Fire regimes and biodiversity: the effects of fragmentation of southeastern eucalypt forests by...

ELSEVIER Forest Ecology and Management 85 (1996) 261-278

Pores~~~ology

Management

Fire regimes and biodiversity: the effects of fragmentation of southeastern Australian eucalypt forests by urbanisation,

agriculture and pine plantations

A. Malcolm Gill * , Jann E. Williams Centrr for Plant Biodiuersity Research, CSIRO Division of Plant Industry, GPO Box 1600. Canberra, A.C.T. 2601, Australia

Abstract

Fragmentation of eucalypt forests has been common in southeastern Australia. Urbanisation, agriculture and the establishment of plantations of the exotic tree Pinus radiaru are major agencies of fragmentation. The study of the effects of these agencies on adjacent forested land has lacked a suitable framework. By constructing generalized trophic-level diagrams for each fragmenting system-farm, urban area and pine plantation-the major potential impacts on adjacent forested land can be examined. Urban areas, for example, have a relatively large non-native predator biomass (especially cats and dogs) compared with the original forest, whilst farms support a relatively large biomass of exotic herbivores. In pine plantations, by way of contrast, the biomass of native or exotic herbivores and predators is relatively small. Landscape fires are an integral part of the ecology of native eucalypt forests but are kept out of suburban areas, farms with improved pastures and standing plantations as much as possible. To explore the potential impacts on biodiversity of fire regimes in forests at the edges of urban areas, farms and plantations, we constructed and sought evidence for, a series of scenarios (each a compound hypothesis). Urban interface scenario: ‘There is a low frequency of unplanned fiie in forest remnants. To prevent losses of life and property in adjacent urban areas, regular frequent prescribed burning is practiced. Regular frequent prescribed burning reduces biodiversity’. Support for the first two parts of this scenario was strong although the frequency of fires, prescribed or unplanned, may be a function of distance from the urban edge, the size of management unit and the nature of the fuels. Urban predators may be expected to reduce vertebrate biodiversity, especially after fires. Agricultural interface scenario: ‘Clearing for agriculture leaves only small forest remnants which become fire free. Fire-free fragments eventually decline in plant species biodiversity’. Forest fragments in rural areas vary widely in size and occur as roadside remnants, farm woodlots, Travelling Stock Reserves, State Forests and designated conservation reserves. The circumstances of burning vary widely. Grazing from domestic stock, especially combined with fire, may negatively affect biodiversity. Absence of fire may also reduce biodiversity. Pine interface scenario: ‘Pines spread from plantations to neighbouring forest areas, reducing native plant and animal species diversity. This situation can be reversed by a prescribed burning regime that has fires intense enough to cause pine death and frequent enough to prevent pine seed set’. There was considerable support for this scenario although at this stage the spread of pines may be only in areas peripheral to plantations. Because fires have effects on

* Corresponding author. Tel: 01 l-61-6-246-51 16; Fax: 01 l-61-6-246-5249.

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262 A.M. Gill. J.E. Wiilium.v/ Forest Ecology und Murqemrnt 85 C/Y%) 261-275

biodiversity, crops, lives and property in fragmented forests and adjacent areas, integrated management-across landuses and jurisdictions-is recommended.

Keywords: Fire; Biodiversity; Fragmentation; Forests

1. Introduction: three forms of forest fragmentation

Inroads into eucalypt forests began with the Euro- pean colonisation of Australia 200 years ago. Gar- dens and farms supplied food to the first settlement of Sydney while building materials and firewood came direct from the forest. A whole new chapter in the ecology of a continent was opened.

Sydney’s forests, like most of Australia’s temperate forests, were found in the wetter coastal and subcoastal areas. In the drier country marginal to the forests, woodlands predominated where soils were richer and deeper while heaths and scrubs were found on the poorer, shallower, soils. Agriculture first developed in the forest and woodland zones, particularly on arable soils. Steep rocky country was left to forestry or “waste”. People clustered on the coast where port facilities were available and the climate was equable. Eventually, large cities developed such that Australia now has one of the most urbanised populations in the world. With agricultural and urban development, the native forests, typically composed of the hardwooded eucalypts, became fragmented. To supply softwood to the cities, plantations of pines were established. In this paper, we examine the impact of adjacent urbanisation, farming and pine forestry on eucalypt-forest biodiversity- particularly through changed fire regimes. We are not concerned here with the biodiversity extant before European settlement, nor the effects of fragmentation per se on biodiversity.

The setting for our contribution is southeastern Australia where eucalypt forests cover about 16 mil- lion ha (Resource Assessment Commission, RAC, 1992a). The area closely resembles that covered by two southeastern biogeographical zones of Eucalyp- tus (“0” and “P”) with high species richness (Gill et al., 1985). It includes large areas of New South Wales. Australian Capital Territory, Victoria and South Australia. These forests “support a rich and varied fauna of mammals and birds... forming the most important refuge for wildlife in Australia”

(Tyndale-Biscoe and Calaby, 1975). The clearance of eucalypt forests and woodlands in southeastern Australia has been extensive (RAC, 1992b) with Wells et al. (1984) reporting 60-100% of fores& being cleared in rural local government areas in the States considered.

As well as widespread clearing of eucalypts, dra- matic changes have occurred in the flora and fauna of southeastern Australia since the arrival of Euro- peans. For example, Marlow (1958) reported that 42% of the native marsupial fauna of NSW was either extinct or rare. Extinction varied, however. between nil in rainforest, 8% in eucalypt forest. 13% in eucalypt woodland and 43% in the “‘plains”. Introduced plants are now a significant part of the flora and feral animals are common. From 21-25% of the floras of NSW, Vie. and SA are regarded as naturalized (Hnatiuk, 1990). The combined effects of clearance, species extinctions and introductions have left a highly modified forest estate.

In this paper, forest “fragments” are considered to be any forest area with adjacent “artificial” landscapes, such as those created by farms, plantations and urban areas. “Remnants”’ are considered to be areas entirely surrounded by artificial landscapes. Here. “ biodiversity” is the richness of native species of vascular plants and vertebrate animals unless oth- erwise noted.

2. The shift from native-ferest ecosystems to artificial ecosystems

It is likely that most of today’s forest fragments have been exposed to a number of influences, many of which may not be appreciated without detailed ecological and historical investigation. Clearing of forests took place for dairy farms, fat-lamb and beef-cattle properties. Drier forest areas were repeat- edly burnt for the “green pick” (Banks, 1989; Pyne. 1991 p. 199ff). Wetter forests were generally unsuitable for grazing and more difficult to burn. Wood- land areas on the inland side of the eucalypt forests

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were particularly suited to grazing of sheep and the growing of cereals.

Forest modification, as distinct from fragmentation, can occur before or after fragmentation. Re- moval of natural materials is a common beginning. Timber, fenceposts, firewood, unusual plants, flow- ers, soil, rocks and gravel are examples of products commonly removed. Other changes may occur due to dumping, grazing and fertilizing. Clark and O’Loughlin (19861, referring to the Lane Cove Val- ley in suburban Sydney, NSW, noted that “Weed invasion and rubbish dumping are obvious in their effect on the bush but the impact of the systematic freelance mining of the bush over the years for everything it has to offer is not as readily recognized. ’ ’ Additionally, feral animals are now present in many fragments. In some cases, existing fragments may bear little or no resemblance to the original vegetation. Calder (1986) observed that on the Mornington Peninsula, a popular Victorian tourist area now supporting many small farms and some remnants of native vegetation, “there is no original ‘bush’ on the peninsula.. [as] all the surviving forests, woodlands and scrubs and heaths are secondary or tertiary communities” (p.36). For example, native communities of Drooping She-oak (Casuarina stricta) and Manna Gum (Eucalyptus uiminalis ssp. pryoriana) have been cleared away and replaced by dense scrubs of native tea tree (Leptospennum laevi- @urn) (p.33). Furthermore, the native Banksia inte- grifoh has been eliminated by insect borers in some areas while an exotic shrub, Polygala myrtifolia, has became rampant in others. These examples highlight the range of influences that may impact on the biodiversity of fragments.

Clearing for agriculture has had regional impacts on forest remnants. Serious salting of the land has resulted from the clearance of forests and other woody vegetation for agriculture particularly in southwestern Western Australia (Macfarlane et al., 1993). In forest remnants in Victoria, up to 30km from the coast, airborne saltspray has caused dieback of trees where the seaward vegetation has been cleared (Offor, 1992). This implies that the various original vegetation types between the now remnant patches and the coast previously intercepted the salt spray, thereby protecting the forest. Another regional “dieback” problem of particular research concern in

eastern Australia in the 1980s was that of “rural dieback” of eucalypts. Recently Davidson and Davidson (1992) considered the interactions between eucalypt performance, improved pastures, eucalypt- feeding beetles (with pasture feeding larvae), other herbivorous insects, cattle and sheep, partial clearing and insect predators, but concluded that “rural dieback is caused by sheep and cattle... not insects or disease” (p.138). Some authors do show that there is mortality of mature trees by insects (Mackay et al., 19841, perhaps exacerbated by agricultural practice, but where seedling establishment is prevented by high stocking rates of sheep and cattle, decline of the rural eucalypt (and other plant species) is inevitable.

In farming areas, pastures have been improved usually by the sowing of subterranean clover and other exotic pasture plants and by the spreading of superphosphate and other fertilizers. Much of the southeastern agricultural area has been improved for grazing by the introduction of exotic pastures (AUSLIG, 1990). While these pastures are not fired intentionally, burning of native pastures has been a regular feature in some parts of the same zone in order to reduce quantities of unpalatable materials and to inhibit weeds, especially woody ones. Where crops have been grown stubble burning may have taken place (Gill, 199 1). At the drier margins of the woodland pastoral zone in many areas of NSW and southeastern Queensland, overgrazing has virtually eliminated the grassy component of the pastures and woody weeds have taken over.

Some farms have been planted to Pinus rudiatu but most softwood plantations have followed forest clearing. It is likely that P. rudiata was introduced in the 1840’s or 1850’s but the first reliable record for P. rudiatu in Australia was in 1857 (Lavery, 1986). P. rudiatu was grown in Victorian nurseries by the 1860’s and in plantations by the 1880’s (Carron, 1985) but much of the growth in area of plantation took place in the 1960’s and 70’s (RAC, 1992b). The rapid expansion of pine plantations coincided with the rapid growth of public interest in conservation. Major concerns associated with the expansion of pine plantations were the “destruction” of native forests and the dubious economics of the venture (Routley and Routley, 1974). Today, most of the 681000 ha of radiata pine grown in plantations in Australia (Lewis and Ferguson, 1993) is in south-

264 A.M. Gill. J.E. Williams/ Forest Ecology and Manugement 85 (1996) 261-278

eastern Australia (RAC, 1992b). P. rudiata is a common tree of windbreaks on intensively managed farms, particularly in southern Victoria. Urban development may follow a farming phase or occur directly in forest. It rarely takes place in pine plantations although this has occurred in Canberra, ACT. The urban area may include houses on traditional “quarter-acre blocks”, retirement-holiday units, semi-rural “hobby farms”, factories and shopping centres. Each has its particular significance in relation to native vegetation at the “urban-wildland interface”. Areas too steep, too wet or too rocky may have been cleared for sporting arenas or become sites for refuse; some end up as conservation reserves. The reserves themselves may then become the sites for drainage works, conduits for power supply and sites for freeways. Our considerations of urban areas is largely focused on suburbs.

There is only a small literature in Australia on how artificial ecosystems function. We introduce this literature here while the interactions between the forest and the artificial ecosystems are discussed later. 1. In urban areas, gardens vary widely in species

composition from mostly exotic to mostly native. Those gardens with more native plant species support more bird species (Munyenyembe et al., 1989). As gardens in Canberra age, the structural diversity, and number of bird species. increases for up to about 12 years from establishment (Munyenyembe et al., 1989). Recher (1972) noted that any lack of vegetative cover by plants in gardens up to a metre in height adversely affected populations of small native birds while the greater frequency of open spaces aided predation of these species by omnivorous (currawong, Strepera uer- sicolor) and carnivorous native birds (kookaburra, Dacelo gigas).

2. In an agroecosystem. Davidson and Davidson (1992) considered many of the interactions between different feeding guilds, including inverte- brates and their habitats in order to approach an optimum mix of native and exotic species on farms.

3. In pine plantations, the diversity of native plant and animal species seems to be influenced strongly and positively by the extent of native and other non-pine vegetation associated with the plantation

(Suckling and Heislers. 1978; Friend, 1982a). Species richness of native mammals (Friend, 1982b) and native birds (Cowley, 1971: Gepp. 1976; Driscoll, 1977; Friend, 1982a) were less in pine plantations than in eucalypt forest. If all growth stages of a pine plantation were considered together, however. the total numbers of species of native birds may be “almost the same” as that of adjacent eucalypt forest (Gepp, 1976). Bird fauna in mature plantations (cf. immature) more closely resemble those of eucalypt communities nearby (Friend, 1982a).

3. Generalized trophic systems for native forests, urban areas, farms and pdaRtations

A diversity of new ecological circumstances in fragmented forests can result from the establishment of new ecosystems at their margins. The new ecosystems, in the form of suburbs, farms and pine plantations, can affect biodiversity-both native and exotic -in different ways. Because of the wide variety of circumstances associated with different forest types, different urban and agricultural ecosystems, different land-use histories etc., we have attempted to capture the essence of native forest ecosystems and the artificial systems that have fragmented them by constructing generalized trophic diagrams.

Trophic diagrams represent the amounts and flows of energy between trophic levels in ecological systems. The trophic levels in our diagrams (Figs. l-4) are the standing crops of vegetation, the litter pool, vertebrate herbivores, vertebrate predators and fire. Fire is considered here to be a general consumer (Gill, 1975) capable of removing the litter layer and part of the standing biomass. The native forest system is regarded as a closed system. The other ecosystems are regarded as open in that they import energy (e.g., animal food) or other inputs affecting energet- its (water, fertilizer) and export materials such as urban rubbish, timber or animal products. Standing crops of vegetation are the bases of trophic diagrams. Proportions of exotic species cf. native are depicted where appropriate. As a convenience, the animal components of the diagrams are in fresh weights (kg ha-‘) while the plant components are in dry weights (t ha-’ ). Another convenience is to scale the values

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Forest

P Bl kg/ha

Pzpredaiors (FW) H.herblvores (!=Wj B=plant biomass (DWI L-litter

L 15ffha

Fig. 1. Generalized trophic level diagram for a native eucalypt forest. The plant materials are scaled according to the logarithm of

dry weight in t ha-’ while the animal matter is scaled according to the logarithm of live weight in kg ha-‘. Occasional flows are indicated by dotted lines. Only above-ground plant parts and

vertebrate animals are considered. ‘F’ is tire.

at each level using logarithms. In the diagrams for the artificial systems (Figs. 2-4), we are concerned with major departures from levels and flows of the native forest system (Fig. 1) in order to explore the effects of the new ecosystems when in juxtaposition with the forest.

While the native forest and the urban systems may be regarded as having relatively stable plant biomasses, the plantations and farms show great life-cycle and seasonal variations in biomasses, re- spectively. We have nominally set the weight of the native forest at 400 t ha-’ and the average weight of

Suburban

m 5Oz/ha* l-1 P=predators (nv) H.herbkores (Fw) kplant biomass (DW) Later

L 2ffha

Fig. 2. Generalized trophic level diagram for a suburban area. The plant materials are scaled according to the logarithm of dry weight in t ha-’ while the animal matter is scaled according to the logarithm of live weight in kg ha- ’ . Black areas indicate exotic

biomass and the hatched areas native biomass. “ex” represents an external source of enrichment. Broken arrows indicate that all the material from the source pool does not flow to the indicated sink alone.

(a) Agricultural

P el kg/ha

p-w WI H=hwbivoma (Fw) Bqbmt biomass (DWJ Ldltter

w Agricultural

P et kg/ha

P=predators (FW) H&erblvores (FW) Bzplent biomass (DW) Later

Fig. 3. Generalized trophic level diagram for (a) a farm with

improved pastures and (b) a farm with native pastures. The plant materials are scaled according to the logarithm of dry weight in t ha- ’ while the animal matter is scaled according to the logarithm of live weight in kg ha- ’ . Black areas indicate exotic biomass and the hatched areas native biomass. “ex” represents an external source of enrichment. Broken arrows indicate that all the material from the source pool does not flow to the indicated sink alone. Occasional flows of energy are indicated by dotted lines. ‘Fire’ is fire.

the exotic plantation, over its life-cycle, at 150 t ha-’ dry weight (after Grierson et al., 1992). The litter accumulation of 15 t ha- ’ is for a native forest in Canberra (Hutchings and Oswald, 1975) while that for the pine plantation is set at 10 t ha-’ for a 35 year old forest in the same area (Jabs, 1988). The standing crop of the exotic pasture is set at a yearly average of 3 t ha-’ (after Moore et al., 1993 for the Southern Tablelands of NSW) with a notional 1 t ha-’ of litter. For the suburban area, the plant standing crop value is set at a notional 5 t ha-’ of native and exotic plants; effective areas of primary production in this system are substantially reduced

266 A.M. Gill, J.E. Williums / Fores Ecology and Mmqqemenr 85 f 1996~261-278

Pines

P P=predators (NV) ~1 kg/ha Hxherbivores (PWJ

Explant biomass (DWJ Later

H 3kg/ha

Fig. 4. Generalized trophic level diagram for a pine plantation.

The plant materials are scaled according to the logarithm of dry weight in t ha- ’ while the animal matter is scaled according to the logarithm of live weight in kg ha-‘. Black areas indicate

exotic biomass and the hatched areas native biomass. “ex” represents an external source of enrichment. Broken arrows indicate that all the material from the source pool does not flow to the

indicated sink alone.

by buildings and roads but trees can add substantially to the weight. Litter values in the urban system are notional at 1 t ha-‘.

Native herbivore biomass for native forest has been set at 30 kg ha-’ fresh weight (with advice from P. Catling, CSIRO Division of Wildlife and Ecology, Canberra). Pine plantations, portrayed as “biological deserts” during the public debates on pine plantations in the 1970s (Cowley, 1971; Routley and Routley, 1974; Friend, 19801, have very low masses of herbivores. Minor nominal amounts are indicated for the biomass of herbivores in gardens and in pines in Figs. 2 and 4. The live weight of exotic vertebrate herbivores in the pasture systems (Fig. 3) has been set at 500 kg ha-’ (about 7 sheep/ha), a more conservative value than that given by Moore et al. (1993) for elite lamb production on the Tablelands of NSW. The point, however, is that this value is an order of magnitude greater than that for herbivores for any other system.

Predators are abundant in urban areas in the form of exotic (i.e., non native) dogs (average fresh weight of 15 kg) and cats (average fresh weight 3 kg) fed from external sources (external influences in genera1 are designated by “ex ” in the Figures). Their mass has been set at 50 kg ha-’ fresh weight on the basis of an informal survey in Canberra. There seems little doubt that this value is vastly greater than that for all the other systems considered here. For example, the

predators in the pasture systems (e.g., foxes, dogs, cats, native snakes) are assumed to have an average biomass per hectare of less than 1 kg (Fig. 31. Predators in gardens and pine plantations are also assigned a nominal weight of < 1 kg ha- ’ in the Figures. People and motor vehicles have not beerr considered as “predators” for the purposes at hand. The large value of the predator biomass in the urban system, cf. the other systems, is the main point to be noted at the predator-level of the diagrams.

Because garden materials and animal and wood products are often transported to sites away from their origins, arrows from the biomass boxes to the litter boxes in Figs. 2-4 are shown with breaks in them.

Fire is shown as a sink of unspecified size for the litter and for the standing crop in the eucalypt forest. Fire is not included in the pine diagram however because plantation managers in eastern Australia aim to exclude it when a crop is present. Burning is not carried out in improved (i.e., with introduced speciesi pastures (Fig. 3a) although it may be in native pastures (Fig. 3b).

The trophic diagrams illustrate the marked differ- ences between the different systems. They indicate biomasses but not the species’ compositions of those biomasses, giving an idea of broad structure but little idea of mechanism. They provide a static picture, not a dynamic one. They give little idea as to how these systems vary locally or how they function at a species level. While our understanding of artificial- ecosystem function at a species level is poor we may use trophic-level generalizations (Figs. l--4) to consider some of the biological interactions between fragmented eucalypt forests and new ecosystems. In addition, there are fluxes of non-biological items across interfaces that require consideration.

4. Interfaces between eucalypt forest and urban, farm or pine communities

Interfaces between fragmented eucalypt forests and artificial ecosystems may incorporate gradients of various steepnesses in ecosystem components. For example, many exotic predators and plants of the urban areas may decline in density sharply in the interface with the forest. On the other hand, birds

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such as native currawongs, may be frequent in the urban areas and show only minor changes in relative abundance at the interface at certain times of year. What may start as a sharp gradient across interfaces can become flatter and more pervasive through the forest as animals (such as the cat, Felis cutus) and plants (such as the indigenous horticultural wattle, Acacia baileyuna) become feral or weedy, respec- tively. In this section we draw attention to the main impacts of the changed land use on the original forest, rather than the reverse. Fire matters are considered later.

Urban areas support large numbers of exotic plant species (despite a modest biomass, Fig. 2), a form of biodiversity inappropriate to conservation of native species in fragments and remnants. Gardens are often extended illegally into surrounding forests, and dumping of garden refuse also occurs there. Further- more, exotic species-and occasionally native species such as the shrub or small tree Pittosporum undulutum (see below)-“escape” from their urban habitats and may spread widely. Native birds such as the currawong eat berries of exotic and native species and spread the seeds. In Victoria and NSW P. undulatum, spread by the introduced European blackbird (Turdus merula) in Melbourne (Gleadow and Ash- ton, 1981) and the native currawong in Sydney (Buchanan, 19891, has become rampant in native forests (Gleadow and Ashton, 1981; Adamson and Fox, 1982; Clark and O’Loughlin, 1986) probably due to reduced fire frequencies (Gleadow and Ash- ton, 1981; Adamson and Fox, 1982). Other dispersal agents of exotics into fragmented forests include people, vehicles and storm water. Enhanced nutrient levels in classically nutrient-poor soils (e.g., Clements, 1983 in Sydney) due to run-off from gardens, roads and roofs and overflow from septic tanks (Bliss et al., 1983) enhance the chances of the establishment of exotics (Clements, 1983). Increased storm flows from the hard surfaces of urban areas can affect erosion and sedimentation in fragmented forests.

Predators abound in urban systems (Fig. lb) in the form of dogs and cats. Other “predators” are motor vehicles, the introduced fox and people. The effects of suburban “predators” vary with circum- stance and species but they extend beyond suburbia into adjacent areas. Dogs, cats and foxes consume a

variety of native and exotic animals and plants including native possums @runner et al., 1991; Kinn- ear, 1993). Cats can transmit diseases that are a threat to native wildlife (see Brunner et al., 1991; Potter, 1991). The Koalas ( Phascolurros cinereus), arboreal folivores, can be attacked by dogs or become the victims of motor vehicles when they leave forest remnants (Koala Preservation Society, Port Macquarie; personal communication). Lyrebirds (Menuru nouaehollundiae) in forest remnants in the Dandenong Ranges near Melbourne-surrounded by “a sea of subdivisions” (Gilpin, 1980)-were increasingly and adversely affected by predators, pre- sumed exotic, up until at least 1980 (Lill, 1980). Wombats (Vombatus ursinus), native herbivores found in the same area, are often killed by cars. A reduction in the number of herbivores may lead to the proliferation of the scrambling native wiregrass, Tetrarrhena juncea, which in turn may lead to a decline in the foraging area of the native lyrebird (D.H. Ashton, LaTrobe University, personal communication).

Two major influences of urban areas on the margins of fragmented forests, identified above, were nutrient enrichment of soils and the establishment of exotics. Agriculture may also cause nutrient emich- ment and the establishment of exotics on adjacent forest areas but the mechanisms are very different. In the agricultural situation nutrients may come from fertilizer or soil drift (Muir, 1979) rather than from runoff or sewerage disposal. Deposition of manure and urine by the herbivores in agricultural systems (Fig. 3) can also contribute to nutrient enhancement if the native forests are unfenced (Hobbs et al., 1993). Higher nutrient levels in soils and leaves of trees may enhance the success of herbivorous insects and cause tree decline (Landsberg et al., 1990). Nutrient enrichment, together with soil disturbance, can promote the establishment of exotics (Hobbs and Atkins, 1988; Hobbs, 1989) spread by domestic stock and other agencies. Unlike the urban situation, “exotics” in this case are likely to be pasture grasses and herbaceous weeds although blackberries are a major problem in many areas (e.g., Cochrane et al., 1962 in South Australia; Goldney, 1987a p.8, in New South Wales).

Many studies have drawn attention to the adverse impact of grazing by domestic animals on conserva-

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tion value of remnant forests and woodlands in southeastern Australia. The main impacts are considered to be: the increased abundance and richness of exotic plants (Bennett, 1990; Prober and Thiele, 1993); and the decline in species richness of native plant (Bennett, 1990; Prober and Thiele, 19931 and animal species (Bennett, 1987; Goldney. 1987b). “Grazing’ ’ , like fire, is more usefully described in terms of regimes, that is the intensity, frequency and season of grazing. Unfortunately these details are often unavailable, thus limiting comparisons between studies. However, Bennett (1990), was able to refer to the effects of “sustained grazing” on remnant forests in southwestern Victoria as: “characterised by an abundance of introduced grasses and weeds and a paucity of native shrubs and herbs. Such disturbed habitat is unsuitable for [a number of] small terrestrial mammals.. that depend upon the natural forest understorey for food and shelter.”

Herbicides have become increasingly important in agricultural systems in southeastern Australia in recent decades, particularly in cropping areas where there has been a move toward minimum tillage systems. With the increased use of herbicides comes the increased risk of their drift beyond cropping boundaries. Fensham and Kirkpatrick (1989) considered the use of herbicides as “particularly haz- ardous” to the viability of native, especially grassy, vegetation of small reserves in the agricultural Mid- lands region of Tasmania.

Pine plantations are depauperate not only in animals in general (Fig. 1) but also in native plant species. In the pine forests of the ACT relatively few species of plants are found, many of which are exotics. Effects of pine plantations on eucalypt forest remnants are mainly in the form of pine spread (see later, below) but fertilizer drift does occur and herbicides may also have an impact. Stream flows may be reduced. Some pine plantations may support populations of large herbivores such as kangaroos, walla- bies and wombats which could affect margins of eucalypt forest communities by their herbivory.

Fragmentation of native forests necessarily leads to an increase in the amount of forest edge. This can result in either temporary edges, such as in a clearcut, or maintained edges, as in an agricultural field. Edges typically have a different microclimate (Chen et al., 1992; Young and Mitchell, 1994) and can

support different species than forest interiors, a phenomenon known as the “edge effect”. Overseas studies (Ranney et al., 1981; Brothers and Spingam, 1992; Fraver, 1994) have demonstrated different suites of plant species in forest edges and that edge- oriented species may have higher reproductive out- put. which over time could lead to changes in regional plant species composition. Little evidence has been found for distinct edge formations of native species in vegetation fragments in the wheatbelt in Western Australia (Hobbs et al., 1993) or open forests in South Australia (Taylor. 1994). This may be because of the relatively recent isolation of the remnants compared to the Northern Hemisphere, but more likely that light is generally not a limiting factor in, at least, the wheatbelt vegetation compared to northern forests (Hobbs et al.. 1993). The abundance and cover of non-native species, however. were consistently greater at the edge of remnants in Western Australia (Hester and Hobbs. 1992) and South Australia (Taylor, 1994). Woodland communities in WA were generally more susceptible to invasion than shrublands (Hobbs, 1989; Hobbs and Atkins, 19881, a phenomenon thought to be related to higher soil nutrient levels in this vegetation type.

Edges can also be a focus for predators, especially in agricultural areas. Several studies in the northern hemisphere have demonstrated elevated predation rates, at least of birds, near edges of habitat islands (Angelstam, 1985; Wilcove, 1985; Andren and An- gelstam, 1988). In Western Australia, the interface between forest or shrubland and pasture is the main habitat of the introduced rabbit (Orycrolagus cctnicu- /us), the introduced fox’s (Vulpes uulpes) primary food item (Christensen and Burrows, 1986). Conse- quently foxes are often concentrated in these areas and make nightly forays into forested areas to prey on native animals. Whether or not the predation causes the local extinction of the prey species is dependent on a number of factors including the population density of the predator and prey.

Fluxes of nutrients, exotic plants, water and animals across the margins of artificial ecosystems into native forests may all have an adverse impact on native-species biodiversity, in particular. or conservation values, in general. Fluxes are greatest at the margins suggesting that edges should continue to be a major focus of attention for the conservation of

A.M. Gill, J.E. Williums/ Forest Ecology and Management 85 (1996) 261-278 269

remnants and fragments of native forests (Harris, 1988). As well as size, shape and orientation of fragments, the type of forest ecosystem, position in the landscape, topographic features and nature of the surrounding matrix will contribute to the dynamics of forest remnants and fragments.

5. Fire-biodiversity scenarios

Fires have had a ubiquitous occurrence throughout southeastern Australia and, indeed, Australia in general (Gill et al., 1981). Their incidence has varied widely in time (Pyne, 1991) and space (Walker, 198 1). To protect forest values, lives and built assets, large-area prescribed burning assisted by aerial ignition (Cheney, 1978) was introduced in the 1960’s. By reducing fuels using fires burning under mild weather conditions, potential fire intensities are also reduced thereby enhancing the chances of successful fire control. The widespread use of prescribed bum- ing over the past 30 years has fuelled a continuing debate on its effectiveness for the protection of life and property, on the one hand and on its impact on conservation values, on the other (Gill and Brad- stock, 1994). There are many facets to the debate. Because landscape fire can have positive and nega- tive effects at various times on attributes of remnants, farms, plantations and urban interfaces, it has an ambiguous image.

Each fire has effects that vary according to: the plants and animals present and their life stages at the time of the fire; spatial patterns of fire intensities and unburnt ground (Gill and Bradstock, 1995); the context of the burnt area (e.g., remnant); and post-fire weather (e.g., Specht, 1981). Realized biodiversity will be influenced by the complement of species at a site, past fire regimes, fragmentation of the vegetation, regional diversity of exotics and land-use contexts of the area.

Of particular significance to the biodiversity of animals in Australian forests is the need of many species for tree hollows. These are usually lost in plantations, urban areas and in intensively managed agricultural systems. Approximately 18% of Aus- tralian species of land birds use hollows for nesting some of the time while 11% have an obligate re- quirement for them (Saunders et al., 1982). In a

comprehensive review of native fauna known to utilise tree-hollows in southern temperate eucalypt forests, Gibbons (1994) listed 13 species of arboreal marsupials, 9 species of scansorial mammals, 16 species of microchiropteran bats and 46 bird species. Competition for hollows may be increased in the presence of introduced species of birds, particularly in urban areas.

Fires have an impact on hollows by destroying them or by causing scars that precede them. An increase in the occupancy of hollows by the Com- mon Brushtail Possum (Trichosurus uulpecula) after fire was noted by Inions et al. (1989), although they were unable to determine if possums were attracted to epicormic growth or were responding to an alter- ation of hollow configuration. Fires under particularly dry conditions may allow hollow trees to be ignited while windy weather may allow fires to completely destroy trees. Fragmenting agencies and associated changes to the fire regime are likely to have an impact on tree hollows and their associated fauna.

Fire histories in southeastern Australia are inextri- cably linked to the human history of the region. Most contemporary fires are caused directly or indirectly by people although lightning-caused fires are still a most important factor in country areas (Dovey, 1994). Unplanned fires in southeastern Australia are usually vigorously suppressed. Management-initiated fires bum large areas. Thus considerations of fires and their effects in southeastern Australia invariably in- volves considerations of fire management either through fire suppression or through fire ignition. To provide a focus, we have formulated a number of scenarios (or compound hypotheses) which we ad- dress below.

5.1. Urban

Our urban-interface scenario is that: There is a low frequency of unplanned fire in forest remnants. To prevent losses of life and property in adjacent urban areas, regular frequent prescribed burning is practiced. Regular frequent prescribed burning reduces biodiversity.

That fires can create a danger to lives and property at the urban interface is an historical fact in southeastern Australia. The most recent evidence for

270 A.M. Gill. J.E. Williams/ Forest Ecology md Manugrmrnt 8.5 (1996) 261-278

this was the occurrence of fires in Sydney in January 1994. There are many other examples applicable to Sydney, Melbourne, Adeiaide, SA, and Hobart, Tas- mania (e.g., Cheney, 1979). In the 1994 Sydney fires (Gill and Moore, 1994) forest remnants only a few hundred metres wide carried fire into houses at suburban Como-Jannali. Oft-repeated calls for more prescribed burning followed the losses of life and property there.

While the recent Sydney example is only one of a long series of similar examples, remnant forest in urban areas have not all been frequently burnt as a result. Corbett (1972) noted that Cumberland State Forest in Sydney, a 39 ha reserve, had not been burnt in historical time (to 1972), nor has this changed in the last 20 years except for a modest strategic bum- ing program of recent introduction (A.L. Yates, personal communication). This State Forest reserve is one part of a larger remnant under the control of a number of authorities (see Forestry Commission of New South Wales, 1984). From studies of the vegetation of another Sydney reserve, Adamson and Fox (1982) concluded that there had been a “reduction of fire”. Clark and O’Loughlin (1986) noted a lower contemporary fire frequency in yet a third Sydney remnant, Lane Cove National Park. Gleadow and Ashton (198 1) suggested that the invasion by P. undulatum into forest remnants in Melbourne was assisted by a lower fire frequency. Variations and departures from this theme emerge in other areas. Petersen (1983) reported the use of limited-area prescribed burning, trittering and slashing in small reserves in the upper Lane Cove Valley. presumably above the National Park. In Queen’s Domain, a woodland remnant in Hobart, Tasmania, Kirkpatrick (1986) found mowing and burning to be fuel-reduction measures while unplanned fires were also reported. In King’s Park, Perth (Western Australia)-a 250 ha woodland remnant-both prescribed and unplanned fires have been frequent (Baird. 1977). Fuel-reduction burning in the 1950’s and 1960’s and a severe wildfire in 1989 were thought to have accelerated the conversion of some of the vegetation in Kings Park from a eucalypt-dominated to a sheoak-dominated woodland (Bell et al., 1992). In parts of Sydney’s National Parks fires too frequent to maintain all native plant species have occurred (Sid- diqi et al., 1976). Prescribed burning is now an

integral part of the management of Sydney’s larger National Parks (e.g., New South Wales National Parks and Wildlife Service, 1994) and in recent years some smaller remnants of native vegetation have also come under the jurisdiction of the New South Wales National Parks and Wildlife Service.

The frequency of unplanned fires may be affected by the risk of ignition (from people) and the ability to suppress fires. On a point basis, ignition may be considered to have taken place if a fire. ignited elsewhere. spreads to the point of concern. Thus areas burnt as well as numbers of ignitions are important in deciding the causes of changes in fire frequencies across the landscape. We have no data on the size of fires in suburban remnants but. because of the availability of fire-fighting facilities and water, we may assume that fires will be small on most occasions provided that there is ready access to the fires.

There are few data on number of fire starts per unit area but McRae (1994) has modelled them against distance from the urban interface in Can- berra. At the urban edge, the modelled density was 1.4 fires kmm2 yr--’ while at lkm, there were 0.2 fires kmm2 yr-‘. Many fires near houses bum areas less than a hectare in size (A.M. Gill, personal observation).

If frequent prescribed burning is introduced in forest fragments in an urban landscape changes in biodiversity may be expected. Because there are correlations between plant community structure and the abundances and richness of birds (Recher, 1969) and small mammals (Newsome and Catling, 1981). we can assume that the simplification of forest structure by oft-repeated prescribed burning (Catling, 199 1) will reduce the numbers and diversity of verte- brates. Such a conclusion is strengthened by the observation that reductions in cover will favour predators and that predators are numerous at urban edges (Fig. 1). It is not expected that the fires per se will affect vertebrate populations directly, but rather that changes in habitat in conjunction with the abundance of predators will have the major impacts. Further, it is thought that exotic animal species are advantaged by the simplification of forest structure (Catling, 199 1).

Dominant plant species have an impact on the species richness of the rest of the plant community

A.M. Gill, J.E. Williams/Forest Ecology and Management 85 f 1996) 261-278 271

(Specht and Specht, 1989). This effect is sometimes alleviated by fire which may reduce the cover of the dominant and allow the expression of species present as seed or other propagules in the soil-seed pool. If the dominant species persist at high cover values, it is thought that some species may die out (Adamson and Buchanan, 1974, Adamson and Fox, 1982; Keith and Bradstock, 1994). In the enhanced soil nutrient conditions of forest remnants in Sydney, a number of exotic plant species are important dominants. These include the prostrate garden weed Tradescantia albi- j7ora and the garden shrubs, Ligustrum spp. (Adam- son and Buchanan, 1974; Adamson and Fox, 1982). These and the native shrub Pittosporum undulatum are out-competing native species, the Pittosporum apparently leading to the demise of Xanthorrhoea in parts of Sydney (Adamson and Fox, 1982) but perhaps a temporary effect in the first instance on understorey species of remnants near Melbourne (Gleadow and Ashton, 1981). Because spread of Pittosporum seems to be associated with a reduction in fire frequency (Gleadow and Ashton, 1981; Adamson and Fox, 1982) artificially increasing fire frequency seems a reasonable counter measure. However, being a gully species in the Sydney area, the conditions when such vegetation may bum may be more severe than those usually desirable for fuel-reduction fires. Hand weeding by community groups is an option for removing Pittosporum and woody exotics in remnants (Adamson and Buchanan, 1974; Adamson and Fox, 1982). This option is inappropriate where the soil-seed pool of native species requires stimulation by fire to persist.

of the vegetation in some cases (e.g., grasses in King’s Park, Perth, reported by Baird, 1977 and the broom and gorse in the Adelaide Hills reported by Cochrane et al., 1962) thereby leading to increased fire frequencies. It appears that fire management of remnants in urban areas will have to consider the nature and extent of invasions by exotic plant-species and their tolerances to fire regimes in relation to those of the native species present-along with the many other considerations necessary in such critical situations.

Forms of fuel reduction such as raking, slashing, trittering, weeding and mowing may be considered appropriate in certain situations. The use of machin- ery is sometimes limited by the terrain. “Mowing” of grassy woodlands in Hobart seemed to have an adverse effect on plant-species richness while “bum- ing” enhanced it (Kirkpatrick, 1986). The perma- nency of such changes is unknown. Petersen (1983) implied that hand removal of plants had less adverse impact on the plant biodiversity of some Sydney vegetations than prescribed burning (at an unspecified frequency). She noted that hand removal of plants was a practical way of reducing the shrub component of the fuel in strategic locations while hand raking without exposure of the soil reduced the litter component.

Keith and Bradstock (1994) suggest that a uari- able fire frequency, within set limits, is necessary if plant species diversity is to be maintained in heathlands of native species in the Sydney area. If fires are too frequent, some common species like the shrub Banksia ericifolia will be eliminated from Sydney vegetation (Siddiqi et al., 1976). If fires are too rare, over-dominance and species loss may result.

There was considerable support for the scenario outlined at the beginning of the section but it would be unwise to assume that this scenario is general. The few data available suggest that there may be more fires (management initiated or unplanned) in larger fragments than in smaller ones. We expect fewer planned fires in wetter eucalypt forests. The fires of January of 1994 in Sydney promoted a wave of prescribed burning in Sydney itself and elsewhere. Perhaps these prescribed fires are the forerunner of a more frequent application of fire to Sydney’s forest remnants.

5.2. Agriculture

There is no easy solution to the questions related The agricultural-interface scenario is that: Clear- to the fire management of forest fragments in urban ing for agriculture leaves only small forest remnants areas even if conservation of native species biodiver- which may become fire free. Fire-free fragments sity was the only objective. Exotic plants are fire eventually decline in plant species richness. tolerant in many cases so not easily removed by fire Rural landscapes may support fragments of euca- regimes. Exotic plants can also increase flammability lypt forests and woodlands on roadsides, within farms

272 A.M. Gill, J.E. Williams/Foresr Ecology and Munagemenr 85 (1996) 261-278

and in “Travelling Stock Reserves” (or “Camping Reserves” in Goldney, 1987b) as well as on reserves designated for conservation. Even small cemeteries may support native plant populations of significance (Prober and Thiele, 1993). With such a diverse range of ownership (and therefore management objectives), more diverse circumstances than those in the scenario seem likely. We have already alluded to the burning of native pastures by farmers in some areas but there is general avoidance of the practice once pastures have been improved using exotic pasture species (e.g., in the New England area of New South Wales; Mackay et al., 1984). Fensham and Kirk- patrick (1989), referring to the Midlands of Tasma- nia, reported an “animosity to burning in agricultural areas”. In an area of both exotic and native pastures in Western Victoria, formerly supporting forests. woodlands and grasslands, Trabaud et al. ( 1993) estimated that the average fire-return interval was longer than the period of European settlement. Based on known large fires only, this figure should be regarded as being indicative of a decline in fire frequency rather than being definitive.

Protection of pastures and other farm assets from the spread of unplanned fire may lead to a local policy of fuel reduction by burning on road and railway easements. Such burning has been traditional along railway lines but this has stopped where lines have been closed. However, frequent burning may have had particular significance for the conservation of plant species in woodland and grassland situations given the importance of the role of frequent fire in reducing dominance of the native grass Themedu australis (reviewed by Lunt, 1991).

In NSW, travelling stock routes have been a feature of rural landscapes for decades (McKnight. 1977). Along these stock routes blocks of native vegetation have been set aside (known as “Reserves for Travelling Stock”) where stock could be held overnight. As road transport has replaced droving and farming methods have improved, the stock reserves have declined in importance for agriculture but have gained in importance for conservation (Hib- berd, 1978; Prober and Thiele, 1993). Many of these reserves, scattered throughout NSW (McKnight, 1977; Goldney, 1987b p.260) were periodically burnt (Hibberd, 1978; A.M. Gill personal observation).

Within farms there seem to be a range of actions

in relation to fires. In a remnant area of woodland on the Central Tablelands of NSW, a conservationist- farmer has a long unburnt area, fenced from stock. within his farm (Breckwoldt, 1983). City people with “bush blocks” sometimes duplicate this situation for their entire, but small, properties. In the Adelaide Hills. Cochrane et al. (19621 noted the historical complexities that can arise. There, most landowners in this mixed farming area kept horses for cartage and cows for milking before 1930. These animals kept in fields surrounding “gardens or orchards” controlled scrub growth in conjunction with burning off. With the advent of mechanization, however “exotic scrub now flourishes in many once well- grazed fields”. The effects of burning plus grazing can have a marked effect on plant species populations and even persistence (e.g., Leigh and Holgate, 1979). With a preponderance of herbivores in agricultural systems (Fig. 31, including the introduced rabbit, adverse effects on native plant species may reach extremes.

Reserves for conservation may be few and occupy only small areas in the agricultural areas because the soils have been fertile and arable and cleared for farming. The fire histories of such reserves have rarely been documented. One exception, however, is a Victorian reserve where there has not been a major fire for “at least 90 years” (Withers and Ashton, 1977). There, a shrubby woodland originally dominated by Eucalyptus ovuta is being taken over by small trees of Casuarina spp. As the eucalypts are dying out, they are not being replaced. With the reintroduction of fire, however, regeneration of eucalypts may possibly recur (Withers, 1978).

While there is support for the scenario outlined above, there are many different fire histories present in the different forms of rural remnants. such as those on roadsides, Travelling Stock Reserves, and on farms. Moreover, these histories are likely to vary in the different social and environmental contexts in which these reserves are found.

5.3. Pinus radiata plantations

The pine-plantation interface scenario is that: Pines spread to neighbouring forest areas, markedly reducing native plant and animal species diversity. This situation can be reversed by a prescribed bum- ing regime that has fires intense enough to cause

A.M. Gill. J.E. Williams / Foresr Ecology and Management 85 (1996) 261-278 273

pine death and frequent enough to prevent pine seed set.

In many areas of southeastern Australia pines are spreading into surrounding eucalypt forests. Exam- ples are to be found in South Australia (van der Sommem, 1978 quoted by Burdon and Chilvers, 1994); Victoria (Mink0 and Aberli, 19861, the ACT (Burdon and Chilvers, 1977; Dawson et al., 1979) and NSW (A.M. Gill, personal observation). In some areas at least, pines are beginning to dominate the eucalypts, prevent eucalypt recruitment to the canopy class and create a deep shade unknown in the original forest (Burdon and Chilvers, 1994). The deep shade is likely to have an adverse affect on the understorey plants while the dominance of pines is likely to reduce the abundance and species richness of the fauna as discussed above. Without interven- tion, permanent changes detrimental to the native flora and fauna of the community seem likely.

Where plantations are adjacent to the wetter types of eucalypt forests, pine invasion is “minimal or absent” (Chilvers and Burdon, 1983). Pine invasion into forests will be at its most serious in the “dry sclerophyll’ ’ type studied by Burdon and Chilvers [;;;;i, Dawson et al. (1979) and Minko and Aberli

Pines can disperse widely. Minko and Aberli (1986) found seedlings up to 1.5 km from their likely source. When the plants produce further seeds, spread can be extended. Wind is the most likely agent of spread (Mink0 and Aberli, 1986) but bird dispersal is a possibility as cockatoos have been observed carry- ing cones in flight (B. Gepp, personal communication). The spread of pines into eucalypt forests appears relatively recent despite their wide dispersal capabilities and the fact that pines have been grown in southeastern Australia for over 100 years. Dawson et al. (1979) suggested that the decline of both wildfires and rabbits in the 1950’s resulted in reduced pine control.

Pinus radiata is fire sensitive and has a canopy seed store only, making plantations an asset that can be readily destroyed by fire. Management authorities in eastern Australia seek to prevent fire in P. rudiatu plantations unless it is to bum slash after harvesting. Protection of the plantations may involve prescribed burning of surrounding forests for litter reduction. Fuel-reduction fires may inadvertently reduce the

spread of pines into adjacent eucalypt forests. How- ever, the frequent nature of these fires could itself impact on forest diversity.

Pryor (1991) suggested the planned use of fire was an inexpensive and effective way to achieve control of feral pines. Burning before a canopy seed-store has been initiated would seem essential. Studies in the ACT (Burdon and Chilvers, 1977) suggest that it took about 12 years for seedling pines to mature while the equivalent time in northeastern Victoria was about 21 years (Mink0 and Aberli, 19861. Repeated burning at these intervals or less should prevent the establishment of reseeding pines if the fires are intense enough to kill the pines. To kill the pines, by fire, the foliage should be completely killed (scorched) or the plants ring-barked at the base. With such damage, they have no viable buds from which to resprout.

Burrows et al. (1989) found that a fire intensity of 200 kW m- ’ , which is within the limits set for safety (Cheney, 1978), would kill pines up to 10 m tall. If it took 10 years to reach this height then a IO-year interval between management fires would suffice to prevent further spread of the species. Fires at this interval, however, may not satisfy fuel-reduction objectives.

The challenge is to implement a fire regime around pine plantations that will protect the fire-sensitive plantations, stop the spread of pines into the eucalypt forest and maintain forest biodiversity. Frequent fuel-reduction burning may not be the most effective approach. Despite precautions, unplanned fires have caused millions of dollars worth of damage to pine plantations in southeastern Australia (e.g., Keeves and Douglas, 1983). Additionally, in some forests, frequent, low intensity fuel-reduction fires have been implicated in a reduction in the richness of native animal species and an enrichment of that of exotic animals (Catling, 1991). Therefore, while fuel-reduction fires may eliminate pines, they may also affect forest biodiversity. If, however, no fires occur pines may eventually dominate and reduce the abundance of native plants in the area. Fires at intervals that prevent pines from forming a canopy seed-store should control the spread of pines into adjacent eucalypt forests.

The scenario outlined at the outset of this section is supported to the extent that pines have been

274 A.M. Gill. J.E. Williuvw/Forest Ecology and Managemrnr 8.5 (1996) 26/-278

observed spreading into a number of forests in different parts of southeastern Australia and that it is likely that these invasives can be controlled by a program of management burning. If allowed to persist, the abundance of pines will increase while eucalypts will decline (Burden and Chilvers. 1994). Given the effects of plantations on animal species richness, it is unlikely that pine stands resulting from invasion of eucalypt forest will retain the same diversity of animals and plants found before invasion. There has, however, been no study of this phenomenon.

6. Discussion and conclusions

Landscapes in southeastern Australia, forested or wooded with eucalypts in the past, are now fragmented by many agencies including urbanisation, agriculture and pine plantations. These three fragmenting land uses vary widely in their extent. Urban- ization has mostly begun near the coastal fringe and extended inland; agriculture has spread through the woodlands and into forests, its inroads into forests restricted largely by impoverished soils, steep slopes or non-arable soils; pine plantations have commonly replaced forest. Forests and woodlands remain in places as remnants (enclosed by suburbs, farms or pines) or fragments (with portions of their perimeters occupied by suburbs, farms or pines). These remnants and fragments are sometimes managed for their conservation value, sometimes not.

Farms, forests, plantations, cities and the easements between them are not isolated landscape elements. Native plants and animals may occur in “artificial” environments as well as “natural” ecosystems. Interactions occur between “natural” and “artificial” systems. Without a suitable framework, the study of the interaction between landscape elements has remained problematic. The trophic diagrams outlined in the first section of this paper represent one attempt to formalise the description of landscape elements. Using this framework, components can be compared and contrasted and interactions between systems examined. The spatial arrangement and context of the landscape elements will also be of relevance. For example, the proximity of a remnant to a more extensive forest fragment

may be important for the dispersal of plants and animals.

Fires do not respect boundaries between land uses. Neither do pines, cats. birds or drainages. People and vehicles move between and through all zones. Indeed roads and tracks represent another form of fragmentation, not discussed here, that can disassociate habitat and provide corridors for the movement of predators and the dispersal of weeds. Furthermore, management costs associated with forest interfaces are disproportionately high cf. the interior of remnants. This is especially true where the interface is created by urban development. Planning so that the juxtaposition of landuses optimizes cross-border values is therefore desirable. Similarly. management in an integrated fashion-across landuses and jurisdictions-is desirable.

The conservation of biological diversity is receiv.- ing increasing attention within government and the general community. At the international level Aus- tralia has ratified the “Convention on Biological Diversity” which deals at the global level with the full range of biological diversity conservation. its sustainable use and the fair and equitable sharing of the benefits arising from this use. At the national level the National Strategy for Ecologically Sustain- able Development (Commonwealth of Australia. 1992a), Intergovernmental Agreement on the Envi- ronment (Commonwealth of Australia, 1992bl and the recently released draft National Strategy for the Conservation of Australia’s Biological Diversity (Commonwealth of Australia, 1994) emphasise the conservation of biological diversity and maintenance of ecological processes and systems. While these agreements and strategies do not have a legislative basis. signatories (both State and Federal Govem- ments) commit themselves to their implementation, This paper suggests that for biodiversity conservation in the eucalypt forests of southeastern Australia. an important biodiversity resource, attention must be paid to the landuse context of the remnant or fragment and to the numerous issues surrounding the management of fire regimes.

Acknowledgements

Denise Elias diligently dug for references in the library, a great help. We thank Tony Norton who

A.M. Gill. J.E. Williams/ Forest Ecology and Management 85 (1996) 261-278 275

was particularly encouraging. Eddie Pook and Peter Moore provided logistic and moral support.

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