal prey than their competitors (Jones and Barmuta 1998, 2000). Competition in the Tasmanian guild has occurred over many generations, with character displacement evident in the trophic structures needed for killing prey in spotted-tailed quolls that co-occur with devils and (previously) thylacines in

Tasmania, but not in quolls on Australia’s south-eastern mainland (Jones 1997). With devils and eastern quolls now in severe decline (McCallum and Jones 2006; Fancourt et al. 2013; see Chapter 9), it is likely that spotted-tailed quolls will change in abundance or behaviour, with unknown effects on species at lower levels in the food web.

There is evidence for a keystone function of devils in suppressing populations of feral preda-tors in Tasmania (Lazenby and Dickman 2013), and in trophic cascades that influence populations of smaller mesopredators, as well as in disease ecology. Sightings of feral cats have increased in forested areas of Tasmania where disease-induced devil decline has occurred; the severe and sus-tained declines in eastern quoll populations follow these changes (Hollings et al. 2014). Where devil density is high, cat density is low, and seropreva-lence of the protistan parasite Toxoplasma gondii in the Tasmanian pademelon (Thylogale billardierii), a species susceptible to toxoplasmosis, is high (Holl-ings et al. 2013). The capacity of devils to kill smaller competitors may have helped, until recently, to prevent establishment of foxes in Tas-mania (see Chapter 9). In particular, with few foxes and many devils, devils may have been able to limit fox reproduction through intraguild kill-ing of dependent young at dens (see Chapter 9). If that is the case, the timing of the devil’s current decline, coinciding with the recent incursion of foxes, may be catastrophic.

In contrast to the very long tenure of dasyurids in the Australian region (Wroe et al. 2004), the dingo is a relative newcomer that was brought to the continent by Asian seafarers at least 3500 years ago (Corbett 2006), perhaps earlier (Oskarsson et al. 2012). This relatively recent arrival has generated debate about whether it is a ‘native’ species (Wood Jones 1925; Dickman and Lunney 2001). Here, how-ever, we take the view that the dingo is native: it is distinct from canids elsewhere and can be distin-guished on genetic, morphological and behavioural criteria from domestic dogs that have become feral (Elledge et al. 2008; Déaux and Clarke 2013; New-some et al. 2013a).


Chris

Insert 'several' between 'on' and 'genetic', and also delete '... that have become feral.'

Chris


The dingo may have contributed to the demise of the thylacine, the Tasmanian devil and the flight-less native hen (Gallinula mortierii) after its arrival on the Australian mainland (Corbett 2006; see Chapter 4), although recent models indicate the importance of increased human influence and den-sity in the landscape (Prowse et al. 2014). Some native animals that are apparently naïve to preda-tion by newly introduced predators display appro-priate avoidance behaviour towards dingoes, suggesting perhaps that dingoes have been present long enough for native prey to have adapted (Pur-cell 2010; Carthey and Banks 2012). Owing to its harassment and depredation of livestock, the dingo is controlled intensively over much of the continent by poison baiting; boundary fences are used as an additional control measure to exclude dingoes from large areas of south-eastern and south-west-ern Australia. Where uncontrolled, dingoes remain common and ubiquitous across all terrestrial habi-tats and may be locally abundant where prey is easy to access (Letnic and Dickman 2006) or at sites where they have access to supplementary food (Newsome et al. 2013b).

The dingo has powerful and pervasive effects on suites of other species, and hence can be consid-ered with much justification to be strongly interac-tive. In the first instance, dingoes hunt and kill a range of medium to large (4–100 kg) herbivores and the young of even larger prey (Letnic et al. 2012). This depredation reduces the numbers and impacts of herbivores on plants, thus altering patterns of primary productivity, plant species composition and the amount of structural cover that is available to other biota (Letnic et al. 2009). The direct and strongly suppressive effects of dingoes on large kangaroos have been suspected for some time (e.g. Krefft 1871; Lucas and Le Souëf 1909). More recent reports describe inverse numerical associations between this carnivore and larger Macropus spp. and support the view that dingoes effectively keep kangaroo numbers in check (Caughley et al. 1980; Corbett and Newsome 1987; Letnic and Crowther 2013). Packs of dingoes can chase and kill even the largest kangaroos (Thomson 1992), but in some

areas juveniles are taken preferentially (Robert-shaw and Harden 1986). Indeed, the pack structure of dingoes may be integral to their role as a strongly interactive predator (Wallach et al. 2010). Sociality not only allows dingoes to take larger prey than they could do alone, but may also confer a competi-tive advantage over solitary predators (Corbett 1995a; see Chapter 4).

A large body of circumstantial evidence sug-gests that dingoes also suppress the numbers and occupancy of large introduced herbivores over vast areas of Australia. In arid regions, for example, feral goats (Capra hircus), wild donkeys (Equus asinus) and feral pigs (Sus scrofa) show inverse spa-tial relationships with the dingo and persist locally only where dingoes are sparse or controlled (New-some 1990; Wilson et al. 1992; Wallach et al. 2010). In tropical regions there is some evidence that din-goes keep populations of feral pigs in check (Cor-bett 1995b), and analyses of bounty records from Queensland indicate a reciprocal numerical rela-tionship between the two species over a period of more than 30 years (Woodall 1983). Although the suppressive effects of dingoes on feral pest species are generally welcome, the dingo can pose a direct threat to some endangered species with very small populations. For example, a remnant population of the threatened northern hairy-nosed wombat (Lasiorhinus krefftii) is thought to be at risk of dingo predation (Tisdell and Nantha 2007). This species was probably driven to low numbers by early losses of habitat and persecution from settlers, so that any level of predation now may be unsustainable.

Inverse numerical or spatial relationships between dingoes and large herbivores certainly suggest top-down suppression, but do not exclude other potential (albeit improbable) explanations for the observed patterns. Comparisons of kangaroo numbers on either side of the dingo fence in south-eastern Australia provide strong evidence for sup-pression. This fence, constructed to exclude dingoes from the sheep rangelands of the south-east, sepa-rates essentially dingo-free habitat ‘inside’ the fence from habitat ‘outside’ the fence where din-goes are controlled only sporadically (Fleming et al.



2001). Aerial surveys by Caughley et al. (1980) showed that densities of the red kangaroo (Macro-pus rufus) were up to 166-fold greater inside the fence than outside; those of the emu (Dromaius novaehollandiae) also were up to 26-fold higher inside than outside the fence. Caughley et al. (1980) examined several hypotheses that could poten-tially account for these disparate densities and con-cluded that the difference in dingo predation provided the only plausible explanation. Although some authors have questioned the generality of this conclusion (Dawson 1995; Newsome et al. 2001), the pattern has been robust over a range of scales and over many different sampling years (Caughley and Grigg 1982; Pople et al. 2000; Letnic et al. 2009, 2012).

The suppression of kangaroo populations by dingoes predictably ‘releases’ vegetation from her-bivory and allows greater coverage of both grasses (Letnic et al. 2009) and other plants (Wallach et al. 2010). These results have been replicated in studies where kangaroo populations were reduced by other means, such as culling or exclusion fencing (Morgan and Pegler 2010; Dickman et al. 2014), and support the idea that dingoes engender trophic cascades (i.e. they are a keystone species). Similar findings have been obtained following the removal of top carni-vores in other systems (e.g. Fortin et al. 2005; Frank 2008). It may be that predation by dingoes is not the sole explanation; dingoes may also reduce the activ-ity or range of habitats used by herbivores simply by their presence. Thus, they could maintain their cascading effects by behavioural means in addition to direct predation (Caughley 1964; Parsons and Blumstein 2010). Interactive models and empirical data suggest further that the strength of trophic cascades varies with rainfall, dingo numbers and dingo social stability (Wallach et al. 2010; Choque-not and Forsyth 2013; Letnic and Crowther 2013). Irrespective of the interaction pathway, and of vari-ation in the magnitude of their effects, it is reasona-ble to conclude that dingoes will often reduce grazing pressure by herbivores and hence facilitate increased vegetation cover. More vegetation may in turn lead to further effects on species that either use

plant cover or prefer more open areas, although these indirect effects have yet to be quantified. Similarly, in the United States coyotes may suppress jackrabbits (Lepus spp.), resulting in increased veg-etation cover, which benefits sage-grouse (Centro-cercus urophasianus) (Mezquida et al. 2006). The capacity of dingoes to suppress the abundance of large herbivores may also be beneficial to primary producers. Overabundant herbivores can be extremely damaging not only for biodiversity, but also for enterprises such as cropping, grazing and forestry (e.g. Biedenweg et al. 2011).

An increasing body of evidence shows that din-goes, in addition to having negative effects on large herbivores, can also suppress the numbers and activity of non-native mesopredators – especially the red fox and feral cat. The impacts of foxes and cats on native fauna in Australia have been much studied and are known to be dramatically negative at the population level for a wide range of native reptiles, birds and mammals (e.g. Kinnear et al. 1988, 1998, 2010; Dickman 1996a, b; Smith and Quin 1996; Jones et al. 2003; Saunders et al. 2010). If din-goes have little direct effect on these relatively small prey species but suppress mesopredators, they could potentially provide net benefits for many native species (Purcell 2010). This possibility, and the evidence that supports it, has stimulated much interest and underpins recent calls to rein-troduce the dingo to areas from which it is cur-rently excluded (Dickman et al. 2009).

Dingoes include foxes and cats in their diet, but generally at very low frequency (Newsome and Coman 1989; Cupples et al. 2011). However, this may not provide an accurate measure of the effect of the dingo for several reasons: foxes and cats may be killed on encounter by dingoes but not eaten (O’Neill 2002; Moseby et al. 2012); the smaller carni-vores overlap to some degree with dingoes in their consumption of small to medium-sized prey spe-cies (Cupples et al. 2011; Glen et al. 2011); and they may be restricted in numbers, habitat occupancy and time of activity in the presence of the larger carnivore (Wallach et al. 2010; Brawata and Neeman 2011; Brook et al. 2012; Wang and Fisher 2012). In



other words, suppression of fox and cat popula-tions may occur via intraguild predation, intragu-ild killing (an extreme form of interference competition), via exploitation competition, and via reduced access to essential resources because of fear of encounter with dingoes. These kinds of interactions occur commonly between dominant and subordinate carnivores in other systems (Crooks and Soulé 1999; Ritchie and Johnson 2009), and can result in dramatically curtailed meso-predator activity. Not all studies agree; Allen et al. (2011, 2013) argue that dingoes do not have demon-strably suppressive effects on mesopredators, although the interpretations of these authors are disputed (Letnic et al. 2011a).

The population-level effects of dingoes are more obvious on foxes than on feral cats, as demonstrated by reduced fox abundance or activity in the pres-ence of dingoes at local, regional and continental scales (Mitchell and Banks 2005; Johnson and Van-DerWal 2009; Letnic et al. 2011b, 2012 and references therein). There is some evidence that these effects vary between habitats. For example, suppression of foxes by dingoes appears to be stronger in open, arid areas than in some forest habitats (Catling and Burt 1995; Letnic et al. 2009), and perhaps reflects the reduced opportunity for foxes to escape or avoid detection where cover is reduced. Suppressive effects of dingoes on feral cats have been reported in some studies (e.g. Wallach et al. 2010; Brook et al. 2012; Kennedy et al. 2012) but not others (Catling and Burt 1995; Letnic et al. 2009). It is possible that cats can avoid moment-to-moment encounters with dingoes at ground level by climbing trees, rocks or other structures. Certainly, all reports of negative associations between dingoes and feral cats have been made in relatively open habitats, but the mechanisms of interaction between these carni-vores remain to be fully elucidated.

In contrast to their suppressive effects on medium-sized to large herbivores and on meso-predators, dingoes generally have limited direct effects at the population level on small prey (< 4 kg) and on very large prey (> 100 kg). Dingoes include small prey in their diet, sometimes at high frequency

(e.g. Newsome and Corbett 1975; Cupples et al. 2011; Glen et al. 2011; Newsome et al. 2014). However, in many situations small prey derive indirect benefit from the presence of dingoes because of the sup-pressive effects of the dingo on mesopredators. That is, even though dingoes may eat some small prey, the rate of increase of the prey exceeds the sum of the per capita rate of predation by both the dingo and the mesopredator (Letnic et al. 2012). Very large herbivores (> 100 kg) may also be hunted by dingoes (e.g. Thomson 1992), but attacks are relatively infre-quent. Although dingoes may affect very large prey through fear effects, this possibility remains to be demonstrated, and there is currently no evidence for effects of predation on very large herbivores at the population level (Freeland and Choquenot 1990; Purcell 2010; Brook and Kutt 2011).

These findings suggest that the effects of din-goes on prey populations scale with prey body size. Small and medium-sized prey (< 4 kg) may benefit via the suppressive effects of dingoes on mesopredators; populations of larger species (4– 100 kg) such as foxes and kangaroos are often suppressed; and very large species (> 100 kg) are probably unaffected by dingo presence. Letnic et al. (2012) integrated these effects into a simple concep-tual model to depict the shifting effects of dingoes on other mammals of varying size (Fig. 13.3). Dif-ferent effect curves were posited to accommodate shifts in the effects of dingoes under conditions of high and low environmental resources; thus, the suppressive effects of dingoes on mesopredators should be greater during periods of resource short-age than resource abundance owing to the greater intensity of predatory and interference interactions that can be expected when resources are scarce (Fig. 13.3). The conceptual model is underpinned by a substantial body of empirical and theoretical data (Letnic et al. 2012).

Strong interactors: maintaining and restoring ecosystem functionTop predators are often controlled and maintained at low numbers, and have been exterminated in


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many parts of the world (Ripple et al. 2014). How-ever, recognition of their pervasive and often ben-eficial effects on species diversity and ecological function has led to calls for them to be reintroduced where it is possible to do so (Hayward and Somers 2009; Ritchie et al. 2012). Recently, Ripple et al. (2014) called for the formation of a Global Large Carni-vore Initiative to coordinate local, national, and international research, conservation, and policy, and promote coexistence with people. They sug-gested that such an effort be modelled in part on the Large Carnivore Initiative for Europe, a non-profit scientific group affiliated with the Interna-tional Union for the Conservation of Nature.

Globally, seven large species of Carnivora are known to function as highly interactive top-order predators: the African lion, leopard, Eurasian lynx, sea otter, mountain lion, grey wolf and the dingo (Ripple et al. 2014). For example, the flagship pro-gram that has elucidated the beneficial ecological effects of a top predator has been the return of the wolf to the Yellowstone region in the northern United States. Here, wolves were reintroduced in 1995 after an absence of several decades (Smith et al. 1999). Their return has caused a dramatic recovery of Yellowstone’s degraded ecosystems. Large herbi-vores, which were previously overabundant, have been reduced by wolf predation. It has also been suggested that remaining grazers have had to modify their foraging behaviour to avoid areas of high predation risk (Ripple and Beschta 2003), although this has been questioned by Kauffman et al. (2010, 2013). Although it is debatable whether the driving force for ecosystem change has been changes in abundance, behaviour and/or distribution of grazers (Beschta and Ripple 2013; Kauffman et al. 2013), the reintroduction of wolves has allowed Yel-lowstone’s vegetation to recover. Trees such as aspen and cottonwood (Populus spp.), which had not estab-lished new seedlings in many decades, have regen-erated in areas where they have been released from overgrazing (Ripple and Beschta 2003, 2007). Wolves have also influenced the behaviour of smaller preda-tors, such as the coyote (C. latrans), with cascading benefits for smaller prey animals (Switalski 2003; Ripple et al. 2013). Cascading effects of re-establish-ing wolves have been observed in other areas, such as Banff National Park in Canada (Hebblewhite et al. 2005) and Grand Teton National Park, United States (Miller et al. 2012).

Case study – reintroducing the dingoDingoes are controlled in most areas of Australia where livestock – especially sheep – are run, with 1080 baiting, targeted shooting and trapping, and in the continent’s south-east by use of an exclusion fence. The fence runs for over 5500 km and keeps dingoes from entering the mostly sheep-grazing lands of western New South Wales and adjacent

15 60 3000 35000 150000

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ct s

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Negatively affected by dingoes

Positively affected by dingoes

Fig. 13.3. Conceptual model showing the population-level effects of dingoes on other species of mammals of varying body mass and under different environmental conditions. In general, small and medium-sized mammals are affected positively by dingoes owing to the suppressive effects of dingoes on mesopredators (red foxes and feral cats) that would otherwise depredate them. Larger mammals fall prey to dingoes and thus are affected negatively by dingo presence, but very large mammals seldom fall prey and are relatively unaffected by dingoes. These effects are moderated by environmental conditions. When resources are scarce (solid line), dingoes exert stronger suppressive effects on mesopredators, providing greater benefits for small mammals. Their depressive effects on large mammals, such as kangaroos, are also greater owing to the reduced availability of alternative prey. When resources are abundant (dotted line), the positive effects of dingoes on small mammals diminish, primarily due to the increase in mesopredator populations and the reduced ability of dingoes to suppress them. Conversely, the suppressive effects of dingoes on large mammals are reduced owing to the increased availability of alternative prey such as locusts or irruptive rodents. Figure from Letnic et al. (2012).


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regions of Queensland and South Australia (Letnic et al. 2012). Much of this vast region has been degraded by overgrazing in the past and by cur-rently high total grazing pressure from sheep, goats, rabbits and kangaroos (Crabb 2004). Native vegetation communities have been fragmented and depleted, soils are eroded and, in the New South Wales part of the region alone, some 38% of the native mammals that were present at the time of European settlement are now wholly or regionally extinct (Dickman et al. 1993). Could reintroduction of the dingo to these conservation wastelands (sensu Dickman et al. 2009) help to restore ecologi-cal processes and improve the conditions needed for small native species to flourish again?

Here, we take the Western Division of New South Wales as a case study area. Bounded by the borders with Queensland, Victoria and South Australia to the north, south and west, respectively, and to the east by a line near the 375 mm rainfall isohyet, the region covers 325 000 km2. Dingoes have been very scarce in the Western Division since 1930 due to the dingo fence and continuing control operations to kill any animals that penetrate the fence (Glen and Short 2000). Dingoes once ranged throughout this region and could, potentially, be reintroduced.

In a review of the environmental consequences that might follow from re-establishing dingoes in the Western Division, Dickman et al. (2009) argued that the suppression of fox and potentially cat popu-lations would be a key benefit. Using a rank-scoring system to identify species at risk from the two meso-predators, the authors estimated that 70 of 80 species of native vertebrates that are listed as threatened at the state level would have some chance of recovery if dingoes were present. These included all 16 spe-cies of listed mammals, 33 of 41 listed birds and 21 of 23 listed species of reptiles. Extending the analy-ses for the Western Division, and using threatened species listed in the schedules of the Environment Protection and Biodiversity Conservation Act 1999, Letnic et al. (2009) calculated further that maintain-ing dingo populations in low rainfall (< 350 mm) rangeland areas would benefit threatened mammals over 2.4 million km2 of the continent.

Other benefits of reintroducing dingoes include the potential to reintroduce medium-sized native mammals such as potoroids that have become regionally extinct, reduced populations of kanga-roos and introduced pest species, maintenance of native vegetation communities, and the restoration of ecological interactions and ecosystem services – such as soil disturbance – that are performed by native species (Dickman et al. 2009). Many of these benefits have been realised already in small reserves from which foxes and cats have been removed (Vieira et al. 2007; Finlayson et al. 2008; see Chapter 15), providing some confirmation of the expected effects of suppressing the two species of mesopredator.

Dickman et al. (2009) also noted four or five large-bodied native species in the Western Divi-sion that could be at risk if dingoes were reintro-duced, but suggested that any impacts of dingoes would probably not exceed the combined effects of foxes and cats. In contrast, Allen and Fleming (2012) proposed that 75 of the 80 threatened verte-brates in the Western Division would be at direct risk of dingo predation if this carnivore were rein-troduced. However, their lists of at-risk species include many that have not been found in the diet of the dingo because they are too small (< 1 g), fos-sorial, arboreal or otherwise inaccessible to forag-ing dingoes, and many that maintain strong populations in other areas where dingoes occur. Nonetheless, the large discrepancy in estimates by Dickman et al. (2009) and Allen and Fleming (2012) suggest that any program of dingo reintroduction be undertaken with caution.

Three steps are probably needed to initiate the reintroduction of dingoes to the Western Division of New South Wales. First, pilot trials should be run to quantify the effects of dingoes on native ver-tebrates, as well as selected plants and inverte-brates, and thus confirm whether dingoes do indeed have the beneficial effects on most compo-nents of the rangeland biota that are predicted. Such benefits need to be marked and demonstrable if dingoes are to be reintroduced more broadly. Pilot trials could be carried out most easily in



reserved areas such as Sturt National Park, within fenced conservation reserves such as Scotia Reserve in New South Wales or Arid Recovery in South Australia (e.g. Moseby et al. 2012), or in non-reserved but dingo-protected areas outside the dingo fence. Second, means of protection would need to be found to protect livestock herds and farm incomes from free-ranging dingoes. In fact, such means are well known and have been employed in other parts of the world for centuries. Some of these approaches have been discussed in detail elsewhere (Fleming et al. 2001; O’Neill 2002; Dickman et al. 2009), with the most cost-effective method almost certainly being the deployment of guardian animals (see Chapter 14; van Bommel and Johnson 2012). Third, because dingoes are so universally despised in Australia’s rangelands, proponents of dingo reintroduction and landhold-ers would need to engage and collaborate so that community concerns could be addressed. Methods of community engagement are well known (e.g. Dickman 2013) and not discussed further here. Legislation also would need to be repealed or rescinded to allow any reintroduction attempts to succeed, as dingoes must now be killed in all parts of the Western Division under both the Wild Dog Destruction Act 1921 (NSW) and the Rural Lands Protection Act 1998 (NSW).

In describing some of the potential benefits of reintroducing dingoes to western New South Wales, and sketching steps that could be taken to achieve this goal, it is also worth noting the costs of continuing to manage the Western Division under a ‘business as usual’ model. Assuming that din-goes have the net benefits to biodiversity that could be quantified at step one, above, not proceeding with a reintroduction program could be cata-strophic (Dickman et al. 2009). We could expect further losses of native species, ecological processes and ecosystem function, continuing losses of soil due to salinisation and erosion, and accelerating use of poisons and other means to reduce the impacts of an ostensible ‘pest’. In light of our cur-rent understanding of the importance for biodiver-sity and ecosystem functioning of strongly interactive species, such as the dingo, and the

inappropriate management of much of the range-lands, we argue that there should be no further delay in establishing pilot reintroduction trials.

Reintroducing other strongly interactive speciesAlthough we have focused on the dingo, other native carnivores such as quolls and the Tasmanian devil might also be considered for reintroduction to parts of their former ranges (Ritchie et al. 2012). In these situations, however, care would need to be taken to ensure that the processes that caused the species’ original demise, such as habitat distur-bance, hunting or impacts from invasive predators, were no longer operating. On the one hand, sub-stantial ecological benefits could be expected to flow from the re-establishment of such species, as we have noted above for the dingo. On the other, carnivores such as the eastern quoll, western quoll (Dasyurus geoffroii) and the Tasmanian devil are at increasing risk in their current ranges (Fancourt et al. 2013; see Chapter 9), so that reintroductions could be critically important means of bolstering their populations. Similar steps to those proposed for the dingo could be taken to reintroduce these other native carnivores. If we are concerned about the integrity of Australia’s natural systems and the fate of the continent’s endemic, iconic and strongly interactive species, we propose that carefully planned reintroduction programs be set up for them as a matter of priority.

AcknowledgementsC. R. Dickman was supported by an Australian Research Council Professorial Research Fellowship, M. E. Jones by an Australian Research Council Future Fellowship, and A. S. Glen by Capability Funding from Landcare Research. We thank R. Shine for helpful discussion about strongly interac-tive reptiles and C. Hughes for helpful comments on a previous draft.

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Strongly interactive carnivore species: maintaining and restoring ecosystem function.

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