Proceedings of the conference held 8-9 October 1994, Footscray, Melbourne
Biodiversity Series, Paper No. 8
Department of the Environment, Sport and Territories, 1996
14. Long-term effects of fuel reduction burning on ivertebrates in a dry sclerophyll forest
Forest Ecology Section, Research Division, State Forests of NSW
Low-intensity fires are extensively used in managed sclerophyll forests to stimulate regeneration, manipulate wildlife habitat and in particular, to reduce fuel levels with the intention of minimising the extent and severity of wildfires. Recent, extensive wildfires have led to calls for increased use of fuel-reduction burning, however the consequences for local and regional invertebrate biodiversity of repeated burning over long time periods and wide areas is poorly understood. This is of considerable concern given that invertebrates constitute 95 per cent of known species of fauna in Australia, thereby making a substantial contribution to our National biodiversity. Results from two long-term fire studies are presented, and impacts on terrestrial invertebrates illustrated by using data from ant communities. Ant communities were found to decline in richness and change in their composition in the years after fire. The ant fauna could be divided into groups of species which reached their peak abundance at different stages in the vegetation succession related to their specific habitat requirements. The use of regular widespread fires for fuel reduction however is likely to result in extensive habitat simplification, truncation of successional patterns and the associated loss of regional biodiversity. These predictions were tested by an experimental study examining the effects of 20 years of frequent fuel reduction burning. It appeared that ant species richness had not been affected by regular burning; however, an analysis of community structure revealed that particular species associations were recognisable, with distinctive communities characterising burnt and unburnt sites. Regularly burnt areas were characterised by ant species with broad environmental tolerances and dietary requirements, with a 'loss' of up to ten rare and highly specialised species. High species richness in these areas was maintained by the presence of additional generalist and 'disturbance indicator' species, compensating for the previously identified time-since-last-fire effect. The loss of many specialist species is of concern as these species generally constitute a substantial fraction of the overall richness of communities. Widespread and frequent use of this management practice may therefore have serious implications for the conservation of regional biodiversity unless appropriate strategies are implemented.
Key words: fuel reduction, repeated burning, low-intensity fires, ants, invertebrate biodiversity, habitat simplification, Australia.
Infrequent, periodic, forest fires (bushfires) are an integral part of the modern physical environment of Australian dry sclerophyll forests. These fires generally result in complete or partial destruction of the understorey and litter layer, and partial or complete defoliation (but not death) of the tree canopy. High intensity wildfires can however result in substantial loss of life and property, reduce the commercial value of production forests, and have marked impacts upon native plant and animal communities. For these reasons, low intensity prescribed fires have become an important management tool used to reduce fuel loads and thereby lessen the frequency and severity of unplanned fires (see Luke & MacArthur 1978; Grant & Wouters 1993). Loss of life and property due to recent bushfires in the dry coastal forests of eastern Australia has once again led to urgent calls for a substantial increase in the use of fuel reduction burning; however little consideration is being given to the ecological consequences of such actions.
While sclerophyll forests, woodlands and heaths are dominated by plant species with adaptive responses to fire that enable them to survive exposure to periodic burning (see for example Gill 1981; Noble & Slatyer 1981), the impact of such fires on terrestrial invertebrates is poorly understood. The consumption of some or all of the leaf litter by flame, causes short-lived but substantial rises in soil temperature during fire; and post-fire changes in the surface radiation budget, mean that soil and litter fauna are substantially affected by fire in the short term (Bornemissza 1969; Springett 1979; Moulton 1982), with recovery from a single fire taking up to three to five years (Metz & Farrier 1973; Seastedt 1984; Neumann & Tolhurst 1991). Given the patchy nature of low-intensity fuel-reduction burns, and the protection afforded by small habitat refuges and within the soil, it has been suggested that periodic fires used for fuel management purposes have few long-term effects on most soil and litter invertebrates (Majer 1980; Campbell & Tanton 1981; Abbott et al. 1984). The consequences for local and regional biodiversity of repeated burning over long time periods and wide areas is, nevertheless, poorly understood. Given that invertebrates constitute the bulk of our National biodiversity (New 1984; CONCOM 1989), this is a serious omission in our knowledge concerning the impact of management practices on our native fauna. This paper therefore investigates the effect of frequent low-intensity burning on terrestrial invertebrate biodiversity in a dry sclerophyll forest.
Australia's dry sclerophyll forests are located primarily along the east coast from north Queensland to southern Victoria, in eastern Tasmania and the south-west of Western Australia (Figure 1). The natural interval between successive fires varies considerably across the geographic range of this broad forest type, primarily in response to climate. The research reported in this paper concerns dry sclerophyll forests of the mid north coast of New South Wales where the natural fire interval has been estimated to be between five and 15 years (Walker 1981); however, intervals of two to three years in some areas are not uncommon now due to human influence. The findings from two complementary experiments are described: firstly, a study that looks at changes in species richness and community structure in the time period between successive fires (the fire interval), and secondly, a study designed to experimentally test these results by repeatedly shortening the fire interval with the frequent use of low-intensity fire.
To illustrate the findings of these experiments, data are presented for ant communities, a group which represent a major component of the forest invertebrate fauna. Ants in Australia are ubiquitous, abundant and highly active, making them one of the most important animal groups in terms of energy flow (Brown & Taylor 1970; Rogers et al. 1972). The diversity and structure of ant communities is often correlated with the composition of other components of the invertebrate fauna (Majer 1983) leading to their increasing use as bio-indicators in a management context (Yeatman & Greenslade 1980; Majer et al. 1984; Andersen & McKaige 1987; Burbidge et al. 1992; Andersen 1993). Preliminary results for a small number of other invertebrate groups are also presented to illustrate some of the difficulties in interpreting the results of such studies.
In order to investigate changes that occur after fire (experiment 1), an area of coastal forest was chosen within the Myall Lakes National Park where a high fire-frequency has resulted in a complex mosaic of post-fire vegetation successional stages, with forest patches ranging from two to 14 years post-fire at the time of the study. Sites typifying six forest understorey stages (2.2, 3.9, 6.0, 7.0, 8.3, and 14.2 years post-fire), and with similar past fire frequencies, were chosen for investigation. One control site was established and monitored for six years in order to assess the accuracy of the other sites as a "time-since-fire" sequence (see York 1989, 1994).
Data from a long-term fuel reduction burning trial (experiment 2) were used to test hypotheses generated by the above study. Twelve 0.1 ha. study plots were established within Bulls Ground State Forest, 20km south-west of Port Macquarie. Six plots were randomly allocated as burning treatments, the remaining six as control plots from which fire was excluded (6 x 2 randomised block design). The plots were located within similarly treated forest blocks of approximately 1 ha. and separated by cleared buffer areas to protect them from wildfire. Fuel reduction burning was implemented (on treated plots) in autumn whenever fuel build-up permitted, generally every three years (1970, 1973, 1977, 1980, 1983, 1986, 1989). At the time of sampling (1991) the burnt plots had experienced 20 years of fuel-reduction burning (seven fires) and it had been two years since the last fire (see York 1993b).
For both studies, ants (and other epigaeic invertebrates) were sampled using nine pitfall traps arranged within a 10m x 10m grid and left open for seven days during fine weather in February of the year of sampling. Material was returned to the laboratory and examined with a binocular microscope where ants were identified to genus using the key published in Greenslade (1979). A reference collection was established with final verification of species completed at the Australian National Insect Collection in the CSIRO Division of Entomology, Canberra. All analytical procedures were performed using the SPSSx statistical package (see SPSSX Inc. 1983) on the VAX 11-785 computing facilities at the University of New South Wales.
The results for the two experiments are reported separately, and then discussed together in terms of their relevance to the existing fire regime in this dry sclerophyll forest environment.
14.4.1 Experiment 1 - Myall Lakes
In excess of 25,000 individuals from 40 ant species were caught and identified in this study, with the analysis of data reported here concerning changes in ant species richness and community composition with time since fire. Analysis of the chronosequence data revealed that mean values of ant species richness varied from 8.8 to 18.8, with comparable levels of variability between forest understorey classes (see York 1994). Using regression procedures it was found that there was a significant relationship between ant species richness and time since last fire, with a slow decline in ant species richness evident in the years after fire (Figure 14.2). Sixty-one percent of the variance in ant species richness was explained by time in the following exponential regression model:
Ant species richness = 20.34 e-0.067 (years since fire)
r = 0.781 r2=60.9% n = 24 P < 0.001
Changes in ant species richness at the control site show comparable trends to those described above (Figure 14.2), and regression equations fitted to each data set independently have regression coefficients that are not significantly different (F=2.62, DF=1,44 P=0.12).
Examination of frequency of capture of ant species in pitfall traps shows that sites differing in time since last fire not only differ in ant species richness, but also in their constituent species (Table 1). There would seem to be three general groupings of species:
- firstly, a small group of species which are abundant across all time periods;
- secondly, a group of species which while present across all time periods, are not as abundant; and
- thirdly, a group showing a replacement of species over time.
The first group is composed of only three species, the second of nine species, while the third is composed of 29 species and comprises the majority (71 per cent) of overall species richness (see Table 1).
To illustrate patterns evident within these groups, changes in the relative abundance of a few species were examined (Figure 14.3). The closely related species of Iridomyrmex and Ochetellus are abundant across all time periods, however the relative abundance of the three species changes over time. Iridomyrmex sp. A has a major peak in abundance at the youngest site and a minor peak at 8.3 years after fire. The peak abundance of Ochetellus sp. A at the <3.9 year old site coincides with low abundance of species Iridomyrmex sp. A. As the abundance of Ochetellus sp. A declines, species I. sp. B reaches its peak abundance. The minor peak in abundance of species A occurs at a time when the abundance of the other two species is low. At the oldest site all three species reach their lowest levels of abundance.
Two species of lower abundance also show changing patterns with time. Melophorus sp. A reaches its peak abundance between three and six years after fire, and then declines rapidly. Monomorium sp. A is most abundant in the first few years after fire, but levels of relative abundance decline more slowly over time. Two species of Rhytidoponera show quite different patterns of abundance over time. Rhytidoponera metallica is most abundant at the youngest site, then declines steadily in abundance to about seven years after fire, and remains at this level. Rhytidoponera victoriae is not abundant initially, reaches its peak abundance between three and six years after fire, then declines rapidly. It was not caught at the two oldest sites.
A number of species were only caught at the oldest sites (>6 years since last fire); these included Lordomyrma punctiventris, Sphincotomyrmex steinheili and Mesoponera australis.
14.4.2 Experiment 2 - Bulls Ground
Regular measurements of 'fine fuel' (leaves, bark, twigs <2.5cm diam. and miscellaneous material) have been made on these study plots from 1969 to 1991. Figure 14.4 illustrates the nature of changes that have taken place over that 21 year period; it presents a contrast between the stable conditions for litter and soil dwelling invertebrates existing on unburnt plots, and the dramatically fluctuating and unstable environment present on frequently burnt plots.
On control plots, mean litter biomass had increased from 14.7 t ha-1 in 1969 to 17.5 t ha-1 in 1991, but has fluctuated from 14.7 t ha-1 to 22.9 t ha-1 during that period. On burnt plots, mean litter biomass varied between 3.9 t ha-1 and 19.8 t ha-1 over the 21 year period. The effect of repeated fuel reduction burning therefore is an episodic loss of 46-73 per cent of litter biomass, but by three years post fire, litter levels have usually reached and often exceeded that achieved before the previous fire (range=72-136 per cent). At the time of the invertebrate survey, litter biomass on burnt plots averaged 9.3 tonnes ha-1. This represented, on average, 74 per cent of pre-fire levels and 53 per cent of levels on control plots, where the mean value was 17.5 tonnes ha-1.
In excess of 55,000 individuals from 25 taxonomic groups of terrestrial invertebrates were collected during a single sampling period in 1991, representing a 'snapshot' in time of the effects of over 20 years of frequent fire. Numerically, the most abundant groups were: springtails (33.1 per cent), ticks & mites (23.9 per cent), ants (23.1 per cent), bugs (4.2 per cent), beetles (4.0 per cent), bees & wasps (2.8 per cent), insect larvae (2.7 per cent), flies (2.6 per cent) and spiders (2.2 per cent), with these groups making up 98.6 per cent of the total number of organisms caught.
The numbers collected from several groups were insufficient to comment about possible effects of frequent burning. These were the pseudoscorpions, harvestmen, centipedes, millipedes, diplurans, termites, embids, booklice, lacewings, caddisflies, moths and butterflies. For these groups, the trapping method used was not the most appropriate and has probably contributed to the low capture rate. While the low numbers collected for several other taxa precluded statistical analysis, frequent burning appears to have led to a reduction in the numbers of amphipods, cockroaches and earwigs, and an increase in the numbers of grasshoppers & crickets, and thrips.
For nine taxonomic groups there were sufficient data to permit statistical testing. A two-way analysis of variance (ANOVA) procedure was used to investigate treatment effects (burnt/unburnt), plot effects (location), and possible interactions between the two. Both significant treatment and plot effects were detected for a number of taxa, with ants being the only group to show interaction between the main effects. A number of groups showed a significant decrease in abundance following frequent burning (see Figure 14.5):
Springtails (Collembola) were numerically the most abundant organism found during this trapping program, accounting for about 33 per cent of individuals. The effect of treatment was significant (P=0.036) with approximately 15 per cent less individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was low.
Ticks & Mites
Ticks and mites were numerically the second most abundant group of organisms, accounting for about 24 per cent of individuals. The effect of treatment was significant (P=0.001) with approximately 31 per cent less individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was high, with a significant location effect (P<0.001).
For terrestrial isopods, the effect of treatment was significant (P=0.004) with approximately 38 per cent less individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was high, with a significant location effect (P=0.034).
For beetles, the effect of treatment was significant (P<0.001) with approximately 32 per cent less individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was high, with a significant location effect (P<0.001).
Although pitfall trapping is not the optimal method of sampling flies, in excess of 1,400 individuals were caught. The effect of treatment was significant (P<0.001), with approximately 58 per cent less individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was high, with a significant location effect (P=0.002).
For insect larvae, the effect of treatment was significant (P=0.005) with approximately 35 per cent less individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was low.
A number of groups showed a significant increase in abundance following frequent burning (see Figure 14.6).
For spiders, the effect of frequent burning was significant (P=0.002) with approximately 33 per cent more individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was low.
For the bugs (Hemiptera), the effect of treatment was significant (P=0.034) with approximately 77 per cent more individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was high, with a significant location effect (P=0.002).
Ants represented (numerically) the third most abundant group trapped (12,895 individuals), accounting for 23 per cent of the total catch.
The effect of treatment was significant (P<0.001), with approximately 250 per cent more individuals caught (on average) on frequently burnt plots. Variation in relative abundance between plots within each treatment was high, with a significant location effect (P=0.001). The interpretation of this result is confounded by significant interaction between the main effects (P<0.001). Figure 14.7 however clearly shows the substantial treatment effect; it also indicates localised plot effects with replicate 2 (burnt), for example, having significantly higher ant numbers than all other burnt plots.
A total of 43 ant species were caught, with richness on unburnt plots ranging from 15-21 species (mean=18.2, s.e.=1.0) and from 16-22 on frequently burnt plots (mean=19.5, s.e.=1.1), (see Table 2). Mean richness values were not significantly different between treatments (t-test: t=0.92, n=6, P=0.377).
Of the 43 species recorded overall, 36 (84 per cent) were recorded on unburnt plots and 33 (77 per cent) recorded on frequently burnt plots. While the different combination of species found on individual plots may partly reflect an artefact of sampling, an inspection of the data (Table 2) suggests that the composition of communities shows a real difference between treatments (fire histories). This is reflected firstly by a number of species which show particular patterns of relative abundance, and secondly by others which were only detected on plots experiencing a particular treatment.
Frequently burnt communities were dominated by Rhytidoponera metallica, Tetramorium sp. 1 and Notoncus sp. 1, which were present in 94 per cent, 67 per cent and 56 per cent of traps respectively, compared with 24 per cent, 15 per cent and 20 per cent of traps on unburnt plots. On unburnt plots the ant communities were dominated by Pheidole sp. 2 and Bothroponera sp. 1, which were present in 61 per cent and 13 per cent of traps respectively, compared with 33 per cent and 2 per cent of traps on unburnt plots. Within the 43 ant species identified overall, ten species are unique to unburnt plots and seven species unique to frequently burnt plots (Table 2). Species restricted to unburnt plots were: Crematogaster sp. 2, Discothyrea sp. 1, Heteroponera imbellis, Hypoponera sp. 1 & 2, Strumigenys purplexa, Technomyrmex sp. 1, Leptogenys sp. 1, Prolasius sp. 2 and Camponotus intrepidus. Species found only on burnt plots were: Melophorus sp. 1, Brachyponera lutea, Campomyrma patiens, Myrmecia gulosa, Myrmecia sp. 1, Monomorium sp. 1 and Colobostruma sp. 1.
|Iridomyrmex sp. 1||+||+||+||+||+||+||+||+||+||+||+||+|
|Paratrechina sp. 1||+||+||+||+||+||+||+||+||+||+||+||+|
|Pheidole sp. 2||+||+||+||+||+||+||+||+||+||+||+||+|
|Prolasius sp. 1||+||+||+||+||+||+||+||+||+||+||+|
|Crematogaster sp. 1||+||+||+||+||+||+||+||+||+||+||+|
|Tetramorium sp. 1||+||+||+||+||+||+||+||+||+||+|
|Notoncus sp. 1||+||+||+||+||+||+||+||+||+||+|
|Bothroponera sp. 1||+||+||+||+||+|
|Pheidole sp. 1||+||+||+||+||+||+||+||+||+|
|Stigmacros sp. 1||+||+||+||+||+||+||+||+|
|Iridomyrmex sp. 2||+||+||+||+||+||+||+||+||+||+|
|Meranoplus sp. 1||+||+||+||+||+||+||+||+||+|
|Camponotus sp. 1||+||+||+||+||+||+||+||+|
|Camponotus sp. 2||+||+||+||+||+||+||+||+||+||+|
|Camponotus sp. 3||+||+||+||+||+|
|Solenopsis sp. 1||+||+||+||+||+||+||+||+||+|
|Leptomyrmex sp. 1||+||+||+||+||+|
|Cerapachys sp. 1||+||+||+|
|Paratrechina sp. 2||+||+||+|
|Paratrechina sp. 3||+||+|
|Polyrhachis sp. 1||+||+|
|Ponera sp. 1||+||+|
|Melophorus sp. 1||+||+|
|Myrmecia sp. 1||+|
|Monomorium sp. 1||+||+||+||+|
|Colobostruma sp. 1||+|
|Crematogaster sp. 2||+||+|
|Discothyrea sp. 1||+|
|Hypoponera sp. 1||+|
|Hypoponera sp. 2||+||+|
|Technomyrex sp. 1||+||+||+||+|
|Leptogenys sp. 1||+|
|Prolasius sp. 2||+||+|
|Species richness mean (± s.e.)||18.2 (1.0)||19.5 (1.1)|
+ indicates presence in at least 1 of the 9 pitfall traps at that replicate
Table 3 Habitat and dietary preferences of ant species trapped only on either frequently burnt or unburnt plots (experiment 2 - Bulls Ground).
|Treatment||Ant Species||Habitat/dietary preferences|
|Unburnt (control)||Hypoponera sp. 1||Cryptic predator of Collembola.|
|Hypoponera sp. 2||Found in litter, under stones & in rotten logs.|
|Strumigenys purplexa||Cryptic predator of Collembola.Litter dweller in moist environments.|
|Technomyrmex sp. 1||Cryptic predator of arthropod eggs.(control)|
|Heteroponera imbellis||Cryptic predator in litter. Uncommon.|
|Leptogenys sp. 1||Predacious on termites.|
|Prolasius sp. 2||Favours moist environments. Uncommon.|
|Camponotus (myopores)||Favours moist environments.|
|Crematogaster sp. 2||Nests in trees. Feeds in understorey vegetation.|
|Discothyrea sp. 1||Cryptic predator. Litter dweller in moist environments.|
|Frequently burnt||Melophorus sp. 1||Specialists in hot dry environments.|
|Brachyponera lutea||Favours dry environments. Widespread.|
|Campomyrma patiens||Common in dry environments.|
|Myrmecia gulosa||Coloniser of disturbed habitats.|
|Myrmecia sp. A||Coloniser of disturbed habitats.|
|Monomorium sp. A||Common seed harvester, especially of grasses.|
|Colobostruma sp. A||Nocturnal forager in litter & on vegetation.|
The dry sclerophyll forests of the north coast of NSW experience regular burning. These burns are caused through the controlled use of fire for forestry and conservation management purposes, through the largely uncontrolled use of fire via otherwise well-intentioned 'burning-off' practices, and through arson and wildfire. While it is recognised that various burning prescriptions do reduce the severity of unplanned fires (Gill et al. 1987; Grant & Wouters 1993), there is growing concern that repeated low-intensity burning may have a negative influence on plant and animal communities. Frequent firing may remove vegetation species that rely on seed production for their persistence (Gill 1981; Bradstock & Myerscough 1981; Benson 1985), often leading to dominance by monocotyledons and ferns (Pidgeon 1938). Fire frequency becomes a significant factor for plant species requiring a long period of time (relative to the interval between fires) to reach reproductive maturity (see Fox & Fox 1986). There is little information however on the effects of fire frequency on forest invertebrates but Abbott et al. (1984) suggest that periodic low intensity fires have few permanent effects on most of the invertebrate taxa present in the litter and soil of the Jarrah forest. The 'patchy' nature of low-intensity fires maintains a mosaic of small 'islands' of unburnt litter and vegetation, theoretically enabling surviving invertebrate populations to recolonise regenerating burnt areas (Bornemissza 1969; Campbell &Tanton 1981).
The inherent variability in natural fire regimes results in a mosaic of habitats with vegetation at different stages of floristic and structural post-fire succession. Previous investigations have pointed to a direct correlation between the structural complexity of the environment and the diversity of fauna in several groups; for example: birds (MacArthur & MacArthur 1961), desert lizards (Pianka 1967) and insects (Balling & Resh 1982; Greenslade & Halliday 1983; Andersen 1986). There is concern however that repeated low-intensity fires reduce the spatial and structural heterogeneity and may have long-term consequences for the survival of invertebrate populations (see Collett et al. 1993). Invertebrates are now recognised as critical elements in the maintenance of forest ecosystems and they contribute the bulk of the biodiversity in these environments (Mattson 1977).
This study has investigated aspects of the impact of repeated disturbance on terrestrial invertebrate communities; it also provides information concerning the use of these invertebrates as monitoring tools for ecologically sustainable management. The results of two long-term burning studies are presented, and data for ant communities used to illustrate the likely effects of fire on invertebrates. Ant communities were shown to decrease in their species richness over time after fire. As richness declined with time since fire, community composition and structure also changed. The ant fauna can be divided into groups of species which reach their peak abundance at different stages in the vegetation succession related to specific habitat requirements. Some ant species have been shown to have flexible habitat requirements while others are more specific (see Elmes 1971; Doncaster 1981; Samways 1983). As the habitat changed over time, there was a gradient over which the success of a species varied. More tolerant species were abundant right across the range of habitats, while for others there was a species replacement or succession. The youngest sites had an ant fauna which was clearly distinguished from older sites by the abundance of Meranoplus sp. A, Iridomyrmex sp. A, Myrmecia gulosa, Camponotus intrepidus and Rhytidoponera metallica. The proliferation of the lower vegetation layers in the first few years after fire provides a food source for other invertebrates. These in turn may themselves serve as a source of protein-rich food for ants (see Dindal & Metz 1977) or provide carbohydrate-rich honey-dew (Way 1963). Secondly, these plants also may act as a source of nectar (Keeler 1980) or provide seeds for harvesting ants (Ashton 1979; O'Dowd & Gill 1984). Adults of Myrmecia gulosa feed largely on nectar and honey-dew (Brown & Taylor 1970), and this species was most abundant at young sites. The high frequency of myrmecochory in sclerophyll shrubs species (Berg 1975; Beattie &Culver 1981; Buckley 1982) suggests that seeds may be a plentiful food resource at this stage in the vegetation succession. Berg (1975) identified Aphaenogaster longiceps and species of Rhytidoponera, Monomorium, Melophorus, Iridomyrmex and Pheidole as seed collectors, and these are all well represented at the youngest sites.
These early successional species are also favoured by the high levels of insolation at the ground surface. This conclusion is supported by a substantial change in community composition observed at the 3.9 year old site at which the mean percentage light at ground level has dropped from 21.3 to 6.5 per cent (York 1989). The pattern of soil insolation has been found to be a key factor controlling the colony size and dispersion patterns of ants (Brian 1956; Pontin 1963; Greaves 1973; Sudd et al. 1977). It has both a direct effect on the temperature of the nest-site, and an indirect effect on the food supply, via the vegetation. Increased levels of shading and changes in surface and nest temperatures with increasing vegetation cover over time reduce ant species richness (Goldstein 1975; Greenslade & Mott 1979), and alter community composition as the environmental conditions become sub-optimal for certain species (Welch 1978; Elmes & Wardlaw 1982).
Actual species richness however does not decline quickly over time since fire, and is maintained by the gradual increased abundance of species better adapted to these changed environmental conditions. Species such as Melophorus sp. A and B, Iridomyrmex sp. B, Monomorium sp. A and Crematogaster sp. A gradually increase in abundance in the first few years after fire. Crematogaster sp. B (a tree/shrub dwelling species) showed its maximum abundance six to seven years after fire, when vegetation in the mid-sized, tall and very-tall shrub layers were at a peak. Species from genera known to generally dwell in leaf litter (Mesoponera, Heteroponera, Mayriella and Sphinctomyrmex), showed their highest abundances at the older sites. Similar patterns in habitat-controlled abundance has been reported for spiders (Uetz 1975; Merrett 1983; Bultman & Uetz 1984) and beetles (Greenslade 1964; Refseth 1980).
As a consequence of the observed decline in ant species richness over time after fire, it might be suggested that a repeatedly short fire interval (a high fire frequency) would not negatively affect biodiversity. In theory, frequent fires would maintain an environment that supports the greatest species richness. It is clear however that the use of a single index, species richness, is inappropriate for management planning without considering the scale at which it applies. In this dry forest environment there is a replacement (succession) of species over time in response to changing habitat conditions. High regional biodiversity is maintained through the preservation of a mosaic of forest patches which each support communities of a particular composition (see York 1994). Large scale frequent burning would reduce this environmental heterogeneity and potentially reduce regional biodiversity. While small populations of invertebrates may survive single fires in islands of unburnt litter and vegetation (Campbell & Tanton 1981), large areas of optimal habitat for mid- and late-successional species would never develop under a regime of frequent fire. York (1994) has shown that a high proportion (60 per cent) of the ant biodiversity in these forests is comprised of rare species, many with specialist habitat requirements. The probability of extinction for these species in a fire-prone environment would therefore be high, with serious implications for the maintenance of regional biodiversity.
Given that many of the dry sclerophyll forests of eastern Australia have a long but poorly documented history of high-frequency burning, the validity of these conclusions can only be appraised through long-term experimental investigation involving fire exclusion. Data from such an experiment were presented here, involving a 20-year study of frequent fuel-reduction burning with matched burnt/unburnt sites. Dry sclerophyll forests support a rich terrestrial invertebrate fauna which utilise the forest floor litter as habitat for feeding and shelter (New 1984). Over the 20 years of measurement in this study the litter environment has provided a relatively stable habitat with litter biomass fluctuating (on average) between about 14 and 23 tonnes per hectare in the absence of fire. Low-intensity autumn fires every three years resulted in an episodic loss of between 45 and 75 per cent of (dry) litter biomass on frequently burnt plots. Coastal blackbutt forests have however an annual litterfall rate of six to eight tonnes ha-1 (Birk & Bridges 1989) and so in these environments low-intensity burning has only a short-term impact on the amount of forest floor litter. Within three years after fire litter levels on burnt plots have reached (on average) 14.9 tonnes ha-1, ranging between 12.3 and 19.8 tonnes ha-1.
A diverse terrestrial invertebrate fauna was recorded in these forests with over 55,000 individuals from 25 broad taxonomic groups collected. It was apparent that the abundance of many of these groups had been affected by the repeated burning, with the numbers of organisms likely to be dependent on a stable litter layer substantially reduced. These included ticks and mites, isopods (slaters), springtails, amphipods, beetles and insect larvae. A number of groups were positively affected by fire: spiders, ants and hemipterans (bugs). As well as this response (or lack of it) to frequent burning, it was apparent that the abundance of many taxonomic groups fluctuated considerably between plots within the same treatment. This was the case for ticks and mites, isopods, bugs, beetles, flies and ants. This is seen as a likely response to spatial heterogeneity in habitat elements – a pattern frequently reported in the literature (see Neumann & Tolhurst 1991; Collett et al. 1993). This has implications for future sampling strategies and the usefulness of invertebrates as monitoring agents.
Ants have frequently been used in Australia to measure the effects of disturbance (fire, mining etc.) and subsequent rates of habitat recovery (see Majer et al. 1982, 1984; Andersen 1993). In this study it was apparent that frequent firing has caused a significant increase in the numbers of ants foraging on the ground surface. Other researchers have suggested that habitat simplification after fire is partially or wholly responsible for this frequently observed phenomenon, whereby ants are more active and easier to capture (Anderson & Yen 1985; Andersen 1988). In this study however the numbers caught on burnt plots were dramatically higher (>250 per cent) and had persisted for two years after fire, suggesting that the treatment effect was real and not purely an artefact of sampling. This is supported by analysis of community structure (see below).
If the observed increase in abundance was due to increased capture rate, then you would also expect to see an increase in the number of ant species (richness) on burnt plots as more uncommon species are collected. Given the previously described decline in species richness over time after fire, you would also expect the unburnt plots (20 years since last fire) to have substantially lower richness than the frequently burnt plots (two years since last fire). This is not the case, with mean species richness on burnt and unburnt plots not significantly different (19.5 and 18.2 respectively). These results suggest that as far as overall richness of ant communities is concerned, there has been no impact due to 20 years of frequent burning. Richness however is only one descriptor of a community, and it does not take into account the species present (composition) and their relative abundance. These two additional elements describe the structure of a community. In this study species assemblages were examined to identify whether any species characterised the unburnt and frequently burnt ant communities; either by their numerical dominance, or because they were found exclusively on plots receiving a particular burning treatment. These species could then be regarded as indicators of habitat conditions which had developed in response to the fire regime.
Frequently burnt plots were characterised by communities numerically dominated by Tetramorium sp.1, Notoncus sp.1 and Rhytidoponera metallica. Species of the genera Tetramorium and Notoncus are typical of dry, open forest habitats and are generally omnivorous in their dietary preferences. Rhytidoponera metallica has a broad diet, flexible foraging times, and is tolerant to a wide range of physical conditions (Greenslade 1979; Andersen 1991). It is a generalist which is characteristic of disturbed environments and was five times more abundant on burnt plots than unburnt plots in this study. Unburnt plots were characterised by low numbers of these species, but relatively high numbers of Pheidole sp. 2, Technomyrmex sp.1 and Bothroponera sp.1. Many species of the genus Pheidole are seed harvesters and their absence from burnt sites may be indicative of the loss of understorey plant species which regenerate from seed. Species of Pheidole are characterised by their flexible foraging times and broad food range which often includes seeds (Taylor & Brown 1985). Species of Technomyrmex and Bothroponera are usually typical of moist habitats and in this study are responding to the higher litter levels and associated moister soil conditions on unburnt plots.
Within species-rich communities it is typical for the majority of species to be of low abundance (see Krebs, 1985). Many are often referred to as uncommon or rare and they may have specific habitat requirements or low competitive abilities. These species may be useful as indicators of particular habitat conditions, and as specialists, may be particularly sensitive to habitat disturbance. When the distribution of uncommon or rare species is examined in this study, it is apparent that certain species are characteristic of either burnt and unburnt plots. Ten species were unique to unburnt plots (Crematogaster sp. 2, Discothyrea sp. 1, Hypoponera sp. 1 & 2, Strumigenys purplexa, Leptogenys sp. 1, Heteroponera imbellis, Technomyrmex sp. 1, Prolasius sp. 1 and Camponotus myopores) and these tend to have specific habitat and feeding requirements. Several are cryptic predators which are found in litter and rotting logs (Hypoponera spp., Strumigenys purplexa, Discothyrea sp. 1, Leptogenys sp. 1 and Heteroponera imbellis) (see Taylor & Brown 1985). Others are more numerous and have greater dietary flexibility (Technomyrmex sp. 1, Prolasius sp. 1); however, they still have specific environmental requirements and are indicative of moist habitats.
Seven species were unique to frequently burnt plots (Melophorus sp. 1, Brachyponera lutea, Campomyrma patiens, Colobostruma sp. 1, Myrmecia gulosa, Myrmecia sp. 1, Monomorium sp. 1. There is one specialist predator (Colobostruma sp. 1), but the others have broad dietary requirements typical of species with a generalist habit and broad environmental tolerances. One species (Rhytidoponera metallica) is well-known as a coloniser of disturbed habitats. Other species, such as Monomorium sp.1 and Tetramorium sp. 1, are known to harvest seed from grasses; their increased abundance on burnt plots probably reflects a compositional change in the understorey plant community from a shrub to grass dominance.
Programs in land assessment and applied resource management increasingly require an environmental indicator; this is an organism (or group of organisms) that reveals important aspects of the structure and function for some part of the ecosystem without exhaustive study of that part (Cornaby 1977). To form part of a standard evaluation system, ideal indicator taxa must meet a special set of criteria (Majer 1983; Greenslade 1984). Firstly they must be functionally important, preferably at both producer and consumer trophic levels, so that they are indeed reliable indicators of the total ecosystem. Secondly, they must be sufficiently abundant, diverse and widespread to be used in a wide variety of habitats. Thirdly, they must be practical to use in the sense of being readily sampled and processed. Finally, they must provide interpretable results, not only by being sensitive to environmental change but also by giving insights into the nature of, and mechanisms underlying, such change.
A wide range of terrestrial invertebrate groups has been used as bio-indicators in Australia. Examples include spiders (Mawson 1986), springtails (Greenslade & Greenslade 1987), termites (Nichols & Bunn 1980; Greenslade 1985), beetles (Greenslade 1985; Yen 1987) and ants (Majer 1984, 1985; Whelan et al. 1980; Yeatman & Greenslade 1980; Majer et al. 1984; Andersen &McKaige 1987). As a major component of the forest floor arthropod community in Australia, ants would appear to be ideal candidates for use as bio-indicators. With their many interactions with soil (Petal 1978), vegetation (Auld 1983; O'Dowd & Gill 1984; Westoby et al. 1982; Andersen 1987) and other arthropods (Briese 1982; Skinner &Whittaker 1981; De Baar 1985), ants play a vital role in energy flow and nutrient cycling in forest ecosystems. Ants have been used to monitor land rehabilitation projects (Majer 1978; Andersen 1993) as well as vegetation succession (Brian et al. 1976; Fox & Fox 1982), and appear to be ideal environmental indicators in studies of habitat disturbance or recovery, particularly in fire-prone areas (see Majer 1980, 1984; Andersen 1986).
Ecologically sustainable management is concerned with maintaining existing patterns of species and communities, maintaining productivity, maintaining natural processes, and preserving regional diversity (Lacey et al. 1990; York 1993a). In order to achieve these aims, faunal communities can be described by the use of several indices. The number of species (species richness) is a common measure, and the maintenance of species diversity is a common conservation goal (Woodward 1993). Declines in richness often indicate a simplification of community structure and, by implication, interference with natural processes and lowered productivity. Richness measures however do not take into account the composition of the communities that they describe. Two communities may have similar richness, but be composed of different species. These species may have different roles and be responding to specific environmental conditions. For example, in a study of mine rehabilitation Greenslade and Majer (1993) found that although the numbers of Collembola (mostly decomposers) sampled from forest and rehabilitated bauxite mines were relatively similar. The former habitat was characterised by native species, while cosmopolitan species tended to dominate the mined areas. This important point would have gone unnoticed had a knowledge of the species not been taken into account.
While this approach is applicable for well known taxa, it is often impossible for lesser known invertebrates. With these groups a common approach has been to use a higher level of taxonomic organisation. For example identification has only been to Order (e.g. Coleoptera, Acarina, etc), with community descriptions based on the abundance of individuals within each order or family (see Neumann & Tolhurst 1991; Collett et al. 1993). This approach provides a relatively quick but low-resolution measure for comparison and implies that overall trends shown by arthropod taxa give a cumulative estimate of the responses of individual species within each taxon (Neumann & Tolhurst 1991). Although repeated burning was shown (in this study) to significantly effect the abundance of a number of taxa (recorded at a coarse taxonomic level), interpretation of these results was somewhat ambiguous. As is its intention, frequent burning had resulted in a reduction in amounts of litter on the forest floor. Did the reduction in abundance of
some groups merely reflect a reduction in the amount of habitat? Similarly, was the increase in abundance of other groups merely reflecting increased trappability in a structurally simplified environment? These questions cannot be easily resolved without data at a reasonably fine level of taxonomic resolution. For many groups data at the level of genus or species are required in order to develop functional groupings of species which can be ecologically and meaningfully interpreted (see Andersen 1990; Burbidge et al. 1992).
In this study the effect of plot location within the forest area was shown to be a statistically significant factor influencing numerical abundance of many faunal groups, indicating that most taxa were responding to site-specific environmental conditions irrespective of the treatment. The implications of this result for future research and monitoring are substantial. If the number of individuals is to be the parameter by which communities are compared, then large sample sizes are required. This applies particularly to ticks and mites, isopods, beetles, and flies. Otherwise within-treatment effects will swamp any between-treatment effects, reducing the cost effectiveness of these groups of organisms for environmental monitoring. These effects were less pronounced for spiders, springtails, bugs, and ants.
More stable measures of community attributes would seem to be appropriate for the large-scale research and monitoring programs currently being developed by land management agencies. A finer scale of taxonomic identification would be the logical approach, and research into the use of RTU's (Recognisable Taxonomic Units) will help to overcome taxonomic difficulties (see Beattie 1993; Oliver & Beattie 1993). In this study, the RTU approach was tested using one largely homogenous group: the ants. Ants are members of a single family (Formicidae) and are readily identified to sub-family and genus using published keys. Within a genus, individual species can be identified and if no described name is available, a numeric code is assigned (eg. Notoncus sp.1). For ants, species richness showed less variability between plots within the same treatment than abundance data, thereby suggesting that lower sample sizes would be required to detect a real difference (if it occurs). Although many species remain undescribed, sufficient information on diet and habitat preference is known at the generic level that spatial patterns of relative abundance can be interpreted ecologically. In this study it was clear that ant communities were responding to habitat changes resulting from frequent firing. A knowledge of the habitat and dietary requirements of the constituent species permitted an interpretation of these changes.
It is clear from studies of single fire events that the scale and intensity of burning will influence the survivorship of invertebrate populations. Little is known however concerning the effects of an increase in the frequency and extent of fires. This paper has shown that for ants, the community changes in its composition in the years after fire. From an understanding of the roles that these species play within the ant communities studied, it was possible to conclude that these changes are in response to changes in habitat suitability for the constituent species (York 1989). It was hypothesised that frequent fire would eliminate those habitat specialists that required conditions which only developed later in the vegetation succession. This hypothesis was experimentally tested with a long-term frequent burning trial. Frequent low-intensity fire resulted in a change in ant community composition and structure. There was a loss of habitat and dietary specialists, an increase in the number of generalist species, and a dramatic increase in the abundance of a disturbance indicator species. This suggested that, although species richness may be maintained at a local scale, the widespread use of this management practice will result in the loss of invertebrate biodiversity on a regional basis.
This scale dependence has implications for the design of forest management strategies involving fuel-reduction burning. The use of frequent broad area burning is likely to have a much greater effect on biodiversity than a more diverse approach involving limited fuel reduction burning, strategic firebreak protection, alternative methods of hazard reduction (eg. grazing, slashing) and 'no burn' areas (see Ridley 1993). The maintenance of habitats specifically for their conservation value is therefore essential for the preservation of forest biodiversity. This is currently a tenet of 'multiple use management', but its success has not been tested in practice. In contrast, it is likely that areas which experience regular arson and extensive burning off are likely to suffer substantial losses in biodiversity. Strategies which involve restrictions to the spatial extent and frequency of fires, and the adoption of alternative methods of fuel reduction, may need to be implemented in spatially restricted or isolated forest patches. This applies particularly to areas of remnant urban bushland where the effects of fire may be compounded by other environmental pressures.
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