Proceedings of the conference held 8-9 October 1994, Footscray, Melbourne
Biodiversity Series, Paper No. 8
Department of the Environment, Sport and Territories, 1996
11. Fire effects on vertebrate fauna and implications for fire management and conservation
Barbara A. Wilson
School of Biological and Chemical Sciences, Deakin University
Studies to determine the effects of fire on fauna in Australia have investigated patterns and rates of recolonisation after wildfire and compared populations in areas of different post-fire successional ages. The impacts of fire on some species, groups and communities have been studied while for others we have little knowledge. The response patterns of small mammals are well documented, while there is little information for reptiles and amphibians. Proposed models to describe responses and predict impacts are examined. The long-term succession of small mammal species after wildfire, and the implications of current fuel reduction practices are considered. The incorporation of ecological burning practices into fire management planning is examined, and future directions for fauna fire ecology research and management are discussed.
Key words: fire effects, vertebrate fauna, fire management, conservation, recolonisation, post-fire succession.
Fire has an important role in Australian ecosystems. It influences a wide range of environments, and fire prone vegetation such as sclerophyll forests and heathlands are subjected to recurrent fires (McArthur 1968 1972; Gill 1975; Luke & McArthur 1978).
Fire is employed extensively in Australian ecosystems for management purposes such as regeneration of areas following timber harvesting, and to decrease fuel loads in order to control the spread of wildfires (Luke & McArthur 1978; Rawson et al. 1985). Whether 'wildfire' or 'managed', fire has major implications for fauna taxa and communities.
Reviews of the interactions of Australian fauna and fire have highlighted the paucity of studies in this area compared to those investigating the responses of flora (Catling & Newsome 1981; Recher & Christensen 1981; Suckling & McFarlane 1984; Meredith 1988; Friend 1993). The impact of fire on some species, groups and communities has been studied while for others we have little knowledge. The aims of this paper are to briefly review the effects of fire on vertebrate fauna and examine models to describe responses and predict impacts.
A case study of the long-term succession of small mammals following wildfire is presented, and the implications of fuel reduction burning regimes for this community is assessed. The paper also addresses future research directions for vertebrate fauna fire ecology, and management requirements.
11.3.1 Responses of vertebrates to fire
The effects of fire on fauna vary depending on the fire regime, including the intensity and frequency of fire and the season of occurrence. High intensity wildfires may result in the incineration of many animals and decimation of faunal populations. Nevertheless small pockets of unburnt vegetation often remain and act as refuges for fauna. Low intensity fires, on the other hand, may leave up to 40 per cent of an area unburnt; this results in survival of many animals which are able to move into unburnt refuges or shelter in burrows, under rocks or in tree hollows. The frequency of fire will affect the re-establishment of populations and several short interfire periods may substantially reduce population numbers. The season a fire occurs in will affect the intensity of the fire; in south-east Australia autumn burns are hotter than spring burns. Fire season also has variable effects on populations, for example spring fires significantly interrupt breeding activities of some species.
Factors such as the shelter and food requirements of taxa, and their behavioural patterns will effect their responses. For example those animals that are fossorial or utilise tree hollows may avoid the acute effects of fire as it passes through an area and may also be protected in the immediate post-fire period. Species that are capable of moving rapidly can seek shelter in gullies and creeks and also avoid the fire. In the early post-fire recovery period animals that are sedentary may be subjected to low food resources, while other taxa may be capable of migrating to unburnt patches.
Our knowledge of the effects of fire on vertebrate taxa and communities is variable with regard to the different taxa and groups. The impacts of fire on birds and mammals has been more extensively documented (Table 1) than for reptiles, while there is a paucity of data for amphibians. Information on the effects of fire is more extensive for small mammals than for arboreal mammals, and there are no data for fire effects on bats.
|Mammals||Catling 1986; Catling & Newsome 1981; Christensen 1980; Christensen & Maisey 1987; Christensen et al. 1981; Cockburn 1981a,b; Cockburn et al. 1981; Fox 1982, 1983, 1990; Fox & McKay 1981; Humphries & Tolhurst 1992; Kemper 1990; Lunney 1987; Lunney & Ashby 1987; Lunney et al. 1987; Masters 1993; Newsome & Catling 1983; Newsome et al. 1975; Posamentier & Recher 1974; Recher et al. 1974; Wilson et al. 1990|
|Birds||Brooker & Rowley 1991; Christensen & Kimber 1975; Cowley 1974; Christensen et al. 1981; Fox 1978; Kimber 1974; Loyn et al. 1992; Meredith et al. 1984; Recher & Christensen 1981; Smith 1985; Woinarski 1990; Wooler & Calver 1988|
|Reptiles||Bamford 1986; Bratihwaite 1987; Caughley 1985; Cheal at al. 1979; Cogger 1964; Fox 1978; Humphries 1992; Newsome et al. 1975; Trainor & Woinarski 1994|
|Amphibians||Bamford 1986, 1992|
There have been a substantial number of studies of the responses of small mammals to fire. Some studies have investigated patterns and rates of recolonisation after wildfire (Recher et al. 1974; Newsome et al. 1975; Catling & Newsome 1981; Newsome & Catling 1983; Catling 1986; Fox 1982, 1983; Heislers 1974). Others have examined the effects of fuel reduction or controlled burns (Cowley et al. 1969; Leonard 1972; Heislers 1974; Christensen & Kimber 1975). The effects of fire have also been investigated by comparing small mammal populations in areas of different post-fire successional ages (Fox & McKay 1981; Cockburn et al. 1981). Most studies have been short term (less than ten years) however, and there are few which have investigated repeated burning.
High intensity wildfires result in the incineration of many animals and decimation of small mammal populations (Recher et al. 1974; Newsome et al. 1975; Catling & Newsome 1981; Newsome & Catling 1983). Nevertheless wildfires can be patchy, and small pockets of unburnt vegetation often remain. Hemsley (1967) reported survival of Isoodon obesulus (southern brown bandicoot) and Potorous tridactylus (long-nosed potoroo) after severe wildfire in Tasmania. Low intensity fires such as those used in fuel reduction burning practice can leave up to 25 per cent of an area unburnt (Hodgson & Heislers 1972; Christensen & Kimber 1975). Many animals survive such fires by moving into unburnt refuges ahead of the fire or sheltering in burrows, under rocks or in tree hollows. Species such as Antechinus stuartii (brown antechinus) and Rattus fuscipes (bush rat) suffered no immediate effects after low intensity fuel reduction burning in dry sclerophyll forest (Leonard 1970). After single fuel reduction burns A. stuartii and R. fuscipes declined in abundance but populations recovered in one and two years respectively (Humphries 1992). Antechinus swainsonii (dusky antechinus), however, was eliminated after a low intensity burn in wet sclerophyll forest (Leonard 1972).
The life history of many small mammal species have been studied (see Lee & Cockburn 1987) and most south-eastern Australian species breed seasonally during spring and summer. Thus spring fires may interrupt the breeding season more so than those occurring in autumn. Many dasyurid species such as the Antechinus spp. are particularly susceptible to fire from August to December after mating and the male "die-off" period. At this time population numbers are extremely low, consisting only of pregnant females or lactating females with young in the pouch or nest.
Very little work has been done on the effects of the frequency of fires on small mammals. The effects of two consecutive wildfires (1972, 1980) on small mammals were studied at Nadgee Nature Reserve. The 1972 wildfire decimated small mammal communities (Newsome et al. 1975). Populations of Mus musculus (house mouse), which was not present before the fire increased rapidly, peaked at years two to three and then crashed. The rodents R. fuscipes and Rattus lutreolus (swamp rat) peaked in biomass after five years and then declined. Antechinus stuartii regained post-fire levels at year five, while
A. swainsonnii did not reappear until the sixth year. After the second fire (1980), all species were recaptured in low numbers in complex habitats, but were absent in least complex habitats (Catling 1986). Species diversity and abundance increased as habitats aged. In the Wombat Forest, Victoria, studies are in progress to investigate the effects of repeated fuel reduction burning (three and ten year rotations) on small mammals (Tolhurst & Flinn 1992; Tolhurst, this volume).
Small mammal species exhibit varied responses to wildfire. The different recolonisation responses of species to fire are considered to be closely related to the successional changes of the vegetation, as both structural and floristic factors play a major role in determining habitat selection (Braithwaite & Gullan 1978; Fox & Fox 1981). Species such as Pseudomys gracilicaudatus (eastern chestnut mouse) which prefers open, floristically rich vegetation recolonises early in the post-fire recovery period, while
A. swainsonii which requires dense ground cover, exhibits low population numbers up to six years after fire (Newsome et al. 1975; Recher et al. 1980; Fox l982, l983; Lunney et al. 1987). In the spinifex grasslands of central Australia species such Pseudomys desertor, P. hermannsburgensis, Mus domesticus, Notomys ridei and Sminthopsis youngsoni are more abundant on older succession areas (11-15 years) while S. hirpites and N. alexis are more abundant on the one to four year aged areas (Masters 1993).
Although fire is acknowledged as impacting on arboreal mammals (Fox 1978; Smith & Lindenmayer 1988; Lindenmayer et al. 1991) information on the effects of fire on the group is lacking. High intensity fires kill species such as Pseudocheirus peregrinus (common ringtail possum) (Newsome et al. 1975; Fox 1978) and survivors of species such as P. perigrinus and Petauroides volans (greater glider) move to unburnt creek and wet gully refuges. Lunney (1987) recorded a substantial shift from logged to unlogged forest of possums and gliders, with unburnt gullies providing important refuges. Species such as P. volans and Petaurus australis (yellow-bellied glider) were confined to these refuges. There is evidence that species such as Acrobates pygmeus (feathertail glider) may favour early successional stages following fire (Braithwaite 1983, Braithwaite et al. 1983). In woodlands of northern Australia Kerle (1985) has suggested that high fire frequency is responsible for low population densities of the northern brushtail possum (Trichosurus arnhemensis).
Large hollow-bearing trees provide shelter and nesting sites for arboreal mammals, and fire regimes affect the number of suitable den trees and hollows (Smith & Lindenmayer 1988; Inions et al. 1989; Lindenmayer et al. 1991). Nesting sites are a major limiting factor for Leadbeater's possum (Gymnobelideus leadbeateri), which is dependent on large hollow bearing trees (Smith & Lindenmayer 1988; Lindenmayer et al. 1991). In the 1939 regrowth montane forests of Victoria loss of old dead hollow-bearing trees from deterioration, and potential nest trees from logging, threaten the future populations of this species (Smith & Lindenmayer 1988; Lindenmayer et al. 1991).
Large burrowing mammals such as the wombat (Vombatus ursinus) probably escape the immediate fire in deep insulated burrows and search out regenerating grasses in the post-fire recovery period (Newsome et al. 1975). The western grey kangaroo (Macropus fuliginosus) and the western brush baby (M. irma) both favour frequently burnt forest (Christensen & Kimber 1975). Macropods such as the woylie (Bettongia penicillata) and the Tammar wallaby (Macropus eugenii) in Western Australian forests have been shown to be 'fire dependent' species (Christensen 1980; Christensen et al. 1981; Christensen & Maisey 1987). The food and cover plant species that they depend on are in turn dependent on certain fire regimes for regeneration and maintenance. Fire can also increase the availability of fungal food resources by facilitating the animals ability to find sporocarps (Christensen 1980). For the long-footed Potoroo (Potorous longipes) which feeds almost exclusively on hypogael or underground sporocarps, the fire regime may have a major impact on dispersion of food resources (Scotts & Seebeck 1989).
The rufous hare wallaby (Lagorechestes hirsutus) favours early post-fire succession spinifex habitat (Burbidge & Johnson 1983). This preference is related to the nutritional value and digestibility of the regrowth (Lundie-Jenkins 1993, Lundie-Jenkins et al. 1993). The species appears to be dependent upon two specialised forms of habitat, dense spinifex for movement and daytime shelter, and more open areas for feeding (Lundie-Jenkins 1993). Fire is clearly implicated in creating the diversity of feeding and sheltering habitats for this species (Lundie-Jenkins 1993, Lundie-Jenkins et al. 1993).
The intensity of fire influences bird mortality rates and the degree to which vegetation structure is altered in the short term. Low intensity fires result in low death rates especially when structural changes to vegetation are minimal (Cowley 1974; Christensen & Kimber 1975; Wooller & Brooker 1980). Loyn et al. (1992) found that bird populations were not significantly affected after single fuel reduction burn treatments. Wooller and Calver (1988) found few changes in species composition after low intensity fires, however the abundance of many species were halved. Birds most affected are those occupying the habitat levels affected by the fire. In general after a low intensity fire the vegetation structure re-establishes rapidly, there is an initial decrease in both numbers and species followed by a substantial increase within one to two years.
High intensity fires can cause substantial bird mortality and have major affects on structural attributes of vegetation and consequently bird numbers (Hemsley 1967; Newsome et al. 1975; Fox 1978; Wegener 1984; Recher et al. 1985). Declines in bird populations after fire result from the changes to vegetation structure, producing reduced cover and shifts in prey availability (Recher et al. 1985). Christensen and Kimber (1975) found an initial decrease in both species and numbers followed by an increase within one year. The understorey birds were most affected but returned in increased numbers within three years. Similarly after the 1983 Ash Wednsday fires 86 per cent of species were recorded within two years, and the abundance of populations continued to increase until four years post-fire (Reilly 1991).
Fire can also affect nesting and breeding activities of birds. After extensive fire in heathlands Brooker and Rowley (1991) found that three species (splendid fairy-wrens, Malurus splendus; western thornbills, Acanthiza inornata; yellow-rumped thornbills, A. chrysorrhoa) were still able to nest in regenerating shrubs in the immediate post-fire year, although breeding was delayed. However, at least two species (white-browed scrubwren (Sericornis frontalis) and white-cheeked honeyeater (Phylidonyris nigra), did not nest for two years, and one (inland thornbills A. apicalis) for five years. Fire free frequencies of at least ten years were recommended to maintain these heathland bird populations.
In some cases there are increases in bird species after fire. In the savannah woodlands of northern Australia Woinarski (1990) found short-term effects of fire involved a substantial increase in the number of species. The increase was due to arrival of species which are attracted to recently burnt sites because of increases in abundance or availability of food, such as carcasses and seed-fall. Many species in the area are nomadic and this habit enables them to locate increased resources in recently burnt patches (Woinarski 1990). Similarly influxes of species that feed on exposed soil and eucalypt nectar were observed after fire in the Wombat Forest (Loyn et al. 1992)
Some species may be dependent on fire, or particular fire regimes. The ground parrot (Pezoporus wallicus) inhabits three major habitat types: swampy sedgelands, graminoid heathland and diverse shrub heathland (Meredith 1983, 1984). In south east Australia the species occurs in heathlands with fire ages of between three and 20-25 years (Meredith et al. 1984). It was proposed that the parrot was not fire adapted as such, but did require fire in heathlands where sedge seed production was variable with time since fire (Meredith et al. 1984). In Queensland densities of the species were greatest at sites where the graminoid-heathland vegetation was unburnt for five to eight years, and this was related to those sites containing the highest number of food species and standing crop of seeds (McFarland 1990, 1991). Declines in number of parrots in old heathlands reflected the fall in food availability (McFarland 1991).
There is evidence of long-term changes in temperate Australian bird communities following post-fire succession changes. In mallee woodland communities there is a rapid increase in both bird numbers and number of species between 0 and 15years post-fire, after which density decreases but species richness continues to increase (Meredith 1983). Low species numbers and density in early stages are related to low structural diversity of vegetation. The peak in density (15 years) corresponds to a high level of productivity. Some species were restricted to vegetation older than 20 years post fire (Meredith 1984). Very frequent fires which inhibit vegetation attaining mature structural features would result in loss of species dependent on such habitat structures (Meredith 1984). Woinarski (1990) found relatively limited long-term changes in bird communities across a post-fire successional gradient of savannah woodlands in Northern Australia. However there was an increase in diversity of birds, and in the density of species which feed in shrubby understorey as it changed from predominantly grasses to shrubs, with time after fire.
Most information on reptiles and fire has been from studies in mallee woodlands, heathlands and northern Australian savannah forests where reptilian diversity is high (Cogger 1969, 1989; Fox 1978; Caughley 1985; Bamford 1986; Braithwaite 1987; Woinarski 1989; Trainor & Woinarski 1994). Few studies have been undertaken in southern temperate areas (Coulson 1990; Lunney et al. 1991; Humphries 1992).
Intense wildfire in Mumbulla State Forest had variable effects on populations of three lizard species (Lunney et al. 1991). There was little impact on the water skink (Eulamprus heatwloei), a gully species requiring partial shade. Populations of the delicate skink (Lamphropholis delicata) were higher in the post-fire period, possibly reflecting the extremely cryptic nature of the species. The grass skink (Lamphropholis guichenoti) exhibited differential survival in habitats. Survival was higher in the gullies than on the ridges where intense heat occurred during the fire. The ground temperatures were close to, or greater than, maximum for the species, prey was scarce and predation pressure higher in exposed ridges during the early post-fire period.
The effects of spring and autumn fuel reduction burns also varied for three lizards in the Wombat Forest in Victoria (Humphries 1992). The abundance of Sphenomorphus tympanum (southern water skink) was relatively stable after fires probably due to the fact that its primary habitat of fallen logs and branches had not been significantly affected. There was a higher abundance of both Leilopisma coventryi (Coventry's skink) and Nannoscincus maccoyi (McCoy's skink) immediately after burning. However this result was considered to reflect the increased detectability of the species as a result of cover removal, rather than an actual increase in population numbers (Humphries 1992). Abundance of both species declined after five months post-fire and Coventry's skink increased in abundance 28 months after fire, following the rapid recovery of the litter layer.
Seral responses of reptiles have been investigated (Cogger 1969; Cheal et al. 1979; Caughley 1985; Bamford 1986; Braithwaite 1987). Cogger (1969) found that the mallee dragon (Ctenophorus fordi) was at higher density in ten year, post fire regrowth than in unburnt areas. In mallee areas Caughley (1985) found similar species numbers in sites aged four, seven, 25 and 60 years since burn, although relative abundance of species was markedly different. Woinarski (1989) found similar changes in lizard communities in broom brush communities. Bamford (1986) found no relationship between total species, number of captures and time after fire at his study sites in heathland, woodland habitats in Western Australia. He concluded that the effect of fire overall on reptiles was negligible, although a small number of species did exhibit clear post-fire seral responses. Some species were absent from early succession areas, while others were present in increased numbers, apparently favouring the more open ground.
The response of lizards to fire was investigated in tropical savannah habitat in Kakadu National Park (Braithwaite 1987). Different species selected habitats created by fire of different intensities, and Braithwaite (1987) found these were correlated with the fire characteristics and post-fire vegetation structure. An experimental approach (Trainor & Woinarski 1994) in such communities found only one lizard species exhibited a short-term response to fire by decreasing in abundance. Two species were more abundant in burnt areas and one in unburnt areas. Differences in relative abundances between treatments (ages) were attributed to differences in vegetation structure. Habitat partitioning appeared to play a more significant role than direct affects of fire. Many species were directly associated with a gradient of moisture availability, one group with seepage and another at the dry end of the gradient.
Fire is acknowledged as a major threat to endangered species such as Delma impar (striped legless lizard) which inhabits predominantly grassland communities (Coulson 1990). While individuals may survive by burrowing, the loss of vegetative cover may threaten any survivors due to heat stress or vulnerability to predators. Acute effects of fire on the species have been recorded. At Derrimut grasslands in Victoria six individuals were found dead after a fire (Coulson 1990).
Many reptiles are resilient to short-term affects of fire, due to their fossorial habits e.g. some snakes and lizards (Friend 1993). Other species which are arboreal or surface dwelling would not be so protected. The degree to which these different microhabitats are modified by the fire event, and the rate of its recovery in the post-fire period will determine the impact of the fire and fire regime on different species.
There is almost no work on this group. Many frog species may be affected by fire in predominantly indirect ways because of their utilisation of pools, ponds and subterranean shelter, sites essentially protected from direct effects of fire (Friend 1993). Bamford (1986, 1992) investigated the responses and adaptations of three frog species to fire and their relationship to vegetation structure, litter density and potential food supply. He found no relationship between time since fire and number of species or total abundance of frogs. However the abundance of two species (Limnodynastes dorsalis and Myobatrachus gouldii) was greater in long unburnt areas. These abundance changes were not related to litter and vegetation changes, or prey abundance. The differences were proposed to be the result of nett movement away from burnt areas after fire in these nomadic species. Frogs appear to be influenced more by proximity to water than by time since fire (Bamford 1986).
Responses of vertebrates to fire in different habitats
Most research on the effects of fire on vertebrates has concentrated on fauna in forests, heathlands and woodlands (Table 2). Less work has been undertaken in mallee ecosystems and very little in temperate grasslands. In the tropical northern savannahs where fires occur frequently there are few studies on the effects of fire on fauna (Begg et al. 1981; Kerle 1985; Braithwaite 1987; Woinarski 1990). The effects of fire regimes are poorly understood (Trainor & Woinarski 1994). Fire is also frequent in arid spinifex grasslands, and fire driven succession of fauna has been documented (Masters 1993). However in both of these communities there is evidence that rainfall and moisture patterns have strong effects on the distribution and abundance of fauna (Masters 1993; Trainor & Woinarski 1994).
|Forest (temperate)||Christensen & Kimber 1975, Cowley 1974, Christensen et al. 1981, Kimber 1974, Loyn 1985, Loyn et al. 1992, Lunney 1987, Lunney & Ashby 1987, Lunney at al. 1987, Tolhurst & Flinn 1992, Wooller & Calver 1988|
|Tropical savannah forests & woodlands||Begg et al. 1981, Kerle 1985, Braithwaite 1987, Woinarski 1990, Trainor & Woinarski 1994|
|Heathlands, woodlands||Bamford 1986, Brooker & Rowley 1991, Catling 1983, Catling et al. 1982, Christensen 1980, Christensen et al. 1981, Cockburn 1981a b, Cockburn et al. 1981, Fox 1978, Fox 1982 1983 1990, Fox & McKay 1981, Kemper 1990, Meredith et al. 1984, Newsome & Catling 1983, Newsome et al. 1975, Recher et al. 1974, Smith 1985, Wilson et al. 1990|
|Mallee-woodlands heathlands||Caughley 1985, Cheal et al. 1979, Cogger 1964, 1989|
|Arid spinifex grasslands||Burbidge 1985 Lundie-Jenkins 1993, Masters 1993|
|Temperate grasslands||Coulson 1990|
It is essential that the responses of vertebrates in different habitats and communities be documented and understood. The responses of species and communities in the central arid grasslands are likely to be considerably different than in the wet tropics, or southern temperate grasslands, where each ecosystem exhibits extremely different climates, landscapes and habitats. Similarly species that occur in temperate south-eastern coastal heathlands compared to their counterparts in sub-tropical Queensland.
Models for predicting fire impacts and succession patterns
There has been a substantial amount of work investigating the responses and life-history patterns of plant species with relationship to fire (Connell & Slatyer 1977; Noble & Slatyer 1979, 1980, 1981; Specht et al. 1958; Gill & Groves 1981). A number of models have been developed to describe the responses, and applied to predicting post-fire succession (Connell & Slatyer 1977; Noble & Slatyer 1979, 1980, 1981).
There are gaps in our knowledge of fauna responses, and responses in different habitats. There are also few appropriate models to assess impacts and predict post-fire succession (Gill 1989; Tolhurst 1989; Wilson 1990; Friend 1993). One model has been developed for post-fire responses of small mammals by Fox (1982, 1990). He proposed a habitat accommodation model whereby animal succession occurs in response to changes in vegetation. Animals enter the succession as their specific requirements are met and are replaced or decline in abundance as conditions become suboptimal. In general this model is supported by results of several studies documenting post-fire succession of small mammals in temperate heathlands (Fox 1982; Newsome & Catling 1979; Catling and Newsome 1981; Wilson et al. 1990). There is also evidence of post-fire succession in arid grasslands (Masters 1993).
Friend (1993) developed a generalised model for predicting the impact of fire on small vertebrates in temperate mallee woodlands and heathlands. He analysed the shelter and dietary requirements of species, as well as their reproductive patterns. These attributes explained general patterns of early, mid and late-successional species. For mammals pyric response patterns of species were closely related to their shelter, food and breeding requirements. For reptiles there is a strong relationship between shelter and foraging requirements of species and their abundance in successional ages. Species abundances and distributions of amphibians were more closely linked to moisture regimes than to fire regimes.
Population viability models (PVA) can also be applied to predict affects of fire and fire regimes on fauna populations (see Possingham & Gepp 1995). Such models are able to assess extinction and survival probabilities under a range of pressures such as catastrophes, demographic variation, and inbreeding depression. Models can also be employed to assess the effects of a variety of management actions such as burning regimes on populations of species.
These, and similar models are important as they provide a basis for assessing responses of faunal taxa, particularly those for which data are minimal. They can be employed to predict effects of fire events, and particular fire regimes, based on the data presently available. New information and data on species and community responses to fire can be incorporated into the models, leading to improvement of the models. Such models can assist in the implementation of ecologically suitable fire regimes within the fire planning process.
11.3.2 Use of fire in fauna management
The use of fire as a management tool for Australian fauna has not been extensive, primarily due to lack of information on species requirements. It is now becoming more prominent particularly with regard to conservation of rare species. There has been a substantial decline in the mammal species of the central and western deserts of Australia since European settlement and changes in fire regimes is one factor implicated (Morton 1990; Burbidge & McKenzie 1989). It is believed that burning patterns used traditionally by aborigines produced mosaics of vegetation of different successional ages which provided a diversity of habitats for mammals. This pattern changed as aborigines settled and has lead to large, intense wildfires which eliminate vegetation such as spinifex and reduce structural and floristic diversity (Morton & Andrew 1987; Burbidge & Johnson 1983). Traditional burning regimes are now being reintroduced in attempts to conserve rare mammals such as the rufous hare wallaby (Lagorechestes hirsutus) (Johnson 1982; Burbidge & Johnson 1983; Lundie-Jenkins 1993, Lundie-Jenkins et al. 1993).
In Western Australian forests, Bettongia penicillata (the woylie) and Macropus eugenii (Tammar wallaby) were identified as being fire dependent (Christensen 1980; Christensen et al. 1981). The food and cover plant species that they depended on were in turn dependent on certain fire regimes for regeneration and maintenance. Fire management plans based on the species requirements have been employed to conserve and manage these mammal species (Christensen & Maisey 1987). Sections of vegetation have been burnt on a rotation system, rotation time was increased (eight to 12 years) and burning season altered to provide appropriate regimes and areas for both species. Small areas of thickets (Gastrolobium bilobum, Melaleuca viminea) were hand lit to provide intense heat for good seed germination.
The ground parrot (Pezoporus wallicus) is a vulnerable species (Meredith et al. 1984) in eastern Australia. The species occurs in heathlands with fire ages of between three and 20-25 years and it was proposed that this was related to sedge seed production which was variable with time since fire (Meredith et al. 1984). Although regular burning was thought to be necessary for maintenance of the populations, it was shown that too frequent burning, (e.g. three or more fires each less than six to eight years apart) lead to degradation of the heath and absence of the parrot.
Many species of pseudomyine rodents (e.g. heath mouse, New Holland mouse) are dependent on floristically-rich vegetation of early post-fire age (Cockburn 1978; Cockburn et al. 1981; Fox & Fox 1978, 1984; Fox 1982). Cockburn (1978) proposed that a decrease in the frequency of fires may have been a major factor contributing to their decline by increasing the probability of large fires and areas of uniform successional age. The use of fire to maintain suitable habitats has been recommended (Cockburn 1978; Cockburn et al. 1981; Fox & Fox 1978, 1984; Fox 1982).
11.4.1 Fire regimes and small mammal populations in Easterns Otways, Victoria
The Eastern Otway Ranges is approximately 100 km southwest of Melbourne. The area has predictable, reliable rainfall throughout the year, with hot summers and cool winters. The vegetation communities are predominantly fire-prone dry sclerophyll communities (open forest, woodland, heathland, shrubland) with less extensive areas of riparian open forest, fern gullies, damp open forest and coastal sand dune (Land Conservation Council 1985; Meredith 1986). The diversity of the small mammal communities in the area is high, 14 species have been recorded (Kentish 1983; Wilson et al. 1986). Several rare or uncommon species occur in the area including Antechinus minimus (swamp Antechinus), Pseudomys novaehollandiae (New Holland mouse) and Sminthopsis leucopus (white-footed dunnart) (Land Conservation Council 1985; Kentish 1982; Wilson 1986).
Although fire records for the study area are incomplete there is a history of regular summer fires. The Ash Wednesday wildfire in February 1983 was of high intensity and burnt 40,000 hectares of the Otway Ranges. Aspects of the recolonisation of small mammals between 1983 and 1988 have been reported (Wilson 1990, 1991; Wilson & Moloney 1985; Wilson et al. 1990). A summary of the recolonisation process over a ten year period (1983-1992) is presented here and the implications of the present practice of fuel reduction burning for the small mammal species and communities is assessed.
Recolonisation of small mammals in the study area was assessed at 22 study sites (15 burnt, three unburnt, four partially burnt). Overall changes in the abundance of native mammals and M. musculus in the study area (n=22 sites) are shown in Figure 11.1. The abundance of native mammals and M. musculus was low in the first year after the fire (0.6 and 3.2 respectively). Native species abundance remained low until Autumn 1986 (year 4) and continued to increase in years 5 and 6. There was a large decline in abundance of native mammals in 1991, nine years after the fire.
Source: after Wilson 1990, Aberton, J. 1991-92 pers. comm.
All twelve native small mammal species recorded in pre-fire studies in the study area were captured in the post-fire study. One capture only was recorded for three native species Perameles nasuta (long-nosed bandicoot), Petaurus breviceps (sugar glider), P. tridactylus (long-nosed potoroo). Other species recorded in low numbers were A. swainsonii and A. minimus.
The different seral responses of species on burnt sites are shown in Figure 2 which plots the relative abundance (abundance/maximum abundance), and the time required for each species to achieve its peak abundance. Species responses are divided into early- (Fig. 2a), mid- (Fig. 2b) and late- (Fig. 2c) succession. Mus musculus peaked in the third year and clearly occupies the early seral stages. Pseudomys novaehollandiae (New Holland mouse), first captured in low abundance (year 3), increased in year 4 and then declined (Wilson 1990, 1991). Antechinus stuartii reached maximum abundance in the fifth year. Although S. leucopus was present in low numbers throughout the study it appeared to achieve maximal abundance in mid-succession. The two Rattus species increased in abundance up to year 6 and declined markedly at year 9. I. obesulus exhibited a substantial increase in relative abundance from years 4 to 6 and was very low from years 9 to 10. Antechinus minimus was eliminated from most sites it occupied pre-fire. It was captured in the first year on partially burnt shrubland, one individual only was captured three years after the fire and the species was not captured again until 1993-1994, ten years post-fire (Fig. 2c). The species thus exhibits a late successional position.
Figure 11.2: Relative abundance (abundance/maximum abundance) of species on burnt sites, showing seral responses and the time required for each species to achieve its peak abundance. Fig 11a, early-, Fig11b mid- and Fig11c late-succession species
Source: after Wilson 1990, Aberton, J. 1991-92. pers. comm.
Implications of fuel reduction burning for small mammals of the Eastern Otways
Fuel reduction burning is currently carried out as an integral part of the region's fire protection policy (Department of Conservation & Environment 1988). Burning is conducted within particular areas to provide firebreaks for major fires and to provide protection to assets, property and human lives. The frequency of burning is determined on the basis of the priority for fuel reduction, and the rate of fine fuel accumulation. Four zones have been designated with different priorities for burning, fire frequency and fine fuel load (Table 3). The implications of the fuel reduction burning practices for small mammals can be considered in light of the information on their population densities and seral responses.
Priority zone 1 (Table 3) has a burning frequency of four years and is likely to reduce populations of all native small mammal species. Late successional species that require a long fire-free period are unlikely to survive. It is thus important to determine the species present and to consider whether translocation of susceptible species out of this zone may be a management option. Priority 1 zone is a very small area (approximately 70 hectares). If susceptible small mammal species are not present in this zone, it is unlikely that the burning regime will have a significant effect overall on the populations of the Eastern Otways.
Priority zone 2 with a frequency of six years would present problems for late successional species particularly A. minimus and A. swainsonii. The distribution of these species within this zone should be determined. Priority zone 2 (approximately 1300 hectares) represents a large area of the eastern Otways. If the whole area was burnt every six years other species could also be affected as population densities may gradually become reduced. The effects of such a repeated burning regime are not known. It is recommended that the area be burnt in a planned mosaic, resulting in a range of successional ages (six to 20 years) within specific floristic vegetation groups.
The burning frequencies in zones 3 and 4 depend on fuel accumulation rates. There is no clear indication of how often burning will occur, it could be at very low intervals (two to four years) in areas that accumulate fuel at a fast rate. In these areas the species that may be at risk should also be determined.
The plan does not define any areas that are to remain unburnt. It is clear that late successional species such as A. minimus and A. swainsonii require old vegetation, possibly eight to 20 years. Suitable habitat should be identified and left unburnt for these late successional species.
A fire management plan of the type outlined has significant implications for the New Holland mouse (P. novaehollandiae), an endangered species listed in 1991 under the Flora and Fauna Guarantee Act, Victorian Government (1988). It has a restricted, disjunct distribution in Victoria. Population studies in the Eastern Otways since 1981 have shown that the species has a patchy distribution, a low population density and is prone to local population extinctions (Kentish 1983;Wilson et al. 1990; Wilson 1991, 1994). The sites where P. novaehollandiae occurs are located on flat to undulating terrain in woodland and low-open forest with heathy understorey. The species exhibits a micro-habitat preference for vegetation of high floristic diversity and with low, dense vegetation cover (Wilson et al. 1990; Wilson 1991). The age of the vegetation where P. novaehollandiae occurs ranges from three to 20 years, most sites being of early successional age (three to four years). The species may survive in old patches, but probably only at very low densities (Wilson 1991). It is not clear what features of the early successional stage are important. A high floristic diversity may provide a variety of plants to produce seeds for this species (Watts & Braithwaite 1978; Cockburn 1980).
Table 3. Fuel reduction burning priority zones (Department of Conservation, Forests and Lands, Geelong Region Fire Protection Policy 1988)
|Zone 1||Near townships settlements andplantations.
|Zone 2||A strategic corridor of a wide continuous belt of vegetation.
|Zone 3||An area where fuel reduction is necessary to prevent destruction of natural and cultural values.
|Zone 4||Burning in this zone is needed to protect and conserve natural and cultural values and achieve other biological objectives
Source: Department of Conservation, Forests and Lands, Geelong Region Fire Protection Policy 1988
An area of approximately 2,300 hectares, located east of the Anglesea River, represents critical habitat for the species (Wilson 1994). A number of processes that represent threats to the survival of this species in the area were identified including potential land clearance, recreational pressures and inappropriate fire regimes. Fire regimes need careful investigation and design, as large extensive fires could wipe out the fragmented populations. Use of small patch burning may be the only way to create and increase the area of suitable patches of preferred early successional habitat. An understanding of the spatial structure of populations is required to enable the appropriate patch sizes and distances between them to be determined.
Finally the effects of the fuel reduction burning regime should be monitored. It is only by monitoring, that any detrimental changes induced by fire regimes can be identified. Changes to fire regimes can then be implemented. The effects of fire on small mammals cannot be treated in isolation. A range of other disturbance factors are present in the study area including the presence of an open-cut mine, introduced predators and the cinnamon fungus (Phytophthora cinnamomi) which effects the understorey vegetation, and habitats of small mammals (Newell & Wilson 1993; Wilson et al. 1990, 1994). All these factors have, or have the potential to exert pressures on the local populations. Although in historical times small mammals adapted well to the fire regimes of the area, it is clear that conservation of the communities now needs a well thought out and judicious fire plan.
As a result of changes to the field structure of the Department of Conservation and Natural Resources in 1993, the Eastern Otways is now part of the South-West area (one of five areas for the state). A new fire plan "Otway Fire Protection Plan" for the area is in draft form (Wouters, M. 1994 pers comm). One objective of the plan is to protect major habitat areas for significant species. The plan incorporates burning zones which allow for ecosystem management, and where environmental constraints necessitate different frequencies of burning. It includes no burn areas, and provides for development of sub-plans for areas where significant fire management issues are defined – such as ground parrot and New Holland mouse habitats. The approval of the plan and the implementation of such features will provide a solid basis for ecological burning regimes, protection of threatened fauna, and management of habitats.
This paper has highlighted the fact that there remains a paucity of data on the responses of particular taxa, species and ecosystems. Our understanding of the responses of reptiles and amphibians is basic and more work is required on the fire responses of these groups. Although there is much information on fire effects on birds and mammals in general, there are gaps in our knowledge for some groups for example arboreal marsupials
There are few data on the effects of repeated burning or frequent fires on fauna, and a major impediment to our understanding of fire effects is time. Long-term monitoring of sites is essential to allow examination of the impact of different fire regimes. There are gaps in our knowledge of fire responses of fauna in different ecosystems. More work is required in sytems such as grasslands, mallee and northern savannas.
Models of effects of fire on fauna, and their responses to fire regimes may be employed to make predictions in situations where primary data is lacking. Requirements for shelter, and food, together with life history patterns need to be incorporated into models (Fox 1982; Friend 1993). PVAs can be employed to predict population responses to different fire regimes and fire management practices.
In addition to modelling species responses there is a need to model community responses, including fire effects on vertebrate food resources (eg. invertebrates, vegetation, seeds, fungi) and habitat changes such as vegetation structure and plant cover. Models need to be developed at different spatial levels ranging from individual sites, to landscapes and ecosystems. PVAs can be employed to model meta-populations, to predict the effects of loss of particular populations on the whole species population.
The application of fire to fauna management, particularly endangered species is a problem for wildlife managers. There is a need to determine the effects of fire on rare and endangered fauna species. Uncontrolled wildfires and unplanned fires may eliminate critical habitat and populations. On the other hand the absence of particular successional aged habitat (e.g. early succession for some species) may result in population declines and threats to species. There is a need to assess effects of present fire present management practices on fauna and to incorporate ecological burning practices into fire management plans.
There is a need to determine the effects of fire disturbance events in concert with other disturbance factors such as introduced predators, land clearance, habitat fragmentation, logging, grazing, drought, Greenhouse.
Incorporation of fire research results into management practice does not always follow. There is a need for improved communication between researchers and fire managers. This volume, and the conference that preceeded it, are good examples of such communication. However a more practical and immediate pathway to improved communication is for researchers and managers to collaborate on projects. Ideally both groups should be involved in determining the aims and design of the project, the collection and analysis of results, as well as the determination of conclusions. Shared ownership of such projects are more likely to result in implementation of results into management plans.
I acknowledge many people with whom I have had stimulating discussions on fire affects on fauna, and fire management particularly J. Barnett, G. Friend, C. Meredith, A. McMahon, K. Tolhurst, M. Wouters. Contributions and data for the case study in the Eastern Otways from D. Moloney and J. Aberton are acknowledged. The study was supported by grants from the Department of Conservation & Environment, Victoria, Deakin Foundation, Deakin University, M.A. Ingram Trust. Work was undertaken under scientific permits issued by the Department of Conservation & Environment, Victoria, and ethics approval from the Deakin University Animal Ethics Committee.
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