Proceedings of the conference held 8-9 October 1994, Footscray, Melbourne
Biodiversity Series, Paper No. 8
Department of the Environment, Sport and Territories, 1996
3. The impact of fire intensity on litter loads and understorey floristics in an urban fringe dry schelrophyll forest and implications for management
Robyn Adams and Dianne Simmons
Applied Australian Ecological Research Unit,
School of Aquatic Science and Natural Resources Management, Deakin University
At the urban-forest interface in south-eastern Australia, there is often a conflict between the management of vegetation to reduce fire hazard, and management of vegetation for conservation. In this study, the accumulation of litter and the floristic composition of the understorey were monitored at two sites where fuel reduction burning had been carried out at low intensity and at moderate intensity. Litter accumulated more rapidly following the low intensity fire, and at both sites it had reached hazardous levels in less than five years. There were significant differences in vegetation floristics following fires of low or moderate intensity, and these observed floristic differences have implications for the conservation of native vegetation. Traditional methods of fire prevention such as fuel reduction burning continue to be emphasised, in spite of a considerable body of evidence which suggests that education of the public and emphasis on householder responsibility can be more effective in reducing the losses of lives and assets during wildfire.
Keywords: fire intensity, fuel dynamics, fuel reduction burning, understorey floristics, urban-forest interface.
Open forest dominated by eucalypts, or dry sclerophyll forest (Christensen et al. 1981), is the predominant vegetation type at the urban-forest interface surrounding the major cities of Melbourne, Adelaide and Hobart in south-eastern Australia. Losses of lives and assets due to wildfire in these areas have been high, and past strategies for wildfire control have relied primarily on fuel reduction by prescribed burning (McArthur 1966; Williams et al. 1994), though information about forest fuel loads at the urban-forest interface is limited (Gill et al. 1987). In the fragmented remnant forest on the urban fringe, conservation values can be high. Fire management which emphasises the use of prescribed fires to maintain low fuel loads, is controversial due to the potential to result in unacceptable ecological impacts (Boura 1994; Catling 1991; Fensham 1992). This study examines the effects of two fires of low and moderate intensity on litter accumulation, and the floristics of understorey vegetation in stringybark/box forest, one of the extensive forest types in Victoria experiencing wildfires more commonly than might be expected (Tolhurst et al. 1992).
The application of fires prescribed primarily for fuel reduction in areas with high conservation value is controversial because of the possible long term ecological impacts of fire regimes. These regimes are believed to differ from pre-European or past natural regimes (Adamson & Fox 1982), though it should be noted that there are inherent contradictions in the concept that pre-European Aboriginal cultural practices are synonymous with natural processes (Taylor 1990). There have been many changes in the Australian vegetation associated with the degrading processes of European settlement. In the forested areas of south-eastern Australia it is not clear whether pre-European fire regimes involved relatively frequent low intensity fires, or less frequent higher intensity fires (eg Adamson & Fox 1982; Clark 1983; Head 1989; Meredith 1988).
The impacts of prescribed fires may be significant for the long-term management of reserves set aside for conservation (Catling 1991;Williams et al. 1994). Gill (1975, 1977) suggests that the effects of fire should be considered in terms of a fire regime, made up of the components of fire frequency, intensity and season. If prescribed burning is part of the management strategy of an area, especially in reserves set aside to conserve plant communities, the effects of all components of a fire regime need to be understood. Plant species and communities may have responses to fire which are more site-specific than has been appreciated in the past (Williams et al. 1994), and data from many sites and vegetation types needs to be accumulated to allow generalisations about fire effects.
Effects of intensity are of particular interest, because most fuel reduction burning is prescribed at low intensity (<350 kW/m), whereas most wildfires are of moderate or high intensity (4000-60000 kW/m) (Cheney 1981; Gill et al. 1987); most prescribed high intensity fires are associated with slash burns (Williams et al. 1994).
This study considers the effects of fire intensity on litter accumulation and initial post-fire floristic composition of the understorey. The effects of intensity, season and frequency have been reported for a number of vegetation communities (Catling 1991; Gill et al. 1981), but the effects of the different components are often difficult to separate because of their interaction. For example summer-autumn fires are often more intense than spring fires, due to climatic conditions and dryness of the fuel; and in areas subject to frequent fires, fuel loads may be lowered, which would also reduce intensity (Fox & Fox 1986). The effects of fire intensity can also be confounded with the seasonal effect of post-fire seedling establishment conditions (for example Shea et al. 1979).
3.3.1 Site description
The study site covers about five hectares, located in dry sclerophyll forest approximately 40 km to the north-east of Melbourne, Victoria. The yellowish gradational soils in the area are weakly-structured and shallow, with stony loam surfaces, and bedrock is generally encountered before 100 cm. Soils are acid, with pH 5, low in organic matter, and low in nutrients, especially phosphorus. Annual rainfall is about 700 mm, and the last major fire occurred in January 1962 (Barber et al. 1984). The overstorey is dominated by Eucalyptus macrorhyncha, E. goniocalyx and E. polyanthemos.
The site is part of a wider study of litter dynamics in the area, and was initially expected to be burned as one unit. The site was initially assessed as homogenous and sampled as one unit for litter load and floristics prior to burning. The area was subsequently burned in two sections due to time and resource constraints; as a result of significant differences in climatic conditions on the days of burning, the two sections burned at low and moderate intensities. Opportunistic sampling of this site allows some inferences about the impacts of low and moderate intensity on floristics and litter accumulation at this site to be made.
The site is bisected by a narrow cleared easement, and fuel reduction burns were carried out either side of the easement one week apart in April 1985. On one side of the easement, the fire was of moderate intensity (approx. 2000 kW/m), and on the other side, it was of low intensity (approx. 350 kW/m), due to different climatic conditions at the time of burning. On the low intensity site, flame heights did not exceed 0.5 m, there was no crown scorch, and some small patches of vegetation did not burn. On the moderate intensity site, flame heights were approximately 3m, all understorey vegetation was reduced to ash, and there was severe crown scorch, followed by leaf fall several weeks later.
3.3.2 Litter levels
Equilibrium (steady-state) litter levels were estimated by sampling litter (fine fuels <6mm diameter for woody material) down to mineral soil immediately before burning in April 1985 using the method of Simmons and Adams (1986). Litter samples were then collected from ten 0.5 x 0.5 m quadrats randomly located along a 30 m transect placed near the centre of the burned areas in February 1988 (2.7 years after burning), in November 1989 (4.5 years after burning), February 1990 (4.7 years after burning), and in November 1992 (7.5 years after burning). Litter sampling was carried out in both spring (November), when litter levels would be near their seasonal minima, and summer (February), when accumulated litter levels would be near their seasonal maxima (Woods & Raison 1983). All samples were oven dried at 105°C for 24 hours, and sorted into components: leaves (intact and fragmented), woody material (sticks <6 mm diameter), graminoid material (grasses and lilies) and comminuted fragments, which are largely decomposing leaves and humus (Birk 1979). Species present at the site shed minimal amounts of bark, and bark in the samples was negligible. Differences between litter samples at the two sites were tested using a t-test (Sokal & Rohlf 1969). Accumulation of litter was modelled from the equations of Raison et al (1983) using the FRISK – Fire Risk Assessment Package (Simmons et al. 1988).
3.3.3 Floristic composition of the understorey
Overstorey species were apparently homogenous over both sites, and only the understorey species were used in subsequent floristic analyses. The understorey species were surveyed immediately before burning, and the sites were monitored following burning. A detailed floristic survey was carried out in October 1987 (2.5 years after burning), November 1989 (4.5 years after burning), and October 1992 (7.5 years after burning). Results of analyses of floristic data collected after 2.5 and 4.5 years only are presented, as senescence of some Acacia and other species had commenced at 7.5 years. Floristic change associated with time since fire will be reported elsewhere.
Sampling was timed to coincide with the peak flowering period of the smaller herbs and seasonal geophytes (mainly orchids and lilies) which are abundant in the area. Floristic sampling was initially carried out using a stratified design, but as the initial analysis indicated homogeneity within the burned areas, subsequent sampling was limited to ten randomly located circular quadrats of 3 m radius sampled at each site. This quadrat size was chosen following construction of species/area curves for the sites. All quadrats were located at least 20 m from the fire boundary to minimise edge effects. In each quadrat, cover values for litter, bare ground and each species present were visually estimated using the Braun-Blanquet scale (Kershaw & Looney 1985). A careful search was then made in the area and several species represented by only one or few individuals, and not sampled in quadrats, were located. Nomenclature follows Ross (1993).
All species were categorised according to their primary mode of regeneration after fire, using Purdie (1977), Fox and Fox (1986), Wark et al. (1987) and Molnar et al. (1989). Differences between total cover, and cover of seed and vegetative regenerators following fires of low and moderate intensities, were tested using the Wilcoxon test for paired samples. The Wilcoxon two-sample test (Hodges & Lehmann 1970) was used for comparisons of cover values for bare ground and individual species. The log-likelihood ratio (G statistic) for contingency tables (Sokal & Rohlf 1969) was used to test departures from proportional representation of the numbers of species from each regeneration type in the areas following low and moderate intensity fire.
Patterns within the floristic data were investigated using both classification and ordination techniques, using the PATN pattern analysis
package (Belbin 1991). Floristic data for understorey species in each quadrat were clustered using the Bray-Curtis association measure with fusion using Ward's method, an intensely clustering technique (Belbin 1993).
|Time since fire
|Predicted litter load
|Actual litter load
on low intensity site (t/ha)
|Actual litter load
on moderate intensity site (t/ha)
|2.7 (seas max)||10.1||9.9 (0.63)||4.1 (0.71|
|4.5 (seas max)||13.8||8.6 (0.71)||6.8 (0.72)|
|4.7 (seas max)||14.2||13.9 (0.54)||9.1 (1.12)|
|7.5 (seas min)||17.3||24.5 (0.90)||17.6 (1.29)|
The model assumes that the steady state fuel load is 20.4 t/ha (measured at seasonal maximum) and k = 0.25.
Standard errors of samples are shown in brackets.
3.4.1 Litter levels
The equilibrium litter level at the site immediately before burning, during the period of seasonal maximum, was 20.4 t/ha. This litter level was slightly higher than the litter levels recorded in the general area in August 1985, at a time when litter levels were likely to be near their seasonal minimum (Simmons & Adams 1986). Litter accumulation in the low intensity area fits the models of the expected pattern of litter accumulation following fire, and litter accumulation is slower than predicted by models in the moderate intensity area (Table 1).
After 2.7 years, the total litter level at the low intensity site had reached a 9.9 t/ha, which was more than double the litter level of 4.1 t/ha following the moderate intensity fire, and there were significant differences between the proportions of all litter components at the two sites (Table 2). After 4.5 years (seasonal minimum) the total fuel loads were 8.6 t/ha and 6.8 t/ha , and there were no significant differences between total litter loads or any of the litter components at the two sites. After 4.7 years (seasonal maximum), total litter loads were 13.9 t/ha on the low intensity area and 9.1 t/ha on the moderate intensity area, and there were significant differences only between total litter, and proportions of leaves and twigs at the two sites. After 7.5 years, total litter loads were 24.5 t/ha on the low intensity site and 17.6 t/ha on the moderate intensity area, and there were significant differences between total litter, and the proportions of all components except leaves.
3.4.2 Floristic composition of the understorey
Survey of the sites prior to burning indicated that vegetation in the area was uniform, and typical of the understorey in the wider study area which has not been burned since 1962. The site was floristically poor, and the understorey consisted of a thick litter layer with occasional shrubs, under a eucalypt canopy approximately 15m tall.
In the first few months following the moderate intensity fire, Marchantia polymorpha, a common colonist on burnt soil (Cochrane 1963), was abundant. Caladenia menziesii, an orchid known to flower prolifically following fire, but rarely in other years (Wark et al. 1987), was also observed at this site in 1985, but has not been sighted since. Pelargonium australe was the dominant species on wetter areas during the first few months after fire, but declined in the second year. There was very little weed invasion, which is probably due to the low levels of soil disturbance, dry conditions over summer, and low nutrient status of soils at the site.
Table 3 lists the 82 understorey species recorded in quadrats, their primary mode of regeneration, and average cover values in the study area 2.5 ,and 4.5 years following fire. There were no significant differences between the total number of plant species in the areas following a low or a moderate intensity fire. The only significant differences between cover values of individual species 2.5 years after fire, when most species had low cover values, was for a significantly higher cover of Lomandra nana (P<0.05) and the grasses Chionochloa pallida and Poa rodwayi (P<0.005) at the low intensity site.
There was no significant difference between the total cover of seedling regenerators following a low or a moderate intensity fire at 2.5 or 4.5 years after fire, but the total cover of vegetative regenerators was significantly higher at the low intensity site after 2.5 and 4.5 years (Tables 4 and 5). The total cover had started to decline after 7.5 years at both sites, and the cover of bare ground was lower, following the low intensity fire (Table 6). There was more variation in cover of individual species within quadrats following the low intensity fire than following the moderate intensity fire. This reflects the mosaic of burnt and unburnt patches in the low intensity area compared to the more even pattern of burning following fire of moderate intensity (Christensen & Kimber 1975). Bare ground in the moderate intensity area was ten to twenty times that in the low intensity area for the first three to five years following fire.
Table 7 shows the number of species in each regeneration category found only in the area following the low intensity fire, only in the area following the moderate intensity fire, and in both areas after 4.5 years. There are significant departures from proportional representation in each category after 4.5 years (P<0.005), with fewer vegetative regenerators, and more seedling regenerators than expected (assuming proportional representation) at the moderate intensity site, compared with the low intensity site.
There was minimal invasion by weeds at both sites. After 2.5 years, five weed species (7 per cent) were recorded and after 4.5 years 13 weed species (17 per cent) were recorded, though the total cover of weeds was less than 5per cent. There were no clear differences observed in the extent of weed invasion between the low and moderate intensity sites.
Cluster analysis resulted in quadrats being divided into two groups corresponding closely to the areas burned by low or moderate intensity fire (Figure 3.1). The quadrats sampled in 1987 when regeneration was minimal were most distinct, and quadrats from the moderate intensity site in 1989 were most similar to those from the low intensity site sampled in 1987. Cluster analysis indicates that quadrats can be clearly differentiated on their overall floristic composition, following fires of differing intensity.
|2.7 years||3.5||0.27||34.9||1.8||0.40||44.6 **|
|4.5 years||3.0||0.23||35.2||2.1||0.37||31.3 ns|
|4.7 years||5.6||0.40||40.0||3.8||0.48||41.5 **|
|7.5 years||9.1||0.60||36.2||7.9||0.44||45.1 ns|
|2.7 years||2.2||0.20||22.4||1.2||0.08||28.5 ***|
|4.5 years||2.7||0.43||31.2||2.2||0.35||32.4 ns|
|4.7 years||3.9||0.30||28.0||1.8||0.21||20.2 ***|
|2.7 years||0.5||0.16||5.2||0.03||0.01||0.7 **|
|4.5 years||0.06||0.13||7.5||0.7||0.23||9.8 ns|
|4.7 years||0.5||0.27||3.6||0.6||0.21||6.7 ns|
|7.5 years||4.3||0.52||17.1||1.4||0.36||8.0 ***|
|2.7 years||3.7||0.42||37.4||1.1||0.37||26.1 ***|
|4.5 years||2.2||0.24||26.1||1.8||0.29||26.5 ns|
|4.7 years||4.0||0.37||28.8||2.9||0.54||31.7 ns|
|7.5 years||6.2||0.41||24.5||4.0||0.50||22.6 **|
ns - not significantly different for P=0.05
Figure 3.1: Dendrogram resulting from clustering using Bray-Curtis association measure and fusion by UPGMA, based on all species recorded. Quadrats sampled from the low intensity site after 2.5 years (filled circle) and after 4.5 years (open circle) and from moderate intensity site after 2.5 years (filled triangle) and after 4.5 years (open star).
3.5.1 Appropriate Design of Field Studies
This study represents opportunistic monitoring of fires of different intensity which were essentially unplanned.The results in this study assume that the sites were equivalent before the fires, and that the only differences between the sites are due to the impact of fires of different intensity. This assumption is reasonable given that as far as can be acertained, sites have the same history, and appeared homogenous during a preburn survey, where no differences in understorey floristics, or litter levels, were detected. However, matching of sites is always difficult, as pre-burn floristic surveys may not detect variation in factors such as the distribution of soil stored seed for example. The extent of ecological heterogeneity is generally unknown and warrants further investigation and emphasis (Williams et al. 1994).
Appropriate replicated designs for assessing the impact of fire and fire regimes are desirable. Green (1993) emphasises approaches that are statistically appropriate, with ecological theory having a very limited role in his suggested designs. For example, it is implied that entities contrasted in his suggested experimental designs are discrete, whereas continuous variation between plant communities and high levels of heterogeneity are frequently observed. Some studies which are considered statistically appropriate may not be ecologically appropriate, especially in terms of site history. Stratification of sample sites into small areas to enable replication is also common, and small experimental sites may produce results which give little indication of impacts on ecological processes.
In the published literature much of our information about the effects of fire on vegetation is derived from opportunistic sampling following unplanned fires, and in a strict sense, many of these are statistically inadequate, with many studies using pseudoreplication (Hurlburt 1984). In the urban fringe and other areas of fragmented habitat, appropriate replication is often simply not possible due to the limited sites available, and lack of control over sites over time. Lack of resources is a common problem in many studies, and in larger studies, many carefully planned designs are not followed. For example, assessment of short-term impacts of fuel reduction burning on invertebrates following fuel reduction burning (Collett et al. 1993) was made at only one site rather than on the five replicate areas initially planned (Tolhurst et al. 1992). When resources are severely limiting, this is simply unavoidable.
However, results from a large number of sites and from a wide range of communities should allow some generalisations to be made, and comparisons with larger scale well designed studies such as Tolhurst et al. (1992) may be made from more limited data. Small studies can support, or question, the results of more complex larger scale studies, and can highlight ecological heterogeneity (or homogeneity) at many levels (Williams et al. 1994).
3.5.2 Litter levels
Pre-fire litter levels on the site were used as an indicator of steady-state levels in the litter-accumulation model. The model is in good agreement with litter levels in the first few years following fire, but is closer to levels on the moderate intensity site after 7.5 years. The fuel accumulation model used has been widely accepted as an accurate predictor of litter accumulation following fire in a range of forest types in Australia (Raison et al. 1983); and the discrepancy between predicted and observed values following fires of different intensity requires explanation. The lower than predicted litter levels at the moderate intensity site implies higher rates of decomposition, lower rates of litter fall at this site, or loss of litter by other means. Raison et al. (1983) suggest that decomposition rates may be reduced for several years following low intensity fires, thus allowing rapid litter reaccumulation. Decomposition rates would also be expected to be low in the arid conditions on the forest floor at the moderate intensity site, with its large patches of bare ground (Woods et al. 1983). The higher proportion of leaves, and lower proportion of the comminuted fraction, following the moderate intensity fire compared to a low intensity fire, suggests that decomposition rates are low at the moderate intensity site. Leaf fall was initially high following crown scorch on the moderate intensity site, but litter fall may be subsequently reduced while the tree canopies recover (O'Connell et al. 1979), and this may also contribute to the lower than predicted litter levels at the moderate intensity site.
Following fires of moderate intensity, erosion losses can be high where soil is exposed. The loss of vegetation allows raindrops to strike the soil with greater impact than sites with some vegetation or fragmented litter cover (Blong et al. 1982; Leitch et al. 1983; Raison et al. 1983). Surface runoff can be very high following fires (Brown, 1972), and wind can also redistribute leaves, even without rain (Blong et al. 1982). At the moderate intensity site the understorey was completely reduced to ash, and following a heavy fall of rain within hours of the fire, the soil surface was bared over large areas. Thus the lower litter level on the moderate intensity site is likely to be the combined result of high levels of surface wash of litter from the soil surface, and an overall reduction in litter fall. The significantly greater proportion of bare ground (49.0 per cent) 2.5 years after the moderate intensity fire compared to the low intensity fire (12.8 per cent) suggests that moderate intensity fire may result in a considerable risk of erosion.
If litter is lost by surface wash rather than decomposition, this may involve a considerable loss of nutrients (particularly phosphorus) from the site, as crown scorch and massive leaf fall immediately following fire prevent the withdrawal of nutrients usually associated with abscission (O'Connell et al. 1979). Losses of nitrogen and phosphorus due to erosion, especially of ash, following fires can also be substantial (Leitch et al. 1983). The impact of nutrient losses and lowering of site fertility, especially on sites which are already nutrient poor, on regrowth of vegetation are not known, but are likely to be important (Leitch et al. 1983).
|Species regeneration primarily from seed||Mean Cover (%)|
|2.5 years||4.5 years|
* denotes alien and non-indigenous species
+ denotes species present at site, but not recorded in a quadrat
- denotes species not recorded
|Total cover values||Low intensity||Moderate intensity||Significance|
n.s. not significant at P=0.05,
|Total cover values||Low intensity||Moderate intensity||Significance|
n.s. not significant at P=0.05,
|Time since fire (years)||Total vegetative cover (%)||Cover of bare ground (%)|
n.s. not significant at P=0.05,
|Species||Only low intensity||Only moderate intensity||Present both sites||Total|
|Seedling regenerators||11 (9.1)||12 (7.5)||18 (24.5||41|
|Vegetative regenerators||6 (7.9)||2 (6.5)||28 (21.5)||36|
3.5.3 Implications for Conservation
Litter levels can be used as an indicator of fire hazard (Gill et al. 1987). There is some variation in the desirable fuel loads aimed for by fuel reduction burning, with suggested hazardous levels from as low as five t/ha (Fensham 1992), about eight t/ha (Gill et al. 1987) and in the range 10-12 t/ha (Raison et al. 1983). Logic suggests that fuel reduction will reduce fire intensity, but in extreme fire weather, even very low fuel loads will give rise to fire intensities >3,000 kW/m that will be uncontrollable using conventional methods (Raison et al. 1983).
Fuel reduction burns are usually prescribed at low intensity. Following these fires, litter levels build up rapidly, and if the primary aim of burning is fuel reduction, frequent burns are necessary for adequate protection from wildfire. Rapid re-accumulation of litter is due to both incomplete combustion of the comminuted fraction, litter fall, and rapid regrowth of vegetation following release of nutrients from the litter layer after fire. At the low intensity site the accumulation of litter to hazardous levels of more than eight t/ha (Gill et al. 1987), occurred in less than 2.7 years. If fuel loads are to be kept below 8 - 12 t/ha, many high risk areas would need to be burned every three years (Simmons & Adams 1986). The results of this study indicate that a fire of moderate intensity results in slower re-accumulation of litter, but at the cost of the potentially increased erosion, and possible nutrient losses. Even with slower litter accumulation, the return of litter levels to hazardous levels still occurs within four to five years. If areas are burned with a frequency sufficient to keep litter loads below hazardous levels for the purpose of fire control, this is likely to have significant effects on vegetation floristics (Fox & Fox 1986; Nieuwenhuis 1987), and result in nutrient losses (Raison et al. 1983).
In remnant vegetation where fire has been excluded for a long period and fuel levels are often high, it may be difficult to achieve a low intensity burn. Gill et al. (1987) suggest that there are very few days per year considered optimal for low intensity fuel reduction burning. Many burns are carried out for pragmatic reasons, and contrary to stated policies (Brown et al. 1991). Where hazard reduction has a high priority there is a tendency to burn on days which are not optimal, and which can result in fires of higher intensity than is desirable. Managers can choose to burn when climatic conditions will give a low intensity fire, but then may not achieve removal of understorey nor good preparation of a seed bed. Where higher intensity fire is prescribed to achieve a vegetation management objective, the close proximity of assets, and the public perceptions of safety may limit the intensity which can be practically applied. The high proportion of bare ground following moderate intensity fire also poses a high erosion risk.
Litter accumulation following fire is well understood in many vegetation types (Raison et al. 1983; Simmons & Adams 1986). Litter returns to pre-burn levels within a few years in most vegetation types, and fuel reduction needs to be frequent and cover a significant proportion of an area if it is to achieve its objectives of lowering fire intensity on high fire danger days (Rawson et al. 1983). In spite of a number of selected anecdotal reports, the evidence for effectiveness of fuel reduction on fire suppression capabilities for more than a few years following fire, under high fire danger conditions is inconclusive (Buckley 1992; Raison et al. 1983). Fire management on public land in Victoria continues to emphasise a primary strategy of broadscale fuel reduction.
The primary purpose of broadscale fuel reduction burning was initially to reduce damage to the standing crop in forests during high intensity wildfire (McArthur 1966; Raison et al. 1983). Its objective is now more generally to reduce fuel over a larger area to reduce intensities to a level where they can be suppressed by available technology on most days.The primary objective of fuel reduction is to achieve a reduction in fire intensity to below 3,000 kW/m, and fuel loads need to be maintained at a low level (8 - 12 t/ha) to achieve this (Table 8). We need to ask whether significant reduction in fire intensities over large areas is achievable. For example, if fire intensity is not reduced to below 3,000 kW/m under extreme conditions when fire control is generally a problem, fuel reduction will achieve little, as most losses occur under these conditions (Rawson et al. 1983). Burning frequencies in the order of every two to four years would be required to achieve fuel loads which will result in significantly lowered and fightable intensities. In the urban fringe where losses have been highest, opportunities for strategic burning are limited, and maintenance of low fuel loads will involve significant vegetation modification. If these lowered fuel loads are not achieved, fire prevention works are cosmetic, and not effective under extreme conditions when they are most needed.
3.5.4 Floristic composition of the understorey
Low intensity burns are usually effective in regenerating native understorey species, including hardseeded species, and generally result in the rapid reestablishment of vegetative cover. Vegetative regenerators may have an initial advantage over seedling regenerators following low intensity fires, as there is generally little damage to rootstocks. Lomandra nana, C. pallida and P. rodwayi had the highest average cover values of all species at both post-fire sampling times, and were the only species showing significant differences in cover between the two sites. Lomandra is reported as a fire resistant increaser by Purdie (1977), who suggests that in structurally similar vegetation at Black Mountain, post-burn vegetation appeared to be dominated by C. pallida (syn. Danthonia pallida) during the first two years, when larger regenerating shrubs were still small seedlings. At Black Mountain C. pallida commenced regrowth very rapidly, though it was classed as a fire resistant decreaser, and C. pallida was also sensitive to fire intensity.
Following fires of higher intensity, rootstocks may be damaged, slowing early regrowth of vegetative regenerators. Posamentier et al. (1981) suggest that increased seed germination and seedling survival in heath at Nadgee, compared to results reported at Black Mountain by Purdie (1977), were due to higher fire intensity. Increased intensity reduced the growth, and post fire competition, of vegetative regenerators. It should be noted that the fire intensity applied by Purdie (1977) at one site was in the high range (Cheney 1981), and reached 4000 kW/m. In vegetation in this study, which is structurally similar to that at Black Mountain, there were significant departures from expected proportions of seed regenerating species; but there were no significant differences between the average cover values of seedling regenerators and vegetative regenerators, following fires of different intensities at this early stage. However, the proportions may change as seedlings become established. The contention that more intense fires favour a flammable shrubby understorey, which in turn results in intense fires because of high fuel loads (Meredith 1988), is supported by the higher than expected proportion of shrubby seed regenerators in this study. The application of fire of particular intensities may therefore be used to modify the relative proportions of shrubs and herbs, and subsequent species composition of plant communities.
It is often assumed that prescribed fires need to achieve moderate intensities for good regeneration of hardseeded species. It is likely that soil temperatures are sufficiently elevated by both low intensity and moderate intensity fires to stimulate germination. The later observed differences in the numbers of seedling regenerators at the moderate intensity site can be explained by differences in establishment of seedlings at the two sites. Following low intensity fires, seedling establishment may be minimal or absent, as the new germinants are unable to establish due to competition from the rapidly regenerating mature individuals surrounding them; the soil seed bank may also be depleted following these low intensity fires. Following moderate intensity fire, seedling establishment may be enhanced due to lowered competition for resources from mature plants. Moderate intensity summer-autumn fires are likely to slightly favour seedling regenerators compared to a low intensity fire, and this trend would be accentuated following spring fires. Water stress (and subsequent seedling losses) due to competition from vegetative regenerators would be expected to be higher over the dry summer period following low intensity fire compared to moderate intensity fire. This is consistent with the observation in this study that after 4.5 years there were more seedling regenerators than expected following the moderate intensity fire. If conservation values (as measured by high species diversity) are to be maintained, burning in spring should be avoided, especially during periods of drought. These are times when preparation for an extreme fire season may often include calls for increased strategic fuel reduction burning prior to the fire season.
Purdie and Slatyer (1976) suggest that in open forests, the initial floristic composition of the vegetation immediately following fire will determine the later composition of the postburn community. However, variations in intensity, frequency or the season of burning may lead to shifts in relative dominance of various species. The application of a particular fire regime is likely to result in plant communities floristically different to the initial community, and may be heavily dependent on post-fire climatic conditions, particularly those conditions which result in water stress.
Effective fuel reduction will require very frequent fire, but this is likely to have significant effects on floristics. Higher intensity fires maintain low fuel loads for a longer period, but are associated with a high cover of bare ground, which may pose an unacceptably high risk of erosion. Differences in
fire regime are likely to result in changes in understorey floristics.
|Climatic conditions||Day type||Fuel load to produce 3,000 kW/m||Time to reach fuel laod|
|RH 15%, 35°C, wind 25 km/hr||Mild summer day||12 t/ha||4.4 years|
|RH 10%, 35°C, wind 40 km/hr||Average summer day||10 t/ha||3.3 years|
|RH 10%, 40°C, wind 50 km/hr||Extreme summer day||8 t/ha||2.4 years|
Modelling was carried out using FRISK (Simmons et al. 1988)
3.5.5 Alternative Approaches to Fire Management
Results of studies following the Ash Wednesday fires of 1983 suggest that we should put less emphasis on strategies which rely on stopping the passage of wildfire, and more emphasis on strategies which assist people and their assets to survive the passage of wildfire. Data from the catastrophic wildfires which burned a large area of in south-eastern Australia on February 13 1983, indicated that 90 per cent of houses survived the passage of this intense wildfire, if an able-bodied person was in attendance (Wilson & Ferguson 1984). There are now clear guidelines for planning to maximise the chances of survival of houses in forested urban fringe areas by modifying building construction and surroundings to reduce risk (Wilson & Ferguson 1986). Lowering of fuel loads near houses has a major impact on the probability of a house surviving a wildfire, as fire intensity contributes approximately 50 per cent of the risk to a house (Wilson & Ferguson 1986).
Similarly, emphasis on public education to modify human behaviour should minimise deaths due to wildfire; many lives are lost as a result of poor, and often inexplicable decisions (Krusel & Petris 1992). It is now apparent that in the urban fringe areas, strategies which emphasise landholder responsibility and and preparedness are likely to be more effective in reducing losses of lives and assets than broadscale fuel reduction; this is the basis for the Country Fire Authority Community Fireguard project (Boura 1994).
Many past fire prevention activities (e.g. broad scale fuel reduction, fire breaks) have been ecologically destructive and achieved very little in terms of reduction in the loss of lives and assets under extreme conditions (Davidson 1988; Meredith 1988; Wilson 1988). On the urban fringe, effective fuel reduction will require frequent fire, and frequent fire is likely to significantly modify understorey floristics. Where conservation values are high, it is reasonable to ask exactly what the various fire prevention strategies are expected to achieve and to evaluate them against expected outcomes.
Unfortunately, it is difficult to achieve both optimal fire protection and optimal conservation value in many areas at any one time. If fuel loads are maintained at low levels near houses, the probability of house survival will be maximised, but where there is relatively dense housing in bushland, this will require significant modification, and possibly loss of some native vegetation. If native vegetation is managed for maximum biodiversity in these areas, it is likely that property owners will need to accept that they carry an increased risk of loss of assets.
Community values vary and it is up to individuals and the communities they live in to make choices about levels of risk they are prepared to accept, and the environment they want to live in. Both the community and individuals will need to decide their priorities, given some realistic assessments of risk (both ecologically and from wildfire). Recent attempts at modelling risk and hazard may be useful in quantifying levels of risk. If communities choose to accept a higher risk and hazard from wildfire, to maintain preferred vegetation and particular ecological values, communities must ensure that they do not place others at risk. They might therefore be expected to more actively contribute to fire prevention by participation in community self-help groups, and to fire suppression activities through their local brigades, to ensure rapid suppression of unplanned fires in their area.
Fuel reduction burning is effective in reducing fuel for only a few years, even with complete removal of the litter layer following moderate intensity fire. However, emphasis continues to be placed on repeated low intensity fuel reduction burning which has had limited value in reducing losses at the urban-forest interface (Rawson et al. 1983), but which has high potential to modify vegetation in areas with high conservation value.
Following the fires of 1983 in south-eastern Australia, we now have firm guidelines for reducing the losses of lives and assets during intense wildfires (Wilson & Ferguson 1986; Ramsay & McArthur 1987; Krusel & Petris 1992). Fire managers often continue to ignore research into the modification of social behaviours, which may be a much more productive area for reducing the losses of both lives and assets from wildfires (Edgell & Brown 1975; Healey et al. 1985).
In areas with high conservation values the primary objectives of prescribed burning need to be clearly stated. Fires with a range of regimes, planned to achieve different management objectives such as habitat manipulation and maintenance, may be prescribed. Repeated low-intensity fires have become part of current management, but if vegetation in the urban fringe is to be managed so that conservation values are maintained, some higher intensity fires may be appropriate. High intensity burns may be necessary for management of particular communities to a preferred floristic composition, especially where promotion of seed regenerators is desired.
The results of this study indicate that fires of different intensity can have subtle impacts on community floristics. There were few differences between the average cover values for individual species between the two sites following fire, but cluster analysis showed that there were subtle differences in the composition of the vegetation at the sites following fires of different intensity. This highlights the need for clear objectives to be set whenever an area is deliberately burnt. Fires with an intensity other than that prescribed for a particular management objective, may result in variation in the floristic composition of the subsequent regrowth which will have a significant effect on conservation values of plant communities in the long term.
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