Ecological impacts of firewood collection - Department of ...

Ecological impacts of firewood collection 

— a literature review to inform firewood 

management on public land in Victoria 

Geoff Brown, Arn Tolsma, Simon Murphy, Anne Miehs, 

Ed McNabb and Alan York 

2009 

Picture goes here (in Header and Footer) 

Arthur Rylah Institute for Environmental Research 

and 

Department of Forest and Ecosystem Science, The University of Melbourne

Ecological impacts of firewood collection — a 

literature review to inform firewood management on 

public land in Victoria 

Geoff Brown, Arn Tolsma and Ed McNabb 

Arthur Rylah Institute for Environmental Research 

Department of Sustainability and Environment 

In partnership with: 

Simon Murphy, Anne Miehs and Alan York 

Department of Forest and Ecosystem Science, 

The University of Melbourne 

April, 2009

Published by the Victorian Government Department of Sustainability and Environment 

Melbourne, July 2010 

© The State of Victoria Department of Sustainability and Environment 2010 

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with the provisions of the Copyright Act 1968. 

Authorised by the Victorian Government, 8 Nicholson Street, East Melbourne 

ISBN 978-1-74242-699-0 (online) 

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Contents 

List of tables and figures...................................................................................................................v 

Acknowledgements.......................................................................................................................... vi 

Summary......................................................................................................................................... vii 

1 Introduction.............................................................................................................................1 

1.1 What is firewood CWD?...........................................................................................................3 

1.2 How does CWD differ between forest types?...........................................................................4 

1.3 Which areas are most affected? ................................................................................................5 

2 Ecosystem processes relating to CWD collection.................................................................8 

2.1 Soil and nutrient processes........................................................................................................8 

2.1.1 Nutrient cycling (see also 4.1.1.)...............................................................................8 

2.1.2 Carbon cycling (see also 4.1.2.) ..............................................................................10 

2.1.3 Soil and water quality (see also 4.1.3.)....................................................................11 

2.2 Habitat.....................................................................................................................................11 

2.2.1 Mammals.................................................................................................................12 

2.2.2 Birds ........................................................................................................................13 

2.2.3 Reptiles....................................................................................................................17 

2.2.4 Amphibians .............................................................................................................18 

2.2.5 Invertebrates ............................................................................................................19 

2.3 Flora........................................................................................................................................22 

2.4 Fungi and microbial organisms...............................................................................................22 

2.5 Fire considerations..................................................................................................................23 

2.6 Assessing the habitat quality of logs.......................................................................................27 

3 Harvesting operations...........................................................................................................30 

3.1 Forest types that provide firewood..........................................................................................31 

3.1.1 Mixed-species non-durable forests..........................................................................32 

3.1.2 Box-Ironbark (durable) forests................................................................................33 

3.1.3 River Red Gum (durable) forests ............................................................................36 

3.2 Types of thinning operations ..................................................................................................38 

3.2.1 Firewood fallen........................................................................................................38 

3.2.2 Commercial thinning...............................................................................................38 

3.2.3 Selective harvest......................................................................................................39 

3.2.4 Ecological thinning..................................................................................................40 

4 Ecosystem processes relating to harvesting........................................................................41 

4.1 Soil and nutrient processes......................................................................................................41 

4.1.1 Soil fertility (see also 2.1.1) ....................................................................................41 

iii

4.1.2 Carbon cycling (see also 2.1.2) ...............................................................................43 

4.1.3 Soil and water quality (see also 2.1.3).....................................................................46 

4.1.4 Forest hygiene and health........................................................................................47 

4.2 Tree hollow development .......................................................................................................50 

4.3 Habitat.....................................................................................................................................51 

4.3.1 Mammals.................................................................................................................52 

4.3.2 Birds ........................................................................................................................54 

4.3.3 Reptiles....................................................................................................................55 

4.3.4 Amphibians .............................................................................................................55 

4.3.5 Invertebrates ............................................................................................................56 

4.4 Flora........................................................................................................................................56 

4.4.1 Understorey .............................................................................................................56 

4.4.2 Eucalypt canopy ......................................................................................................60 

4.4.3 Nectar and pollen resources.....................................................................................61 

4.4.4 Cryptogams .............................................................................................................63 

4.5 Fungi and microbial organisms...............................................................................................64 

5 Which communities or species may be affected by firewood activities?..........................65 

5.1 Threatened EVCs and plant species........................................................................................65 

5.1.1 Vegetation communities..........................................................................................65 

5.1.2 Plant species ............................................................................................................67 

6 Knowledge gaps.....................................................................................................................74 

References........................................................................................................................................76 

Appendix 1 .....................................................................................................................................107 

Appendix 2 .....................................................................................................................................117 

iv

List of tables and figures 

List of tables 

Table 1.1 Fuel classes used by management specialists............................................................4 

Table 1.2 EVCs most likely to be affected by firewood harvesting, and log benchmarks........7 

Table 2.1 Mean concentrations of nutrients in debris before burning component of 

experiment (Nutrient concentrations g/kg) (Stewart and Flinn 1985). ...................................10 

Table 2.2 Threatened vertebrate taxa for the key Victorian bioregions of this review, 

compiled from the Atlas of Victorian Wildlife (DSE database), January 2009......................14 

Table 2.3 Examples demonstrating the array of CWD habitat features required to maintain 

saproxylic species diversity. ...................................................................................................20 

Table 3.1 Comparative estimates of biomass and broad average growth rates for three 

firewood forest types (Flinn et al. 2001).................................................................................32 

Table 3.2 Number of stems per hectare by diameter class in each working circle (Victorian 

Environmental Assessment Council 2001).............................................................................34 

Table 3.3 Stocking level, basal area and basal area distribution by species composition for 

DSE work-centres in the Bendigo Forest Management Area (Department of Natural 

Resources and Environment 1998). ........................................................................................34 

Table 4.1 Area, biomass and carbon density (above-ground) of Victoria’s predominant 

firewood forest types (Grierson et al. 1992)...........................................................................44 

Table 5.1 Bioregional conservation status of EVCs likely to be subject to firewood 

harvesting................................................................................................................................69 

Table 5.2 Summary of bioregional conservation status of EVCs likely to be subject to 

firewood harvesting. ...............................................................................................................71 

Table 5.3 EPBC or FFG-listed vascular plant species from forests and woodlands in three 

key bioregions likely to be subject to firewood harvesting. ...................................................71 

Table 5.4 Summary of listed rare and threatened species potentially affected by firewood 

harvesting in three key bioregions. .........................................................................................73 

List of figures 

Figure 1.1 Determination of bioregions. a) Forest Management Areas by firewood volume. 

b) Bioregions most relevant to this study..................................................................................2 


b) Bioregions most relevant to this study................................................................................68 

v

Acknowledgements 

This project was funded by the Division of Natural Resources, Department of Sustainability and 

Environment, Victoria. We thank the following individuals and agencies for their important 

contributions: 

DSE: Natural Resources: Joanne Wallace, Shaun Suitor, Lisa Saxton 

DSE: Arthur Rylah Institute for Environmental Research: Kasey Stamation, Richard Loyn, 

David Cheal 

DSE Land and Fire Management: Mary Camilleri, Jim Allen, Les Vearing (all Bendigo), Paul 

Bates (Maryborough) 

DSE: Statewide Services, Biodiversity Services: Peter Johnson, Peter Morison (both Bendigo), 

Jerry Alexander (Wodonga), Steven Deed (Ovens), Ryan Incoll (Traralgon) 

DSE: Biodiversity and Ecosystems Services: James Todd 

DSE: members of the Ecological Impacts and Opportunities Working Group 

DSE: Information and Business Technology: Sally Dwyer (East Melbourne) 

DPI: staff of the Knowledge Resource Centre (Werribee) 

CSIRO Sustainable Ecosystems (Canberra): Jacqui Stol 

Monash University: Greg Horrocks, Jody Taylor 

NSW DPI Science and Research Division: Brad Law 

vi

Summary 

In this commissioned review we report on the ecological impacts of firewood supply, from: (1) 

collection of ground woody debris – ‘dry’ firewood, and (2) harvesting of living trees for firewood 

– ‘green’ firewood. The review is to inform the development of a Victorian statewide strategy that 

ensures that firewood supply from public land has a sustainable future, and that the associated 

environmental impacts continue to be managed to a high standard. This report reviews existing 

scientific information and also provides information that has emerged since recent Australian 

reviews on firewood (or coarse woody debris (CWD)) availability and distribution, its biodiversity, 

and firewood demand and usage. 

For the purposes of this review we have adopted a broad definition of firewood: the detrital 

biomass encompassing a wide variety of material, including standing dead trees (also called snags 

or stags), stumps, dead branches, whole fallen trees, coarse roots and wood pieces that have 

resulted from the fragmentation of larger dead trees and logs. Firewood also includes residual 

wood generated from harvesting operations. 

We have taken as our primary focus those forests and woodlands of north-central Victoria that 

supply the bulk of the firewood market, although reference is also made to wetter ash-type forests 

of southern and eastern Victoria, even though they are not considered to be traditional sources of 

domestic or commercial firewood, because they provide much of the available ecological 

knowledge. 

To our knowledge there have been no empirical studies on the impacts of firewood supply on flora, 

fauna and ecosystem processes. However, the amount of inferential and correlative evidence is 

sizeable, particularly for fauna in relation to its use of CWD (including dead standing timber) and 

hollow-bearing trees and, to a lesser extent, timber harvesting residue. 

Vegetation 

Nine vegetation communities that are likely to be affected by firewood harvesting are listed under 

Victoria's Flora and Fauna Guarantee Act (FFG), while three ecological communities are listed 

under the Commonwealth Environmental Protection and Biodiversity Conservation Act (EPBC). 

About 60 vascular plant species that occur in forests or woodlands of concern are listed under the 

EPBC Act, FFG Act or both. However, the extent to which these species and communities will be 

affected by firewood harvesting or collection is unknown, as are the minimum levels of CWD that 

should be retained to allow ecosystem functions to operate. There may be some negative effects 

from soil disturbance associated with collection activities. 

Overstorey trees might benefit from the thinning of the canopy associated with firewood 

harvesting, with an increase in growth rates and eventually more profuse flowering. However, 

selective cutting of preferred species might lead to long-term changes in overstorey composition, 

and affect the availability of floral and nectar resources upon which many fauna species rely. 

Understorey species may also benefit from canopy thinning, with increased light and other 

resources leading to increased vigour and flowering. However, some species, such as winterflowering 

orchids, might be disadvantaged by an increased light regime, while weeds may be 

encouraged by soil disturbance, particularly near tracks and roads. 

Vertebrates 

CWD is important for a multitude of Australian vertebrate species; logs are acknowledged by 

many authors as a critical resource for small Australian ground mammals. Logs provide nesting,

sheltering and foraging sites, food sources, particularly for insectivorous or mycophagous (fungusfeeding) 

mammals, facilitate movement, and can be important in the social behaviour of some 

forest-dependent taxa. Key mammal studies include the CWD manipulation research in northern 

Victoria and the study of Yellow-footed Antechinus Antechinus flavipes in a fragmented woodland 

landscape of the South West Slopes region of New South Wales. 

Fallen trees and branches as well as the residual wood from timber harvesting provide vital habitat 

for a range of birds. Twenty-one species of native birds are considered to be threatened by 

firewood collection in Australia; nineteen of these species occur in Victoria. One example, the 

hollow-nesting Brown Treecreeper Climacteris picumnus, forages predominantly amongst 

standing dead trees and logs, gleaning invertebrate prey from fissures and hollows. Studies have 

shown that densities of the Brown Treecreeper increased substantially in River Red Gum forests 

where fallen timber loads exceeded 40 t ha -1 . In Victorian box-ironbark forests in the Goldfields 

bioregion, bird numbers were found to be nine times greater, and bird species diversity three times 

greater, in areas containing piles of CWD. 

Many terrestrial reptile species are dependent on suitable structural heterogeneity in the ground 

strata, typically around CWD, and this has been documented for a number of Australian species in 

a variety of wet and dry forest types — reptiles use logs for a variety of purposes, including 

basking, nesting, shelter, hibernation and foraging. Large logs, which are able to retain moisture, 

may also provide refuge during drought or fire. 

The role of CWD in amphibian occurrence is poorly understood and therefore primarily 

inferential. The value of CWD for amphibians probably lies in its moisture holding qualities and 

its ability to provide refuge from environmental extremes (e.g. fire, temperature). Other qualities 

of CWD include the provision of calling sites for males, oviposition sites, refuge from predation, 

and probably even a contributing determinant of the composition of frog assemblages. 

The mammals of south-eastern Australia include many arboreal and aerial taxa that depend on 

hollow-bearing trees, as well as some facultative hollow users. The presence, abundance and 

taxonomic diversity of mammals have been correlated with the number of hollow-bearing trees, 

and tree size (dbhob) is significantly correlated with occupancy of tree-hollows, in both dead and 

live trees, by mammals. 

Large trees are known to be important for other woodland mammals. Woodland patches in 

southern New South Wales are more likely to support populations of Yellow-footed Antechinus if 

they contain, inter alia, larger trees of select species(Korodaj 2007). In the box-ironbark 

woodlands of central Victoria, gullies, which occupy a very limited area in the ecosystem, are 

known to support significantly greater numbers of some arboreal mammals compared with nongully 

sites. 

In Victoria, tree hollows are considered essential for 47 bird species, 14 of which are listed as 

threatened, which use them primarily for nesting or roosting. Many additional species nest on 

ledges or open hollows (e.g. woodswallows), or use hollows opportunistically. Some bird species 

require highly specific nest hollow characteristics; therefore, a diversity of hollow types is more 

likely to support a diversity of bird species. 

About 10% of the Australian reptile assemblage use hollows in Australia, as either den or nest 

sites, and by some reptiles as sources of prey. Two such threatened taxa in the ‘firewood’ regions 

of Victoria are the Tree Goanna and the Carpet Python, both of which utilise hollows in both large 

logs and large trees. To our knowledge there have not been any empirical studies on the use of 

hollow-bearing trees by frogs, although the number of arboreal frog species in south-eastern 

viii

Australia, principally from the Litoria genus, suggests that hollows are used, if only 

opportunistically. 

Site impacts 

Harvesting of standing forest as part of ‘green’ firewood-related operations is carried out in 

accordance with the Victorian Code of Practice for Timber Production and the specific guidelines 

and prescriptions that are applicable in that location. These, in general, are designed to allow the 

harvesting of firewood (a minor forest product) from GMZ and SMZ providing it is: (1) 

compatible with Forest Management Plan objectives, and; (2) it is for silvicultural, ecological, 

safety, or specific construction and maintenance requirements. Domestic Firewood Permits allow 

the conditional collection of ‘dry’ firewood from the forest floor. 

The long-term ecological condition of a site is influenced by the functional impacts of firewood 

collection and harvesting; by the way in which a site retains (or leaks) its soil, nutrient, carbon and 

water resources. Additionally, the relationship of fire, wind, extended drought and pests and 

disease with firewood management needs to be considered in the context of long-term ecological 

condition. 

Functional impacts will be affected by the spatial and temporal scales of harvesting and collection 

activities, and their intensity. These vary substantially, and consequently collection or harvesting 

may result in effects that are high-impact but localised, to low-impact but broadscale. The 

characteristics of specific sites will also vary considerably and influence the level of impact. 

These impacts have been considered under the following headings: 

Soil fertility 

With appropriate management, the impact of firewood-related harvesting disturbance on soil 

fertility and associated forest productive capacity is likely to be only a minor element of the 

production and supply of sustainable firewood. While soil fertility can be affected by the loss of 

nutrients and carbon as ‘dry’ and ‘green’ firewood is removed, or through associated soil 

disturbance, a key issue is whether this may affect long-term forest health, productivity or other 

ecosystem processes. 

Most nutrient studies have focussed on the wetter forests, and typically include: Leaves;Stembark; 

Stemwood; Subordinate vegetation (understorey, shrubs and ground-layer), and; Litter layer. The 

nutrient content of CWD has been poorly reported, however, generally smaller pieces (i.e. 7 cm diam.). In drier forests it is 

expected that similar trends would be observed. 

Because the concentration of nutrients in wood is small relative to those in other parts of trees, 

collecting or harvesting part of the wood removes a relatively small nutrient store. However, if 

bark and smaller diameter branch and bole material is also removed then the amount of specific 

nutrients removed will increase significantly and could lead to longer-term impacts on some sites. 

Losses of nutrients such as N can be replaced by biological N2-fixation, and P from reserves and 

through the weathering of parent rock, however, Ca may be more problematic. Fire intensity and 

frequency can also be important considerations in nutrient budgets where these are influenced by 

firewood activities. 

Disturbance associated with dry and green firewood removal will likely lead to some small 

decreases in soil organic carbon (SOC) due to oxidation of carbon in residues from the disturbance 

and in soil organic matter. However, this is unlikely to be significant given the limited soil mixing 

and compaction associated with firewood harvesting. The response of biomass carbon to 

ix

harvesting disturbance is most likely to be influenced by the inherent nature of the forest. Partial 

harvesting for firewood will stimulate some growth response, which is usually more rapid and 

vigorous in the more productive wetter forests and slower in drier forests. Collection of dry 

firewood from the forest floor is unlikely to cause any growth response and the net result will be a 

loss of carbon. However, the gradual decay of CWD or consumption by fire will also result in loss 

of carbon over time, with some level of carbon residue. In some management areas the collection 

of naturally fallen wood is not permitted. 

Carbon 

The impact on carbon budgets and Greenhouse gas (GHG) emissions of firewood-related 

disturbance is likely to be a minor element of the production and supply of sustainable firewood. 

Its potential to reduce fossil fuel use and attendant CO2 emissions, is dependent on a number of 

factors, including: forest growth rate, management, harvesting and transport systems, and; the 

efficiency with which firewood is burnt. This must be balanced against carbon losses from any 

reductions in CWD and soil organic carbon. 

Forests sequester carbon in biomass and as below ground carbon. CWD has been recognised as a 

quantitatively important component of the forest’s carbon stocks, equivalent to approx 10-20% of 

the above ground carbon biomass. However, generally little work has been conducted on the 

amount of carbon held in CWD in Australian systems. 

Carbon is ‘lost’ in wood taken off-site as part of the collection and harvesting of dry and green 

firewood. There are different management regimes under which this firewood removal can occur, 

each with a different impact on carbon balances. To affect an understanding of these different 

regimes simulation modelling is required which incorporates the following: forest growth; natural 

mortality; disturbance related mortality; fire impacts; forest product removals; decay rates; SOC 

losses; etc., to keep track of all the key carbon pools. It is important that appropriate time horizons 

for the analysis are used when modelling to explore the influence of carbon balances on net CO2 

emissions, otherwise misleading conclusions may be reached. The task of exploring the carbon 

impact of different firewood options is significant. 

Modelling has indicated that in regard to CO2 emissions firewood may be generally more 

favourable for domestic heating than other sources of domestic heating such as gas and electricity. 

Fuelwood modelling has found that for CO2 equivalent emissions, greenhouse balances are 

dominated by the potential savings due to the offset of fossil fuel emissions. Consequently, the 

type of energy generation that will be replaced by the use of the harvesting residues was critical to 

any evaluation. 

In normal forestry operations there is generally only a slight change, if any, to total soil carbon, 

however, the inclusion of soil cultivation can led to some reduced soil carbon storage, particularly 

in the labile carbon and microbial carbon fractions which make up 13-18% of SOC. Recalcitrant 

carbon, or the ‘stable’ carbon fraction, can make up 69-81% of SOC, with char (charcoal, black 

carbon) comprising about 13-27%. These fractions are considered to be generally inert 

components of the soil. 

Access 

With appropriate access management, the impact of firewood-related harvesting disturbance on 

water quality and forest health is likely to be a minor element of the sustainable production and 

supply of sustainable firewood. The physical disturbance associated with accessing ‘dry’ or 

‘green’ firewood, or with its production can impact on soil condition and water quality, and on 

x

forest health, with both the nature and timing of access significant influences. The factors that are 

most relevant to minimising soil disturbance and compaction are soil moisture content at the time 

of collection or harvest, machinery type, extraction track design and factors specific to soil type. 

Current codes of practice and management procedures ensure that the risk of connectivity between 

sources of sediment and drainage lines is minimised to acceptable levels, and the impact of 

harvesting operations is mainly found to be minimal. Generally, off-coupe road networks have 

been found to be the dominant source of sediment, with the landscape position of roading 

identified as a critical linkage factor together with the nature of road surfacing. 

The principal forest diseases that could impact on firewood operations are Armillaria root rot and 

Phytophthora, being known dieback diseases of mixed-eucalypt forest types. The local risk of tree 

mortality from these diseases will need to be evaluated, bearing in mind local conditions and the 

suitability of remedial techniques to help minimise risk. 

Fire 

The management for both planned and unplanned fire is important to the management of CWD. 

Fire is often the dominant disturbance in forests, and either directly or indirectly responsible for 

much of the creation of CWD from trees, contributing to tree injury, death and collapse, and also 

to the consumption of CWD. Fire management should be an integral part of the planning and 

implementation of any native forest silviculture, and consequently it is necessary to a consideration 

of the amount and nature of firewood which may be collected; as firewood removal impacts on the 

size and amount of woody debris fuels remaining on site. While information on CWD-related 

fauna species is scarce, the limited information indicates that some species are well adapted to fire, 

whilst others are more at risk, depending on fire frequency, timing and intensity. Large pieces of 

CWD have been described as “effective small-scale fire breaks” because of their greater ability to 

survive fire and the protection their larger size provides. 

Harvesting for firewood produces additional fuel loads and changed fuel drying conditions, which 

will likely increase fire risks. Planned fire can be an important consideration in managing these 

higher risks, and used to reduce fuel loads; removing much of the fine elevated fuel and some of 

the litter. However, burning of larger-diameter woody residue could cause substantial tree 

damage. Due to this type of damage, post-thinning burning is not generally recommended where 

wood degrade is likely to be unacceptable (eg. in ash and some mixed species regrowth). Where 

planned burning may be appropriate there are guidelines that assist in its implementation and the 

reduction of ecological impacts. Given adequate management of fuel hazard, any additional fire 

risk associated with harvesting is likely to be small. 

Firewood collection has been proposed as a way of reducing fuel loads and subsequent fire risk. 

The effectiveness of this approach will be influenced in particular by the standard of firewood 

utilisation. Removal of coarse woody material down to a small end diameter (under bark) of 

around 10cm will have little impact on the rate of fire spread - finer fuels (generally < 6mm) are 

more important to the flame height, fireline intensity and rate of spread of a fire. Coarse fuels do 

impact on the total heat output of the fire, which can affect soil heating, plant/tree injury and 

mortality. 

Burning for fuel reduction appears to be generally a more useful approach at the broader landscape 

scale than firewood collection for managing overall fuel hazard. At the smaller scale, CWD fuel 

manipulation by removal (firewood collection) or relocation may be a useful method of managing 

coarser fuel loads around high-value assets. 

xi

During fire suppression of wildfire, particularly in the “first-attack”, “mop-up” and “blacking-out” 

stages, the proximity of CWD and its impact on bulldozer activities and vehicle access can be an 

important consideration, especially on strategic firebreaks. 

Knowledge Gaps 

Most research has concentrated on the moist forests of eastern and south-eastern Australia where 

CWD production is higher, though the impacts on biodiversity and ecosystem processes are 

arguably less than those in woodlands — more research is required in these drier, less productive 

forests. There has been a tendency to utilise anecdotal observations and inferential evidence in the 

absence of empirical data and to conclude that particular taxa are likely to decline if this habitat 

resource was removed. We suggest a variety of key research areas based on a lack of information, 

particularly in dry forests and woodlands. We particularly note the lack of information on the nonvertebrate 

biota, particularly invertebrates, vascular flora and cryptogams, as well as the potential 

effects on ecosystem processes, such as nutrient, carbon and energy cycling, pollination, water 

cycling and filtration, decomposition, soil production and climate regulation. 

xii

1 Introduction 

Ecological impacts of firewood collection in Victoria — a literature review 

Since European settlement there has been extensive clearing of private land. Dry forests in 

particular, such as box-ironbark, have been cut over multiple times since the 1840s (Calder et al. 

1994; Environment Conservation Council 2001a; Newman 1961), as the hardness and durability of 

the timber made it suitable for fuel, structural applications, fence posts and sleepers. 

Ecological consequences of the removal of trees and logs include a shortage of suitable habitat 

logs, a substantial reduction in the number of standing mature or dead trees, and a significant 

increase in the density of small trees (Edgar 1958; Environment Conservation Council 2001a). 

Firewood removal remains a long-standing use of public land and is an important source of heating 

and energy for many people in regional Victoria. However, in recent years there have been 

increases in the number and area of National Parks and other conservation reserves, reducing the 

areas available from which to obtain firewood. This will place additional pressures on those forests 

that remain available, on top of a range of other pressures such as climate change, fragmentation 

and weed invasion. 

The Victorian Government is currently developing a statewide strategy to ensure that firewood 

collection from public land has a sustainable future, and that the environmental impacts from 

firewood collection continue to be managed to a high standard. 

In drafting this review we have taken into account two different methods of obtaining firewood: 

1. collection of ground woody debris – ‘dry’ firewood, and 

2. harvesting of living trees for firewood – ‘green’ firewood 

These two methods are elements of the domestic and commercial firewood supply chain as 

outlined below in Figure 1.1. The paths highlighted in yellow describe the two methods, with the 

path forest stand/material on the ground/processed into firewood describing “collection of dry 

firewood”, and forest stand/trees harvested/residual roundwood not suited to paper 

making/processed into firewood describing “harvesting for green firewood”. Where trees are 

harvested in Victoria’s State forests, DSE is responsible for ensuring compliance with the Code of 

Practice for timber production on public land (DSE 2007) and for the implementation of 

Management Procedures covering timber harvesting operations and associated activities (DSE 

2007). The main environmental focus of DSE is on those requirements associated with timber 

harvesting which either minimise or lead to an acceptable level of environmental impact. 

Firewood collection usually occurs in conjunction with forest management or production activities 

(such as timber harvesting or silvicultural operations, or other operation such as road works). 

All coupes in which tree falling is planned to occur for the production of firewood must be 

included in the Wood Utilisation Plan (WUP) or an approved Timber Release Plan (TRP). Coupes 

designated for the collection of firewood that was fallen during a previous harvesting activity do 

not require inclusion in the WUP. Domestic firewood permit or licence holders are not permitted 

to fall trees for domestic firewood. Tree felling is undertaken by appropriate DSE officers or by 

contractors engaged by DSE in accordance with the Management Procedures (DSE 2007) 

Domestic firewood collection requires a Domestic Firewood Permit or Forest Produce Licence, 

which amongst other things has conditions requiring compliance with either, (1) Standard 

Operation Procedure: Domestic Firewood collection in non-Timber Release Plan coupes (DSE 

2006), or (2) Standard Operation Procedure: Domestic Firewood collection in approved Timber 

Release Plan coupes (DSE 2006), depending on who has been responsible for the coupe. Persons 

Arthur Rylah Institute, DSE, and School of Forest & Ecosystem Science, The University of Melbourne 1


harvesting/collecting firewood for commercial purposes from coupes managed by DSE must hold 

a Forest Produce Licence. 

Figure 1.1 Domestic and commercial firewood supply chain, from forest stand (includes 

woodland) to end user (adapted from Sylva Systems (2007)). 

In this commissioned review we report on the ecological impacts of firewood collection and 

harvesting in Victoria. This report reviews existing scientific information and also provides 

information that has emerged since recent Australian reviews on firewood (or coarse woody debris 

(CWD)) availability and distribution, its biodiversity, and firewood demand and usage 

(Department of Natural Resources and Environment 2002; Department of Sustainability and 

Environment 2003e; Driscoll et al. 2000; Grove and Meggs 2003; Sylva Systems Pty Ltd 2007; 

Woldendorp and Keenan 2005; Woldendorp et al. 2002). 

The licensed collection and harvesting of firewood on public land in Victoria accounts for 

approximately 20% of all firewood in the state (Department of Natural Resources and 

Environment 2002). This review considers the impact of these operations on the flora, fauna and 

key ecological processes. Our primary focus is on the value for biodiversity, particularly 

threatened taxa, that is held by CWD on the forest floor as well as live and standing dead timber, 

and those Victorian bioregions or broad vegetation communities that provide the bulk of firewood, 

through legal collection. We also examine the effects that thinning of the canopy might have on 

subordinate vegetation strata. 



To our knowledge there have been no empirical studies on the impacts on flora, fauna and 

ecosystem processes of the collection or harvesting of firewood. However, the amount of 

inferential and correlative evidence is sizeable, particularly for fauna in relation to its use of CWD 

(including dead standing timber) and hollow-bearing trees and, to a lesser extent, timber harvesting 

residue — it is primarily this information that we draw upon for this report. We also note the lack 

of information on the non-vertebrate biota, particularly invertebrates, vascular flora and 

cryptogams, as well as the potential effects on ecosystem processes, such as nutrient, carbon and 

energy cycling, pollination, water cycling and filtration, decomposition, soil production and 

climate regulation. 

In this report nomenclature for the vertebrate fauna follows Van Dyck and Strahan (2008) for 

mammals and bats, Christidis and Boles (2008) for birds, and Wilson and Swan (2008) and Cogger 

(2000) for the herpetofauna. 

1.1 What is firewood CWD? 

Coarse Woody Debris (CWD) occurs naturally as a result of branch or tree death, or as a byproduct 

of harvesting. CWD encompasses a broad range of material, including standing dead trees 

(aka ‘stags’ or ‘snags’), stumps, dead branches, whole fallen trees, coarse roots and wood pieces 

derived from the disintegration of larger stags and logs (Woldendorp and Keenan 2005). 

CWD loads are dependent the rate and timing of tree death, limb fall and trunk decline in forests 

and woodlands which are influenced by several factors, including the size and density of trees, 

wood durability, and disturbance regimes. Fire is one of these disturbances as are extended 

drought, pests and disease, wind damage, and various land use. For example, if silvicultural 

operations are undertaken, the amount of harvesting residue left in situ. 

It varies substantially between sites according to forest type and site history, and is considered to 

be as important as the living overstorey, leaf litter and soil components in maintaining biodiversity 

and conserving biodiversity (Australian and New Zealand Environment and Conservation Council 

2001b). However, it also represents an easy source of firewood. 

For the purposes of this review we have adopted a broad definition of firewood CWD, following 

Woldendorp and Keenan (2005) who describe CWD as detrital biomass encompassing a “wide 

variety of material, including standing dead trees (also called snags or stags), stumps, dead 

branches, whole fallen trees, coarse roots and wood pieces that have resulted from the 

fragmentation of larger snags and logs”. The definitions and size thresholds utilised to define 

CWD vary widely among researchers, making results between different studies and ecosystems 

difficult to interpret (Meggs 1996; Woldendorp and Keenan 2005); the largest discrepancy in the 

definition of the form of CWD appears to be whether standing dead trees and stumps are included 

in overall assessments. 

There is considerable variation in the minimum CWD diameter thresholds adopted by land 

managers and researchers. While these range from 1–25 cm (reviewed in Harmon et al. 1986), the 

minimum diameter threshold of 10 cm appears to be the most commonly utilised in forestry or 

wildlife work. Fire research generally categorises minimum diameter CWD in the range 2.5– 

7.5 cm. A maximum CWD diameter threshold is also apparent in some fire-fuel studies where 

CWD is characterised by its burning time (Table 1.1) (Woldendorp and Keenan 2005). 



Table 1.1 Fuel classes used by management specialists. 

Fuel class is related to time-lag, the time it takes for a fuel to lose 63 percent of the moisture content under a 

particular set of conditions (Maser et al. 1979). 

Fuel size Fuel time-lag class Range of time-lags Definition 

< 0.6 cm 1 - hour 0 - 2 hours Dead herbaceous and woody fuels 

less than 0.6 cm in diameter and the 

uppermost 0.6 cm of needles and 

leaves on the forest floor. 

0.6 – 2.5 cm 10 - hour 2 - 20 hours Dead fuels from 0.6 to 2.5 cm in 

diameter and litter from 0.6 to 2.5 

cm below the surface of the forest 

floor. 


diameter and litter from 2.5 to 10.2 

cm below surface of the forest floor. 


diameter and litter 10.2 cm below 

surface of the forest floor. 

> 20.3 cm > 10000 - hour > 2000 hours All dead fuels larger than 20.3 cm in 

diameter or more than 30.5 cm 

below surface of the forest floor. 

1.2 How does CWD differ between forest types? 

Few data exist on the amount of CWD expected to occur under 'natural' conditions in the various 

woodlands and forests from which firewood is sourced. Woldendorp and Keenan (2005) have 

summarised the findings of a literature review (to 2002) of CWD and fine litter quantities, and 

provide a breakdown, albeit in broad terms, of CWD loads by state and broad forest type. The 

latter category includes ‘woodland’ and ‘open forest’, the two forest types that account for the bulk 

of firewood collected in Victoria. The constituent vegetation communities of these two categories 

are very diverse; however, they serve to underscore the relative differences in CWD loads across 

the landscape. Mean mass of CWD for woodland was 18.9 t ha -1 , and for open forest 50.4 t ha -1 

(Woldendorp and Keenan 2005). 

Two recent Victorian studies serve to underscore the influence of forest type on CWD loads. 

Mac Nally et al. (2002a; 2000a; 2002c) have provided recent evidence for the post-European 

settlement depletion of CWD in floodplain forests of northern Victoria. Historical levels of CWD 

in this ecosystem were probably in the order of 125 t ha -1 ; current mean loads are approximately 

19 t ha -1 . 

Volumes of fallen logs across four age-classes of Mountain Ash Eucalyptus regnans forests of the 

Central Highlands of Victoria were reported by Lindenmayer et al. (1999) to be approximately 350 

m 3 ha -1 . This converts to 210 t ha -1 using the volume-mass conversion approach adopted by 

Mac Nally et al. (2002c), more than the 134.1 t ha -1 reported for Australia-wide ‘tall open forest’ 

by Woldendorp and Keenan (2005). To illustrate further, the lowest biomass estimation (0.2 t ha -1 ) 

was recorded in 5-year-old Mountain Ash forest that had been clear-felled and burnt, while the 



highest (1089 t ha -1 ) was from 63-year-old Messmate Stringybark Eucalyptus obliqua forest that 

had regenerated after fire (Woldendorp and Keenan 2005). In general, older forests contain a 

higher biomass of CWD than younger forest (Woldendorp and Keenan 2005). 

The Victorian Department of Sustainability and Environment (2008c) makes available EVC 

benchmarks — “An EVC benchmark is a standard vegetation-quality reference point relevant to 

the vegetation type that is applied in assessments. It represents the average characteristics of 

mature and apparently long-undisturbed stands of the same vegetation type. EVC benchmarks 

have been developed to assess the vegetation quality of the EVCs at the site scale in comparison to 

the ‘benchmark’ condition. Each EVC benchmark contains a range of information necessary for 

conducting a vegetation quality assessment” (2008c). One element of these benchmarks is the 

average amount of logs ha -1 — this amount is an estimate, using total log length (not volume), by 

experienced DSE ecologists. This measure is a conservative underestimate (to accommodate 

variability), yet it is the only state-wide measure available for CWD in most Victorian forest or 

woodland EVCs. Despite its obvious limitations, it nonetheless serves to provide a useful 

comparison of CWD loads in those forests and woodlands from which most of Victoria's firewood 

is obtained (Table 1.2). 

Substantial differences exist in the amount of CWD expected to occur in particular EVCs, with log 

benchmarks ranging from 50 to 300 linear metres of logs per hectare (Table 1.2). In previous 

research, total log lengths in 41 Grassy Dry Forest sites ranged from 61 to 696 m ha -1 (average 334 

m ha -1 ), while that in 85 Box-Ironbark Forest sites ranged from 13 to 530 m ha -1 (average 116 m 

ha -1 ), reflecting the substantial degree of disturbance in the latter (Arthur Rylah Institute for 

Environmental Research, unpublished data). 

1.3 Which areas are most affected? 

In Victoria, the proportions of retail firewood types vary dramatically. River Red Gum and box 

species dominate the northern floodplain forests and central Victorian woodlands (Goldfields 

bioregion), account for approximately 92% of all sales in the state (Driscoll et al. 2000). 

The Bendigo Forest Management Area, which incorporates much of Victoria’s box-ironbark forest 

and woodlands, accounts for approximately 21,000 tonnes of firewood per annum, more than 

twice the amount of firewood than the next most significant FMA (Midlands) (Department of 

Natural Resources and Environment 2002). The Mid Murray FMA, which roughly corresponds to 

the Murray Fans bioregion, is dominated by floodplain forests and box-ironbark woodlands. This 

FMA provides the third-most significant volume of firewood per annum in Victoria (4,000 tonnes, 

North East Catchment Management Authority 2004). 

Then primary foci of this report then are the forests and woodlands of north-central Victoria, from 

which the largest proportion of Victoria's firewood is obtained. Ash-type wet forests of southern 

and eastern Victoria are not considered to be traditional sources of domestic or commercial 

firewood, as their burning properties are poorer than most other eucalypt species. Also, the nonsawlog 

component of ash-type forests is normally used for domestic pulping and by the export 

woodchip markets, although small quantities of Mountain Ash Eucalyptus regnans have recently 

been sold by VicForests for firewood (Sylva Systems Pty Ltd 2007). In relation to ecological 

impact, green-firewood collection on ash-type coupes would follow the harvest of sawlogs and 

pulpwood and be taken from that component which is largely consumed by slash-burning as part 

of coupe regeneration. These high intensity fires remove residual debris and expose mineral earth 

seedbeds suitable for the sowing and growth of ash species (Flint and Fagg 2007). Consequently, 

removal of some of this residue as firewood, prior to burning, will have marginal ecological 



impact in the overall context of the harvest and regeneration operations. For these reasons ashtype 

wet forests will not be considered specifically in this literature review of Victoria’s main 

firewood species. However, there will be considerable reference to this forest type through the 

reviewed literature, as it provides much of the available knowledge. 



Table 1.2 EVCs most likely to be affected by firewood harvesting, and log benchmarks. 

Log benchmarks, derived from Department of Sustainability and Environment (2008c), provide a standard 

reference point relevant to the vegetation type that is applied in assessments; it represents the average 

characteristics of a mature and apparently long-undisturbed stand of the same vegetation type. 

EVC No. EVC Name Logs (m ha -1 ) 

16 Lowland Forest 200 

20 Heathy Dry Forest 200 

21 Shrubby Dry Forest 200 

22 Grassy Dry Forest 200 

23 Herb-rich Foothill Forest 200 

24 Foothill Box Ironbark Forest 200 

45 Shrubby Foothill Forest 200 

47 Valley Grassy Forest 200 

55 Plains Grassy Woodland 100 

56 Floodplain Riparian Woodland 300 

61 Box Ironbark Forest 200 

66 Low Rises Woodland 200 

68 Creekline Grassy Woodland 300 

69 Metamorphic Slopes Shrubby Woodland 150 

70 Hillcrest Herb-rich Woodland 150 

71 Hills Herb-rich Woodland 150 

80 Spring Soak Woodland 50 

103 Riverine Chenopod Woodland 50 

106 Grassy Riverine Forest 300 

127 Valley Heathy Forest 200 

128 Grassy Forest 200 

151 Plains Grassy Forest 200 

168 Drainage-line Aggregate 200 

169 Dry Valley Forest 200 

175 Grassy Woodland 150 

177 Valley Slopes Dry Forest 200 

198 Sedgy Riparian Woodland 200 

282 Shrubby Woodland 150 

295 Riverine Grassy Woodland 200 

641 Riparian Woodland 200 

652 Lunette Woodland 100 

659 Plains Riparian Shrubby Woodland 200 

663 Black Box Lignum Woodland 150 

679 Drainage-line Woodland 300 

704 Lateritic Woodland 150 

793 Damp Heathy Woodland 100 

803 Plains Woodland 100 

813 Intermittent Swampy Woodland 200 

814 Riverine Swamp Forest 200 

815 Riverine Swampy Woodland 100 

816 Sedgy Riverine Forest 200 

818 Shrubby Riverine Woodland 100 

823 Lignum Swampy Woodland 100 



2 Ecosystem processes relating to CWD collection 

Dead and dying trees and logs are key habitat components for a broad variety of plants, animals 

and microorganisms. These components of nearly all terrestrial ecosystems also play a crucial role 

in ecosystem processes and contribute to species richness. The myriad ecological functions of 

dead wood have been neatly summarised recently by several authors (e.g. Grove and Meggs 2003; 

Grove et al. 2002; Harmon et al. 1986; Lindenmayer et al. 2002; McComb and Lindenmayer 

1999) and include, inter alia, nutrient cycling and energy flow, carbon storage, soil conditioning, 

substrate for saproxylic (pertaining to dead or decaying wood) and epixylic (living on the surface 

of wood) organisms, refuge from environmental extremes, moisture reservoir, as well as habitat 

for fauna (e.g. provision of nesting, denning, shelter and feeding sites). 

CWD adds complexity to the forest floor, in so doing affecting the function of terrestrial systems. 

This has an important temporal dimension across forest stands; the stage of dying and decaying 

logs influences the occurrence of and use by fauna; animal taxa that use a particular stage of log 

(or tree) decay in one seral stage may differ from those that can use the same type of log (or tree) 

in another seral stage (McComb and Lindenmayer 1999). The availability of logs varies with 

stand age in Victorian box-ironbark forests; older growth forests have been found to contain more 

than three times the density of logs and nine times the volume of logs than stands of young 

regrowth (Venosta 2001). 

The long-term ecological condition of a site is also influenced by the functional impacts of 

firewood harvesting (Freudenberger et al. 2004). In other words, changes to the way in which a 

site retains (or leaks) its soil, nutrient, litter and water resources after disturbances such as 

harvesting, will affect the function of a site. Sustainable firewood harvesting will depend on a 

site’s capacity for tree regeneration; the capacity for tree regeneration is essential for sustainable 

firewood harvesting, thus the satisfactory regeneration of canopy trees relies on the ecosystem 

functioning properly. 

The value of CWD (including dead and standing material) for biodiversity and ecological 

processes is recognised in the official listing in Victoria of (1) the loss of coarse woody debris 

from Victorian native forests and woodlands as a key threatening process (Department of 

Sustainability and Environment 2008b) and (2) the loss of hollow-bearing trees from Victorian 

native forests as a key threatening process (Department of Sustainability and Environment 2008b). 

The removal of dead wood and dead trees is also officially listed as a key threatening process in 

NSW (Department of Environment and Climate Change 2008). 

In the following sections we elaborate on the ecological functions associated with CWD, and 

possible alterations to ecosystem processes associated with CWD collection. However, the spatial 

and temporal scales of harvesting and collection activities, and their intensity, vary substantially. 

Consequently, collection or harvesting may result in effects that are high-impact but localised, to 

low-impact but broadscale. It is not our intent to differentiate between these two extremes, but the 

reader must nonetheless be mindful of the implications of scale. 

2.1 Soil and nutrient processes 

2.1.1 Nutrient cycling (see also 4.1.1.) 

CWD is a structurally and chemically heterogeneous substrate (Brown et al. 1996a). It consists of 

a number of layers including the outer and inner bark, the sapwood and heartwood (Mackensen et 

al. 2003). The inner bark usually decomposes most quickly. It contains the cambium and phloem 



and is rich in sugars (Brown et al. 1996a). The sapwood usually decays faster than the heartwood. 

Heartwood is the largest component of CWD and decomposes more slowly than the other wood 

components, in part because it has lower carbon:nutrient ratios (Mackensen et al. 2003). Many 

nutrients occur in freshly fallen CWD in low concentrations, but as decomposition proceeds and 

carbon is lost via respiration, the concentration of nutrients may increase (Brown et al. 1996a). In 

many cases the carbon (C):nitrogen (N) ratio decreases as decay proceeds (Mackensen et al. 

2003). There are approximately seventeen elements essential to higher plants, but C, hydrogen (H) 

and O (oxygen) make up most of undecayed CWD. The elements phosphorus (P), magnesium 

(Mg), iron (Fe) and sodium (Na) are more concentrated in sapwood while manganese (Mn) is 

more concentrated in the heartwood. Nitrogen content is higher in the sapwood than it is in the 

heartwood (Brown et al. 1996a). Unfortunately, CWD has largely been ignored by soil scientists, 

even though it is a significant source of soil organic matter (McKenzie et al. 2000). 

Decomposing CWD enables a large proportion of the nutrients accumulated by living trees to be 

returned to the soil for reabsorption by flora and other organisms (Franklin et al. 1987; Grove et al. 

2002). The amount of nutrients returned is dependent on the input of CWD into the system. 

Decomposition is brought about by fungi and micro-organisms, often assisted by invertebrates. As 

decomposition progresses CWD enriches the soil (Bull et al. 1997; Sollins et al. 1987). For 

example, mulga Acacia aneura log mounds have been described as fertile patches within the semiarid 

woodlands of eastern Australia. Their soils differ from surrounding soils in having 

significantly greater amounts of mineralizable nitrogen as well as more organic carbon and total 

nitrogen. They appear to be more suitable for the growth of perennial herbs (Tongway and 

Ludwig 1989). A study in a mixed Tasmanian forest showed that CWD had significantly higher 

concentrations of Nitrogen than soil in half of the study sites (McKenny and Kirkpatrick 1999). 

While leaf litter decomposition has been widely studied, the break down of CWD has received 

little attention (Brown et al. 1996b). As most woods are high in polymeric material and low in 

soluble substrates (Harmon et al. 1986), there is an expectation of slow rates of mass loss and 

mineralisation of nutrients from woody material compared with leaf litter components (Brown et 

al. 1996b). Nutrient concentrations tend to be higher in bark than in wood pieces and smaller 

pieces of CWD (i.e. 3-5 cm diameter) tend to be higher in nutrients than larger ones (i.e. 10-15 cm 

in diameter). A decay study of CWD in Western Australian forests found Nitrogen was the only 

nutrient to be immobilised over the 5-year study period (Brown et al. 1996b) and research 

conducted in a pine plantation in the ACT found that only 12% of the original Nitrogen was 

released in eight years of decay exhibiting the length of time it takes for nutrients to be returned to 

the soil. Research examining the nutrient content of CWD in an open eucalypt forest in northeastern 

Victoria found CWD had the ability to contribute to soil nutrients (Stewart and Flinn 

1985). Some nutrient concentrations in woody debris from this study are outlined in Table 2.1. 

With improved water and nutrient conservation, CWD should help provide better conditions for 

seedling germination and plant growth. However, drier forests such as Box-Ironbark are often low 

in nutrients with low water-holding capacity (Muir et al. 1995), and localised increases in moisture 

and nutrients might favour weed species. In Box-Ironbark and Heathy Dry forests, the author 

(AT) has observed piles of branches acting as 'run-on' zones for the accumulation of water and 

nutrients, encouraging high localised cover of the short-lived weeds Large Quaking-grass Briza 

maxima and Hair-grass Aira spp. 

CWD adds complexity to ecosystems (Harmon et al. 1986; Lindenmayer et al. 2006), but its 

removal for firewood will simplify those ecosystems by reducing CWD-associated taxa and 

ecosystem pathways and functions. Long-term site productivity, particularly in lower-nutrient 

sites, may be reduced (Davidson et al. 2007; Harmon et al. 1986), with implications for forest 

sustainability. The degree to which nutrient cycling would be affected depends heavily on the 



intensity of firewood harvesting and collection. Nonetheless, codes of forest practice should 

recognise the importance of CWD as part of ecosystem function (Lindenmayer and McCarthy 

2002). Some DSE Forest Management Plans place conditions on licenced firewood cutters that do 

not permit the collection of naturally fallen wood or the harvest of dead standing trees (DSE 2001) 

2.1.2 Carbon cycling (see also 4.1.2.) 

Forests sequester carbon in biomass and through plant residues in the soil, with the accumulation 

of above ground carbon generally reflecting forest growth and productive capacity. Below ground, 

carbon accumulation is affected by root growth and soil-carbon balances. Soils are expected to 

increase in carbon, dependent on soil type, and then reach stability. Disturbances to CWD such as 

firewood collection lead to direct losses of carbon from the system followed by a process of reaccumulation 

during forest recovery. 

Table 2.1 Mean concentrations of nutrients in debris before burning component of 

experiment (Nutrient concentrations g/kg) (Stewart and Flinn 1985). 

Debris size Nitrogen Phosphorus Potassium Calcium Manganese 

≤6 mm 6.10 0.37 4.77 4.79 1.27 

6-30 2.05 0.10 1.12 1.94 0.45 

30-70 1.66 0.10 0.94 1.23 0.45 

≥70 1.26 0.06 0.45 1.26 0.32 

In addition to its role in nutrient cycling, CWD represents a large and long term store of carbon, 

which is gradually released through its decomposition (Brown et al. 1996b; Grove et al. 2002). 

Guo et al. (Guo et al. 2006) estimated that only 42% of carbon was released during eight years of 

the decay of Pinus radiata logs. During the decomposition process, microbes turn organically 

bound carbon (which accounts for approximately 50% of the organic material) into carbon dioxide 

(Mackensen and Bauhus 1999). The decay rate is slower in dry forests and is predicted to exceed 

25-30 years in most cases (Mackensen and Bauhus 1999). However, Barrett (2002) found the 

carbon turnover times in three Australian biomes to be reduced and more rapid in drier areas (i.e. 

23 years in tall forests, 4 years in arid shrubland, 3 years in open woodland). The amount of CWD 

in some areas is equivalent to approx 10-20% of the above ground carbon biomass, indicating that 

dead wood can represent a significant amount of carbon in forests (Delaney et al. 1998). 

Roxburgh et al. (2006) found in a temperate eucalypt forest in NSW that the mean carbon stock of 

CWD was 52.2 ± 15.6 tC/ha (tonnes of carbon equivalent per hectare). This made up 

approximately 19% of the total above-ground biomass. Guo et al. (2006) found CWD partly offset 

soil carbon losses after alterations in land use and removing CWD from sites may well reduce soil 

carbon. Unfortunately, little work has been conducted on the amount of carbon held in CWD in 

Australian systems. 

Modelling can be used to explore these losses on net CO2 emissions. The AGO’s FullCAM model 

(Paul et al. 2003) was developed to track carbon flows in a range of ecosystems. Paul et al. (2003) 

report that for remnant woodlands with a maximum aboveground biomass of about 77 t DM ha -1 

(tonnes of dry matter per hectare) three case studies were simulated over a 100 year period: (1) No 

firewood collection – dead wood resulting from tree death and litterfall was left on the ground to 

decompose; (2) Firewood collection – 80% of fallen dead wood was manually collected every five 



years, the rest remaining on-site to decompose; (3) Intense firewood collection – both dead trees 

and fallen wood were collected each year. The modelling found that the woodland systems were 

degrading because old dying trees were not being replaced, and there was a release of between 

about 30 and 60 t CO2 ha -1 . When firewood was collected every five years from the ground, net 

emission of greenhouse gas increased by 20.0 t CO2 ha -1 , and when firewood was collected each 

year from both the ground and dead trees, an extra 26.6 t CO2 ha -1 was emitted. The impact of 

harvesting in managed native forests on the net amount of CO2 emitted is explored in Section 4.1.2. 

2.1.3 Soil and water quality (see also 4.1.3.) 

The removal of CWD from the forest floor can expose the soil to wind and water, potentially 

leading to an increase in soil erosion and sedimentation (New South Wales Department of 

Environment and Climate Change 2003). CWD may help control the downslope movement of 

water, soil and litter on hillsides, reducing erosion and helping to capture sediment and organic 

matter (Harmon et al. 1986). On steeper slopes CWD is an important component of ‘surface 

roughness’, slowing down overland flow and aiding infiltration. Groves (run-on zones with CWD 

and plants) in Mulga Acacia aneura woodlands had a mean water infiltration rate that was 154% 

higher than in inter-zones (Berg and Dunkerley 2004). 

Additionally, the combination of CWD, saproxylic invertebrates such as termites, and decay fungi 

have been shown to increase soil and water quality by creating degraded-wood barriers and 

infiltration zones. For example, Mulga log mounds (where mounds develop around dead or fallen 

mulga trees due to termite activity and earth movement) have a higher water filtration rates and 

higher nutrient contents than soils away from the mounds (Tongway and Ludwig 1989). 

Bioturbation of soil can be observed around decayed pieces of CWD where fauna such as echidnas 

and fungi-eating macropods forage for food. This is often important to the breakdown of litter and 

soil-mixing processes and the development of macropores and good soil structure, which are 

important to water infiltration. 

However, soil and water quality are more likely to be affected by vehicle and machinery access to 

firewood areas than by removal of CWD. Surfaced roads are not normally constructed to these 

areas due to resource limitations and vehicles can damage tracks, compact soil and significantly 

impact on water quality. In some areas firewood collection is not permitted in winter to reduce 

damage to wet tracks (Department of Sustainability and Environment 2004a). These issues are 

reviewed in more detail in Section 4.1.3, dealing with harvesting impacts. 

2.2 Habitat 

Several reviews of the relationships between Australian fauna and key habitat and structural 

components of woodland and forest ecosystems, generally including dead and dying trees and 

logs, have been prepared in the last decade (Department of Natural Resources and Environment 

2002; Driscoll et al. 2000; Freudenberger et al. 2004; McElhinny et al. 2006). Each of these 

reviews has generally summarised the documented associations between the major vertebrate 

groups and vegetational structural attributes or complexity, though most studies reviewed have 

concerned birds and mammals (both ground and arboreal). By comparison, relatively few studies 

have investigated the habitat requirements of bats, reptiles, frogs or invertebrates, let alone 

microorganisms. 

There are few Australian (or Victorian) empirical studies that have as their focus the value of 

CWD for fauna, though the most notable recent exception has been the experimental study of 



vertebrate biodiversity in relation to CWD loads on the River Red Gum floodplains of northern 

Victoria (Mac Nally 2006; Mac Nally et al. 2002a; Mac Nally and Horrocks 2002; Mac Nally and 

Horrocks 2008; Mac Nally et al. 2002b; Mac Nally and Horrocks 2007; Mac Nally et al. 2001; 

Mac Nally et al. 2000a; Mac Nally et al. 2002c). In that study, the amount of manipulated CWD 

was found to influence the densities of select vertebrates (Yellow-footed Antechinus Antechinus 

flavipes, Brown Treecreeper Climacteris picumnus); both species responded strongly, through 

elevated densities, to increased loads of CWD. 

In another recent manipulation study, this time to determine whether faunal habitat was enhanced 

by coarse woody debris in semi-arid grasslands and woodlands of Terrick Terrick National Park, 

north-central Victoria, strategically placed fence-posts were used to mimic natural accumulations 

of fallen timber (Michael 2001; Michael et al. 2004). In that study there was evidence of seasonal 

and spatial usage of these refuges by several vertebrate species, including the threatened Fat-tailed 

Dunnart Sminthopsis crassicaudata and Curl Snake Suta suta. 

The key bioregions of this review support an array of threatened Victorian and Australian 

vertebrate fauna (Department of Sustainability and Environment 2007; Department of the 

Environment Water Heritage and the Arts 2008), many of which are dependent on or utilise logs or 

tree hollows (Table 2.2; Appendix 1). The categories for national and state conservation status for 

threatened vertebrate fauna follow those of the International Union for Conservation of Nature and 

Natural Resources (IUCN 2008). Taxa listed under the Victorian Flora and Fauna Guarantee Act 

1988 (FFG, Department of Sustainability and Environment 2008b) statutory lists of threatened 

taxa are also acknowledged. 

2.2.1 Mammals 

The relationship between CWD and the occurrence of many mammal species has been 

documented for a variety of ecosystems, though, for the purposes of this review, we have generally 

restricted our focus to Australia and, where information is available, Victoria. 

Driscoll et al. (2000) reported nearly a decade ago that international investigations of the 

relationship between CWD and mammals comprised a handful of correlational studies, no 

definitive experimental work and contrasting results, though there was evidence to show that some 

species are influenced by the existing range of woody debris. In the USA, voles (Bowman et al. 

2000) and shrews (Butts and McComb 2000) are known to be positively influenced by the cover 

and type of CWD. 

Since the review by Driscoll et al. (2000) several studies in different ecosystems of south-eastern 

Australia have demonstrated the importance of CWD for some Australian terrestrial mammals. 

These studies include the CWD manipulation research in northern Victoria mentioned above (Mac 

Nally et al. 2002a; Mac Nally and Horrocks 2008; Mac Nally et al. 2001; Michael 2001; Michael 

et al. 2004) and the study of Yellow-footed Antechinus Antechinus flavipes in a fragmented 

woodland landscape of the South West Slopes region of New South Wales (Korodaj 2007). 

Korodaj found that at the trap-site scale, greater structural complexity best explained occurrence 

patterns of A. flavipes, and that there was a strong association with hollow-bearing logs. 

Logs are acknowledged by many authors as a critical resource for small Australian ground 

mammals. Lindenmayer et al. (2002) and McElhinny et al. (2006) summarised the importance of 

logs as nesting, sheltering and foraging sites for many mammals, including many species that 

occur in south-eastern Australian forests and woodlands (e.g. Bush Rat Rattus fuscipes, Agile 

Antechinus Antechinus agilis, Dusky antechinus A. swainsonii, Eastern Quoll Dasyurus viverrinus, 



Mountain Brushtail Possum Trichosurus cunninghami, Short-beaked Echidna Tachyglossus 

aculeatus). Logs provide a large proportion of available shelter sites for many mammal species; 

for instance, in south-eastern Queensland, the Short-beaked Echidna typically preferred loghollows 

and depressions under the roots of fallen trees as shelter sites, disproportional to their 

availability (Menkhorst 1995; Wilkinson et al. 1998), and partially decayed logs in Tasmanian wet 

sclerophyll forest are important nest-sites for Pygmy-possums Cercartetus spp. (Duncan and 

Taylor 2001). 

Lindenmayer et al. (2002) also note the value of logs as important food sources, particularly for 

insectivorous or mycophagous (fungus-feeding) mammals. Logs are sites where hypogeous 

(underground fruiting) mycorrhizal fungi develop and become an important source of food for 

several forest-dwelling mammal taxa, like the Bush Rat, Southern Brown Bandicoot Isoodon 

obesulus, and Mountain Brushtail Possum Trichosurus cunninghami (as T. caninus) (Claridge 

1988; Claridge and Barry 2000; Claridge et al. 2000; Claridge and Lindenmayer 1998). A similar 

relationship has been found in the USA between small terrestrial mammals, fungi and decaying 

logs (e.g. Bowman et al. 2000; Bull 2002). 

Logs also facilitate movement for many terrestrial mammals, providing travel routes along or 

beside logs through undergrowth, and can be important in the social behaviour of some forestdependent 

taxa, such as the Common Wombat Vombatus ursinus and the Mountain Brushtail 

Possum, two species that deposit scats on logs to designate territory boundaries (Halstead-Smith 

1999; Lindenmayer et al. 2002; McElhinny et al. 2006). 

2.2.2 Birds 

Fallen trees and branches as well as the residual wood from timber harvesting provide vital habitat 

for a range of birds. 

Twenty-one species of native birds were considered by Garnett and Crowley (2000) to be 

threatened by firewood collection in Australia; nineteen of these species occur in Victoria. One 

example, the hollow-nesting Brown Treecreeper Climacteris picumnus, forages predominantly 

amongst standing dead trees and logs, gleaning invertebrate prey from fissures and hollows as well 

as from fallen branches on the ground below. Studies by Mac Nally et al. (2001) and Mac Nally et 

al. (2002b) have shown that densities of the Brown Treecreeper increased substantially in River 

Red Gum forests where fallen timber loads exceeded 40 t ha -1 . Other examples include the 

nocturnal Australian Owlet-nightjar Aegotheles cristatus, which roosts and nests in hollows in 

standing and fallen timber (EM pers. obs.), and the Bush Stone-curlew Burhinus grallarius, which 

roosts and forages amongst fallen logs. The Bush Stone-curlew nests beside a fallen log to avoid 

detection, relying on camouflage to avoid predation (Department of the Environment Water 

Heritage and the Arts 2005). Its current range is now largely confined to grassy woodlands (as in 

the Goldfields and Riverina bioregions in Victoria) (Department of the Environment Water 

Heritage and the Arts 2005). 

CWD provides shelter for species that forage in the lower strata. In Victorian box-ironbark forests 

in the Goldfields bioregion, bird numbers were found to be nine times greater, and bird species 

diversity three times greater, in areas containing piles of CWD (Laven and Mac Nally 1998) than 

in areas lacking such features. A range of bird taxa, strongly associated with logs for foraging or 

shelter in box-ironbark and River Red Gum forests from which large volumes of firewood are 

extracted were identified by Laven and Mac Nally (1998). These include: robins Petroica spp., 

Eastern Yellow Robin Eopsaltria australis, thornbills Acanthiza spp. and White-throated 

Treecreeper Cormobates leucophaeus. 



Table 2.2 Threatened vertebrate taxa for the key Victorian bioregions of this review, 

compiled from the Atlas of Victorian Wildlife (DSE database), January 2009. 

MAMMALS 

Victorian (Vict Cons and FFG code) and national (EPBC) threatened status* are shown. 

Genera are arranged alphabetically within Family and Families are arranged taxonomically 

within Order. Only extant non-vagrant Victorian native taxa are included. Appendix 1 

provides a full list of extant vertebrate fauna per bioregion and the use of CWD and hollowbearing 

trees by this fauna. 

Common name Scientific name EPBC Cons. Vict. FFG 

Dasyuridae Swamp Antechinus Antechinus minimus NT L 

Brush-tailed Phascogale Phascogale tapoatafa VU L 

Spot-tailed Quoll Dasyurus maculatus EN EN L 

Fat-tailed Dunnart Sminthopsis crassicaudata NT 

White-footed Dunnart Sminthopsis leucopus NT L 

Common Dunnart Sminthopsis murina VU 

Peramelidae Southern Brown Bandicoot Isoodon obesulus obesulus EN NT 

Eastern Barred Bandicoot Perameles gunnii EN CR L 

Petauridae Squirrel Glider Petaurus norfolcensis EN L 

Macropodidae Eastern Wallaroo Macropus robustus robustus EN L 

Brush-tailed Rock-wallaby Petrogale penicillata VU CR L 

Pteropodidae Grey-headed Flying-fox Pteropus poliocephalus VU VU L 

Rhinolophidae Eastern Horseshoe Bat Rhinolophus megaphyllus VU L 

Vespertilionidae Common Bent-wing Bat Miniopterus schreibersii (group) L 

Southern Myotis Myotis macropus NT 

Greater Long-eared Bat Nyctophilus timoriensis VU VU L 

Muridae Broad-toothed Rat Mastacomys fuscus DD 

Smoky Mouse Pseudomys fumeus EN CR L 

Canidae Dingo Canis lupus dingo NT 

BIRDS 

Megapodiidae Malleefowl Leipoa ocellata VU EN L 

Phasianidae Brown Quail Coturnix ypsilophora NT 

King Quail Excalfactoria chinensis EN L 

Anseranatidae Magpie Goose Anseranas semipalmata NT L 

Anatidae Australasian Shoveler Anas rhynchotis VU 

Hardhead Aythya australis VU 

Musk Duck Biziura lobata VU 

Cape Barren Goose Cereopsis novaehollandiae NT 

Blue-billed Duck Oxyura australis EN L 

Freckled Duck Stictonetta naevosa EN L 

Columbidae Diamond Dove Geopelia cuneata NT L 

Phalacrocoracidae Pied Cormorant Phalacrocorax varius NT 

Ardeidae Intermediate Egret Ardea intermedia CR L 

Eastern Great Egret Ardea modesta VU L 

Australasian Bittern Botaurus poiciloptilus EN L 

Little Egret Egretta garzetta EN L 

Australian Little Bittern Ixobrychus dubius EN L 

Nankeen Night Heron Nycticorax caledonicus NT 



Threskiornithidae Royal Spoonbill Platalea regia VU 

Glossy Ibis Plegadis falcinellus NT 

Accipitridae Grey Goshawk Accipiter novaehollandiae VU L 

Spotted Harrier Circus assimilis NT 

White-bellied Sea-Eagle Haliaeetus leucogaster VU L 

Square-tailed Kite Lophoictinia isura VU L 

Falconidae Grey Falcon Falco hypoleucos EN L 

Black Falcon Falco subniger VU 

Gruidae Brolga Grus rubicunda VU L 

Rallidae Lewin's Rail Lewinia pectoralis VU L 

Baillon's Crake Porzana pusilla VU L 

Otididae Australian Bustard Ardeotis australis CR L 

Burhinidae Bush Stone-curlew Burhinus grallarius EN L 

Charadriidae Inland Dotterel Charadrius australis VU 

Greater Sand Plover Charadrius leschenaultii VU 

Pacific Golden Plover Pluvialis fulva NT 

Pedionomidae Plains-wanderer Pedionomus torquatus VU CR L 

Rostratulidae Australian Painted Snipe Rostratula australis VU CR L 

Scolopacidae Common Sandpiper Actitis hypoleucos VU 

Red Knot Calidris canutus NT 

Pectoral Sandpiper Calidris melanotos NT 

Long-toed Stint Calidris subminuta NT 

Great Knot Calidris tenuirostris EN L 

Latham's Snipe Gallinago hardwickii NT 

Black-tailed Godwit Limosa limosa VU 

Eastern Curlew Numenius madagascariensis NT 

Wood Sandpiper Tringa glareola VU 

Turnicidae Red-chested Button-quail Turnix pyrrhothorax VU L 

Little Button-quail Turnix velox NT 

Glareolidae Australian Pratincole Stiltia isabella NT 

Laridae Whiskered Tern Chlidonias hybridus NT 

White-winged Black Tern Chlidonias leucopterus NT 

Gull-billed Tern Gelochelidon nilotica EN L 

Caspian Tern Hydroprogne caspia NT L 

Cacatuidae Glossy Black-Cockatoo Calyptorhynchus lathami VU L 

Major Mitchell's Cockatoo Lophocroa leadbeateri VU L 

Psittacidae Swift Parrot Lathamus discolor EN EN L 

Elegant Parrot Neophema elegans VU 

Turquoise Parrot Neophema pulchella NT L 

Regent Parrot Polytelis anthopeplus VU VU L 

Superb Parrot Polytelis swainsonii VU EN L 

Cuculidae Black-eared Cuckoo Chalcites osculans NT 

Strigidae Barking Owl Ninox connivens EN L 

Powerful Owl Ninox strenua VU L 

Tytonidae Masked Owl Tyto novaehollandiae EN L 

Sooty Owl Tyto tenebricosa VU L 

Alcedinidae Azure Kingfisher Ceyx azureus NT 

Halcyonidae Red-backed Kingfisher Todiramphus pyrrhopygia NT 

Climacteridae Brown Treecreeper 

(south-eastern ssp.) 

Climacteris picumnus victoriae NT 

Acanthizidae Rufous Fieldwren Calamanthus campestris NT 

Chestnut-rumped Heathwren Calamanthus pyrrhopygia VU L 

Speckled Warbler Chthonicola sagittata VU L 

Meliphagidae Regent Honeyeater Anthochaera phrygia EN CR L 



Painted Honeyeater Grantiella picta VU L 

Purple-gaped Honeyeater Lichenostomus cratitius VU 

Black-chinned Honeyeater Melithreptus gularis NT 

Pomatostomidae Grey-crowned Babbler Pomatostomus temporalis EN L 

Eupetidae Spotted Quail-thrush Cinclosoma punctatum NT 

Campephagidae Ground Cuckoo-shrike Coracina maxima VU L 

Pachycephalidae Crested Bellbird Oreoica gutturalis NT L 

Apostlebird Struthidea cinerea L 

Petroicidae Hooded Robin Melanodryas cucullata NT L 

Estrildidae Diamond Firetail Stagonopleura guttata VU L 

REPTILES 

Cheluidae Murray River Turtle Emydura macquarii DD L 

Broad-shelled Turtle Macrochelodina expansa EN L 

Agamidae Bearded Dragon Pogona barbata DD 

Pygopodidae Pink-tailed Worm-lizard Aprasia parapulchella VU EN L 

Striped Legless Lizard Delma impar VU EN L 

Hooded Scaly-foot Pygopus schraderi CR L 

Scincidae Swamp Skink Egernia coventryi VU L 

Alpine Water Skink Eulamprus kosciuskoi CR L 

Samphire Skink Morethia adelaidensis EN L 

Glossy Grass Skink Pseudechis rawlinsoni NT 

Varanidae Lace Goanna Varanus varius VU 

Boidae Carpet Python Morelia spilota metcalfei EN L 

Typhlopidae Woodland Blind Snake Ramphotyphlops proximus NT 

Elapidae Bandy Bandy Vermicella annulata NT L 

FROGS 

Hylidae Booroolong Tree Frog Litoria booroolongensis CR L 

Large Brown Tree Frog Litoria littlejohni VU NT L 

Growling Grass Frog Litoria raniformis VU EN L 

Spotted Tree Frog Litoria spenceri EN CR L 

Alpine Tree Frog Litoria verreauxii alpina VU CR L 

Myobatrachidae Giant Burrowing Frog Heleioporus australiacus VU VU L 

Giant Bullfrog Limnodynastes interioris CR L 

Baw Baw Frog Philoria frosti EN CR L 

Brown Toadlet Pseudophryne bibronii EN L 

Dendy's Toadlet Pseudophryne dendyi DD 

Smooth Toadlet Uperoleia laevigata DD 

Rugose Toadlet Uperoleia rugosa VU L 

* Status under the Victorian DSE Advisory List (Vic. Cons., Department of Sustainability and Environment 2007): CR – 

Critically Endangered, EN – Endangered, VU – Vulnerable, NT – Near Threatened; Status under the Victorian Flora and 

Fauna Guarantee Act 1988 (FFG) (Department of Sustainability and Environment 2007): L – Listed; Status under the 

Commonwealth Environmental Protection and Biodiversity Conservation Act 1999 (Department of the Environment 

Water Heritage and the Arts 2008): EN – Endangered, VU – Vulnerable. ^Type of hollow: H – enclosed hollow, L – 

may be ledge, crevice or below bark; symbols bracketed if other types of nest-site or roost-site are commonly used. + 

nests in hollows only in Tasmania. 


2.2.3 Reptiles 


Many terrestrial reptile species are dependent on suitable structural heterogeneity in the ground 

strata, typically around CWD, and this has been documented for a number of Australian species in 

a variety of wet and dry forest types — reptiles use logs for a variety of purposes, including 

basking, nesting, shelter, hibernation and foraging (e.g. Brown 2001; Brown and Nelson 1993b; 

Driscoll et al. 2000; Fischer et al. 2003; Henle 1989; Kanowski et al. 2006; Lindenmayer et al. 

2002; McElhinny et al. 2006; Melville and Swain 1997; Sumner et al. 1999). Large logs, which 

are able to retain moisture, may also provide refuge during drought or fire (McElhinny et al. 

2006). 

Many species of oviparous reptiles are known to lay their eggs in or under logs; indeed, some 

skink species demonstrate communal egg-laying, in which large aggregations of eggs are often 

deposited inside or under a log (Couper 1995; Porter 1993; Radder and Shine 2007; Wells 1981). 

At other times of the year, aggregations of some species can be found overwintering deep within 

rotting logs (Lindenmayer et al. 2002). 

Logs are important basking sites for heliothermic taxa — often, logs are used as elevated perches 

for basking, especially in wetter forests where a dense ground cover of vegetation may restrict 

basking opportunities (GB pers. obs.) — and can also play a vital role in the social behaviour of 

some species, exemplified by the territoriality displayed by some log-utilising skinks (e.g. 

Eulamprus spp.). Several reptile taxa (Southern Water Skink E. tympanum, Coventry’s Skink 

Niveoscincus coventryi, Spencer’s Skink Pseudemoia spenceri) in mesic Mountain Ash forest of 

the Victorian Central Highlands are arboreal or extensive users of logs, though empirical studies 

failed to find a significant positive association between skink abundances and total number or 

volume of logs, probably because logs are not a limiting factor in these environments (Brown and 

Nelson 1993a; b). However, counts of the arboreal Spencer’s Skink were significantly correlated 

with both the number of large trees and the number of highly decomposed logs (Brown and Nelson 

1993a; b). This raises the notion that CWD may be a more important habitat component in dry 

forests and woodlands than more mesic environments. 

In the River Red Gum forests of northern Victoria, reptile numbers are relatively low — this may 

be a reflection of historical impoverishment, perhaps as a consequence of broad-scale depletion of 

fallen timber (Mac Nally et al. 2001), or else naturally low occurrence in flood-prone 

environments. Nevertheless, these forests support several reptile taxa, including the threatened 

Inland Carpet Python Morelia spilota metcalfei. This large nocturnal predator is dependent on 

large hollow-bearing logs and trees in some systems, including the floodplain forests of northern 

Victoria (Department of Sustainability and Environment 2003c; Heard et al. 2004), where, 

Driscoll (2000) reports, the predominant choice of rest-sites of radio-tracked pythons were tree 

hollows and logs on the ground. 

A recent investigation of the reptile fauna of the Victorian Riverina found that this assemblage is 

in serious decline, and argued that this is primarily a result of changing land use across different 

spatial scales, including disturbance to structural complexity of vegetation and ground strata; 

specifically, it found the occurrence of total number of blind snakes to have significant positive 

relationship with the amount of coarse litter (Brown et al. 2008). The Riverina bioregion in 

Victoria currently supports a diverse, though diminished, reptile fauna, many species of which 

depend on CWD or hollow-bearing logs and trees (Brown and Bennett 1995; Brown 2002; Brown 

et al. 2008; Brown and Nicholls 1993). These species include, amongst others, the threatened Tree 

Goanna Varanus varius and Carpet Python, Tree Skink Egernia striolata, and several gecko 

species (Brown 2002). 



In the Midlands bioregion (primarily box-ironbark woodlands) of Victoria, historically the source 

of approximately half of the firewood in the state (Driscoll et al. 2000), reptiles are known to rely 

on structural components of the ground-layer and are disadvantaged by its disturbance or removal 

(Brown 2001). Brown (2001) found that reptiles were generally 2.4 times more abundant on 

‘undisturbed’ than ‘disturbed’ sites — where ‘disturbed’ sites had less structural and floristic 

diversity and less CWD — and that this disparity was also reflected in the number of species per 

site. The greater species richness and abundance of reptiles recorded for ‘undisturbed’ sites were 

attributed to the greater structural complexity of the ground strata on these sites. 

In their review of the ecological roles of logs in Australian forests, Lindenmayer et al. (2002) 

provided an extensive, though incomplete, list, sourced from general texts, of south-eastern 

Australian reptiles that utilise logs. This list, comprising nine families and fifty-seven species, 

serves to highlight the diversity of reptile taxa that depend on or utilise logs, as well as the dearth 

of dedicated research on this association. 

The Four-fingered Skink Carlia tetradactyla, a common resident of box and stringybark 

woodlands in north-eastern Victoria and southern New South Wales, is regularly observed in 

association with logs (GB pers. obs.). Recent modelling of the occurrence of this lizard at 

different spatial scales failed to identify a relationship between it and the abundance of fallen 

timber, though this was attributed to the preponderance of fallen timber across the study sites such 

that it wasn’t a limiting resource (Fischer et al. 2003). 

2.2.4 Amphibians 

The role of CWD in amphibian occurrence is poorly understood and therefore primarily inferential 

— we could find no Australian studies that have documented this relationship, although one study 

currently underway in fire-prone stringybark woodlands of south-western Victoria is investigating 

the association between select vertebrate taxa, including frogs, and CWD (Miehs et al. unpubl. 

data), and only a handful of international (American) studies have included amphibians (including 

salamanders) (Bull 2002; Butts and McComb 2000; McCay et al. 2002; Owens et al. 2008). 

It is easy to surmise that the value of CWD for amphibians lies in its moisture holding qualities 

and its ability to provide refuge from environmental extremes (e.g. fire, temperature) (Grove et al. 

2002). Other qualities of CWD, as reviewed by McElhinny et al. (2006) include the provision of 

calling sites for males, refuge from predation, and probably even a contributing determinant of the 

composition of frog assemblages. CWD also provides sites for oviposition — this is reported for 

several south-eastern Australian toadlet species (e.g. Pseudophryne spp., Anstis 2002; Chambers et 

al. 2006; Woodruff 1976a; b). 

In a study of the relationship between terrestrial vertebrate diversity and CWD in riverine 

floodplains of northern Victoria, Mac Nally et al. (2001) used pitfall traps to record frogs (as well 

as other vertebrates). While there was no significant difference in total frog records or species 

richness between River Red Gum sites with different CWD loads, about twice as many species 

were recorded on average at sites with CWD loads > 14 t/ha than sites with very low CWD loads 

(Mac Nally et al. 2001). 

While international studies are not the focus of this review, some hold particular relevance because 

they underscore the associations between herpetofauna and CWD, especially where local data are 

lacking. Two recent studies in the USA revealed the importance of this habitat component; in 

loblolly pine forests of south-eastern USA, where CWD loads were experimentally manipulated, 

increased capture rates of amphibians were related to increased loads of CWD (Owens et al. 



2008), and in Douglas-fir stands of north-western USA, the abundance of two salamander species 

increased with CWD volume (Butts and McComb 2000). 

2.2.5 Invertebrates 

Saproxylic invertebrates are a diverse and dominant functional group that are dependent on dead 

or dying wood during some part of their lifecycle, or upon wood-inhabiting fungi or the presence 

of other saproxylic species (Speight 1989 in Grove 2002a). Examples demonstrating the broad 

array of CWD habitat features required to maintain saproxylic species diversity are provided in 

Table 2.3. 

We could not source any Australian studies that have examined the impacts of firewood harvesting 

or collection on saproxylic invertebrate abundance, richness or distribution. Much of the work on 

the impacts of forestry has focussed on saproxylic beetles in the wet forests of Tasmania; more 

research is required in drier, less productive ecosystems and on other invertebrate groups. 

The dependence of saproxylic invertebrates on CWD and the impacts of forestry on their 

abundance and distribution has been extensively reviewed (Grove 2002b; Grove et al. 2002). 

Most invertebrate taxa have members from this guild (especially Coleoptera (beetles) and Diptera 

(flies)). Saproxylic insects make up a large proportion of the fauna in any forest (Grove 2002b; 

Grove et al. 2002). For example, one hundred and forty eight saproxylic beetle species were first 

identified during the first year of the Warra log-decay project in Tasmania (Grove and Bashford 

2003; Grove and Meggs 2003) and Yee (2005) found more than 350 beetle species associated with 

logs in an intermediate decay stage. Saproxylic species play an important role in the 

decomposition of CWD and are a key food source for other forest-dwelling organisms. Termites 

in particular are reported to have a great influence on the decomposition of CWD (Mackensen and 

Bauhus 1999). They are an important food source for an array of Australian vertebrates including 

frogs, skinks and small mammals (e.g. Craig et al. 2007; Pengilley 1971). 

Coarse woody debris is also a key habitat for generalist ground-dwelling invertebrate predators, 

such as spiders which utilise this habitat and leaf litter directly adjacent to the dead wood yet are 

not strictly dependent on this resource (Buddle 2001; Varady-Szabo and Buddle 2006). CWD can 

also influence the movement of fine litter through the forest and therefore contribute to the 

heterogeneity of the litter layer and patterns of ground cover (Lindenmayer et al. 2002), providing 

habitat for litter-dwelling fauna (Andrew et al. 2000). Other species such as the endangered stage 

beetle from south-eastern Tasmania are soil-dwelling but exhibit a preference for inhabiting the 

upper layer of soil underneath CWD (Meggs and Munks 2003). 



Table 2.3 Examples demonstrating the array of CWD habitat features required to maintain 

saproxylic species diversity. 

CWD feature Details Forest type Authors 

Abundance Stag beetle Lissotes latidens prefers forests 

with > 10 % ground cover of CWD 

Abundance Species richness and abundance of 

Coleoptera greater in plots where slash piles 

were retained rather than removed 

Charring The abundance of saproxylic beetles was 

higher on charred CWD and there was a 

higher species richness of pyrophilic beetles. 

Eleven red-listed species were found on 

charred CWD. 

Charring Charred CWD had a lower abundance of 

beetles than non charred CWD. Interestingly 

pyrophilous insects were almost exclusively 

confined to burned forest but occurred in 

both charred and uncharred CWD. 

Decay stage Less decayed CWD had greater species 

diversity, however web-building species 

were more diverse in more decayed logs. 

Decay stage Two species of tenebrionid species exhibited 

a preference for undecayed CWD reflecting 

field observations that they feed on hard 

wood. 

Decay stage Adult and larvae of the eucalyptus 

longhorned borer exhibited a preference for 

undecayed CWD 

Elevation 

(ground-level 

vs vertical) 

There was a lower abundance and species 

richness of spiders on elevated CWD. Less 

than half of the spider species collected on 

elevated wood were shared with those 

collected from ground-level CWD. 

Fungi One lucanid beetle species was associated 

with soft rot and another two species 

associated with brown rot (out of eight 

lucanid beetle species) 

Shading of 

CWD 

The assemblage of saproxylic beetles found 

in the shade treatments were significantly 

different than the control CWD. Four redlisted 

species were found on naturally 

shaded logs. 

Size Stag beetle Lissotes latidens prefers small 

(


Size One lucanid beetle species was found in 

significantly smaller diameter (i.e. 8 cm) 

CWD than the other seven species. However 

this may be related to rot type. 

Substrate type 

(logs, snags and 

tree tops) 

Abundance and species richness was higher 

in logs, however species composition varied 

between the three substrate types 

Broad-leaved forest in 

central Japan 

Boreal zone in northern 

Sweden and spruce 

dominated forest in 

Norway. 

Araya (1993) 

Hjältén et al. 

(2007) 

Research has highlighted the sensitivity of saproxylic invertebrates to forest management, with 

secondary forests generally supporting lower species abundance and richness than old-growth or 

primary forests. In Europe for example, many saproxylic species have gone extinct (Grove 2002a; 

Grove et al. 2002; Odor et al. 2006) and 542 saproxylic invertebrates have been red-listed in 

Sweden (Jonsell et al. 1998). This diminution of the invertebrate fauna is not only linked to an 

overall reduction in the amount of CWD left on the forest floor and the low dispersal abilities of 

some saproxylic invertebrates, but altered features of individual pieces of CWD (Table 3.3, 

Schmuki et al. 2006). Many saproxylic species exhibit preferences for larger diameter CWD. 

Research in Tasmania’s wet sclerophyll forest suggests that smaller diameter CWD (30 - 60 cm in 

diameter) do not exhibit all the rot types that large (> 100 cm) ones do. Therefore beetles that are 

dependent on particular rot types may not be present in the smaller pieces of CWD. Over the longterm 

this has management implications for the use of CWD as industrial firewood (Yee et al. 

2006; Yee et al. 2001). Araya (1993) found that three out of eight lucanid beetle species captured 

in a Japanese forest preferred a particular type of wood rot. 

Declines in one group caused by firewood removal could have indirect impacts on an array of 

other species and ecosystem processes, due to the co-adapted systems of these invertebrates with 

fungi and other fauna species. For example, the endangered large ant-blue butterfly Acrodipsas 

brisbanensis and its association with the threatened coconut ant Papyrius nitidus are jeopardized 

by the removal of CWD in Broadford for fire wood collection (Department of Sustainability and 

Environment 2003a). 

Many invertebrates also play a pivotal role in ecosystem processes by facilitating the entry of 

decay organisms into the heartwood of living trees. It remains unclear which saproxylic species 

also depend on living trees for part of their lifecycle (Hopkins et al. 2005). Local declines in these 

species could have a marked impact on the creation of hollows in trees, stags and CWD in forests. 

Trees >150 years old in Eucalyptus obliqua forests in Tasmania were found to contain higher 

amounts of decay and higher species richness of beetles and fungi (Hopkins et al. 2005; Yee et al. 

2006). 

Some invertebrates whose larval forms inhabit CWD are reported to be important pollinators in 

Australian forests. Unfortunately, no studies were sourced that examined this relationship in 

Australia. In other areas only limited information is available on invertebrate use, such as the use 

of tree-hollows in Australia (Gibbons and Lindenmayer 2002). 

Not only can firewood collection reduce habitat for particular species at a site, but the collection 

and transport of CWD can potentially alter the natural distributions of invertebrates by introducing 

species into new areas (Driscoll et al. 2000; Todd and Horwitz 1990). 


2.3 Flora 


Nationally, three quarters of people who collect firewood for personal use claim to target fallen 

timber (Driscoll et al. 2000). However, while CWD has been well documented as being important 

habitat for various fauna groups, its contribution to vegetation structure and ecological functioning 

is less well known. Existing research focuses almost exclusively on CWD in wet forests, not in 

the drier forests from which most firewood is collected (Driscoll et al. 2000). Therefore, this 

review can only provide a general overview rather than an in-depth summary. 

Moss cover is significantly higher on logs in older stands of Mountain Ash forest (Lindenmayer et 

al. 1999), where it can form thick mats. Other autotrophic (synthesising their own organic 

substances from inorganic material using light or chemical energy) taxa that commonly occur on 

CWD include lichens, liverworts, ferns, gymnosperms (conifers, cycads) and angiosperms 

(flowering plants) (Harmon et al. 1986). However, while CWD can be a substrate for seedling 

germination in some ecosystems (Harmon et al. 1986; Heinemann and Kitzberger 2006), it appears 

to be a feature restricted to wetter forests. For example, tree seedlings were more abundant on 

fallen logs than on adjacent ground in moist forests in Tasmania, but have not been observed on 

fallen wood in drier forests (McKenny and Kirkpatrick 1999). Where seedling growth occurs on 

CWD, it is generally slower than growth in mineral or organic soils, due to the lower 

concentrations of nutrients in CWD (Harmon et al. 1986). 

In some instances, the mass of small branches and foliage from a fallen branch or tree might act as 

a protective cage against grazing animals (Harmon et al. 1986; Kirkpatrick 1997), improving the 

rate of seedling survival. CWD may also moderate environmental extremes and provide shaded 

microsites for seedlings, particularly in disturbed areas (Harmon et al. 1986). However, increased 

leaf litter (as may be expected immediately following harvesting operations) may also have an 

adverse effect on seedling survival, at least initially, as noted in Jarrah Eucalyptus marginata 

forest (Stoneman et al. 1994). In forests or woodlands subject to firewood harvesting, the amount 

of litter would depend on the severity of the thinning operations, degree of post-harvest burning 

and the canopy size of felled trees. 

2.4 Fungi and microbial organisms 

Fungi and bacteria are highly specialised and perform an important role in ecosystem health and 

function. There are fungi (moulds and staining fungi) that live on the cell contents of dying and 

recently dead wood and those fungi (soft rots, white rots and brown rots) and bacteria that degrade 

already dead wood and break down cellulose and lignin (Harmon et al. 1986). No Australian 

studies were found that examined the relationship between CWD and bacteria. 

CWD hosts a wide range of fungi species that help to break down the wood and thereby eventually 

cycle nutrients back into the soil (Driscoll et al. 2000; Harmon et al. 1986; Lindenmayer et al. 

2002; O'Connell 1997). Nitrogen fixing can also occur in CWD, making it an important source of 

this element (Harmon et al. 1986). CWD may host mycorrhizal fungal species that have symbiotic 

associations with various vascular plant species (Driscoll et al. 2000), and those vascular species 

might be at risk if CWD is continually removed. Partial cutting of European oak-rich forests led to 

a significant decline in the richness of fungi species, particularly basidiomycetes (Norden et al. 

2008), although the authors warned against extrapolating these results to drier forests. 

Fungi are the principal agents of wood decay in terrestrial ecosystems and they provide habitat for 

many organisms and enable the regeneration of forests (Lonsdale et al. 2008). CWD is an 

important substrate for certain fungi (Andersson and Hytteborn 1991). For example, fruiting 



bodies of hypogeous mycorrhizal fungi are often produced in association with tree roots in CWD. 

These sporocarps provide nutrients to invertebrates and mammals (Scotts 1991). Mycorrhizal 

symbiosis is essential for tree growth and establishment. The species richness of CWD dependent 

fungi has been shown to increase with the abundance of substrate in international research (Odor et 

al. 2006). Amaranthus et al. (1994) found truffles were eight times more abundant in mature 

Douglas fir forests than in the surrounding plantations. Of the twenty one truffle species recorded, 

eight species only occurred in or under CWD. 

Different fungal species occupy and utilise CWD of differing host species, decay stages, various 

diameters and lengths (Kuffer and Senn-Irlet 2005; Sylva Systems Pty Ltd. 2002) highlighting the 

importance of maintaining a diversity of CWD. Küffer and Senn-Irlet (2005) found that even fine 

and very fine woody debris served as important refuges for many species in an array of Swiss 

forest types. However, the importance of CWD in fungi conservation may differ according to the 

individual species and the forests in which they inhabit. Claridge et al. (2000) developed a model 

for examining the habitat explainors of seven hypogeous fungal taxa sampled in 136 forested study 

sites in East Gippsland and NSW. Only one of the taxa exhibited a relationship with CWD and 

this was a negative one. Stag abundance also did not feature. Leaf litter depth and diversity of 

potential host eucalypt species were important explanatory variables. In thinning operations it 

would therefore be wise to keep a diversity of tree species and not just to concentrate on large tree 

species that can potentially develop large hollows. 

No information was found during this review relating to the amounts of CWD that are required to 

ensure the maintenance of adequate nutrient recycling and related ecosystem processes, or the 

possible effects of CWD removal in drier forests in Australia. There is also little available 

research on the decomposition of CWD in Australian forests (Mackensen et al. 2003). The small 

amount of research uncovered predominantly focused on the wet forests of Tasmania (i.e. Hopkins 

et al. 2005; Yee et al. 2006; Yee et al. 2001) and south-eastern Australia (East Gippsland) and 

adjacent New South Wales (Claridge and Barry 2000; Claridge et al. 2000). 

2.5 Fire considerations 

Planned burning to meet multiple objectives, including ecological and fuel hazard, and unplanned 

fire, such as wildfire, will impact on forest ecosystem processes at different scales and different 

intensities than disturbances such as harvesting. Fire management must be an integral part of the 

planning and implementation of any native forest silviculture (McCaw et al. 2001), and 

consequently it is critical to any consideration of the amount and nature of firewood which may be 

collected; as firewood removal impacts on the size and amount of woody debris fuels remaining on 

site. Fire is often the dominant disturbance in forests, and either directly or indirectly responsible 

for much of the creation of CWD from trees. Fires can cause or contribute to tree injury, death and 

collapse, and also to the consumption of CWD. 

Unplanned fire 

Unplanned fires by their nature usually show considerable variation in fire intensity across the 

burnt area. Consequently, it is not surprising to find that ecological studies also show that 

unplanned fire impacts are highly variable depending on a range of factors including fire intensity 

and forest type in particular. Higher-intensity unplanned fires (3000-70000 kW/m) will, 

depending on specific intensities and canopy height, have a more direct effect on forest structure. 

They will remove a greater proportion of tree canopy, tree bole bark, and more of the woody 

debris from the forest floor, as well as inducing greater soil heating and plant death and causing 

higher fauna mortality (DSE 2003). 



Stands of many eucalypts are seldom killed by fire, of any intensity, while others are killed by 

moderate fire intensities. For forest trees, bark thickness rather than type is the most important 

factor in protecting the cambium of eucalypts from lethal temperatures (McArthur 1968). During 

drier periods, severe fires are capable of completely killing individual trees, even where these 

species are fire resistant. Bark is drier and burns more readily and sap flow is usually much 

reduced and unable to carry heat away from the cambium (Gill and Ashton 1968, DSE 2003). 

Burning around and up a tree stem may lead to superficial damage and charring, death of part of 

the underlying cambium (forming a fire scar or 'dryside') or death of the stem. Young trees are 

most susceptible, as they have relatively thin bark and their crowns are close to the ground (Incol 

1981). Old trees may also be more susceptible to unplanned fire than trees of intermediate to 

mature age, able only to produce less vigorous epicormics than younger stems. Additionally, tree 

collapse is more likely in older trees that have previously been damaged by fire. Butt damage is 

common in many forests and is often related to the presence of large fuel accumulations near the 

base of trees (Gill 1981). Fire-related tree deaths are important in stand and CWD dynamics. Fire 

can result in significant thinning of stands, provide conditions for regeneration, and also contribute 

to the generation of new CWD. The death and collapse of larger-diameter trees is particularly 

important in relation to CWD because this material can potentially provide habitat for many 

decades, depending on rates of decay and fire consumption. 

In an unplanned fire, much of the pre-existing larger-diameter CWD may suffer little more than 

external charring, depending on moisture status. Some CWD-dependent species are firedependent 

and well adapted to disturbances of this nature. CWD may function as critical shelter 

during fires and provide remnant islands from which fauna and micro flora can colonise 

surrounding areas following burning (Lee et al. 1997; DSE 2003). Some skink species (e.g. 

Nannoscincus maccoyi and Sphenomorphus tympanum) have been observed under CWD after 

prescribed burns (Humphries 1992) and an investigation of invertebrates revealed they could 

shelter under CWD and survive fires, even though the leaf litter associated with it had been burnt. 

The area under CWD was found to reach lower temperatures than the surrounding litter and 

showed less moisture fluctuations (Campbell and Tanton 1981). CWD has also been identified as 

important habitat for maintaining ant species diversity in areas subject to frequent low intensity 

burns (Andrew et al. 2000) and lucanid beetle abundance in forestry clearfell burns (Michaels and 

Bonemissza, 1999). However, excessive soil heating is also reported to be concentrated beneath 

large pieces of CWD, particularly where they intersect (Brown et al. 2003). The security of these 

refuges may therefore be dependent on a number of factors including: CWD size and decay state 

as well as seasonal dryness and fire intensity (Humphries 1992). After a prolonged rain-free 

period, when soil moisture deficit is high (high Soil Dryness Index or Ketch Byram Drought 

Index), there is greater opportunity for larger-diameter CWD to dry, which increases the risk of it 

being consumed by fire. The risk of consumption is increased by CWD having more advanced 

decay, or large pieces of CWD being elevated or intersecting. 

Fire impacts on tree growth, with the radial stem growth of mixed-eucalypt species usually 

reduced following fire. In Messmate (Eucalyptus obliqua) and possibly Silvertop (E. sieberi), 

radial stem growth is likely to be reduced for two to four years following fire, depending on the 

severity of crown damage (Incoll 1981). The evidence also suggests that stand growth lost during 

this period is not regained (Kellas and Squire 1980). When fire intensity is insufficient to kill the 

cambium but sufficient to damage the phloem of a tree, gum veins form (Jacobs 1955). The 

shedding of epicormic shoots may also give rise to gum veins. However, when fire is severe 

enough to kill appreciable areas of cambium, the bark dies and the xylem is exposed to the entry of 

insect and decay organisms, contributing to tree hollow formation. Hollow formation can be 

further exacerbated by subsequent fires, as dry partially decayed wood is readily consumed. 



Few studies on the long-term frequency of unplanned fires are available. Analysis of damp forest 

in east Gippsland (Silvertop Ash) indicates an average fire-free period of 22.6 years over the last 

300 years (Woodgate et. al. 1994). Historically, it is suggested that most of the fires in this remote 

stand originated naturally. Closer to human habitation fire frequency and intensity is usually more 

variable. 

During fire suppression of unplanned fire, particularly in the “first-attack”, “mop-up” or blackingout” 

stages, the proximity of CWD can impact on the effectiveness or speed of different activities. 

The smouldering of CWD for days/weeks can impact on fire control, being a potential point of fire 

escape across containment lines (Tolhurst et. al 1992). Also, bulldozer fire-line construction and 

vehicle access can be hindered by heavy fuels (McCarthy et.al. 2003). 

Where harvesting/thinning produces additional fuel loads and there is an increased rate of drying 

associated with opening the forest canopy, this will likely have an impact on unplanned fire 

behaviour. Buckley and Corkish (1991) found that in east Gippsland regrowth forests dominated 

by Silvertop (Eucalyptus sieberi) and White Stringybark (E. globoidea), thinning significantly 

altered the type, quantity and distribution of fuels. Typically, they found that commercial 

thinning, where not more than 60% of the original basal area was removed, added about 10 t ha -1 a 

of leaf and twig material and about 14 t ha -1 of coarse fuels (2.6-10.0 cm diam.). They also found 

debris from previous harvesting and dead mature trees were critical factors, producing fire of 

higher intensity and duration which caused cambial butt damage on living trees. McCaw et al. 

(1997) reported a fuel loading of 76 t ha -1 of leaf litter and woody fuel 10 cm have not been shown to affect the 

rate of fire spread (Brown et al. 2003), like the finer fuels. Dynamically, these finer fuels are burnt 

in the continuous flaming zone of a fire and are hence important to the flame height, fireline 

intensity and rate of spread of a fire (Burrows 1994). Coarse fuels usually require fine fuels to be 

present before they ignite, and if they do burn they contribute little to the flame front (Luke and 

McArthur 1978). While coarser fuels do not significantly affect the rate of spread of fires (Cheney 

1990; Burrows 1994), their ignition does impact on the total heat output of the fire. Total heat 

output of the fire can affect things such as soil heating, plant death and convective updraughts 

(Burrows 1994). 

Research conducted by Cheney et al. (1980) on heavy fuels in an undisturbed forest in 

Tumbarumba indicates how unplanned fires can impact on CWD. They found that pieces of CWD 

>22.5 cm were not wholly consumed by low or high intensity fires unless they were highly 

decayed. They reported that 56% of CWD in the 10-20 cm diameter class and 26% in the >20 cm 

diameter class was consumed during a moderate intensity wildfire in the long unburned sub-alpine 

1 DSE, Bendigo 



eucalypt forest. The moisture content of larger woody fuels is much slower to respond to 

environmental conditions than fine fuels and therefore smaller diameter classes of CWD are most 

likely to be consumed by fires (Tolhurst et al. 2004). Large pieces of CWD have been described 

as “effective small-scale fire breaks” because of their greater ability to retain moisture and larger 

size (Andrew et al. 2000). 

Planned burning 

Planned burning to meet multiple objectives, including ecological and fuel hazard, can impact on 

CWD. Burning for fuel hazard reduction aims to reduce forest fuels so they are less available to 

unplanned fire and to subsequently influence its intensity and extent. Evidence is overwhelming 

for the effectiveness of this approach in reducing finer fuels, ‘flash fuels’ that contribute to the 

bulk of the flames, and which are typically dead woody material


proximity of debris (CWD) could cause significant fire damaged to Jarrah (Eucalyptus marginata) 

and Marri (Eucalyptus calophylla) if they were less than one metre away. Similar findings have 

been reported by Cheney et al (1990), Buckley and Corkish (19991) in Silvertop (Eucalyptus 

sieberi) regrowth forests, and McCaw et al. (2001) in Karri regrowth. Due to this type of damage, 

post-thinning burning is not generally recommended where wood degrade is likely to be 

unacceptable (eg. in ash and some mixed species regrowth) unless the area lies within a strategic 

burning corridor or area (Sebire and Fagg 1997). 

Sebire and Fagg (1997) and McCaw et al. (2001) have identified some strategies which may avoid 

the need to burn in thinned regrowth. These include either consolidation or broad dispersal of 

harvest areas, fuel reduction burning around thinned areas rather than in them, and selection of 

thinning coupes to avoid high fire hazard areas. Additionally, where ash regrowth is thinned Fagg 

(2006) has indicated that these areas should be located at least 1 km from current clear-felling 

coupes that will have slash-burns. Where fuel reduction burning may be appropriate, Sebire and 

Fagg (1997) have identified a number of factors that should be considered for mixed species 

regrowth. Similarly, Fagg and Bates (2009) have outlined factors influencing burning in boxironbark 

forests. Generally, these factors relate to: 

- positioning of fuels in relation to retained trees 

- timing of burning and lighting patterns 

- acceptable flame height 

- distance between tree crowns and fuel layer 

Given adequate management of fuel hazard, any additional fire risk associated with harvesting is 

likely to be small and should diminish further as fine fuel resulting from thinning breaks down 

within about 2-3 years (Sebire and Fagg (1997), Fagg and Bates (2009)). 

From the literature viewed, where burning for fuel reduction is used appropriately it appears to be 

generally a more useful approach at the broader landscape scale than firewood collection for 

managing overall fuel hazard. At the smaller scale, CWD fuel manipulation by removal (firewood 

collection) or relocation may be a useful method of managing coarse fuel loads around high-value 

assets (eg. very old or culturally significant trees). Where the risk of damage from fire is high, 

consideration should be given to the proximity of fuel and the potential impact to the tree in the 

case of fire, weighed against the impact of physically removing CWD fuel. 

This review is drawn largely from literature reporting on wet forests or drier mixed-eucalypt 

forests, rather than the Box-ironbark and River Red Gum forests where much of the firewood 

collection has traditionally occurred. They need to be viewed in this light. 

2.6 Assessing the habitat quality of logs 

While the primary focus of this review is the value that CWD (down and standing) holds for 

biodiversity, and how this may inform the development of firewood management on public land, 

the value of logs as important habitat elements is recognised in the DSE vegetation ‘net gain’ 

policy, the main goal of which is to achieve a reversal, across the entire landscape of the long-term 

decline in the extent and quality of native vegetation; this is set out in Native Vegetation 

Management: A Framework for Action released in 2002 (Department of Sustainability and 

Environment 2006). 

DSE manages native vegetation and forest products on public land to conserve biodiversity based 

on sustainability principles. This framework focuses on the need to restore the health of the 



environment while at the same time building a sustainable and competitive economy. The 

approach adopted seeks to improve the clarity and flexibility of native vegetation management 

and, in part, improve biodiversity outcomes through better strategic and regional planning, simpler 

regulations, flexible offset arrangements and incentive programs. Such incentive schemes assist 

landholders with their native vegetation management efforts, and result in both environmental and 

commercial gains. 

Programs such as BushTender and BushBroker pay landholders in return for delivering improved 

management of native vegetation under management agreements signed with DSE. EcoTender 

pays landholders for delivering broader environmental benefits through native vegetation-related 

activities. Under these schemes, retention of logs qualifies as a native vegetation gain. The 

amount of “log gain” is assessed by qualified DSE and agency staff, calculated using the DSE 

Gain Calculator and depends on the amount and type of logs currently on the site and the 

landholder commitments to forego any entitlement to remove these for the length of the 

agreement. Landholders establish the price required to deliver these management services either 

through a competitive auction (BushTender, EcoTender) or offset market (BushBroker) 

(http://www.dse.vic.gov.au). 

The assessment of the habitat quality of logs and gains from management requires consideration of 

multiple factors both for quantity (area, quality and time) and value (types and locations). Area 

and quality uses the habitat hectares approach that assesses the quality of the vegetation in 

comparison to a benchmark that represents the average characteristics of a mature and apparently 

long-undisturbed state for the same vegetation type (Department of Sustainability and 

Environment 2004b; Parkes et al. 2003). 

Generating gains depends on the activities (delivering either active improvement or avoidance of 

future impacts) and the ‘starting quality’ of the vegetation – see graph below. Gains are 

recognized for land manager commitments that forego a current use entitlement (e.g. grazing, 

firewood collection) that may otherwise contribute to the decline in vegetation quality over time 

(maintenance gains) and for land manager commitments beyond current obligations under 

legislation (e.g. weed control, supplementary planting) that improves the current vegetation quality 

(improvement gains). Security gains are also generated depending on the level of changed security 

of the site that averts a future risk of loss (e.g. state forest to nature conservation reserve, freehold 

land to on-title agreement). 

Value is assessed using the conservation status of vegetation types (EVCs at the bioregional scale; 

listed floristic communities) and habitat types (including the relative habitat quality of locations), 

and other recognised criteria for significance. 

Under habitat hectares, logs are assessed according to the observed amount and type (large / small) 

of logs per unit area in comparison to the relevant Bioregional EVC benchmark (see also Section 

1.2 How does CWD differ between forest types?). This contributes 5% of the overall vegetation 

quality score (Department of Sustainability and Environment 2004b; Parkes et al. 2003). 

Maintenance gains for logs can be scored where a land manager is currently entitled to remove 

fallen timber and agrees to forego this entitlement (Department of Sustainability and Environment 

2006). 


Habitat 

Condition 

current 

score 


estimated improvement 

due to active management 

estimated continuing decline 

due to current or entitled uses 

10 years 

estimated likelihood that a future 

management change will cause 

decline 

100’s years 

Arthur Rylah Institute, DSE, and School of Forest & Ecosystem Science, The University of Melbourne 29 

?


3 Harvesting operations 

Wood harvesting practices in general are known to have had a detrimental impact on CWD and its 

associated biodiversity in Australian and international forests. Harvesting practices tend to reduce 

veteran trees and simplify stand structure with a bias towards younger trees (Kirby et al. 1998; 

Mac Nally et al. 2002a; Mac Nally et al. 2002c). This results in an overall reduction in CWD 

abundance in the long-term and a greater proportion of smaller size classes in logged stands, 

compared to primary forests or old-growth stands (Andersson and Hytteborn 1991; Grove 2001; 

2002a; b; Harmon et al. 1986; Kirby et al. 1998; Lindenmayer et al. 2002; Woldendorp and 

Keenan 2005; Yee et al. 2001). In production forestry, CWD is often referred to as “waste wood”, 

implying it could be put to better use (Grove et al. 2002; Maser et al. 1988; Yee et al. 2001) than 

fulfilling an ecological role. It is often subjected to mechanical damage from harvesting 

machinery, further altering its nature (Grove et al. 2002; Lindenmayer et al. 2002; Maser et al. 

1988). 

Timber utilisation and the silvicultural regimes that support it are linked to the management 

objectives for State Forest in that region. These objectives are based on state government policies 

relating to natural resource management, the Regional Forest Agreements that coordinate state and 

federal policies, Forest Management Plans and timber resource data that assist in the location and 

scheduling of individual coupes. 

Of particular importance are the Forest Management Plans that delineate areas available for 

harvesting. The Plans sub-divide State Forest into: 

• Special Protection Zone. These generally have special conservation values that may be 

incompatible with timber harvesting and where the precautionary principle dictates that 

harvesting not occur. 

• Special Management Zone. These also have conservation values; however, modified 

harvesting is permitted where it is compatible with the identified values. 

• General Management Zone. This area is available for timber harvesting after 

consideration and management for any conservation, social and economic values that are 

identified for the area. 

In all cases timber harvesting must be carried out in accordance with the Victorian Code of 

Practice for Timber Production (http://www.dse.vic.gov.au/dse/index.htm) and the specific 

guidelines and prescriptions that are applicable in that location. These, in general, require the 

reservation of filter and buffer strips to protect water quality and faunal values, reservation of 

habitat trees and the implementation of strategies that minimize or prevent soil movement within 

and from the harvested area. 

A number of silvicultural systems can be used, dependent on specific management objectives or 

priority, stand conditions (e.g. slope, forest structure, age/size class distribution, seed sources and 

availability, etc), and commercial factors, such as market access, availability and skill of the 

harvesting crew and appropriate machinery configurations. The selection of a silvicultural system 

may often be a compromise between the desirable outcome from a silviculture view and the 

economic and social realities of the task. The ‘triple bottom line’ of social, economic and 

environmental factors should be maximized. Often guidelines do not exist to assist in the 

decision-making process and a successful result relies on the skill of the forester planning and 

supervising the operation. A general principle that is applicable across most silvicultural systems 

is that the simpler the system, the easier it will be to implement; however, the risks associated with 

a simple system may also increase the possibility of failure. Conversely, the more complex the 



system, the more difficult and costly it will be to implement; however, it will also probably have 

more inbuilt safeguards to help ensure its successful implementation. 

Generally, the underlying aim of silviculture applied to Forest Management Areas where firewood 

is produced is to continually improve the existing forest structure by promoting health and growth 

of stands. This is achieved with thinning operations to promote a healthy multi-aged forest with 

more larger trees for conservation purposes and the remainder as a continued timber resource 

(DSE 2006, DSE 2008). Firewood is produced from the residue of harvesting operations, 

including thinning. Suppressed, poorly-formed or unhealthy stems are removed to maximise the 

growth and health of retained trees (thinning from below). These retained trees benefit from 

reduced competition for light, nutrients and water. Forest regeneration occurs primarily from new 

growth sprouting from the cut stumps, known as coppicing. Many of the traditional firewood 

forests are coppice regrowth forests that have been selectively cut-over several times (DSE 2006). 

The planning of harvesting by DSE is carried out using a Wood Utilisation Plan (WUP). This plan 

outlines areas (known as coupes) proposed for harvesting over a rolling 3 year period. The draft 

WUP is prepared in consultation with business units in DSE, such as Fire and Biodiversity. Public 

comment is also encouraged (DSE 2006). 

This review focuses on forests and harvesting operations where the generation of firewood is an 

intended product, rather than those operations where firewood is unintended. A brief review of the 

structure of these forests, their silviculture and the harvesting systems that are used will be 

conducted under forest type headings. More detailed descriptions of the harvesting and 

silvicultural systems are outlined in Section 3.2. 

3.1 Forest types that provide firewood 

The main firewood species and the forest management areas (FMAs) where they occur are 

outlined in Sylva Systems Pty Ltd (Sylva Systems Pty Ltd 2007) and identified as either common 

(non-durable) or durable species. These species are found in the following forest types: 

- Mixed-species (non-durable) forests: Eucalyptus baxteri, E. consideniana, E. dives, 

E. globoidea, E. macrorhyncha, E. muelleriana, E. obliqua, E. sieberi, E. viminalis 

- Box-ironbark (durable) forests: E. melliodora, E. microcarpa, E. polyanthemos, 

E. sideroxylon, E. tricarpa 

- River Red Gum (durable) forests: E. camaldulensis 

Flinn et al. (2001) provided comparative estimates of biomass and broad average growth rates for 

volume, as outlined in Table 3.1. They felt that the estimates for River Red Gum were 

conservatively low and the estimates for Mixed Species would not be realised without future 

attention to thinning, fire protection and disease management. 



Table 3.1 Comparative estimates of biomass and broad average growth rates for three 

firewood forest types (Flinn et al. 2001). 

Forest type Mean annual increment (m 3 /ha/yr) Dry weight 

Sawlog volume Gross bole volume (t/ha) 

Mixed-species 1.80 3.85 340 

Box-ironbark 0.10 0.22 85 

River Red Gum 0.25 0.55 135 

3.1.1 Mixed-species non-durable forests 

Significant statewide variation in soil types, elevations, aspect, climatic conditions, disturbance 

history and management are reflected in the nature of these ‘common species’ forests. Species 

variation, age, stand structure and health, understorey composition and many other forest 

characteristics are affected by these factors. 

Most stands are uneven-aged (more than three age classes present), but the extent to which this is 

so depends on their disturbance history. Past timber harvesting and unplanned fire events in these 

forests have resulted in extensive areas of eucalypt regrowth, particularly in East Gippsland. 

These forests are considered to be fire-prone, and all species have fire adaptive traits rather than 

population lifecycle adaptations found in wetter forests where regeneration is more dependent on 

seed (Ashton 1981). Unplanned fires periodically impact on these forests, so that dense patches of 

fire regrowth of varying ages are characteristic. Dependent on fire intensity, duration and 

frequency, these unplanned fires can result in an overstorey of senescent, dead and degraded trees, 

often referred to as overwood. 

Firewood Fallen (FWF)* 

Objective 

Site 

characteristics 

Prescription 

Thinning from Below (THB)* 

To reduce fire hazard and supply firewood 

Areas where firewood is lying on the ground as a result of natural events or previous 

forest operations 

All fallen timber available for collection 

Objectives To release larger better formed trees and allow them to increase their growth and 

accelerate hollow development, by removing the smaller and poorly formed trees from 

the stand. 

Site 

characteristics 

Uneven-aged stands, or young regrowth stands with trees suitable for use as firewood 

(10-30cm Diameter Breast Height Over Bark (dbhob)) 



Prescription Retention of at least 50% of the pre-harvest basal area, including trees with identified 

habitat values (e.g. hollows) and trees selected for multiple purposes. 

These forests are used to produce a range of wood products, specifically sawlogs, posts and poles, 

chop logs, and firewood. The silvicultural systems that are used to provide a sustainable supply of 

commercial and domestic firewood can be outlined, as follows: 

* Source: Department of Sustainability and Environment (2008e; 2009). 

3.1.2 Box-Ironbark (durable) forests 

Victoria's Box-Ironbark forests have been extensively disturbed since European settlement. The 

discovery of gold in the region, for example in Castlemaine in 1851, initiated dramatic long-term 

changes to the structure of these forests. Original stands of box-ironbark were clearfelled to 

provide timber and fuel for the mining industry and associated settlements. In the 1890s, the rapid 

expansion of the railway system across Victoria made additional demands for heavy construction 

and sleeper timbers from the box-ironbark forests. By the 1920s, when the newly created Forests 

Commission introduced forest utilisation controls, all box-ironbark forests, especially those near 

population centres, had been selectively cut-over several times (Department of Natural Resources 

and Environment 1998). 

The heavy cutting during the latter half of the nineteenth century resulted in seedling and coppice 

regeneration over extensive areas of these forests, while the supervised harvesting and thinning, 

commencing early this century, produced forests containing essentially two size-classes, with 

various strata of regrowth beneath older and larger overwood stems. Typically, over-wood stems 

are uniformly distributed with a total basal area of about 11 metres/ha, whereas regrowth occurs in 

clumps within which basal area may be equivalent to about 10 metres/ha, with individual stems 

often under intense competition (Kellas et al. 1998). Diameter growth in fully or over-stocked 

stands is very low and recruitment into larger size classes relies on reducing competition through 

death or removal of individual trees. Natural self-thinning in box-ironbark forests is slow because 

the trees are tolerant of extreme conditions so they tend to persist through droughts and fires. 

The current forest structure is indicated by data from the Bendigo Forest Management Area and 

Pyrenees Ranges, for predominantly merchantable stands of durable species, where there is an 

average of almost 500 stems per hectare, most being less than 25 cm diameter. However, there is 

considerable variation, as illustrated by the data in Tables 3.2 and 3.3. Table 3.3 provides 

summary stocking, basal area and basal area distribution by species for work centres in the 

Bendigo Forest Management Area. 



Table 3.2 Number of stems per hectare by diameter class in each working circle (Victorian 

Environmental Assessment Council 2001). 

Data from recommended parks and reserves has been excluded. 

______________________________________________________________________________ 

Stocking (stems per hectare) 

______________________________________________________________________________ 

Working Circle < 20 cm 20-40 cm 40-60 cm > 60 cm Total 

______________________________________________________________________________ 

St Arnaud 117 86 18 3 224 

Inglewood-Dunolly 207 85 13 0 305 

Avoca Maryborough 582 82 13 1 678 

Bendigo 451 65 9 0 525 

Castlemaine 607 84 6 0 697 

Rushworth-Heathcote 300 100 12 0 412 

______________________________________________________________________________ 

Table 3.3 Stocking level, basal area and basal area distribution by species composition for 

DSE work-centres in the Bendigo Forest Management Area (Department of 

Natural Resources and Environment 1998). 

Studies on thinning have indicated that removal of competing coppice and the wider spacing of 

trees will lead to improved growth on the remaining individual trees (Kellas et al. 1982). The 



results showed that individual Red Ironbark E. tricarpa trees retain a capacity to respond to 

reductions in competition in fully stocked stands. For regrowth (dbhob 20 cm), the response was slower. For both 

regrowth and overwood trees, total competition from all competitors appears more important than 

that from overwood or regrowth alone. The thinning response of early-1930’s Red Ironbark, in the 

absence of overwood, has been studied near Heathcote. The trial provided an opportunity to better 

understand the thinning response in a small stand situation, particularly in the 25-40 cm (dbhob) 

size class, and also to better understand the effect of the coppice on retained tree growth (Murphy 

and Forrester 2009 in prep.). Murphy and Forester (2009 in prep.) reported that over a ten-year 

period the heaviest thinning, the 33% retention thinning (for largest 100 trees ha -1 ), was the only 

treatment which significantly increased growth (basal area and volume). For this treatment, when 

coppice was retained there was a small and insignificant reduction in tree size compared to when 

coppice was removed. 

These studies show that while the box-ironbark forests have low productive capacity (relative to 

forests in the higher rainfall zones), significant growth responses can be expected with appropriate 

thinning regimes but responses may be limited to less than 10 years, requiring periodic thinning 

for sustained responses. However, impacts on other values would possibly not make this 

appropriate. 

While these forests have a history of low fire frequency, all species have fire adaptive traits. 

Natural disturbance appears generally to be infrequent. Early records indicate that wind damage, 

either through breakage or uprooting of trees (James Clow in Bride 1898; Brough Smyth 1878; 

Mitchell 1839, 2 Ron Hateley per. comm.) may have been a significant local disturbance. The 

areas affected by these tornadoes were reported as 500m or so wide and a few kilometres long, 

with the length apparently determined by topography. These are small areas of disturbance, but if 

‘tornadoes’ were relatively frequent or ‘nested’ then historically the overall disturbance over 

several hundred years could have been significant. 

These forests are used to produce a range of wood products, specifically sawlogs, posts and poles, 

and firewood. The silvicultural systems that are used to provide a sustainable supply of 

commercial and domestic firewood can be outlined, as follows: 

Single Tree Selection (STS)* 

Specific objective To produce sawn timber products and minimise impacts on species 

composition and forest structure. 

Site characteristics Previously thinned sites which contain trees up to 59cm dbhob 

Prescription Retention of at least 50% of the pre-thinning basal area, including all 

trees >60cm dbhob and trees with identified habitat values (e.g. 

hollows). Species composition to be maintained. 

Thinning from Below (THB)* 

Specific objective To release larger better formed trees and allow them to increase their 

growth and accelerate hollow development, by removing the smaller and 

poorly formed trees from the stand. 

2 Ron Hateley, DFES, University of Melbourne 



Site characteristics Regrowth dominated stands with trees suitable for use as firewood (


Objective To produce sawn timber products and minimise impacts on 

species composition and forest structure. 

Site characteristics Uneven-aged sites with scattered mature trees 

Prescription Felling individual mature trees, reducing basal area by 

generally less than 10%, at intervals (generally 10-15 year 

cycle) over the rotation. Identified habitat values (e.g. 

hollows) are retained. 

Thinning from Below (THB) 

Objectives To release larger better formed trees and allow them to 

increase their growth, by removing the smaller and poorly 

formed trees from the stand. 

Site characteristics Stands dominated by young regrowth trees suitable for use 

as firewood 

Prescription Retention of at least 50% of the pre-harvest basal area, 

including trees with identified habitat values (e.g. hollows) 

and trees selected for multiple purposes. CHECK?? 

* Source: Department of Sustainability and Environment (2008d) 

Traditionally, harvesting in Red Gum forests is primarily done through single tree selection; 

however, small-group selection is currently the most widely used silviculture, and aims at leaving 

gaps of generally less than a hectare (Di Stefano 2002). Single tree selection has been criticized as 

it promotes small canopy gaps favouring seedling regeneration in close proximity to established 

Red Gum species (DNRE, 2001). Also, the large zone of influence of River Red Gum can 

negatively impinge on seedling germination and survival (Dexter 1968). Single tree and smallgroup 

selection encourages mixed aged stands and is in marked contrast to the natural condition of 

River Red Gum forests (Di Stefano 2002), which tends to regenerate along the edge of receding 

flood waters resulting in a more even-aged stand development. 

One of the most important findings of Dexter's research was the favourable response of River Red 

Gum ecosystems to seed tree silviculture, whereby intensive clear-felling in clumps was followed 

by seed application. The creation of these larger gaps helps maintain the natural stand age. 

Despite these findings, single tree and small-group selection remain the standard industry practice. 

Seed tree silviculture may be used in some selected areas where there is significant tree mortality 

and poor health, such as following unplanned fire or dieback (Murray Thorson 3 pers. comm.). 

Incoll (1981) reported on thinning trials in 20-26 year old regrowth stands of low productivity 

River Red Gum. Commercial thinning trials that removed competing stems (for firewood and 

posts up to 20 cm dbhob) at a number of intensity levels increased the growth of retained stems. 

Recent measurements indicated that net basal area growth was greatest in moderately thinned 

stands, whereas the diameter growth of the largest 123 trees/ha was greatest in the most heavily 

thinned stands. Variation in branch retention and incidence of stem bends with initial density were 

not measured, although observations suggest that both were more prevalent at low initial densities. 

If the thinning objective is to achieve near maximum diameter growth for stands of about 20 years, 

then thinning to approximately 7-8 m 2 /ha of retained basal area will produce an increase in 

diameter growth without noticeable increase in branch retention or in frequency of stem bends 

(Connell 2005). 

3 Murray Thorson – FIC, Cohuna, Department of Sustainability and Environment, 



Although fire is a natural feature of the River Red Gum forests, the trees are more susceptible than 

many eucalypts to damage by fire. In Aboriginal times it seems that fire was used regularly in the 

forests, maintaining them in a fairly open condition, and undoubtedly contributing to the butt 

damage of many of the stems. Fire of only moderate intensity will kill the cambium near the base 

of the tree, leading to dry sides. Such wounds are often quite rapidly occluded, but enclose 

pockets of dead sapwood; the fires also promote the formation of gum veins. Intense fire round 

the base of a tree may kill the tree or, if more localised, lead ultimately to the hollow, burnt out 

butts that are encountered in many of the larger River Red Gums (Forestry Commission of NSW 

1984). 

The extent, frequency, seasonality and duration of flood events will influence the distribution, 

quality and growth of River Red Gum forests, as illustrated by forested areas on higher-ground 

which are frequently less productive and of poorer quality due to the high moisture stress (Davies 

1953). Whilst flooding is normally vital to the existence of the River Red Gum forests, prolonged 

inundation will ultimately kill trees. This is one of the effects of river regulation which maintains 

higher than natural summer flow levels, leading to the prolonged, or even permanent, flooding of 

some of the lower lying, and usually highest site quality, stands. 

3.2 Types of thinning operations 

3.2.1 Firewood fallen 

The collection of fallen firewood for domestic use occurs in areas where firewood is lying on the 

ground as a result of natural events or previous forest operations. The collection does not involve 

any additional felling of trees, but material will need to be crosscut (DNRE 2001). The fallen 

firewood can either be ‘dry’ or ‘green’ firewood. Green firewood is becoming more common as 

firewood collection is more closely integrated with recent harvesting or contract felling. 

Designated areas are usually set-up that allow car and trailer or light truck access to facilitate 

manual loading. Crosscutting is usually done by chainsaw. The ecosystem processes and impacts 

related to the collection of fallen CWD were covered in Section 2. 

3.2.2 Commercial thinning 

Commercial thinning usually involves the silvicultural treatment of overstocked, mainly even-aged 

regrowth stands to release potential sawlogs from competition. These stands of young trees have 

regenerated either naturally, such as following unplanned fire, or been assisted following a 

previous harvesting operation. Where the primary objective of the thinning treatment is wood 

production, suppressed trees or trees of poor form or quality are removed and dominant and codominant 

trees of good form and quality are retained, so that all the growth potential of the site is 

available to the retained stems. This “thinning from below” results in either a shorter rotation or 

larger trees at harvest. Also, wood that would be otherwise lost through death due to natural 

suppression in the stand is harvested, providing an interim return for the forest owner. Thinning is 

not intended to encourage regeneration, with the stand already considered to be fully stocked. 

While increased wood production is usually the primary goal, thinning may also enhance 

conditions for biodiversity. The specific habitat retention prescriptions will reflect the overall 

management objectives, such as a desire to restore forest structure while providing a sustainable 

timber supply and maintaining forest habitat (e.g. Department of Sustainability and Environment 

2008e). 



Commercial thinning is restricted to stands that meet specific criteria, with the economic viability 

affected by: 

1. Tree size and yield 

2. Site factors 

3. Coupe size and location 

4. Harvesting system and operator experience, and 

5. Thinning system 

Some of these factors are covered in more detail in Sebire and Fagg (1997), Brown et al. (2001), 

and Kerruish and Rawlins (1991). 

Generally, about one-half of the fully stocked live basal area may be removed providing minimum 

basal areas are retained, although a minimum retained basal area may be specified, which is often 

age and forest type related (Sebire and Fagg 1997). Commercial thinning in fully stocked young 

regrowth is normally conducted using an ‘outrow and bay’ method, where a 4.5m strip, or access 

track, is removed and 12m bays retained. This non-selectively removes about 25% of the stand 

and allows machinery access for the selective felling and removal of stems from the bays. 

Commercial thinning methods used in the generation of green firewood, and which satisfy 

silvicultural requirements as well as returning a commercial return for the operator include the 

following elements: 

1. Felling and crosscutting techniques – felling and crosscutting is done either manually 

using chainsaws, mechanically using specifically designed felling machinery (usually 

tracked rather than wheeled) or using a combination of both. Mechanical felling and 

crosscutting has definite advantages over the manual alternative. It is much safer, more 

productive, and offers better tree control both during felling and crosscutting. However, 

with mechanical operations there is considerably greater capital and running costs, and 

manual felling is more flexible across a range of terrain (Kerruish and Rawlins 1991). 

2. Extraction – both shortwood (billet) and longwood (bole length) operations are used 

depending on the configuration of the harvesting system. Depending on the nature of 

the operation, the ease of access and piece-size of this operation can be conducted using 

a truck, tractor or specialised equipment such as skidders or forwarders. Loading onto 

trucks is usually done by crab-grab or grapple loaders (built-on or stand-alone). 

Depending on the system used, woody debris can impact on stand access and machine movement, 

as well as butt damage to retained trees. This debris can be entirely natural or incorporate old 

logging material. Sebire and Fagg (1997) identify 50 t/ha and less than 0.5m diameter as being 

critical indicators for mixed species regrowth. 

3.2.3 Selective harvest 

Selective silviculture involves the selection-felling of marked trees, either individually or in small 

groups, with the objective of producing sawn timber products, whilst minimising impacts on 

species composition and forest structure. The harvest is focused on previously-thinned sites which 

contain trees up to 59cm dbhob, and involves the retention of at least 50% of the pre-thinning 

basal area, including all trees >60cm dbhob and trees with identified habitat values (e.g. hollows). 



Species composition is maintained (Department of Sustainability and Environment 2008d). 

Sawlog and sleeper operations cut trees from 45 cm to 60 cm diameter. Post-cutters harvest trees 

up to 40 cm diameter, mostly for sawing into split posts and other fencing products, and cut 

smaller dimension wood, producing round posts. Firewood is produced as a by-product of the 

sawlog harvesting and post-cutting, from the heads of felled trees and thinning of small stems. 

The current specification is a minimum Small End Diameter Under Bark (SEDUB) of 10cm and 

30cm dbhob. 

In selective harvesting, felling and crosscutting are done manually using chainsaws, and extraction 

of both shortwood (billet) and longwood (bole length) is conducted using truck, tractor or 

specialised equipment such as skidders or forwarders to suit the product being extracted. Loading 

onto trucks is usually done by crab-grab or grapple loaders (built-on or stand-alone). 

3.2.4 Ecological thinning 

Thinning is a silvicultural technique used in forest management to modify tree growth by reducing 

the competition for resources. Where trees are managed commercially, stems that exhibit less 

favourable timber quality potential are removed to reduce competition. When left in a natural state 

trees will 'self-thin' but this process of natural selection can sometimes be unreliable and slow; for 

instance, the Box-Ironbark forests and woodlands of Victoria support a large proportion of trees 

that are multi-stemmed regrowth (or coppice), a consequence of timber-cutting over previous 

decades (Muir et al. 1995). Ecological thinning has the principal aim of forest thinning to increase 

growth of selected trees, favouring development of wildlife habitat (such as hollows) over 

increased timber yields. 

Research programs under way in various parts of the world (e.g. USA, Australia) are aimed at 

providing an alternative approach to forest management where conservation objectives are a high 

priority. Recently (2003), Parks Victoria initiated the Ecological Thinning Trial in box-ironbark 

woodlands of central Victoria (Parks Victoria 2007; 2009), a direct response to ECC 

recommendations for the management of box-ironbark forests and woodlands (Environment 

Conservation Council 2001b). This is a long-term field-based experimental programme that 

aspires to evaluate different methods of ecological thinning and the effects they have on 

components of the box-ironbark forest ecosystem, including select vertebrate fauna and key habitat 

characteristics, with the broad aim of restoring a greater diversity of habitat types to the landscape, 

and therefore allowing the improved functioning and persistence of key communities and species 

populations. 

According to Parks Victoria (2007; 2009), ecological thinning is one of the methods that will be 

used as part of an Ecological Management Strategy to improve the ecological integrity of the 

forests and woodlands and their flora and fauna species instead of just maintaining the status quo. 

In contrast to silvicultural thinning, ecological thinning retains trees of all forms and sizes in a 

patchy distribution (clumps of high tree density are retained within a general mosaic of wider 

spaced trees to support species that favour both or either habitat) and competition is reduced to 

address the low proportion of larger trees in these forests. 



4 Ecosystem processes relating to harvesting 

4.1 Soil and nutrient processes 

4.1.1 Soil fertility (see also 2.1.1) 

When forest produce is removed in the form of ‘dry’ and ‘green’ firewood, either from the forest 

floor or following harvesting of standing forest, nutrients and carbon are removed with this wood. 

Associated soil disturbance can also lead to nutrient losses, either as soil erosion, leaching, or 

through losses of soil organic matter through soil respiration (Attiwill et. al. 1996, O’Connell and 

Grove 1996). 

Nutrient losses 

Strong associations exist between forest types and site characteristics. Species with greater site 

demands are found on better soils and this is reflected in growth rates and the nutrient status. 

While in absolute terms the quantities of nutrients removed in dry and green firewood removals 

will differ between sites, generally they are proportionally similar across most sites. However, 

results obtained for one forest type cannot necessarily be applied to other site types. 

The quantities of nutrients available in the soil for forest growth are affected by a number of 

factors. A key issue is whether firewood management activities can or will lead to a reduction in 

soil nutrient status and whether this may affect long-term health, productivity or other ecosystem 

processes. Typically, to investigate this issue the approach has been to estimate quantities of 

nutrients in the system together with nutrient fluxes and then use a simple input/output model over 

a number of rotations to determine possible impacts. Quantities of nutrients will include total and 

available nutrients contained in each component of biomass. Inputs usually include precipitation 

inputs and nitrogen fixation while losses include those in runoff water, forest product removal and 

fire. Most of these nutrient studies have focussed on the wetter forests, and typically for this 

forest, biomass nutrient content varies as follows (Attiwill et. al. 1996): 

- Leaves account for 1-2% of the total biomass of the trees above-ground, but for 20% of the 

N and P contents; 

- Stembark accounts for 10% of the total mass of the trees above-ground, but for 25-40% of 

the N, P and Mg and up to 60% of the K and Ca contents; 

- Stemwood is nutrient-poor relative to other components of the tree, accounting for almost 

80% of tree biomass but containing 10-20% of the K, Ca and Mg and 30-40% of the N and 

P content; 

- Subordinate vegetation (understorey, shrubs and ground-layer) accounts for


of calcium and magnesium were significantly lower when slash was removed from thinned stands 

(Rosenberg and Jacobson 2004). On low fertility sites where intensive harvesting is practiced 

inter-rotational management of the forest floor and harvesting residues is critical to maintain soil 

fertility (productive capacity) (Hopmans 2009) Long-term studies evaluating the sustainability of 

fast-growing second rotation plantations (Pinus radiata) on podsolised sands have indicated that 

productivity was maintained or improved. This was attributed to the conservation of organic 

matter and nutrients through retention of litter and harvesting residues after the first rotation and 

the exclusion of fire. Where only foliage was retained (log residues and branches removed) 

nutrient accession was not exceeded over 30 years except for N. In contrast, whole-tree harvesting 

including foliage increased nutrient exports above inports and fertilizers are likely to be required 

for this additional removal of nutrients to maintain site productivity in the next rotation. 

In drier native forests forests used for firewood collection, it is expected that similar trends would 

be observed. Because the concentration of nutrients in wood is small relative to those in other 

parts of trees, collecting or harvesting part of the wood removes a relatively small nutrient store. 

On this basis, where only wood >10 cm diameter is removed for dry and green firewood it is not 

expected that the level of nutrient removals would have a detectable impact on forest productivity. 

However, if bark and smaller diameter material is also removed then the amounts of nutrients 

removed will increase significantly and there may be a greater impact. Losses of other nutrients 

such as N will generally be replaced by biological N2-fixation, and P from reserves and through 

the weathering of parent rock (Attiwill et. al. 1996). 

Where dry and green firewood removal is only associated with stem wood (i.e. larger diameter 

wood), the level of nutrient removals is not expected to have a detectable impact on productive 

capacity. Removal of biomass components other than the stem should be avoided, as this is likely 

to impact on site nutrient budgets and consequently on soil organic carbon. Foliage, smaller 

diameter branchwood (


will likely lead to some small decreases in SOC due to oxidation of carbon in residues from the 

disturbance and in soil organic matter. 

Where dry firewood is removed from the forest floor the ability to restore this lost carbon is 

limited, as there is unlikely to be any associated growth response from the forest. However, if this 

dry firewood is left as CWD then it will be exposed to gradual decay, diminishing to some level of 

residual carbon. Unfortunately, knowledge about the decay rates of CWD is limited, and can be 

summarised in two general relationships: (1) a decrease in decay rates with increasing log size, 

and; (2) a decreasing rate of decomposition with increasing wood density (Raison et. al. 2002). 

There are no clear rates for the conversion of CWD into SOC (Mackensen and Bauhus 1999). 

Additionally, fire will readily convert decaying wood into char and release bound nutrients. 

Carbon in char is inert and its contribution to soil fertility is difficult to interpret, although it 

appears to influence soil structure and water infiltration (Bauhus et. al. 2003). 

Where firewood is been derived from harvesting of the forest, then it is likely that there will be 

some growth response. In wetter forests, where growth responses are usually more rapid and 

vigorous, any decrease in carbon is usually relatively short-lived, but in drier forests there will be a 

slower recovery as growth responses are more restrained. The sensitivity of site fertility to the 

effects of disturbance on SOC and other measures of fertility has been studied for disturbance 

factors, such as canopy removal, extraction track use, fire intensity, and soil disturbance in 

clearfell silviculture (Bauhus et. al. 2003). The major impact on these measures was on areas 

where extensive mechanical soil disturbance was used to prepare a seedbed for regeneration, 

compared to where harvesting slash was burnt, where only minor effects were observed. Where 

there is no requirement for the preparation of a receptive seedbed, soil fertility impacts can be 

reduced if soil disturbance is minimised and organic matter conserved by retained harvesting 

debris (Raison et. al. 2002). 

4.1.2 Carbon cycling (see also 2.1.2) 

The impact on carbon budgets and Greenhouse gas (GHG) emissions of firewood-related 

disturbance is a critical element of sustainable firewood production. Its potential to reduce fossil 

fuel use and attendant CO2 emissions, is dependent on a number of factors, including: forest 

growth rate, management, harvesting and transport systems, and; the efficiency with which 

firewood is burnt (Raison et. al. 2002). Additionally, any possible reduction in the use of fossilfuels 

must be balanced against carbon losses from the reduction in CWD and soil organic carbon. 

Carbon storage 

Forests sequester carbon in biomass and through plant residues in the soil, as soil organic carbon 

(SOC), with the accumulation of above ground carbon generally reflecting forest growth and 

productive capacity. The quantities of carbon vary according to a number of factors including soil, 

climate, forest type, stage of stand development and level and type of disturbance. Grierson et al. 

(1992) estimated the above-ground quantities of carbon in Victoria's forests using a series of agedependent 

biomass functions. The forests types that contain firewood species are outlined in Table 

4.1. These forests contribute about 71.8% of the above-ground carbon storage described by 

Grierson et al. (1992). 



Table 4.1 Area, biomass and carbon density (above-ground) of Victoria’s predominant 

firewood forest types (Grierson et al. 1992). 

Forest type Total area (ha) Mean above-ground density (t 

DM/ha) 

Carbon 

Storage 

Biomass Carbon (tonne x 10 6 ) 

Foothill mixed species 2,639,735 477.3 238.6 629.840 

Coastal mixed species 404,050 379.1 189.5 76.567 

Box-ironbark 236,262 149.2 74.6 17.625 

River Red Gum 112,851 417.0 208.5 26.529 

Main firewood spp. 3,392,898 751.886 

Alpine Ash 311,997 394.8 197.4 61.588 

Mountain Ash 181,989 492.1 246.0 44.769 

Shining Gum 13,057 469.8 234.9 3.067 

Mountain mixed species 464,761 423.7 211.9 98.460 

Alpine mixed species 203,224 450.0 225.0 45.725 

Lesser firewood spp. 1,175,028 253.609 

Other (native forests) 1,896,226 42.249 

Below ground, carbon accumulation is affected by root growth and SOC balances. Soils are 

expected to increase in carbon, dependent on soil type, and then reach stability. Disturbances, such 

as fire lead to direct losses of carbon from the system followed by a process of re-accumulation 

during forest recovery (DSE 2003). 

CWD has been recognised as a quantitatively important component of the forest’s carbon stocks. 

The amount of CWD in some areas is equivalent to approx 10-20% of the above ground carbon 

biomass, indicating that dead wood can represent a significant amount of carbon in forests 

(Delaney et. al. 1998). However, generally little work has been conducted on the amount of 

carbon held in CWD in Australian systems. CWD inputs are often from dieing or dead standing 

trees, and often associated with fire, wind damage, or disease. CWD represents a large and long 

term store of carbon, which is gradually released through its decomposition (Brown et al. 1996b; 

Grove et al. 2002). Decay often starts in standing trees and usually increases once trees fall over 

and there is greater contact with the ground (Raison et. al. 2002). During decomposition, microbes 

turn organically bound carbon (which accounts for approximately 50% of the organic material) 

into carbon dioxide (Mackensen and Bauhus 1999). 

Changes in soil carbon are potentially very important for carbon budgets because soil carbon tends 

to be more stable than other carbon pools so that any increases or decreases in soil carbon are 

potentially longer-lasting than changes in other carbon pools. Char (charcoal, black carbon) can 

comprise a significant proportion of SOC, particularly in those forests where wildfire or 

regeneration burning have been significant disturbances. This fraction has been reported to 

contribute 13-27% of SOC, and together with the ‘stable’ carbon fraction can make up 69-81% of 

SOC (Bauhus et. al. 2003, Hopmans et. al. 2005)). These fractions are considered to be inert 

components of the soil, along with, arguably fragmented rocks and mineral aggregates. Fire will 

readily convert CWD into char, particularly decaying wood, providing burning conditions are 

suitable. Other components, such as labile carbon and microbial carbon make up the oxidisable 

organic carbon (13-18%) (Bauhus et. al. 2003). In normal forestry operations there is generally 

only a slight change, if any, to total soil carbon, however, the inclusion of soil cultivation can led 



to some reduced soil carbon storage, particularly in the labile carbon and microbial carbon 

fractions. 

Carbon is ‘lost’ in wood taken off-site as part of the collection and harvesting of dry and green 

firewood. There are different management regimes under which this firewood removal can occur, 

each with a different impact on carbon balances. To affect an understanding of these different 

regimes (e.g. Selective harvesting, no firewood collection; Selective harvesting with firewood 

collection; Selective harvesting with intense firewood collection) simulation modelling is required 

which incorporates the following: forest growth; natural mortality; disturbance related mortality; 

fire impacts; forest product removals; decay rates; SOC losses; etc., to keep track of all the key 

carbon pools. Such an undertaking, especially for different forest types, is beyond the scope of this 

review, but simulations of carbon emissions or GHG balances which have been reported are 

considered in the next Section. 

Greenhouse gas (GHG) emissions 

The task of exploring the impact of different firewood options on GHG balances is significant and 

is considered generally here with reference to published material. Modelling can be used to 

explore the influence of carbon balances on net CO2 emissions, with particular care needed to 

select an appropriate time horizon for the analysis. The AGO’s FullCAM model was developed to 

track carbon flows in a range of ecosystems, accounting for changes in carbon in all forest pools 

including vegetation (above ground and roots), litter, soils, and in carbon taken off-site in wood 

products. Additionally, it also tracks carbon use in the harvest and transport of forest products, and 

account for the decomposition of these products (Paul et. al. 2003). The model was used to 

explore two forest types: (1) unmanaged remnant woodlands (maximum aboveground biomass of 

about 77 t DM ha -1 ), and; (2) managed native forest with selective harvesting (maximum 

aboveground biomass of about 140 t DM ha -1 ). Scenarios involving a number of harvesting and 

firewood collection intensities were modelled over a 100 year period. The modelling found that 

the unmanaged woodland systems were degrading because old dying trees were not being replaced, 

and there was a release of CO2. Firewood collection further increased the net emission of GHG. 

For the managed native forest the FullCAM model indicated that the forest was in a state of near 

equilibrium with respect to increments of tree growth, with a small sequestering of carbon. 

Firewood collection resulted in net emission of GHG. When the firewood was used for domestic 

heating, the net amount of GHG emitted per unit of heat energy produced ranged from 0.03-0.11 kg 

CO2 per kWhr -1 depending on the scenario. This indicated that firewood may be generally more 

favourable for domestic heating than other sources of domestic heating such as gas and electricity 

(which generally produce at least 0.31 kg CO2 per kWhr -1 , excluding solar-, wind- or hydroelectricity) 

Additionally, GHG balances (including non-CO2 gases) have been evaluated for the proposed use 

of fuelwood for electricity generation, involving the use of harvesting residue from wet forest in 

Tasmania. This modelling found that for CO2 equivalent emissions, greenhouse balances were 

dominated by the potential savings due to the offset of fossil fuel emissions (Raison et. al. 2002). 

Consequently, the type of energy generation that will be replaced by the use of the harvesting 

residues was critical to this evaluation. This highlights the importance of assumptions in this 

modelling, particularly in relation to fossil fuel offsets, but also more generally. Both this example 

and the previous example using the FullCAM model contain many assumptions and the results are 

only semi-quantitative. As a consequence, they are useful in comparing contrasting scenarios, but 

not as useful in fully quantifying a particular option. 

In evaluating the GHG balances of different firewood options, it is worth noting that CO2 

emissions from burning wood are 1,687 kg CO2 tDW -1 . Additionally, non-CO2 GHGs are also an 



important consideration (Raison et. al. 2002). This principally involves the GHGs methane and 

nitrous oxide that are released during burning. In uncontrolled burning, such as a forest fire, it is 

estimated that for each tonne of wood (dry weight) 6.1 kg of methane and 0.006 kg of nitrous 

oxide are released, or 128.5 and 17 kg CO2 tDW -1 (Raison et. al. 2002). More efficient burning 

(industrial boilers) significantly influences the emission of methane, but has little effect on nitrous 

oxide emissions. Domestic firewood use is likely to have limited effect on burning efficiency. 

4.1.3 Soil and water quality (see also 2.1.3) 

The physical soil disturbance associated with accessing ‘dry’ or ‘green’ firewood, or with its 

production can impact on water quality. The nature and timing of access can significantly 

influence this impact (Rab et. al. 2005). Harvesting and collection activities associated with 

firewood lead to differing levels of soil physical disturbance including soil movement and 

compaction (Rab 2004). Loss of soil through erosion may reduce productive capacity and impact 

on aquatic values. 

Much of the literature on the impact of harvesting on soil and water quality is focussed on 

harvesting associated with clearfell rather than harvesting for firewood or selective harvesting. It 

is expected that firewood harvesting operations would result in a much lesser impact on water 

quality than the more intense operations associated with clearfell. Generally, firewood-related 

operations will result in reduced areas of extraction track and a lower unit wood volume extracted. 

Water quality impacts usually associated with native forest harvesting operations include: 

1. Operations in the harvested coupe; (felling, snigging/forwarding, processing, loading and 

transporting) resulting in soil disturbance and compaction and associated surface runoff of 

sediment/nutrients 

2. Outside the coupe activities; road and stream crossing construction, maintenance, usage and 

associated increases in sediment loads in surface runoff. 

In-coupe, the factors that are most relevant to minimising soil disturbance and compaction are soil 

moisture content at the time of collection or harvest, machinery type, extraction track design and 

factors specific to soil type (Rab et. al. 2005). In particular, soil trafficability (or resistance to 

compaction) is affected by soil type; with the soil layer supporting traffic loads the critical factor. 

Gravel content along with a soil dryness index (SDI) are good indicators of soil trafficability 

during harvesting in the ‘shoulder-periods’ of Spring and Autumn. During these wetter periods it 

is often necessary to call a halt to forest operations for a winter break, generally using a trigger, 

such as a closure date or soil moisture conditions (e.g. soil saturation). SDI is one possible 

mechanism by which such a halt can be ‘triggered’ (Rab et. al. 2005). It is a soil water balance 

model, which is driven by rainfall and temperature, and is expressed as the nominal rainfall deficit 

from field capacity. It is easily used and can be useful in predicting threshold values where soil 

trafficability is important. 

Sediment and nutrients (e.g. total phosphorus and total nitrogen) are common pollutants of streams 

and water impoundments. Comparatively, in-coupe, the literature indicates that the dominant 

source of sediment/nutrients (pollutants) to streams is often roads/extraction tracks, with the more 

heavily used extraction and access tracks responsible for most of the water runoff and movement 

of soil (Dignan 1999). Croke et al. (1999) conducted rainfall simulator erosion studies within a 

range of forest types in Victoria and NSW and concluded that tracks and snigging areas were 

responsible for most of the runoff and erosion. The general harvest area was found to act more as 

a sink for runoff water and sediment generated on the road/track surfaces, rather than a source of 

sediment. Current codes of practice and management procedures ensure that the risk of 



connectivity between sources of sediment and drainage lines is minimised to acceptable levels, and 

the impact of harvesting operations is mainly found to be minimal. 

The most striking aspect of the literature relating to forest harvesting and water quality under 

modern management prescriptions is the almost universal finding that the road network is the 

dominant source of sediment. The impact of harvesting operations is generally found to be 

minimal (Cornish 2001; Grayson et al. 1993). Motha et al. (2003) used sediment tracing 

techniques and estimated that between 18%-39% of the sediment load from a Victorian forest was 

from unsealed roads while harvest areas contributed only 5-15%. Unsealed roads are identified as 

major contributors to sediment levels (Sheridan and Noske 2007). They found that annual 

sediment load was found to be twenty five times higher on an unsurfaced road on erodable subsoil 

(5373 mg/m² per millimetre of rain) than for a high-quality gravel surface road (216 mg/m² per 

millimetre of rain). The landscape position of roading has been identified as a critical factor in 

determining linkage between the road network and the stream network. Ridge-top roading has less 

direct linkage than roads lower in the landscape, with stream crossings and associated approaches 

being identified as the critical linkage points (e.g. Hairsine et al. 2002). 

Several studies have investigated the relationship between traffic volume and sediment generation 

from unsealed roads. These studies have reported a range of increases in sediment generation due 

to traffic. Grayson et al. (1993) found a 2 fold increase, Croke et al. (1999) report a 4-5 fold 

increase with traffic, Foltz (1996) an 8-12 fold increase depending on gravel quality, Bilby et al. 

(1989) a 20 fold increase, and Reid and Dunne (1984) a greater than 100 fold increase. If traffic 

occurs during a rainfall event, sediment concentrations have been reported to increase immediately 

by 25% (Constantini et al. 1999), 600% (Reid and Dunne 1984) and 2500% (Bilby et al. 1989). 

However, Sheridan et al. (2005) found that for well-maintained gravelled roads of moderate slope 

and length, average suspended sediment generation rates are around 200-300 mg/L, increasing 3 to 

10 fold to a maximum of 900-2000 mg/L with heavy traffic. Following the cessation of traffic, 

generation rates declined exponentially to pre-traffic levels after 50-70 mm of runoff. He found 

that these sediment generation values and traffic related increases are substantially less than 

generally reported previously. The results also showed that if roads are surfaced well and 

maintained correctly, their use in moderately wet conditions should not result in a more erodable 

surface than if used in dry weather. However, use of poorly surfaced roads or tracks should cease 

when the risk of damage to the road or sediment pollution of watercourses is high. During wetter 

periods it is often necessary to call a halt to forest operations, generally using a trigger, such as a 

closure date or soil moisture conditions (e.g. soil saturation). 

4.1.4 Forest hygiene and health 

Forest hygiene and health (and vitality) relates to the general condition of the forest, with reference 

to soundness and vigour; freedom from injury, damage, decay, defect and disease; robustness; 

invasive species and capacity for energetic active growth. The impact of firewood harvesting on 

eucalypt health and its ability to influence the general condition of the forest is the particular focus 

here. 

Insects 

Native eucalypt forests support a wide range of foliage-feeding and wood boring insects. To date, 

research on insects in native forests has generally focused on specific major pest species and their 

impact, distribution, abundance, autecology and control. Also, the impact of disturbance events, 

including wildfire, fuel reduction burning and timber harvesting have been studied, with studies 

focusing on insect recovery, as an indicator of forest resilience to disturbance (Neumann and 

Marks 1976; Neumann 1991, 1992; Collett 1997, Collett and Neumann 2003). Economically 

important outbreaks of insect defoliators usually occur in the simpler ecosystems, examples of 

which are natural single species forests and even-aged stands that regenerate following large 



destructive fires. Insects that degrade timber or cause wood destruction in standing trees seem 

more common in overstocked or fire-damaged stands (Neumann and Marks 1976). No studies to 

date have specifically examined the effects of harvesting for firewood in Victorian forests on 

insect pest species and insect biodiversity, although a few have studied the impact of thinning. 

One study examined the influence of non-commercial thinning upon defoliation by the Gumleaf 

Skeletoniser (Uraba lugens) in river red gum (Eucalyptus camaldulensis) forests (Harris 1974). 

He found that thinning regrowth stands to a density of less than 750 stems per hectare appeared to 

significantly reduce the severity of outbreaks which followed, provided the stumps of thinned trees 

were prevented from coppicing and thinning debris was removed. In older stands of river red 

gum, thinning using patch-cutting, has been used to manage areas of eucalypt die back (Murray 

Thorson 4 pers. comm.). 

A lack of specific study in this area means that information tends to be anecdotal and general 

observation rather than based on scientific assessment. For example, it is a commonly held view 

that losses of wood production from insect attack may be lowered by the early removal of 

suppressed or dying trees or by the prevention of damaging ground fires or of injuries to remaining 

trees during felling and timber extraction. This view is based on observation rather than based on 

scientific assessment. 

Within Victorian native forests there are a wide range of insect pest species that cause damage of 

varying degrees to a wide range of tree species. By far, the majority of these insect pests cause 

damage on an infrequent basis with the immediate effects generally short-term in duration and 

localised in their extent. Examples of these insect species are adults of leaf-chewing Christmas 

beetles (Anoplognathus chloropyrus, A. hirsutus), larvae of the leaf-mining Leafblister sawfly 

(Phylacteophaga froggatti), and larvae and adults of the leaf-feeding Eucalypt Weevil (Gonipterus 

scutellatus) (Collett 1997). While the causes of such outbreaks are not fully understood, factors 

such as availability of food resources, prevailing climatic conditions, foliage nutrient conditions, 

whether host trees are in single or mixed species stands, and population status of predator species 

all appear to play a role (Collett 2001). 

Observations made over many years in Victoria have identified a group of insect species that have 

caused economically and aesthetically significant damage to occur on an ongoing basis. These 

insect species and the principal forest types they impact are, as follows: 

- Spurlegged Phasmatid (Didymuria violescens), Ash and damper mixed-eucalypt forests 

- Mottled Cup Moth (Doratifera vulnerans), mixed-eucalypt forests 

- Mountain Ash Psyllid (Cardiaspina bilobata), Ash forests 

- Southern Eucalyptus Leaf Beetle (Chrysophtharta agricola), Ash forests 

- Wood boring moths (Family Cossidae), Ash forests 

- Dampwood termite (Porotermes adamsoni), Ash forests 

All these insect pest species are considered significant, causing widespread severe defoliation 

damage (e.g. D.violescens and D.vulnerans) or having the ability to cause significant wood 

degradation of wood in standing trees (eg. Phorocantha spp and Cossidae). 

Hygiene 

Within forested areas movement of machinery, vehicles and other equipment can potentially result 

in the transport of weeds and disease. These weeds and diseases can come from both within and 

4 Murray Thorson – FIC, Cohuna, Department of Sustainability and Environment, 



also from outside the forest. Weeds are discussed in Section 4.4 Flora. In relation to diseases, 

native eucalypt forests support a broad range of fungal pathogens that cause tree diseases. To date, 

research has mainly focused on the distribution, life cycle and control of a few specific diseases 

that have, or are likely to have, significant economic impact in native forests. The extent of tree 

disease in a forest is the result of an interaction between a host, a pathogen and the environmental 

factors that affect host response and pathogen virulence. Of the several significant pathogens, both 

native and exotic, that have been recorded (Marks et al 1982, Keane et al 2000), few have been 

studied in sufficient detail to be enable an accurate prediction of the impact that harvesting for 

firewood may have on disease development. Of principal interest to this review are collar rot and 

root diseases, and wood decay. 

The introduced soil-borne pathogen Phytophthora cinnamomi and native Armillaria spp. (notably 

A. luteobubalina), are recognised across Australia as the principal causal agents of collar rot and 

root disease associated with native forest dieback and isolated patch death of eucalypts (Shearer 

and Smith 2000, Shaw and Kile 1991, Kile 2000). 

Phytophthora cinnamomi 

Significant outbreaks of phytophthora-related dieback were recorded in the mid 1950's and 60's, 

and 1971 associated with heavy summer rainfall and autumn droughts (Tregonning and Fagg 

1984), combined with the use of a selection felling silvicultural systems which resulted in reduced 

basal area on affected sites (Marks and Smith 1991). While the pathogen is now widely 

distributed in Victoria, the symptoms of disease are mainly confined to the coastal and foothill 

forests of East and South Gippsland, where significant dieback of eucalypts and losses of 

understorey species have occurred. These forests are significant sources of firewood. 

Research indicates that phytophthora is an introduced primary plant pathogen of native plants. 

This root rot is favoured by the following conditions (Marks et al. 1982): 

1. Saturation of soil for short periods of time, usually after heavy rain or as a result of run-off 

from hill slopes and drains. 

2. Poor, internal soil drainage caused by either poorly developed crumb structure or by clay- 

rich layers close to the surface. 

3. Soils of low fertility containing little organic matter 

4. Soil temperature above 16 o C. 

Combinations of these conditions greatly aggravate disease. For example, heavy summer 

rainstorms can produce severe disease conditions in infertile, sandy soils overlying a clay-pan 

close to the surface. The motile spores (zoospores) of the pathogen infect the roots of susceptible 

species when the soils are wet, and in highly susceptible species spread through the root system 

until it girdles the major roots and stems. As the roots die, the pathogen produces resting spores 

(chlamydospores), which can survive dry soil conditions and can be picked up in gravel taken 

from pits surrounded by infected vegetation. Soil adhering to vehicles, machinery, animals and 

footwear, and infected nursery plants provides a means for long-distance spread. 

In south west Western Australia, management has placed seasonal restrictions on forest operations 

in areas affected by P.cinnamomi, where high soil moisture conditions increase the risk of its 

spread by vehicles and machinery (J. Bradshaw 5 pers. comm.). 

Armillaria luteobubalina 

5 Jack Bradshaw –Silviculturalist (retired), CALM, WA 



This fungus is a native primary pathogen of many species of both native and introduced trees 

(Shaw and Kile 1991, Kile 2000). The conditions that trigger an outbreak are not clear but appear 

related to the presence of a food base (e.g. stump), and/or soil moisture and physiological state of 

the host. Symptoms vary from the occurrence of scattered dead individual trees to distinct patches 

or infection centres up to 20 ha in extent (Edgar et al. 1976). In Victoria, it occurs naturally 

mainly in mixed species eucalypt forests and is an important disease in some Damp and Wet 

Forest types, principally in west-central Victoria. Historically, it has been mainly associated with 

selectively logged areas, with the disease impact being greatest in mature and overmature stands, 

causing crown dieback, reduction in basal area and volume and eventually death. A.luteobubalina 

has damaged approximately 2000 ha of mixed forest in Mt Cole and Wombat State Forests. As 

with P. cinnamomi, this species kills trees and shrubs of any age through the infection of the major 

roots and stem of the plant. It spreads between plants mainly through root to root contact. To 

reduce the chance of contact between healthy trees and the fungus, clearfelling (followed by 

regeneration burning) rather than selective harvesting techniques, may reduce the effects of 

Armillaria in areas prone to infestation. The creation of an ashbed should promote dense and 

healthy seedling regeneration that will allow for disease escape, genetic selection for resistance to 

infection, and drying of the site, thus making conditions less favourable for Armillaria (Smith and 

Smith 2003). 

Wood decay fungi 

There are two principal sources of wood decay formation. Those associated with defective branch 

ejection and wounding, and those linked to stem damage and decay in the major roots. White and 

Kile (1991) have demonstrated that stem wounds inflicted during harvesting operations can lead to 

the development of substantial columns of decay. Decay pathogens are most active in areas of 

high rainfall where their impact can be considerable on wood quality (Wardlaw and Neilsen 1999). 

4.2 Tree hollow development 

The loss of hollow-bearing trees in Victorian native forests (including native forests on private 

land) is listed as a potentially threatening process under the Flora and Fauna Guarantee Act 1988 

(Department of Sustainability and Environment 2003d). Because hollow-bearing trees are likely 

to be affected by firewood harvesting (felling of standing trees), we will commence this chapter 

with a brief introduction to the formation of hollows. 

There are no Australian vertebrates that actually excavate hollows in eucalypt trees in temperate 

forests, although several parrot and marsupial species may chew at hollow entrances to enlarge 

them or keep them open (Richard Loyn pers. obs.). Hollow development in eucalypts is generally 

considered to be a long-term (>100 year) process, though other schools of thought hold that 

hollows are formed in trees of any age through damage to the bark layer caused by fire, lightning 

or wind storm (Vearing 2000). Different species begin to develop hollows at different ages and 

rates (Mackowski 1984; Stoneman et al. 1994). Therefore, older forests have more hollow trees 

than younger forests (Lindenmayer 1996). An example of this was shown by Soderquist (1999) 

who reported that the frequency of hollows increased as tree size and age increased in boxironbark 

forests. 

A range of factors are known to contribute to hollow formation. These include mechanical 

damage during high winds, branch abscission and breakage, lightning strike and fire. Such 

damage can leave an open scar which is susceptible to fungi and insect (predominantly termite) 

attack, thus initiating the decomposition process (Department of Sustainability and Environment 

2003d). Fire can also accelerate enlargement of tree hollows (particularly base hollows) and 

subsequent deterioration and collapse of older trees (EM pers. obs., Department of Sustainability 



and Environment 2003d). These contribute to the abundance of coarse woody debris on the 

forest/woodland floor (see below). 

When fire is severe enough to kill tree cambium, the bark dies and the xylem is exposed to the 

entry of insect and decay organisms. Species such as Messmate Eucalyptus obliqua, Silvertop 

E. sieberi, Red Ironbark E. tricarpa are quite fire resistant, with Messmate having a thick fibrous 

bark persistent to the smallest branches, and Silvertop thick bark that persists to the larger 

branches of the crown. McArthur (1968) and others have reported that bark thickness rather than 

type was the important factor in protecting the cambium of eucalypts from lethal temperatures. 

Other species, such as River Red Gum are thinner barked and quite fire-sensitive, and fire-related 

mortality or damage is more likely when fires occur. In older trees, structural stem failure is more 

likely, as fire-damaged butts are more likely in older trees that have previously been damaged by 

fire. Butt damage is common in many forests and is often related to the presence of large fuel 

accumulations near the base of trees. 

Tree damage by fire and consequent hollow development is influenced by the location of CWD. 

Burrows (1987) found that 92% of Jarrah Eucalyptus marginata and Marri E. calophylla trees 

were damaged by low and medium intensity fires if they were less than one metre away from 

CWD. This was due to the long duration of heating produced by the burning log. Buckley and 

Corkish (1991) found that in East Gippsland Lowland and Damp regrowth forests, debris from 

previous harvesting was a critical factor affecting butt damage. Retained trees suffered severe butt 

damage up to 3.2 m from the old logs that caught alight during post-thinning burning. Standing 

dead trees were also often ignited during post-thinning burning and damaged retained trees. Fire 

was also found to expand the area of damage on trees that suffered mechanical damage during 

thinning. 

4.3 Habitat 

The ecological importance of stand structural complexity has been articulated by Lindenmayer et 

al. (2002), amongst others, and their review of this important habitat component is summarised 

here. The two key reasons that stand structural complexity is important are (1) structurally 

complex forests allow the potential for greater inter-specific segregation of resources and 

microhabitats thereby enabling more species to occur locally, and (2) many types of structural 

attributes can be essential nesting, sheltering and foraging sites for a wide variety of taxa. The loss 

of key elements of stand structural complexity, like large diameter trees, thickets of understorey 

plants and logs, can: (1) eliminate organisms from logged areas that would otherwise occur there, 

(2) prolong the period that logged and regenerated stands are unsuitable habitat for species that 

have been displaced, (3) impair the dispersal and movement of some animals through logged 

areas, and, (4) eliminate within-stand variation in habitat conditions required by some taxa. 

Forests where stand structural complexity has been simplified through intensive management have 

impaired value for biodiversity. 

Hollows are considered essential for a range of fauna, and each species has its own requirements 

for type of hollow (Australian Rainforest Conservation Society 1999; Gibbons and Lindenmayer 

2002; Gibbons et al. 2002); in Victoria, the abundance of arboreal mammals has been correlated 

with densities of hollow-bearing trees in montane ash, River Red Gum and box-ironbark forests 

(Department of Sustainability and Environment 2003d). 

Stand thinning is the most common harvesting method for obtaining firewood from live trees. 

Typically, select trees are removed from a dense regrowth stand in order to achieve a particular 

management objective or different structure of the forest stand — when less desirable trees are 



removed, resources may become available to the remaining trees and their growth and vigour 

increased (see earlier). There are potential ecological benefits associated with the removal of 

smaller trees — larger trees are expected to grow faster and provide for more rapid hollow 

development as well as the development of other habitat components (e.g. increased CWD loads, 

understorey growth). Carefully-controlled ‘ecological thinning’ aimed at increasing average tree 

size has been advocated, particularly within areas dominated by regrowth forest, as a means to 

enhance habitat for hollow-dependent fauna, including the Squirrel Glider (Department of 

Sustainability and Environment 2003b). This means that, for biodiversity, particular habitat 

characteristics are removed or modified. The removal of trees by thinning means that, in the shortmedium 

term, there are fewer resources (e.g. roosting, nesting, basking, shelter and foraging 

structures, food, nest material) available to wildlife. 

4.3.1 Mammals 

The mammals of south-eastern Australia include many arboreal and aerial taxa that depend on 

hollow-bearing trees, as well as some facultative hollow users (Menkhorst 1995; Van Dyck and 

Strahan 2008). The distribution and abundance of such mammals in the landscape is typically 

patchy, reflecting an association with variety in habitat quality and floristic diversity. Arboreal 

mammals depend on the following critical habitat attributes: foliage, flowers, bark and hollows 

(McElhinny et al. 2006), and the removal of trees for firewood (and other reasons) will obviously 

diminish the availability of these crucial resources for fauna. The importance of these resources, 

particularly hollows, has been documented for a variety of south-eastern Australian arboreal and 

aerial mammals (Duncan and Taylor 2001; Friend and Wayne 2003; Gibbons and Lindenmayer 

1997; Gibbons and Lindenmayer 2002; Gibbons et al. 2002; Harper 2005; Kavanagh et al. 1995; 

Kavanagh and Stanton 2005; Lindenmayer 1997; Lindenmayer et al. 1991; Lindenmayer et al. 

2008; Lindenmayer et al. 1998; Lumsden et al. 2002a; b; McElhinny et al. 2006; Menkhorst 1995; 

Soderquist and Mac Nally 2000; Traill 1991; Tzaros 2005; Van Dyck and Strahan 2008). Indeed, 

the presence, abundance and taxonomic diversity have been correlated with the number of hollowbearing 

trees, and that tree size (dbhob) is significantly correlated with occupancy of tree-hollows 

by mammals (McElhinny et al. 2006). Most of these arboreal and aerial species are known to 

utilise hollows in both dead and live trees. 

Several scientific studies have demonstrated the direct association of mammal taxa with tree 

hollows in the forests of Australia, and a couple of examples are provided here. Dickman and 

Steeves (2004) documented the significance of, variously, tree hollows and logs for the Agile 

Antechinus, Brown Antechinus Antechinus stuartii and Bush Rat Rattus fuscipes in forests of 

eastern Australia. 

Tree hollows are also a key habitat component for the Common Ringtail Possum Pseudocheirus 

peregrinus, especially where the ability to construct nests (dreys) in understorey vegetation is 

limited (Lindenmayer et al. 2008). The frequent use by the Common Ringtail Possum of smaller 

diameter trees with fewer cavities in the Victorian Central Highlands is at odds with other findings 

for this species (e.g. Gibbons and Lindenmayer 2002) and may mean that animal size and 

competition for hollows with other (larger) species may be determinants of hollow utilisation. The 

partitioning of hollow-bearing trees by arboreal marsupials in the Central Highlands has been 

documented by Lindenmayer et al. (1991), who argue that the then clear-felling rotations would 

prevent the development of characteristics that make trees suitable nest sites for arboreal 

marsupials. 



Several threatened arboreal marsupials of south-eastern Australian woodlands depend on tree 

hollows, including the Eastern Pygmy-possum Cercartetus nanus (Duncan and Taylor 2001; 

Menkhorst 1995; Tulloch and Dickman 2006), Brush-tailed Phascogale Phascogale tapoatafa 

(Rhind 2004; van der Ree et al. 2006) and Squirrel Glider Petaurus norfolcensis (Beyer et al. 

2008; Department of Sustainability and Environment 2003b; Menkhorst 1995; van der Ree 2002). 

Beyer et al. (2008) also highlighted the issue of the sustainability of suitable den trees, reporting 

an annual loss on their study sites of 3% of den trees, comparable to the annual loss of 4% of den 

trees used by the threatened Leadbeater’s Possum Gymnobelideus leadbeateri in the Victorian 

Central Highlands (Lindenmayer et al. 1997; Lindenmayer et al. 1991). 

Large trees are known to be important for other woodland mammals. Woodland patches in 

southern New South Wales are more likely to support populations of Yellow-footed Antechinus 

A. flavipes if they contain, inter alia, larger trees of select species (Korodaj 2007). In the boxironbark 

woodlands of central Victoria, gullies, which occupy a very limited area in the ecosystem, 

are known to support significantly greater numbers of some arboreal mammals (e.g. Common 

Brushtail Possum Trichosurus vulpecula, Common Ringtail Possum) compared with non-gully 

sites; gullies also revealed 53% more trees with hollows in the upper bole and branches, and 

almost six times more very large trees (Soderquist and Mac Nally 2000), the inference being that 

hollow-bearing trees are probably the limiting habitat characteristic for these mammals. The 

importance of such gullies for birds has also been documented (Mac Nally et al. 2000b). 

Bats comprise over 20% of the Australian mammal species (Van Dyck and Strahan 2008), and 

they play a significant role in several ecosystem processes — insectivory, pollination, seed 

dispersal (Law 1996). Their occurrence is largely determined by several key habitat attributes: 

foliage and canopy spaces, hollows and decorticating bark, and access to water (McElhinny et al. 

2006). Both empirical and inferential evidence exist for the value of tree hollows to insectivorous 

bats in south-eastern Australia; hollow-bearing trees are important as roosting, hibernation and 

maternity sites (Brown et al. 1997; Churchill 2008; Herr and Klomp 1999; Law and Anderson 

1999; Law 1996; Lumsden et al. 2002a; b). 

In the woodlands of the Victorian Riverina, insectivorous bats utilise different habitats for roosting 

and foraging, often commuting large distances between the two types of habitat (Lumsden et al. 

2002a). Two species, the Lesser Long-eared Bat Nyctophilus geoffroyi and Gould’s Wattled Bat 

Chalinolobus gouldii, were found to roost, variously, in trees, fallen and decaying timber and 

under bark, though maternity roosts for both species were predominantly located in large dead 

trees. The spouts of large River Red Gum Eucalyptus camaldulensis trees were especially 

important as roost sites for male Gould’s Wattled Bats (Lumsden et al. 2002a; b). In these 

floodplain forests both species roosted in locations that had greater densities of hollow-bearing 

trees than were generally available, suggesting roost selectivity by these bats; Lesser Long-eared 

bats utilised dead hollow-bearing trees and Gould’s Wattled Bat, large live trees (Lumsden et al. 

2002a). 

Many species of insectivorous bats of wetter forests in south-eastern Australia (e.g. Highlands 

Northern Fall, Highlands Southern Fall, Northern Inland Slopes bioregions in Victoria) are known 

to require hollows in mature trees as roost sites (Law 1996); some of these species (e.g. 

Vespadelus spp.) are also likely to show a high degree of site fidelity, potentially making them 

more vulnerable to logging activities (Brown and Howley 1990; Churchill 2008; Law 1996). 

Mentioned above is the Parks Victoria Ecological Thinning Trial in box-ironbark woodlands of 

central Victoria (Parks Victoria 2007; 2009), a long-term field-based experimental programme that 

will evaluate different methods of ecological thinning and the effects they have on the biotic 

components of the box-ironbark forest ecosystem, including select vertebrate fauna and key habitat 



characteristics. Mammals (particularly arboreal taxa and bats) and birds are key foci of this study, 

and benchmark values for their composition and abundance at the experimental sites in these 

woodlands have been established (Brown and Horrocks 2008). It is too soon to gauge the 

responses of either the biodiversity or desired habitat characteristics to ecological thinning in this 

Trial; this should become apparent in years (decades) to come. 

4.3.2 Birds 

Various estimates have been made of the number of Australian vertebrate species that use tree 

hollows (e.g. 400 species, Ambrose 1982; 303 species, Gibbons and Lindenmayer 2002). In 

Victoria, tree hollows are considered essential for 47 bird species (Department of Sustainability 

and Environment 2007; Emison et al. 1987; Menkhorst 1984), which use them primarily for 

nesting or roosting (Table 3.2). Fourteen of these bird species are listed as threatened (Department 

of Sustainability and Environment 2007). Many additional species nest on ledges or open hollows 

(e.g. woodswallows), or use hollows opportunistically. One species (the endangered Swift Parrot) 

depends on hollows for nesting, but not in Victoria as this migratory species nests only in 

Tasmania. 

The six Victorian owl species (Powerful Owl Ninox strenua, Barking Owl N. connivens, Sooty 

Owl Tyto tenebricosa, Masked Owl T. novaehollandiae, Eastern Barn Owl T. javanica and 

Southern Boobook N. novaeseelandiae) all nest mainly in hollows, though some use is made of 

caves and buildings. The latter four species are also dependent to varying degrees on hollows for 

daytime roosting (Higgins 1999). 

Large forest owls are often considered as ‘umbrella’ species (sensu Simberloff 1998) in the sense 

that they occupy large home ranges, are hollow-dependent and prey heavily on arboreal (hollowdwelling) 

prey (possums and gliders). They have been used in this way in Victoria, where their 

conservation depends largely on retention of extensive tracts of old forest (Loyn et al. 2001) 

including abundant tree hollows. Modelling has shown that the probability of detecting a Powerful 

Owl responded positively to the number of live hollow-bearing trees and the Sooty Owl responded 

positively to the number of dead hollow-bearing trees at the call playback survey site (Loyn et al. 

2002). 

Powerful Owls occur across most of the bioregions covered by this report with the most significant 

concentrations in the Victorian Highlands - Southern Fall; Central Victorian Uplands and, to a 

lesser extent, Goldfields (Victorian Fauna Database, DSE, Emison et al. 1987). Barking Owls are 

scarce in Victoria with clusters of records in the Northern Inland Slopes, Victorian Highlands - 

Southern Fall and Goldfields bioregions (Victorian Fauna Database, DSE, Emison et al. 1987). 

Sooty Owls favour wetter forests in the eastern half of the state (Victorian Fauna Database, DSE, 

Emison et al. 1987). 

Other species that nest in hollows include parrots, cockatoos, owlet-nightjars, kingfishers and a 

small number of passerines (notably treecreepers and Striated Pardalote Pardalotus striatus). 

Dead trees can provide valuable sources of hollows (e.g. Nelson and Morris 1994), but generally 

do not remain standing for as long. Studies in forests of Mountain Ash have shown that hollowdependent 

birds (and several other bird groups) respond more strongly to numbers of live trees 

than dead trees (Loyn and Kennedy in press). Their study also showed that the density of old trees 

was more important than their spatial distribution. 

Some bird species require highly specific nest hollow characteristics (McElhinny et al. 2006). The 

dimensions of a hollow can determine the species that may use it. Therefore, a diversity of hollow 



types is more likely to support a diversity of bird species (McElhinny et al. 2006). Small species 

favour hollows with the smallest entrance they can enter to preclude larger predatory species from 

access. Similarly, large birds such as owls require large hollows. Lower hollows can lead to 

greater risk of predation as reported for the near threatened Turquoise Parrot Neophema pulchella, 

nesting in old hollow fence posts (Quinn and Baker-Gabb 1993). 

Hollows may be used by birds for purposes other than nesting. Some owls and Australian owletnightjars 

use hollows for roosting (HANZAB). Treecreepers often roost in crevices in large 

hollows or fire-scars. Several birds may drink from hollows when they fill with water. 

A number of diurnal bird species take prey opportunistically from tree hollows. For example, 

Laughing Kookaburras Dacelo novaeguineae (Victorian Fauna Database, DSE) and Ravens 

Corvus spp. take nestlings (EM pers. obs.) of smaller, hollow-nesting bird or mammal species. 

Insectivores such as Thornbills Acanthiza spp. and omnivores such the Grey Shrike-thrush 

Colluricincla harmonica often forage in hollows seeking invertebrates (EM pers. obs.). 

4.3.3 Reptiles 

Gibbons and Lindenmayer (2002) estimated that 79 species of reptiles, about 10% of the 

Australian reptile assemblage, use hollows in Australia. Hollows are used by some reptile taxa as 

den or nest sites, and by some reptiles as sources of prey (Greer 2006). 

In the floodplain forests (Murray Fans bioregion) and woodlands of north-central Victoria 

(Goldfields, Riverina, Northern Inland Slopes bioregions), reptiles have declined, primarily, as 

argued by Brown et al. (2008), through the broad-scale loss of native vegetation and changing land 

use. It is not difficult to imagine that the loss of many large trees (and fallen or standing dead 

timber) across these regions have adversely affected the reptile fauna, especially those taxa that are 

arboreal or utilise hollow-bearing trees. 

Two such threatened taxa in these regions are the Tree Goanna and the Carpet Python (Department 

of Sustainability and Environment 2003c), both of which utilise hollows in both large logs and 

large trees (Alexander 1997; Department of Sustainability and Environment 2003d; Greer 2006; 

Greer 1989; Heard et al. 2004; Vincent and Wilson 1999). Other arboreal reptile taxa of these 

regions that utilise hollow-bearing trees include Carnaby’s Wall Skink Cryptoblepharus carnabyi, 

Tree Skink Egernia striolata, Marbled Gecko Christinus marmoratus (Brown and Bennett 1995; 

Brown 2002; Brown and Nicholls 1993). 

The Tree Goanna also occurs in wetter Victorian forests where it utilises hollow-bearing eucalypts, 

typically Mountain Ash Eucalyptus regnans, as do other reptile taxa, including the Black Rock 

Skink Egernia saxatilis, which has been observed 30 metres above ground on large living 

Mountain Ash trees (GB pers. obs.), and Spencer’s Skink Pseudemoia spenceri, which commonly 

occurs in large colonies on large emergent stags and is significantly associated with numbers of 

large Mountain Ash trees (Brown and Nelson 1993b). 

4.3.4 Amphibians 

To our knowledge there have not been any empirical studies on the use of hollow-bearing trees by 

frogs, although the number of arboreal frog species in south-eastern Australia, principally from the 

Litoria genus, suggests that hollows are used, if only opportunistically. Gibbons and Lindenmayer 

(2002) nominate 27 Australian arboreal or semi-arboreal frog species that potentially use hollows, 



and note hollow use by frogs is difficult to detect because of their small size, absence of evidence 

of hollow use, and a lack of knowledge of their ecology. 

In northern Victoria (Murray Fans, Riverina bioregions) Peron’s Tree Frog Litoria peronii is an 

arboreal species that is often found under bark and in fissures of large River Red Gum trees (GB 

pers. obs.) and also in hollows of these floodplain trees (Gibbons and Lindenmayer 2002). 

4.3.5 Invertebrates 

Limited information is available on tree-hollow use by Australian invertebrates (Gibbons and 

Lindenmayer 2002), thus we draw on the following international examples. Harvesting may have 

a negative impact on invertebrate diversity if it results in the future reduction of stags or living 

trees with hollows. Nilsson and Baranowski (1997) found a greater number of red-listed beetle 

species in hollow trees from old-growth beech forest in Sweden than in recently (50 – 100 years) 

disturbed forests. They also observed the red-listed species occurred in low frequencies i.e. only 

every 20 th hollow or dead tree were inhabited by a particular species. This means that a large 

number of stags and hollow trees need to be retained at a site. The retention of trees with large 

girths may also be important for beetle conservation and therefore ecological thinning may benefit 

some species. Rainus (2002) found that species richness for eleven beetle species in Sweden was 

highest in large trees. 

4.4 Flora 

The impact of firewood collection and harvesting can affect plant communities directly, through 

physical damage incurred during the operation, and indirectly, through changes to the ambient or 

edaphic conditions that plants experience, or through the introduction of competitors and disease 

(Driscoll et al. 2000; Penman et al. 2008b). However, despite the volumes of wood harvested 

from Victoria's forests, little research has been undertaken to quantify the effects that this 

harvesting is having on the composition and function of forest ecosystems, requiring us to draw 

heavily on research from other applications. Most of the available research is derived from 

forestry operations, particularly clearfelling of logging coupes, and, to a lesser extent, heavy 

silvicultural thinning, and the effects of these operations are expected to be substantially more 

intensive than those from firewood harvesting. Thus, we must tread cautiously when extrapolating 

results from other research. An ecological thinning trial was recently initiated by Parks Victoria in 

Box-Ironbark and Heathy Dry forests from west-central Victoria (Pigott et al. 2008), and over the 

next few years this should provide data that are pertinent to the forests that are experiencing high 

demand for firewood. 

4.4.1 Understorey 

The canopy formed by overstorey trees in a forest has a major impact on the conditions 

experienced by plants at ground level or in subordinate strata, particularly through the interception 

of light and water and the complexities of plant-plant competition. These effects depend to a large 

extent on the type of forest, as canopy density varies both spatially and temporally according to the 

overstorey species present (Belsky et al. 1989; Kirkpatrick 1997; Messier et al. 1998; Rokich and 

Bell 1995; Stewart 1988; Turton and Duff 1992) and aridity (Specht 1972; Specht and Morgan 

1981). 



The creation of canopy gaps by the partial removal of the overstorey leads to localised changes in 

the ambient environment in which understorey species exist. Gaps or other areas without 

overstorey generally experience higher photosynthetically-active radiation, higher maximum soil 

temperatures, higher minimum ground temperatures and decreased water deficit when compared to 

areas under canopy (Bowman and Kirkpatrick 1986a; Collins et al. 1985) (Bauhus et al. 2001; 

Belsky et al. 1989; Kirkpatrick 1997; Nunez and Bowman 1986; Rokich and Bell 1995; Stoneman 

et al. 1994). Given the potential differences in ambient conditions between gaps and the 

surrounding canopy, the creation of additional gaps by firewood harvesting might elicit local 

responses in understorey vegetation. 

Some forest herbs may be adapted to high-intensity light, while others may need low-intensity 

light to avoid inhibition of photosynthesis. Other herbaceous species display plasticity or 

flexibility, and are able to adjust both physiologically and physically to a wide range of light 

regimes (Collins et al. 1985). Changes in the amount or wavelengths of light reaching the ground 

may affect seed germination, as seeds of some species require varying amounts of light for 

germination while seeds of other species require darkness (Rokich and Bell 1995). The variations 

in temperature, moisture and light in canopy gaps can affect photosynthesis and assimilation in 

forest herbs, influencing growth rates, growth form, and even allocation to sexual and asexual 

reproduction (Collins et al. 1985). 

The degree to which canopy thinning affects the understorey environment, hence drives species 

change, differs substantially depending on forest type, the size and nature of the canopy gaps and 

the individual characteristics of understorey species. In Northern Hemisphere conifer forests, 

thinning leads to a large increase in the amount of light reaching the light-limited understorey, 

causing pronounced (although variable) increases in the cover of herbaceous species, particularly 

grasses (Alaback and Herman 1988; Dodson et al. 2007; Laughlin et al. 2005; Liira et al. 2007; 

McConnell and Smith 1970; Thomas et al. 1999), moving stands closer to older-growth 

composition (Lindh and Muir 2004) and promoting flowering (Lindh 2008). Shade-intolerant 

species display improved establishment and growth (Bock and Van Rees 2002). In broadleaved, 

deciduous forests, ground and shrub layer cover increased significantly with increasing harvest 

intensity (Fredericksen et al. 1999), although shade-tolerant species exhibited reduced growth and 

increased mortality in full sun treatments (Small and McCarthy 2002). However, changes may be 

complex and unpredictable, with the same species showing both increases and decreases in 

response to thinning at different sites (Götmark et al. 2005). In South American Lenga Beech 

Nothofagus pumilio forest, tree seedling survival and growth was highest in canopy gaps in mesic 

forest, but highest under shade in xeric forests (Heinemann and Kitzberger 2006). 

In Australia, eucalypt canopies are persistent but often open, with more light reaching the 

understorey than in many other forest types (Kirkpatrick 1997), reducing the need for understorey 

plants to be shade-tolerant, and foliage cover tends to reduce from humid to arid zones (Specht 

1972; Specht and Morgan 1981). Box-Ironbark and River Red Gum forests, which bear the brunt 

of firewood collection in Victoria (Driscoll et al. 2000), may be considered both relatively dry 

(Muir et al. 1995) and open, and the understorey changes following increased light penetration 

might therefore be smaller, or substantially slower, than that noted in denser forest types. 

At a broader community level, and depending on site conditions, a reduction in the foliage 

projective cover of the overstorey may be compensated by an increase in the cover of the 

understorey, as the covers of the two strata in many forest types tend to be inversely related 

(Specht and Morgan 1981). Previous research suggests that the grassy layer is particularly 

responsive to change. For example, tree thinning in Narrow-leaved Ironbark Eucalyptus crebra 

woodland in Queensland resulted in a significant increase in herbage biomass (Walker et al. 1986). 

Similarly, thinning in Bimble Box E. populnea shrub woodlands led to increasing yields of 



herbage biomass (Walker et al. 1972), as it did in Mulga scrub (Beale 1973), while eucalypt sites 

in central Queensland produced higher pasture yields when tree basal area was lower (Scanlan and 

Burrows 1990). The difference between high- and low-basal area sites was more pronounced at 

sites of lower productivity. In mixed Eucalyptus communities in central Queensland, sites with 

lower tree basal area had increased amounts of grasses such as Black Speargrass Heteropogon 

contortus and Kangaroo Grass Themeda triandra than did sites with higher tree basal area 

(Scanlan and Burrows 1990), while Flooded Gum Eucalyptus grandis plantation sites had higher 

cover of grass under a more severe thinning treatment (Cummings et al. 2007). 

However, changes are species-specific, and depend on individual habitat preferences. In Silvertop 

Stringybark E. laevopinea open-forest in northern New South Wales, Weeping Grass Microlaena 

stipoides was dominant beneath mature forest canopy, while Cane Wire-grass Aristida ramosa 

(and to a lesser degree Grey Tussock-grass Poa sieberiana) was dominant in large open spaces 

(Gibbs et al. 1999). Similarly, the abundance of Weeping Grass was significantly correlated with 

higher tree density in paddocks (Magcale-Macandog and Whalley 1994). Local frequency and site 

occurrence of Slender Wallaby-grass Austrodanthonia racemosa were both positively correlated 

with increasing tree cover, as was site occurrence of Velvet Wallaby-grass Austrodanthonia pilosa 

(Scott and Whalley 1982). 

Observations by early explorers suggested that woodlands such as Box-Ironbark were originally 

open, with a highly diverse grassy to shrubby understorey (Calder et al. 1994). This suggests that 

on-going canopy thinning from firewood harvesting may have a positive effect on overall grass 

cover in that vegetation community. 

The same factors that drive the abundance and richness of grasses can combine to drive the 

abundance and richness of other herbaceous species. In Silvertop Ash Eucalyptus sieberi forest in 

Victoria, thinning promoted the abundance of herbaceous species, particularly Germander 

Raspwort Gonocarpus teucrioides, although changes in total understorey cover, species richness, 

species diversity or lifeform diversity were not significant (Bauhus et al. 2001). Tree thinning in 

Narrow-leaved Ironbark woodland in Queensland led to a significant increase in herbage biomass 

and forb density (Walker et al. 1986). In contrast, maximum forb richness in Flooded Gum 

plantation was associated with higher, not lower, canopy cover (Cummings et al. 2007). However, 

no data are available to suggest what the response of forbs will be to thinning in forests such as 

Box-Ironbark or Red Gum that are most impacted by firewood harvesting. 

Many winter-flowering orchids such as Greenhood Pterostylis, Midge-orchid Corunastylis, 

Mosquito-orchid Acianthus and Gnat-orchid Cyrtostylis prefer moister conditions, often under tree 

cover, and might respond negatively to thinning of the canopy or disturbance of the shrub layer 

associated with firewood harvesting. However, spring-flowering orchids such as Spider-orchid 

Caladenia, Beard-orchid Calochilus, Diuris Diuris and Wax-lip Orchid Glossodia often prefer 

drier conditions, and might respond positively to an increased light regime (pers. comm., Mike 

Duncan, Arthur Rylah Institute). Similarly, in northern hemisphere mixed mesophytic forests, 

flowering of the herb White Snakeroot Eupatorium rugosum was higher in gaps than under the 

shade of the canopy (Landenberger and Ostergren 2002). 

Opportunistic species such as weeds are likely to be favoured in some sites. For example, soil 

disturbance in Mountain Ash forest led to an initial, abrupt increase in ruderal and weed species 

after thinning (Peacock 2008) and clearfelling (Appleby 1998), although the absolute cover of 

weeds remained relatively low. Logging also increased the abundance of weeds (mostly annual 

grasses and short-lived herbs) in Jarrah forest (Burrows et al. 2002). Frequently disturbed areas 

such as roadsides appear to be an important source of propagules (Appleby 1998), suggesting that 

regular disturbance and vehicle access associated with firewood cutting might result in an 



incremental increase in weeds. Weeds are not generally a major feature of Box-Ironbark and 

similar forests, although a high abundance of weeds is sometimes found in flat moister areas 

(Arthur Rylah Institute, unpublished data), which are also habitat for Yellow Box Eucalyptus 

melliodora, a preferred firewood species. Disturbance in these moister areas is likely to be more 

significant in promoting weeds than disturbance on drier slopes or ridges. 

The suppression zone resulting from canopy tree roots may extend well out beyond the canopy 

(Incoll 1979; Lamont 1985; Rotheram 1983), and root competition for water in this zone appears 

to suppress shrub and overstorey regrowth. This suggests that a more open overstorey might 

promote the vigour of understorey species, with attendant benefits to ecosystem processes. 

Nonetheless, responses to canopy thinning by shrubs will be species-specific. In semi-arid Bimble 

Box woodland in New South Wales, shrubs such as Wilga Geijera parviflora and Turkey Bush 

Eremophila deserti grew beneath the canopy, while Mulga and Desert Cassia Senna artemisioides 

grew away from the canopy (Harrington et al. 1981). In mixed Eucalyptus communities in central 

Queensland, sites with lower tree basal area had decreased amounts of some native legumes or 

broad-leaved plants than did sites with higher tree basal area (Scanlan and Burrows 1990). In 

Flooded Gum plantation, the cover, density and richness of shrubs and woody climbers were 

lowest in plots with the least canopy (Cummings et al. 2007), suggesting regeneration in these 

species was better under closed canopy. 

Anecdotal evidence from early explorers suggested that the shrub layer in forests and woodlands 

such as Box-Ironbark was well developed, under widely-spaced trees (Calder et al. 1994), 

suggesting that shrub communities in drier forests and woodlands might respond differently to 

canopy thinning than shrubs in wetter, taller forests. Shrubs might respond positively to thinning 

in drier forests, particularly in areas that are rocky with lower grass cover. Germination cues will 

be important. For example, Spreading Wattle Acacia genistifolia, Golden Wattle Acacia 

pycnantha and Hedge Wattle A. paradoxa all show a strong heat-stimulated germination response 

(Brown et al. 2003), yet persist at low levels (with occasional recruitment) in long-unburnt forest 

(Arthur Rylah Institute for Environmental Research, unpublished data). The soil disturbance 

associated with firewood harvesting might encourage some germination (Franco and Morgan 

2007). 

Finally, the disturbance associated with the firewood harvesting activities will impact directly on 

plant cover, at least initially, through physical damage to the plants. Disturbed Box-Ironbark sites 

had lower overall diversity and cover of understorey and ground layer species (Edwards 1997). 

Four years after logging in Jarrah forest in WA, the total abundance of individual native plants, 

particularly perennial herbs and sedges, was significantly lower in logged forest patches than in 

buffer zones (Burrows et al. 2002). Felling disturbance in Mountain Ash forest led to significant 

decreases in tall shrubs and small trees (Peacock 2008). Damage to flowering plants before they 

have had time to flower and set seed might result in a reduction in the soil seed bank, but no 

literature was found during this review that was relevant to thinning or firewood extraction 

activities. 

Reseeding species might be less disadvantaged than resprouting species by intensive disturbance. 

For example, logging followed by slash burning favoured the germination of reseeding species 

over the growth of resprouting species in dry sclerophyll forest (Penman et al. 2008b) and wet 

forest (Murphy and Ough 1997). Soil disturbance from grading tracks and vehicle movements 

appear to have facilitated the spread of Hedge Wattle in grassy woodland, particularly in areas of 

higher soil moisture (Franco and Morgan 2007). 

A high proportion of species used resprouting as a regeneration mechanism after fire in Box- 

Ironbark forest (Orscheg 2006), as did 60 of 66 species in dry sclerophyll foothill forest in north- 



central Victoria (Tolhurst 1996). By damaging the roots of these resprouting species, intensive 

harvesting activities may leave them at a competitive disadvantage in the early years after such 

harvesting. However, the response to logging disturbance is likely to be different to that after fire 

(Penman et al. 2008a), making inferences difficult to make. Data relating to the potential effects 

of firewood harvesting or canopy thinning on plants with different regeneration strategies were not 

found during this literature review. 

Logging disturbance also results in soil compaction (Edwards 1997; Small and McCarthy 2002), 

which can have a negative influence on plant growth, although the severity would depend to a 

large extent on the intensity and frequency of the disturbance. Changes in macroporosity 

following logging in Mountain Ash forest were not considered severe enough to affect root growth 

(Rab 2004). No data were found during this literature review on the effects of compaction in dry 

forests that would be subject to firewood harvesting. Disturbance to the soil crust may also occur, 

although the impacts may be limited if the soil is relatively dry. 

4.4.2 Eucalypt canopy 

Thinning of overstorey trees during firewood harvesting operations can affect the growth of 

remaining trees and the germination and growth of recruits. 

Reducing the overall density of canopy trees by thinning (such as regularly undertaken in 

commercial forestry) should result in a decrease in competition between the remaining trees for 

water and nutrients, thereby allowing increased growth rates (Department of Natural Resources 

and Environment. 1999). Of preferred firewood species, such increases in growth rates have been 

measured in Red Ironbark (Kellas et al. 1998; Kellas et al. 1982), although increases were more 

evident in regrowth stems than overwood (Kellas et al. 1982). 

Tree species recruitment is limited in some forest types by the presence of an intact canopy. For 

example, adult Alpine Ash (Eucalyptus delegatensis) trees, due to their impact on soil moisture, 

suppress seedlings under their canopies (Bowman and Kirkpatrick 1986a; Bowman and 

Kirkpatrick 1986b), as do adults of Silvertop Ash (Incoll 1979). Similarly, Jarrah seedlings on 

sites without overstorey experience smaller soil and leaf water deficits and higher survival than 

seedlings with the overstorey retained (Stoneman et al. 1994). Few eucalypt seedlings were noted 

in areas where large, mature trees formed a closed canopy, but small seedlings were common in 

open areas (Gibbs et al. 1999). In contrast, regeneration of canopy species in Flooded Gum 

plantation was higher with increased canopy retention (Cummings et al. 2007), suggesting again 

that response will be species-specific. 

In Wandoo Eucalyptus wandoo woodland in Western Australia, shrub seedlings were able to 

establish in the suppression zone around the overstorey trees, but premature death of adult shrubs 

appeared to occur when shrub roots eventually met the large lateral tree root system (Lamont 

1985). Thus, initial establishment by woody recruits following firewood harvesting in some 

forests may not translate to longer-term increases in cover. However, overstorey recruitment in 

Box-Ironbark, Grassy Dry and Heathy Dry Forests is a continual process, and does not appear to 

be inhibited by the relatively low level of shade or the presence of mature trees. Seedlings and 

juveniles of Long-leaf Box E. goniocalyx, Red Stringybark E. macrorhyncha and Red Box 

E. polyamthemos are common in these forests, regardless of canopy density, although they may 

remain suppressed (Arthur Rylah Institute, unpublished data). 

Higher cover of Eucalyptus species has been noted in sites subject to heavy disturbance (Edwards 

1997), suggesting that firewood harvesting may result in an increase in the density of smaller 



stems. The overall stem densities in forests such as Box-Ironbark are also likely to increase 

substantially as a result of coppice growth following cutting (Edgar 1958), and the resultant (albeit 

patchy) denser canopy might affect the understorey as discussed in the previous section. 

It is worth noting that recruits of Red Ironbark E. tricarpa, a preferred firewood species, are less 

abundant in areas where it occurs, and this is not surprising given that flowering and seed 

production in this species tends to be sporadic (Kellas 1991). The germination requirements for 

this species are poorly understood. Seeds seem to be short-lived and lack dormancy, yet while 

they germinate readily under laboratory conditions, they do not germinate readily under field 

conditions (Orscheg 2006). The long-term response of Red Ironbark to continual firewood 

harvesting is therefore of particular interest, but does not appear to have been the subject of 

research. 

Firewood harvesting can affect overstorey diversity directly through the selective cutting of 

preferred species (Shahabuddin and Kumar 2007). In Victoria, River Red Gum E. camaldulensis 

is the most popular firewood consumed (Driscoll et al. 2000). This tends to occur in relatively 

monospecific stands, and cutting would alter forest structure more than composition. However, 

Red Box and Yellow Box, which occur in mixed stands, are the second most common species 

burned (Driscoll et al. 2000), and selective cutting of these species could eventually alter forest 

composition. Almost one third of Victoria's state forest harvest in 1997-1998 came from the 

mixed Box-Ironbark forests of the Bendigo Forest Management Area (Driscoll et al. 2000), 

suggesting that these compositional changes may be concentrated in particular areas. Disturbed 

Box-Ironbark sites are often dominated by a single species (Edwards 1997). The author (AT) has 

observed areas of the Craigie Forest (near Maryborough) where Red Box existed almost entirely as 

new coppice growth on stumps, rather than as mature trees. Selective cutting of Yellow Box is of 

particular concern, as it is generally found in sheltered sites near rivers or flat areas with poorer 

drainage (Viridans, Victorian Flora Database) and gullies, which have a disproportionate 

importance in terms of fauna richness and conservation value (Mac Nally et al. 2000b). 

Interruptions to tree life cycles and life-spans may also occur as a result of selective cutting. Trees 

must reach a particular age before they set viable seed, and this time-to-maturity has been used 

extensively for determining successional processes in plant communities following fire (Noble & 

Slatyer 1981). We note that the Bendigo Forest Management Area specifies a minimum period of 

25 years between sawlog harvesting at a site to allow recruitment across all age classes (DSE 

2008), but we have not researched the equivalent prescriptions for firewood harvesting in the 

various regions. If young trees or coppice growth are continually harvested before they have 

reached a sufficient age to produce viable seed, and if the larger, retained trees eventually become 

senescent, then the resultant long-term effects might be similar to that experienced by forests that 

are overdue for a burn. However, no research was found during this review that related to this 

aspect of harvesting. 

In Victoria, 49 plant communities are recognised as potentially threatened by firewood collection, 

including 23 forest communities, mostly in lower rainfall areas (Driscoll et al. 2000). Our analysis 

indicates that 43 forest or woodland EVCs are potentially affected in the 13 bioregions from which 

most firewood appears to be harvested, of which 32 are considered to be Endangered or 

Vulnerable in at least one of those bioregions (see 2.2.1 Vegetation communities). 

4.4.3 Nectar and pollen resources 

Nectar and pollen represent important resources upon which many vertebrate and invertebrate 

fauna species rely. Of course, they are also important to the honey industry. However, the 



availability of these resources, particularly from canopy species, varies both spatially and 

temporally (Department of Agriculture 1946; Keatley and Hudson 2007; Keatley et al. 2004; Law 

et al. 2000; Wilson 2002). 

Flowering of some eucalypt species can occur any time of year, but in general there is a consistent 

sequential pattern of peak flowering between species (Keatley and Hudson 2007). For example, in 

Box-Ironbark forests, peak flowering of Yellow Box occurred in November to January (depending 

on location), followed by Grey Box Eucalyptus microcarpa in March, Red Ironbark in June/July, 

Yellow Gum E. leucoxylon in July to September, then Red Box in September to November 

(Keatley and Hudson 2007). Flowering in forests around Rushworth commenced and peaked 1 to 

4 months earlier than flowering in forests around Havelock (further to the south-west) (Keatley 

and Hudson 2007). Flowering peaks were also skewed in some species, with production in Red 

Ironbark slowly tapering off after an early peak, but production in Yellow Gum slowly increasing 

to a late peak (Keatley et al. 2004). 

Many eucalypt individuals flower only every second year or so, sometimes en masse, and flower 

abundance may vary substantially due to conditions in the current or previous years (Department 

of Agriculture 1946; Law et al. 2000; Wilson 2002). In July 1997, for example, the winterflowering 

Red Ironbark was flowering in only 3 of 5 geographic areas, while the percentage of 

trees flowering in a particular area ranged from 0% to 42% (Wilson and Bennett 1999). Nectar 

production within and between individual eucalypt trees is equally variable (Law and Chidel 

2008). 

The flowering patterns and differential responses by individual trees to seasonal conditions suggest 

that a substantial degree of genetic diversity (and adaptability) exists within species, and the 

resultant asynchrony in peak flowering patterns ensures that floral resources are available 

throughout the year. However, selective cutting of species, such as the summer-flowering species 

Yellow Box and River Red-gum, may negatively influence the distribution, abundance and timing 

of floral resources (Wilson 2002), with implications for those organisms dependent on them. 

Yellow Box, which generally flowers en masse every second year, is considered by the honey 

industry to be the most valuable nectar-yielding tree in Victoria (Department of Agriculture 1946), 

hence forest management prescriptions for the Bendigo Forest Management Area require the 

retention of all living Yellow Box trees (Department of Sustainability and Environment 2008a). 

Asynchrony in eucalypt flowering may also help reduce hybridisation between species (Keatley et 

al. 2004). 

Little research was found during this literature review that addressed the likely effects of timber 

harvesting on flowering patterns, floral resources or pollination. In northern New South Wales, 

time-since-logging was not correlated with the percentage of the canopy in flower, and logging did 

not generally interrupt flowering cycles (Law et al. 2000). Nonetheless, differences were noted 

between species. For example, Smooth-barked Apple Angophora costata had a greater proportion 

of canopy in flower in recently-logged sites, possible due to reduction in competition, but Grey 

Ironbark Eucalyptus siderophloia and Forest Red Gum E. tereticornis tended to flower poorly 

(Law et al. 2000). In Spotted Gum E. maculata forest, smaller trees in recently logged forest 

produced less sugar, and more diluted sugar, than trees in regrowth or mature forest (Law and 

Chidel 2008). When scaled up to forest stand level, the amount of sugar produced per night in 

recently logged forest was around one-tenth of that produced in mature forest (Law and Chidel 

2008). This corresponds with the observation that nectarivorous birds tend to be more numerous 

in mature forest than regrowth (Loyn 1980; 1985) and respond positively to numbers of retained 

live old trees (Loyn and Kennedy in press). 



The age of a tree is an important factor influencing eucalypt flowering patterns. Larger trees 

flower more frequently, more intensely, and for a greater duration than smaller trees, and have 

bigger canopies that produce more flowers per unit area of canopy (Wilson 2002; Wilson and 

Bennett 1999). For example, in Spotted Gum forest, a large tree was estimated to have 74 000 

flowers, compared to a medium-sized tree with 4 000 flowers (Law and Chidel 2008). This 

suggests that large trees have a disproportionate importance in eucalypt forests. 

In tropical forests, large trees may also, in some instances, contribute more to pollination than 

small trees (Lourmas et al. 2007). However, few data exist on the potential effects of logging or 

selective harvesting on pollination. Pollen dispersal by insects (and presumably larger vertebrates) 

may occur over relatively large distances (Lourmas et al. 2007), and there may be little impact 

from a localised reduction in the quantity of individual species. Hybridisation rates in the 

uncommon Black Gum E. aggregata increased with reduced population size, while seed 

production, germination and seedling survival declined (Field et al. 2008). Species targeted for 

firewood harvesting are considered to be relatively common, and increased hybridization rates and 

reduced seedling performance are probably unlikely. Nonetheless, if populations of uncommon, 

non-target canopy or understorey species are reduced by firewood harvesting, increased 

hybridization (leading to reduced fecundity) in those species remains a possibility. However, no 

research was found that addressed that issue. 

4.4.4 Cryptogams 

Cryptogams are plants that reproduce by spores rather than seeds, and include lichens, bryophytes, 

algae and fungi (Scott et al. 1997). Fungi are dealt with in Section 4.6. Despite being important 

ecologically, their responses to disturbance have been poorly studied in comparison with those of 

vascular plants. 

Lichens in particular play a major role in stabilising the surface and preventing erosion in semiarid 

areas (Scott et al. 1997), but are vulnerable to the physical impacts associated with firewood 

harvesting. In Mulga woodland, for example, soil surface features such as lichens and 

cyanobacterial crusts were not present at firewood sites, despite being common elsewhere (Berg 

and Dunkerley 2004), rendering surfaces vulnerable to erosion by wind and rain. Lichens are also 

used as food by invertebrates such as mites or gastropods (slugs and snails), or as protective cover 

by some insects or their larvae (Scott et al. 1997). In sclerophyll forests, where lichens grow on 

the branches and trunks of many trees (Scott et al. 1997), firewood harvesting would result in a 

localised reduction in the resource that they represent. 

Bryophytes (especially mosses) also help to stabilise soil surfaces, and can form a crust in semiarid 

areas in conjunction with lichens and algae (Scott et al. 1997). They would also be affected 

by the physical disturbance associated with firewood harvesting. However, while they occur in 

large amount on trees and logs in wetter forests (Scott et al. 1997), they do not appear to be a 

major feature of woody material in drier forests (see Section 3 above). 

CWD is also an important substrate for certain lichen and bryophyte species (Andersson and 

Hytteborn 1991). International studies have shown that a diversity of decay stages are also 

important for bryophyte diversity with different guilds forming a successional pathway (Andersson 

and Hytteborn 1991; Odor et al. 2006). Species richness was observed to increase with increasing 

CWD diameter in semi-natural beech forests in Europe (Odor et al. 2006). The limited work that 

has been conducted in Australia has predominantly focused on wet eucalypt forests in Tasmania. 

CWD was found to consist of the greatest number of significantly associated bryophyte species 

than any of the other substrate types. Many of these species were also associated with old growth 



forest logs. Dead trees or stags however only had one bryophyte species positively associated with 

it (Turner and Pharo 2005). Some bryophyte species have an obligatory association with CWD 

and others facultative (Odor et al. 2006). 

Cryptogams (lichens in particular) are sensitive to changes in light and humidity, and destruction 

of habitat, especially the protective cover of vascular plants, represents a major threat to them 

(Scott et al. 1997). However, empirical research data are lacking. 

4.5 Fungi and microbial organisms 

Wood decay is an important contributor to internal tree defect, often in association with termite or 

borer attack. There are two principal sources of wood decay formation; associated with defective 

branch ejection and wounding, and linked to stem damage. White and Kile (1991) have 

demonstrated that stem wounds inflicted during mechanical harvesting operations can lead to the 

development of substantial columns of decay. They also identified that defect developed more 

rapidly from closed wounds than open wounds. Other studies have found that these decay columns 

also often originated from other sources, such as branch stubs and wood-boring insects (Wardlaw 

1996). 

These pathogens are most active in areas of high rainfall where their impact on wood structure can 

be considerable (Wardlaw and Neilsen 1999). While some research and reviews of current 

knowledge of the effect of damage and defect due to thinning and harvesting have been carried out 

(Dudzinski et al. 1992; Kile and Johnson 2000; Old et al. 1991; White and Kile 1991), there is still 

a need for further research (Old et al. 1991; White and Kile 1991), particularly in relation to 

revisiting previously thinned sites and trials (e.g. CSIRO silvertop damage trials). 

Fungi help decompose plant material and form symbiotic relationships with higher plants, and in 

forests and woodlands the number of macrofungi species appears to always exceed the number of 

vascular plant species (Scott et al. 1997). Clearing and alteration of habitats is considered the 

major threat to this group, either directly or through the reduction in host species (Scott et al. 

1997). Damaging processes associated with firewood harvesting could include soil compaction, 

removal of shading cover, loss of mycorrhizal host (plant) or dispersal agent (animal), loss of 

substrate (dead wood of a certain age or size), homogenisation of forest age, or invasion by exotic 

taxa. 



5 Which communities or species may be affected by 

firewood activities? 

The scale of firewood harvesting or collection varies substantially across Victoria, due to the 

occurrence or availability of preferred firewood species and demographic issues. In this section 

we determine the forest types that are most likely to be subjected to firewood activities, and 

identify vegetation communities or species therein that might be of concern in terms of their 

biodiversity status. Note that the degree to which these forest types might be affected by firewood 

collection will vary substantially depending on factors such as their spatial extent or proximity to 

townships, and we do not claim that firewood activities will always impact on them. 

To determine the main Ecological Vegetation Classes (EVCs) that were most likely to be under 

pressure from licensed firewood harvesting and collection, it was first necessary to identify the 

Forest Management Areas (FMAs) from where most firewood was sourced. For example, in 

1999/2000, the Bendigo FMA accounted for 35.7% of all firewood sold (Sylva Systems Pty Ltd. 

2002), followed by Midlands FMA (14.6%) and Mid-Murray FMA (10.5%). All FMAs were 

plotted in GIS, and colour-coded by firewood volumes (Figure 6.1a), providing a visual guide that 

allowed the 13 bioregions which supply most of Victoria's firewood to be identified (Figure 6.1b). 

A list was then compiled of all EVCs within these 13 bioregions. Wetland, heathland, grassland 

and other irrelevant EVCs were deleted, leaving only those forest and woodland EVCs in which 

firewood cutting was likely to occur. 

Forty-three EVCs are considered to be subject to firewood harvesting and collection (Table 6.1), 

although many others are no doubt affected by private or unlicensed collection, including those in 

areas outside the bioregions chosen for the analysis. 

5.1 Threatened EVCs and plant species 

5.1.1 Vegetation communities 

The Bioregional Conservation Status was determined for the 43 EVCs that were most likely to 

experience firewood harvesting in the 13 designated bioregions (see Figure 6.1) 

(http://www.dse.vic.gov.au/dse/nrence.nsf/Home+Page/DSE+Conservation~Home+Page?open) 

(Table 6.1). Note that the total area of individual EVCs and the extent to which they have been 

depleted vary between bioregions, and the same EVC may have a different status in a different 

bioregion. 

EVCs of most concern are those that are considered to be Endangered or Vulnerable. Endangered 

EVCs are those that are contracted to less than 10% of former range; OR less than 10% pre- 

European extent remains; OR combination of depletion, degradation, current threats and rarity is 

comparable to the others. Vulnerable EVCs are those where 10 to 30% of pre-European extent 

remains; OR combination of depletion, degradation, current threats and rarity is comparable. Our 

analysis indicates that 32 forest or woodland EVCs are considered to be Endangered or Vulnerable 

in at least one of the 13 defined bioregions (Table 6.1). Only two EVCs (Heathy Dry Forest and 

Shrubby Riverine Woodland) are considered to be of Least Concern. 

A summary is presented in Table 6.2, which shows that the bioregion Northern Inland Slopes has 

the highest number of endangered EVCs potentially subject to firewood harvesting (10). East 

Gippsland Uplands has no endangered EVCs likely to be affected. 

A list of vegetation communities covered by Victoria's Flora and Fauna Guarantee Act 1988 

(FFG Act) was then consulted to determine if any relevant forest or woodland EVCs were listed 



(http://www.dse.vic.gov.au/dse/nrenpa.nsf/Home+Page/DSE+Plants~Home+Page?open). These 

FFG-listed communities are generally more narrowly defined than EVCs, and may therefore be 

contained within several different EVCs. Similarly, relevant ecological communities included 

under the Commonwealth's Environment Protection and Biodiversity Conservation Act 1999 

(EPBC Act) (http://www.environment.gov.au/epbc) were identified. These ecological 

communities also do not align directly with EVCs, but in contrast to FFG-listed vegetation 

communities are generally broader than EVCs in their definition. Thus, with Commonwealth and 

state protection acting at community scales higher or lower than EVC scale, it is not possible to 

generate a simple table that specifies precisely the different levels of protection given to an 

individual EVC (as is possible with species). 

Three ecological communities listed under the EPBC Act (that are likely to be subject to firewood 

harvesting) occur in and around the bioregions assessed, of which two have affinities with 

vegetation communities listed under Victoria's FFG Act. At a state level, nine vegetation 

communities potentially subject to firewood harvesting are listed under Victoria's FFG Act, 

although only six are likely to be found in the 13 defined bioregions. Six of these FFG-listed 

vegetation communities have affinities with EPBC-listed ecological communities. 

The EPBC-listed Buloke Grassy Woodland occurs in the Riverina and Murray-Darling Depression 

Bioregions, and is dominated by Buloke Allocasuarina luehmannii, and occasionally Slender 

Cypress Pine Callitris gracilis or Grey Box Eucalyptus microcarpa. This ecological community 

has been extensively cleared, although its current exposure to firewood harvesting is unknown. 

Associated FFG-listed vegetation communities are Grey Box-Buloke Grassy Woodland 

Community, Semi-arid Herbaceous Pine Woodland Community and Semi-arid Northwest Plains 

Buloke Grassy Woodlands Community (the latter sometimes with Black Box E. largiflorens or 

Yellow Gum E. leucoxylon. The FFG-listed Semi-arid Herbaceous Pine-Buloke Woodland 

Community and Semi-arid Shrubby Pine-Buloke Woodland Community are also dominated by 

Cypress Pine and Buloke, but tend to occur in the north-west, outside the seven defined 

bioregions. 

Another EPBC-listed ecological community, Box-Gum Grassy Woodland and Derived Grassland, 

is affiliated with three Victorian EVCs (Valley Grassy Forest, Plains Grassy Woodland and Grassy 

Woodland), but not directly with any FFG-listed vegetation community. The overstorey variously 

consists of White Box E. albens, Yellow Box E. melliodora, Blakely's Red-gum E blakelyi and 

various other box or stringybark species. This ecological community has also been extensively 

cleared, and includes several preferred firewood species. 

The EPBC-listed Gippsland Red Gum E. tereticornis subsp. mediana Grassy Woodland and 

Associated Native Grassland, found on the Gippsland Plains, is dominated by Gippsland Red 

Gum, but may also include preferred firewood species such as Yellow Box and Red Box 

E. polyanthemos). Officially-identified threats include timber harvesting and firewood collection. 

In Victoria, the equivalent Forest Red Gum Grassy Woodland Community is listed under the FFG 

Act. 

Red Gum Swamp Community No. 1 and Creekline Grassy Woodland (Goldfields) Community, 

both dominated by River Red-gum E. camaldulensis, are also listed under Victoria's FFG Act. 

The latter may include Yellow Box and Grey Box. The FFG-listed Western Basalt Plains (River 

Red Gum) Grassy Woodland Floristic Community 55-04 is found outside the 13 defined 

bioregions, but may be affected to some degree by past or present firewood harvesting. None of 

these River Red-Gum vegetation communities correspond to an EPBC-listed ecological 

community. 


5.1.2 Plant species 


A list of all plant species recorded in three Victorian bioregions (Goldfields, Victorian Riverina 

and Murray Fans, Figure 6.1), including their conservation and FFG status, was extracted from the 

Victorian Flora Information System (Viridans Biological Database). The ten additional bioregions 

previously used for identification of threatened EVCs were not used, because the inclusion of such 

a broad range of vegetation types (that included coastal, Mallee, and montane to alpine) led to an 

unmanageable number of irrelevant species. In any event, these three bioregions account for up to 

two-thirds of the firewood harvested in Victoria. These data were then vetted to ensure that the list 

contained only those plant species likely to be in vegetation types (forests or woodlands) that were 

likely to be subject to firewood harvesting. Species listed under the Commonwealth's EPBC Act 

were identified as for EVCs. 

Vascular plant species listed under Victoria's FFG Act or the Commonwealth's EPBC Act, that 

occur in forests or woodlands of concern, are presented in Table 6.3, while a full list of species 

with a rare or threatened status is presented in the Appendix. Note that we make no claims about 

the extent to which individual species might be threatened by firewood harvesting activities, nor 

have we considered the mechanisms by which these plant species would be affected; we have 

simply generated a list of endangered species of which we should be mindful. 

Within the forests and woodlands of the three defined bioregions, 58 vascular plant species are 

listed under the EPBC Act, FFG Act, or both. In summary, 23 EPBC-listed and 55 FFG-listed 

species are in forests or woodlands potentially subject to firewood harvesting (Table 6.4). Twelve 

and 36 of these species are considered endangered in Australia and Victoria, respectively. 

However, the total number of potentially threatened species may be substantially higher, as an 

additional 10 endangered and 29 vulnerable species are not listed under the EPBC or FFG Acts 

(Appendix 2). A further 20 species are poorly known, but likely to be endangered, vulnerable or 

rare. 




b) Bioregions most relevant to this study. 



Table 5.1 Bioregional conservation status of EVCs likely to be subject to firewood harvesting. 

Bioregions: CVU Central Victorian Uplands; GO Goldfields; HN Highlands - Northern Fall; HS Highlands - Southern Fall; MF Murray Fans; NIS Northern Inland Slopes; VR 

Victorian Riverina; WI Wimmera; LM Lowan Mallee; EGU East Gippsland Uplands; EGL East Gippsland Lowlands; OR Otway ranges; WP Warrnambool Plain. 

Conservation Status: E Endangered; V Vulnerable; D Depleted; R rare; L Least concern. 

EVC No. EVC Name CVU GO HN HS MF NIS VR WI LM EGU EGL OR WP 

16 Lowland Forest L L L D V 

20 Heathy Dry Forest L L L L L L L L 

21 Shrubby Dry Forest L V L L L V L L L 

22 Grassy Dry Forest D D L L D D D L L D 

23 Herb-rich Foothill Forest D D L L L D L L D V 

24 Foothill Box Ironbark Forest V 

45 Shrubby Foothill Forest L D L L L D 

47 Valley Grassy Forest V V V V E V D D 

55 Plains Grassy Woodland E E E E E E E E E 

56 Floodplain Riparian Woodland E E E E D E V E 

61 Box Ironbark Forest V D V V V D 

66 Low Rises Woodland E V E E 

68 Creekline Grassy Woodland E E E E E E E E 

69 Metamorphic Slopes Shrubby Woodland D D 

70 Hillcrest Herb-rich Woodland D D D 

71 Hills Herb-rich Woodland V D V 

80 Spring Soak Woodland E E V 

103 Riverine Chenopod Woodland E V V E D 

106 Grassy Riverine Forest D D 



127 Valley Heathy Forest V E E V E V 

128 Grassy Forest V V E 

151 Plains Grassy Forest E E 

168 Drainage-line Aggregate E V E E 

169 Dry Valley Forest V V V V 

175 Grassy Woodland E V D D E E E E D D E E 

177 Valley Slopes Dry Forest L L R R 

198 Sedgy Riparian Woodland D V V E 

282 Shrubby Woodland R L L 

295 Riverine Grassy Woodland V V E V D 

641 Riparian Woodland E E E 

652 Lunette Woodland E E 

659 Plains Riparian Shrubby Woodland V 

663 Black Box Lignum Woodland E 

679 Drainage-line Woodland E 

704 Lateritic Woodland E V 

793 Damp Heathy Woodland D 

803 Plains Woodland E E E E E E E 

813 Intermittent Swampy Woodland D D V V 

814 Riverine Swamp Forest D D 

815 Riverine Swampy Woodland V V E V 

816 Sedgy Riverine Forest D V 

818 Shrubby Riverine Woodland L 

823 Lignum Swampy Woodland V V V V D 



Table 5.2 Summary of bioregional conservation status of EVCs likely to be subject to 

firewood harvesting. 

Bioregions: CVU Central Victorian Uplands; Gold Goldfields; H-NF Highlands - Northern Fall H-SF Highlands - 

Southern Fall; MF Murray Fans; NIS Northern Inland Slopes; VR Victorian Riverina. 

Bioregion Endangered Vulnerable Depleted Rare Least Concern 

Central Victorian Uplands 8 5 4 1 5 

Goldfields 8 6 6 0 2 

Highlands - Northern Fall 4 2 2 0 5 

Highlands - Southern Fall 4 5 2 0 7 

Murray Fans 5 4 5 0 1 

Northern Inland Slopes 10 3 1 0 3 

Victorian Riverina 8 11 5 0 1 

Wimmera 9 5 4 0 1 

Lowan Mallee 2 1 3 0 0 

East Gippsland Uplands 0 3 2 1 5 

East Gippsland Lowlands 2 1 2 1 4 

Otway Ranges 2 1 3 0 2 

Warrnambool Plain 3 2 1 0 0 

Table 5.3 EPBC or FFG-listed vascular plant species from forests and woodlands in three key 

bioregions likely to be subject to firewood harvesting. 

EPBC (Australian Threatened) Status: E Endangered; V Vulnerable. FFG Status: f = listed. Victorian (Rare or 

Threatened) Status: e endangered; v vulnerable; r rare. 

DICOTYLEDONS Common name EPBC FFG Vic 

Acacia deanei subsp. deanei Deane's wattle f e 

Acacia omalophylla Yarran Wattle f e 

Allocasuarina luehmannii Buloke f 

Brachyscome chrysoglossa Yellow-tongue Daisy f v 

Brachyscome gracilis Dookie Daisy f v 

Brachyscome muelleroides Mueller Daisy V f e 

Cullen tenax Tough Scurf-pea f e 

Discaria pubescens Australian Anchor Plant f r 

Dodonaea procumbens Trailing Hop-bush V v 

Eucalyptus aggregata Black Gum f e 

Eucalyptus alligatrix subsp. limaensis Lima Stringybark V f e 

Eucalyptus froggattii Kamarooka Mallee f r 

Euphrasia collina subsp. muelleri Purple Eyebright E f e 

Euphrasia scabra Rough Eyebright f e 

Geijera parviflora Wilga f e 



Glycine canescens Silky Glycine f e 

Glycine latrobeana Clover Glycine V f v 

Goodenia macbarronii Narrow Goodenia f v 

Grevillea floripendula Ben Major Grevillea V f v 

Hibbertia humifusa subsp. erigens Euroa Guinea-flower V f v 

Lepidium pseudopapillosum Erect Peppercress V f e 

Olearia pannosa subsp. cardiophylla Velvet Daisy-bush f v 

Philotheca difformis subsp. difformis Small-leaf Wax-flower f e 

Ptilotus erubescens Hairy Tails f 

Pultenaea graveolens Scented Bush-pea f v 

Pultenaea lapidosa Stony Bush-pea f v 

Santalum lanceolatum Northern Sandalwood f e 

Swainsona adenophylla Violet Swainson-pea f e 

Swainsona galegifolia Smooth Darling-pea f e 

Swainsona recta Mountain Swainson-pea E f e 

Swainsona sericea Silky Swainson-pea f v 

Swainsona swainsonioides Downy Swainson-pea f e 

Thesium australe Austral Toad-flax V f v 

Westringia crassifolia Whipstick Westringia E f e 

Zieria aspalathoides subsp. aspalathoides Whorled Zieria f v 

MONOCOTYLEDONS Common name EPBC FFG Vic 

Acianthus collinus Hooded Mosquito-orchid f v 

Caladenia audasii McIvor Spider-orchid E f e 

Caladenia cruciformis Red-cross Spider-orchid f e 

Caladenia fulva Tawny Spider-orchid E f e 

Caladenia ornata Ornate Pink-fingers V v 

Caladenia rosella Little Pink Spider-orchid E f e 

Caladenia sp. aff. fragrantissima (Central Bendigo Spider-orchid f e 

Victoria) Caladenia toxochila Bow-lip Spider-orchid f v 

Caladenia versicolor Candy Spider-orchid V f e 

Caladenia xanthochila Yellow-lip Spider-orchid E f e 

Calochilus richiae Bald-tip Beard-orchid E f e 

Dianella amoena Matted Flax-lily E e 

Diuris dendrobioides Wedge Diuris f e 

Diuris palustris Swamp Diuris f v 

Diuris punctata var. punctata Purple Diuris f v 

Diuris tricolor Painted Diuris f e 

Prasophyllum hygrophilum Swamp Leek-orchid f e 

Prasophyllum sp. aff. fitzgeraldii A Pink-lip Leek-orchid f e 

Prasophyllum subbisectum Pomonal Leek-orchid E f e 

Pterostylis despectans Lowly Greenhood E f e 

Pterostylis woollsii Long-tail Greenhood f e 

Thelymitra epipactoides Metallic Sun-orchid E f e 

Thelymitra mackibbinii Brilliant Sun-orchid V f e 



Table 5.4 Summary of listed rare and threatened species potentially affected by firewood 

harvesting in three key bioregions. 

Species that are Endangered or Vulnerable in Victoria, or listed under the Commonwealth EPBC Act, are not 

necessarily listed under Victoria's FFG Act, and vice versa. 

Category Australia Victoria 

Endangered 12 36 

Vulnerable 11 18 

Total EPBC-listed (Australia) 23 

Total FFG-listed (Victoria) 55* 


6 Knowledge gaps 


Our report confirms the findings of several relatively recent reviews on the ecological impacts of 

firewood harvesting at the national and state level (e.g. Australian and New Zealand Environment 

and Conservation Council 2001a; Driscoll et al. 2000; Grove et al. 2002; Lindenmayer et al. 2002) 

— that few studies have examined the direct impacts of firewood removal and harvesting on the 

diversity of flora and fauna and ecosystem processes, such as soil nutrient turnover. Notable 

exceptions include the investigations of saproxylic invertebrates at the Warra field site in 

Tasmania (e.g. Grove 2002b; Grove and Bashford 2003; Yee 2005; Yee et al. 2006) and the 

vertebrate and invertebrate work carried out in the Victorian River Red Gum forests in northern 

Victoria (e.g. Ballinger et al. 2003; Mac Nally 2006; Mac Nally et al. 2002a; Mac Nally and 

Horrocks 2008; Mac Nally et al. 2001). 

Most research has concentrated on the moist forests of eastern and south-eastern Australia where 

CWD production is higher, though the impacts on biodiversity and ecosystem processes are 

arguably less than those in woodlands — more research is required in these drier, less productive 

forests. There has been a tendency to utilise anecdotal observations and inferential evidence in the 

absence of empirical data and to conclude that particular taxa are likely to decline if this habitat 

resource was removed (Driscoll et al. 2000; Lindenmayer et al. 2002). Driscoll et al. (2000) 

summarised the major gaps in knowledge about the impacts of firewood harvesting in Australia, 

though their focus was broader than ours and included the extent (i.e. amount of firewood used, its 

geographical source and the tree species taken) of harvesting firewood across Australia. We 

suggest the following as key research areas because information for each is lacking, particularly in 

dry forests and woodlands: 

• The historical and current abundances of CWD, rates of accumulation and decay and 

sustainable rates at which to harvest it in different vegetation communities 

• The abundance of CWD required to conserve particular fauna species, particularly 

terrestrial taxa that utilise CWD 

• The features of CWD (e.g. decay stage, presence of hollows etc.) required to conserve 

particular wildlife species 

• The impacts of firewood removal on less-researched taxa, such as invertebrates, fungi, and 

cryptogams (algae, lichens, mosses, ferns) 

• The impact of fire on CWD decomposition is not well known. Some research suggests that 

charring slows down decomposition while other work suggests it increases it (Mackensen 

and Bauhus 1999) 

• The mechanisms that accelerate the development of firewood timber species and the 

characteristics deemed desirable for biodiversity (e.g. hollows) 

Little information exists on the contribution by CWD to vegetation structure and processes, and 

existing research focuses almost exclusively on wet forests. Research areas that should be 

addressed for drier forests include: 

• The inter-relationships between CWD, mycorrhizal fungi and understorey plant species 

• The role of CWD in providing microsites for seedling germination and survival 



• The contribution of CWD to weed establishment and abundance 

The physical disturbance associated with timber harvesting has been researched to a small degree 

in wetter, commercial forests, but little information exists on the effects of harvesting disturbance 

in drier forests. Research should determine: 

• The degree of soil compaction associated with firewood harvesting, and the implications 

for seedling germination and survival, and plant growth rates 

• Effects of soil compaction on water infiltration, run-off and erosion 

• Differential impacts of disturbance on reseeding and resprouting species 

The canopy formed by overstorey trees has a major impact on the conditions experienced by 

subordinate strata. Those conditions are likely to be altered, at least in the short-medium term, by 

the thinning associated with firewood harvesting, and this has implications in turn for fauna 

species. However, little research exists on the effects of thinning, either for commercial or 

ecological reasons, particularly in relation to drier forests. Areas for potential research include: 

• Effects of canopy thinning on understorey vegetation composition and structure, including 

weeds 

• Effects of canopy thinning on threatened species, such as winter-flowering orchids 

• Effects of canopy thinning on flowering and nectar production 

• Effects of canopy thinning on overstorey recruitment 

Nectar and pollen represent important resources, not only for recruitment and persistence of plant 

species, but also for fauna. However, the degree to which timber harvesting will affect these 

resources is largely unknown. Research should determine: 

• Peak flowering and sugar production times for all key firewood tree species 

• The effects of firewood harvesting on the volume and timing of nectar production, total 

nectar availability and quality 

• Pollination distances for key firewood species and the likely effects of decreases in mature 

tree density. This includes hybridisation rates, and changes in seed production and 

viability. 


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Appendix 1 



Vertebrate taxa for each of the select Victorian bioregions, compiled from the Atlas of Victorian Wildlife (DSE database), January 2009. 

Victorian (Vict Cons and FFG code) and national (EPBC) threatened status* are shown, along with use of CWD and hollow-bearing trees (B = basking, F = foraging, N = nesting, S = shelter). Genera are arranged alphabetically 

within Family and Families are arranged taxonomically within Order. Only extant non-vagrant Victorian native taxa are included. 

Murray 

Fans 

Victorian 

Riverina Goldfields 

Bioregion 

Central 

Victorian 

Uplands 

Northern 

Inland 

Slopes 

Highlands 

Northern 

Fall 

Highlands 

Southern 

Fall EPBC 

Cons. 

Vict. FFG Logs/CWD 

MAMMALS 

Ornithorhynchidae Platypus Ornithorhynchus anatinus √ √ √ √ √ √ √ 

Tachyglossidae Short-beaked Echidna Tachyglossus aculeatus √ √ √ √ √ √ √ 

Dasyuridae Agile Antechinus Antechinus agilis √ √ √ √ √ √ FNS FNS H 

Yellow-footed Antechinus Antechinus flavipes √ √ √ √ √ √ FNS FNS H 

Swamp Antechinus Antechinus minimus √ NT L 

Dusky Antechinus Antechinus swainsonii √ √ √ FNS 

Brush-tailed Phascogale Phascogale tapoatafa √ √ √ √ √ √ VU L FNS FNS H 

Spot-tailed Quoll Dasyurus maculatus √ √ √ √ √ √ EN EN L FNS FNS H 

Fat-tailed Dunnart Sminthopsis crassicaudata √ √ √ √ √ NT FNS 

White-footed Dunnart Sminthopsis leucopus √ √ NT L FNS 

Common Dunnart Sminthopsis murina √ √ √ √ √ VU FS 

Peramelidae Southern Brown Bandicoot Isoodon obesulus obesulus √ √ √ EN NT 

Eastern Barred Bandicoot Perameles gunnii √ EN CR L 

Long-nosed Bandicoot Perameles nasuta √ √ √ √ √ 

Phascolarctidae Koala Phascolarctos cinereus √ √ √ √ √ √ √ 

Vombatidae Common Wombat Vombatus ursinus √ √ √ √ √ √ √ 

Petauridae Leadbeater's Possum Gymnobelideus leadbeateri √ √ FNS H 

Yellow-bellied Glider Petaurus australis √ √ √ FNS H 

Sugar Glider Petaurus breviceps √ √ √ √ √ √ √ NS FNS H 

Squirrel Glider Petaurus norfolcensis √ √ √ √ √ EN L NS FNS H 

Pseudocheiridae Common Ringtail Possum Pseudocheirus peregrinus √ √ √ √ √ √ √ NS [H] 

Greater Glider Petauroides volans √ √ √ √ √ FNS H 

Acrobatidae Feathertail Glider Acrobates pygmaeus √ √ √ √ √ √ √ FNS FNS H 

Phalangeridae Mountain Brushtail Possum Trichosurus cunninghami √ √ √ √ √ NS FNS H 

Common Brushtail Possum Trichosurus vulpecula √ √ √ √ √ √ √ FNS FNS H 

Potoroidae Long-footed Potoroo Potorous longipes √ √ 

Macropodidae Western Grey Kangaroo Macropus fuliginosus √ √ √ √ 

Eastern Grey Kangaroo Macropus giganteus √ √ √ √ √ √ √ 

Eastern Wallaroo Macropus robustus robustus √ √ EN L 

Red-necked Wallaby Macropus rufogriseus √ √ √ 

Tammar Wallaby Macropus eugenii √ 

Brush-tailed Rock-wallaby Petrogale penicillata √ 

Black Wallaby Wallabia bicolor √ √ √ √ √ √ √ VU CR L 

Pteropodidae Grey-headed Flying-fox Pteropus poliocephalus √ √ √ √ VU VU L 

Little Red Flying-fox Pteropus scapulatus √ √ √ √ 

Rhinolophidae Eastern Horseshoe Bat Rhinolophus megaphyllus √ √ √ VU L 

Emballonuridae Yellow-bellied Sheathtail Bat Saccolaimus flaviventris √ L S H 

Molossidae Freetail Bat (eastern form) Mormopterus sp. EG √ √ √ √ √ √ SN H 

Southern Freetail Bat (long penis) Mormopterus sp. 1 √ √ √ √ √ √ √ SN H 

Hollowbearing 

trees 

Type of 

hollow^



White-striped Freetail Bat Tadarida australis √ √ √ √ √ √ √ SN H 

Vespertilionidae Gould's Wattled Bat Chalinolobus gouldii √ √ √ √ √ √ √ NS NS H 

Chocolate Wattled Bat Chalinolobus morio √ √ √ √ √ √ √ NS NS H 

Eastern False Pipistrelle Falsistrellus tasmaniensis √ √ √ √ √ NS NS H 

Common Bent-wing Bat Miniopterus schreibersii (group) √ √ √ √ √ L 

Southern Myotis Myotis macropus √ √ √ √ √ √ NT NS NS H 

Lesser Long-eared Bat Nyctophilus geoffroyi √ √ √ √ √ √ √ NS NS H 

Gould's Long-eared Bat Nyctophilus gouldi √ √ √ √ √ √ √ NS NS H 

Greater Long-eared Bat Nyctophilus timoriensis √ VU VU L NS NS H 

Inland Broad-nosed Bat Scotorepens balstoni √ √ √ √ √ NS NS H 

Eastern Broad-nosed Bat Scotorepens orion √ √ √ NS NS H 

Large Forest Bat Vespadelus darlingtoni √ √ √ √ √ √ √ NS NS H 

Southern Forest Bat Vespadelus regulus √ √ √ √ √ √ √ NS NS H 

Little Forest Bat Vespadelus vulturnus √ √ √ √ √ √ NS NS H 

Muridae Water Rat Hydromys chrysogaster √ √ √ √ √ √ √ 

Broad-toothed Rat Mastacomys fuscus √ √ DD 

Smoky Mouse Pseudomys fumeus √ √ EN CR L 

Bush Rat Rattus fuscipes √ √ √ √ √ √ 

Swamp Rat Rattus lutreolus √ √ √ √ 

Canidae Dingo Canis lupus dingo √ √ √ NT 

BIRDS 

Casuariidae Emu Dromaius novaehollandiae √ √ √ √ √ √ √ 

Megapodiidae Malleefowl Leipoa ocellata √ VU EN L 

Phasianidae Stubble Quail Coturnix pectoralis √ √ √ √ √ √ √ 

Brown Quail Coturnix ypsilophora √ √ √ √ √ √ √ NT 

King Quail Excalfactoria chinensis √ √ EN L 

Anseranatidae Magpie Goose Anseranas semipalmata √ √ √ √ NT L 

Anatidae Chestnut Teal Anas castanea √ √ √ √ √ √ √ N [H] 

Grey Teal Anas gracilis √ √ √ √ √ √ √ N [H] 

Australasian Shoveler Anas rhynchotis √ √ √ √ √ √ √ VU 

Pacific Black Duck Anas superciliosa √ √ √ √ √ √ √ N [H] 

Hardhead Aythya australis √ √ √ √ √ √ √ VU 

Musk Duck Biziura lobata √ √ √ √ √ √ √ VU 

Cape Barren Goose Cereopsis novaehollandiae √ √ NT 

Australian Wood Duck Chenonetta jubata √ √ √ √ √ √ √ N H 

Black Swan Cygnus atratus √ √ √ √ √ √ √ 

Plumed Whistling-Duck Dendrocygna eytoni √ √ √ √ √ √ 

Pink-eared Duck Malacorhynchus membranaceus √ √ √ √ √ √ √ N [H] 

Blue-billed Duck Oxyura australis √ √ √ √ √ √ √ EN L 

Freckled Duck Stictonetta naevosa √ √ √ √ √ √ EN L 

Australian Shelduck Tadorna tadornoides √ √ √ √ √ √ √ N N [H] 

Podicipedidae Great Crested Grebe Podiceps cristatus √ √ √ √ √ √ √ 

Hoary-headed Grebe Poliocephalus poliocephalus √ √ √ √ √ √ √ 

Australasian Grebe Tachybaptus novaehollandiae √ √ √ √ √ √ √ 

Columbidae White-headed Pigeon Columba leucomela √ √ 

Diamond Dove Geopelia cuneata √ √ √ √ √ √ NT L 

Bar-shouldered Dove Geopelia humeralis √ 

Peaceful Dove Geopelia striata √ √ √ √ √ √ √ 

Wonga Pigeon Leucosarcia melanoleuca √ √ √ √ √ 

Brown Cuckoo-Dove Macropygia amboinensis √ √ 

Crested Pigeon Ocyphaps lophotes √ √ √ √ √ √ √ 

Common Bronzewing Phaps chalcoptera √ √ √ √ √ √ √



Brush Bronzewing Phaps elegans √ √ √ √ √ 

Podargidae Tawny Frogmouth Podargus strigoides √ √ √ √ √ √ √ 

Eurostopodidae White-throated Nightjar Eurostopodus mystacalis √ √ √ √ √ √ 

Spotted Nightjar Eurostopodus argus √ √ √ √ √ 

Aegothelidae Australian Owlet-nightjar Aegotheles cristatus √ √ √ √ √ √ √ NS NS H 

Apodidae White-throated Needletail Hirundapus caudacutus √ √ √ √ √ √ √ 

Fork-tailed Swift Apus pacificus √ √ √ √ √ √ √ 

Anhingidae Darter Anhinga novaehollandiae √ √ √ √ √ √ √ 

Phalacrocoracidae Little Pied Cormorant Microcarbo melanoleucos √ √ √ √ √ √ √ 

Great Cormorant Phalacrocorax carbo √ √ √ √ √ √ √ 

Little Black Cormorant Phalacrocorax sulcirostris √ √ √ √ √ √ √ 

Pied Cormorant Phalacrocorax varius √ √ √ √ √ √ √ NT 

Pelecanidae Australian Pelican Pelecanus conspicillatus √ √ √ √ √ √ √ 

Ardeidae Cattle Egret Ardea ibis √ √ √ √ √ √ √ 

Intermediate Egret Ardea intermedia √ √ √ √ √ √ √ CR L 

Eastern Great Egret Ardea modesta √ √ √ √ √ √ √ VU L 

White-necked Heron Ardea pacifica √ √ √ √ √ √ √ 

Australasian Bittern Botaurus poiciloptilus √ √ √ √ √ √ √ EN L 

Little Egret Egretta garzetta √ √ √ √ √ √ EN L 

White-faced Heron Egretta novaehollandiae √ √ √ √ √ √ √ 

Australian Little Bittern Ixobrychus dubius √ √ √ √ √ EN L 

Nankeen Night Heron Nycticorax caledonicus √ √ √ √ √ √ √ NT 

Threskiornithidae Yellow-billed Spoonbill Platalea flavipes √ √ √ √ √ √ √ 

Royal Spoonbill Platalea regia √ √ √ √ √ √ √ VU 

Glossy Ibis Plegadis falcinellus √ √ √ √ √ √ √ NT 

Australian White Ibis Threskiornis molucca √ √ √ √ √ √ √ 

Straw-necked Ibis Threskiornis spinicollis √ √ √ √ √ √ √ 

Accipitridae Collared Sparrowhawk Accipiter cirrhocephalus √ √ √ √ √ √ √ 

Brown Goshawk Accipiter fasciatus √ √ √ √ √ √ √ 

Grey Goshawk Accipiter novaehollandiae √ √ √ √ √ √ VU L 

Wedge-tailed Eagle Aquila audax √ √ √ √ √ √ √ 

Swamp Harrier Circus approximans √ √ √ √ √ √ √ 

Spotted Harrier Circus assimilis √ √ √ √ √ √ NT 

Black-shouldered Kite Elanus axillaris √ √ √ √ √ √ √ 

Letter-winged Kite Elanus scriptus √ √ √ 

White-bellied Sea-Eagle Haliaeetus leucogaster √ √ √ √ √ √ √ VU L 

Whistling Kite Haliastur sphenurus √ √ √ √ √ √ √ 

Black-breasted Buzzard Hamirostra melanosternon √ √ √ 

Little Eagle Hieraaetus morphnoides √ √ √ √ √ √ √ 

Square-tailed Kite Lophoictinia isura √ √ √ √ √ √ √ VU L 

Black Kite Milvus migrans √ √ √ √ √ √ √ 

Eastern Osprey Pandion cristatus √ 

Falconidae Brown Falcon Falco berigora √ √ √ √ √ √ √ N L 

Nankeen Kestrel Falco cenchroides √ √ √ √ √ √ √ N L 

Grey Falcon Falco hypoleucos √ √ √ √ EN L 

Australian Hobby Falco longipennis √ √ √ √ √ √ √ 

Peregrine Falcon Falco peregrinus √ √ √ √ √ √ √ N L 

Black Falcon Falco subniger √ √ √ √ √ √ √ VU 

Gruidae Brolga Grus rubicunda √ √ √ √ √ √ VU L 

Rallidae Eurasian Coot Fulica atra √ √ √ √ √ √ √ 

Dusky Moorhen Gallinula tenebrosa √ √ √ √ √ √ √ 

Buff-banded Rail Gallirallus philippensis √ √ √ √ √ √ √ 

Lewin's Rail Lewinia pectoralis √ √ √ √ √ VU L



Purple Swamphen Porphyrio porphyrio √ √ √ √ √ √ √ 

Australian Spotted Crake Porzana fluminea √ √ √ √ √ √ 

Baillon's Crake Porzana pusilla √ √ √ √ √ √ VU L 

Spotless Crake Porzana tabuensis √ √ √ √ √ √ 

Black-tailed Native-hen Tribonyx ventralis √ √ √ √ √ √ √ 

Otididae Australian Bustard Ardeotis australis √ √ √ CR L 

Burhinidae Bush Stone-curlew Burhinus grallarius √ √ √ √ √ √ √ EN L FNS 

Recurvirostridae Banded Stilt Cladorhynchus leucocephalus √ √ √ 

Black-winged Stilt Himantopus himantopus √ √ √ √ √ √ 

Red-necked Avocet Recurvirostra novaehollandiae √ √ √ √ 

Charadriidae Inland Dotterel Charadrius australis √ √ VU 

Double-banded Plover Charadrius bicinctus √ √ √ √ 

Greater Sand Plover Charadrius leschenaultii √ VU 

Red-capped Plover Charadrius ruficapillus √ √ √ √ √ 

Oriental Plover Charadrius veredus √ 

Black-fronted Dotterel Elseyornis melanops √ √ √ √ √ √ √ 

Red-kneed Dotterel Erythrogonys cinctus √ √ √ √ √ √ √ 

Pacific Golden Plover Pluvialis fulva √ √ NT 

Masked Lapwing Vanellus miles √ √ √ √ √ √ √ 

Banded Lapwing Vanellus tricolor √ √ √ √ √ √ √ 

Pedionomidae Plains-wanderer Pedionomus torquatus √ √ √ √ √ √ VU CR L 

Rostratulidae Australian Painted Snipe Rostratula australis √ √ √ √ VU CR L 

Scolopacidae Common Sandpiper Actitis hypoleucos √ √ √ VU 

Ruddy Turnstone Arenaria interpres √ √ 

Sharp-tailed Sandpiper Calidris acuminata √ √ √ √ √ 

Red Knot Calidris canutus √ √ NT 

Curlew Sandpiper Calidris ferruginea √ √ 

Pectoral Sandpiper Calidris melanotos √ √ NT 

Little Stint Calidris minuta √ 

Red-necked Stint Calidris ruficollis √ √ √ √ √ 

Long-toed Stint Calidris subminuta √ NT 

Great Knot Calidris tenuirostris √ EN L 

Latham's Snipe Gallinago hardwickii √ √ √ √ √ √ √ NT 

Asian Dowitcher Limnodromus semipalmatus √ 

Bar-tailed Godwit Limosa lapponica √ √ √ 

Black-tailed Godwit Limosa limosa √ VU 

Eastern Curlew Numenius madagascariensis √ √ NT 

Little Curlew Numenius minutus √ 

Red-necked Phalarope Phalaropus lobatus √ √ 

Ruff Philomachus pugnax √ 

Wood Sandpiper Tringa glareola √ √ √ VU 

Common Greenshank Tringa nebularia √ √ √ √ √ 

Marsh Sandpiper Tringa stagnatilis √ √ √ √ √ 

Turnicidae Red-backed Button-quail Turnix maculosus √ 

Red-chested Button-quail Turnix pyrrhothorax √ √ √ VU L 

Painted Button-quail Turnix varius √ √ √ √ √ √ √ 

Little Button-quail Turnix velox √ √ √ √ √ √ NT 

Glareolidae Oriental Pratincole Glareola maldivarum √ 

Australian Pratincole Stiltia isabella √ √ NT 

Laridae Whiskered Tern Chlidonias hybridus √ √ √ √ √ √ NT 

White-winged Black Tern Chlidonias leucopterus √ √ NT 

Silver Gull Chroicocephalus novaehollandiae √ √ √ √ √ √ √ 

Gull-billed Tern Gelochelidon nilotica √ √ √ √ EN L



Caspian Tern Hydroprogne caspia √ √ √ √ √ √ NT L 

Cacatuidae Sulphur-crested Cockatoo Cacatua galerita √ √ √ √ √ √ √ N H 

Little Corella Cacatua sanguinea √ √ √ √ √ √ √ N H 

Long-billed Corella Cacatua tenuirostris √ √ √ √ √ √ √ N H 

Gang-gang Cockatoo Callocephalon fimbriatum √ √ √ √ √ √ √ N H 

Yellow-tailed Black-Cockatoo Calyptorhynchus funereus √ √ √ √ √ √ √ N H 

Glossy Black-Cockatoo Calyptorhynchus lathami √ VU L N H 

Galah Eolophus roseicapilla √ √ √ √ √ √ √ N H 

Major Mitchell's Cockatoo Lophocroa leadbeateri √ √ √ √ VU L N H 

Cockatiel Nymphicus hollandicus √ √ √ √ √ √ N H 

Psittacidae Australian King-Parrot Alisterus scapularis √ √ √ √ √ √ √ N H 

Australian Ringneck Barnardius zonarius zonarius √ √ √ √ √ √ √ N H 

Musk Lorikeet Glossopsitta concinna √ √ √ √ √ √ √ N H 

Purple-crowned Lorikeet Glossopsitta porphyrocephala √ √ √ √ √ √ √ N H 

Little Lorikeet Glossopsitta pusilla √ √ √ √ √ √ N H 

Swift Parrot Lathamus discolor √ √ √ √ √ √ EN EN L N + H 

Budgerigar Melopsittacus undulatus √ √ √ √ √ √ N H 

Blue-winged Parrot Neophema chrysostoma √ √ √ √ √ √ √ N H 

Elegant Parrot Neophema elegans √ √ √ VU N H 

Turquoise Parrot Neophema pulchella √ √ √ √ √ √ √ NT L N H 

Blue Bonnet Northiella haematogaster √ √ N H 

Pale-headed Rosella Platycercus adscitus √ √ √ N H 

Crimson Rosella Platycercus elegans elegans √ √ √ √ √ √ √ N H 

Eastern Rosella Platycercus eximius √ √ √ √ √ √ √ N H 

Regent Parrot Polytelis anthopeplus √ √ √ VU VU L N H 

Superb Parrot Polytelis swainsonii √ √ √ √ √ VU EN L N H 

Red-rumped Parrot Psephotus haematonotus √ √ √ √ √ √ √ N H 

Mulga Parrot Psephotus varius √ N H 

Scaly-breasted Lorikeet Trichoglossus chlorolepidotus √ √ N H 

Rainbow Lorikeet Trichoglossus haematodus √ √ √ √ √ √ N H 

Cuculidae Fan-tailed Cuckoo Cacomantis flabelliformis √ √ √ √ √ √ √ 

Pallid Cuckoo Cacomantis pallidus √ √ √ √ √ √ √ 

Brush Cuckoo Cacomantis variolosus √ √ √ √ √ √ √ 

Horsfield's Bronze-Cuckoo Chalcites basalis √ √ √ √ √ √ √ 

Shining Bronze-Cuckoo Chalcites lucidus √ √ √ √ √ √ √ 

Black-eared Cuckoo Chalcites osculans √ √ √ √ √ √ √ NT 

Strigidae Barking Owl Ninox connivens √ √ √ √ √ √ √ EN L N H 

Southern Boobook Ninox novaeseelandiae √ √ √ √ √ √ √ NS H 

Powerful Owl Ninox strenua √ √ √ √ √ √ √ VU L N H 

Tytonidae Eastern Barn Owl Tyto javanica √ √ √ √ √ √ NS H 

Masked Owl Tyto novaehollandiae √ √ √ √ √ EN L NS H 

Sooty Owl Tyto tenebricosa √ √ VU L NS H 

Alcedinidae Azure Kingfisher Ceyx azureus √ √ √ √ √ √ √ NT 

Halcyonidae Laughing Kookaburra Dacelo novaeguineae √ √ √ √ √ √ √ F FN H 

Red-backed Kingfisher Todiramphus pyrrhopygia √ √ √ √ √ NT 

Sacred Kingfisher Todiramphus sanctus √ √ √ √ √ √ √ N H 

Meropidae Rainbow Bee-eater Merops ornatus √ √ √ √ √ √ √ 

Coraciidae Dollarbird Eurystomus orientalis √ √ √ √ √ √ √ N H 

Menuridae Superb Lyrebird Menura novaehollandiae √ √ √ √ √ 

Climacteridae Red-browed Treecreeper Climacteris erythrops √ √ √ √ √ F FNS H 

Brown Treecreeper (south-eastern 

ssp.) Climacteris picumnus victoriae √ √ √ √ √ √ √ NT F FNS H 

White-throated Treecreeper Cormobates leucophaeus √ √ √ √ √ √ √ F FNS H 

Ptilonorhynchidae Satin Bowerbird Ptilonorhynchus violaceus √ √ √ √ √



Maluridae Superb Fairy-wren Malurus cyaneus √ √ √ √ √ √ √ 

Variegated Fairy-wren Malurus lamberti √ √ √ √ 

White-winged Fairy-wren Malurus leucopterus √ √ 

Splendid Fairy-wren Malurus splendens √ 

Southern Emu-wren Stipiturus malachurus √ 

Acanthizidae Inland Thornbill Acanthiza apicalis √ √ √ √ 

Yellow-rumped Thornbill Acanthiza chrysorrhoa √ √ √ √ √ √ √ 

Striated Thornbill Acanthiza lineata √ √ √ √ √ √ √ 

Yellow Thornbill Acanthiza nana √ √ √ √ √ √ √ 

Brown Thornbill Acanthiza pusilla √ √ √ √ √ √ √ 

Buff-rumped Thornbill Acanthiza reguloides √ √ √ √ √ √ √ N L 

Chestnut-rumped Thornbill Acanthiza uropygialis √ √ √ √ √ √ N H 

Southern Whiteface Aphelocephala leucopsis √ √ √ √ √ √ N H 

Rufous Fieldwren Calamanthus campestris √ NT 

Shy Heathwren Calamanthus cautus √ √ 

Striated Fieldwren Calamanthus fuliginosus √ √ √ 

Chestnut-rumped Heathwren Calamanthus pyrrhopygia √ √ √ √ √ VU L 

Speckled Warbler Chthonicola sagittata √ √ √ √ √ √ VU L FN 

White-throated Gerygone Gerygone albogularis √ √ √ √ √ √ 

Western Gerygone Gerygone fusca √ √ √ √ √ √ √ 

Brown Gerygone Gerygone mouki √ √ 

Pilotbird Pycnoptilus floccosus √ √ √ √ √ 

White-browed Scrubwren Sericornis frontalis √ √ √ √ √ √ √ 

Large-billed Scrubwren Sericornis magnirostris √ √ 

Weebill Smicrornis brevirostris √ √ √ √ √ √ √ 

Pardalotidae Spotted Pardalote Pardalotus punctatus √ √ √ √ √ √ √ N 

Striated Pardalote Pardalotus striatus √ √ √ √ √ √ √ N N H 

Meliphagidae Spiny-cheeked Honeyeater Acanthagenys rufogularis √ √ √ √ √ √ 

Eastern Spinebill Acanthorhynchus tenuirostris √ √ √ √ √ √ 

Red Wattlebird Anthochaera carunculata √ √ √ √ √ √ √ 

Little Wattlebird Anthochaera chrysoptera √ √ √ √ √ √ √ 

Regent Honeyeater Anthochaera phrygia √ √ √ √ √ √ EN CR L 

Pied Honeyeater Certhionyx variegatus √ 

Blue-faced Honeyeater Entomyzon cyanotis √ √ √ √ √ 

White-fronted Chat Epthianura albifrons √ √ √ √ √ √ √ 

Orange Chat Epthianura aurifrons √ √ 

Crimson Chat Epthianura tricolor √ √ 

Tawny-crowned Honeyeater Glyciphila melanops √ √ √ √ 

Painted Honeyeater Grantiella picta √ √ √ √ √ √ √ VU L 

Yellow-faced Honeyeater Lichenostomus chrysops √ √ √ √ √ √ 

Purple-gaped Honeyeater Lichenostomus cratitius √ √ VU 

Fuscous Honeyeater Lichenostomus fuscus √ √ √ √ √ √ √ 

White-eared Honeyeater Lichenostomus leucotis √ √ √ √ √ √ √ 

Yellow-tufted Honeyeater Lichenostomus melanops √ √ √ √ √ √ 

Yellow-plumed Honeyeater Lichenostomus ornatus √ √ √ √ √ 

White-plumed Honeyeater Lichenostomus penicillatus √ √ √ √ √ √ √ 

Singing Honeyeater Lichenostomus virescens √ √ √ √ √ √ √ 

Yellow-throated Miner Manorina flavigula √ √ 

Noisy Miner Manorina melanocephala √ √ √ √ √ √ √ 

Bell Miner Manorina melanophrys √ √ √ √ √ √ 

Lewin's Honeyeater Meliphaga lewinii √ √ 

Brown-headed Honeyeater Melithreptus brevirostris √ √ √ √ √ √ √ 

Black-chinned Honeyeater Melithreptus gularis √ √ √ √ √ √ √ NT



White-naped Honeyeater Melithreptus lunatus √ √ √ √ √ √ √ 

Scarlet Honeyeater Myzomela sanguinolenta √ √ √ 

Little Friarbird Philemon citreogularis √ √ √ √ √ √ √ 

Noisy Friarbird Philemon corniculatus √ √ √ √ √ √ √ 

New Holland Honeyeater Phylidonyris novaehollandiae √ √ √ √ √ √ √ 

Crescent Honeyeater Phylidonyris pyrrhopterus √ √ √ √ √ √ 

Striped Honeyeater Plectorhyncha lanceolata √ √ √ 

White-fronted Honeyeater Pumella albifrons √ √ √ √ √ √ 

Black Honeyeater Sugamel niger √ √ √ √ √ 

Pomatostomidae Chestnut-crowned Babbler Pomatostomus ruficeps √ 

White-browed Babbler Pomatostomus superciliosus √ √ √ √ √ √ 

Grey-crowned Babbler Pomatostomus temporalis √ √ √ √ √ √ √ EN L 

Eupetidae Spotted Quail-thrush Cinclosoma punctatum √ √ √ √ √ √ NT 

Eastern Whipbird Psophodes olivaceus √ √ √ √ √ 

Neosittidae Varied Sittella Daphoenositta chrysoptera √ √ √ √ √ √ √ 

Campephagidae Ground Cuckoo-shrike Coracina maxima √ √ VU L 

Black-faced Cuckoo-shrike Coracina novaehollandiae √ √ √ √ √ √ √ 

White-bellied Cuckoo-shrike Coracina papuensis √ √ √ √ √ √ √ 

Cicadabird Coracina tenuirostris √ √ √ √ √ 

White-winged Triller Lalage sueurii √ √ √ √ √ √ √ 

Pachycephalidae Grey Shrike-thrush Colluricincla harmonica √ √ √ √ √ √ √ 

Crested Shrike-tit Falcunculus frontatus √ √ √ √ √ √ √ 

Crested Bellbird Oreoica gutturalis √ √ √ √ NT L 

Gilbert's Whistler Pachycephala inornata √ √ √ √ √ 

Olive Whistler Pachycephala olivacea √ √ √ √ √ √ √ 

Golden Whistler Pachycephala pectoralis √ √ √ √ √ √ √ 

Rufous Whistler Pachycephala rufiventris √ √ √ √ √ √ √ 

Oriolidae Olive-backed Oriole Oriolus sagittatus √ √ √ √ √ √ √ 

Australasian Figbird Sphecotheres viridis √ 

Artamidae Black-faced Woodswallow Artamus cinereus √ √ √ √ √ N N L 

Dusky Woodswallow Artamus cyanopterus √ √ √ √ √ √ √ N N L 

White-breasted Woodswallow Artamus leucorynchus √ √ √ √ √ N N L 

Masked Woodswallow Artamus personatus √ √ √ √ √ √ N N L 

White-browed Woodswallow Artamus superciliosus √ √ √ √ √ √ √ N N L 

Pied Butcherbird Cracticus nigrogularis √ √ √ √ √ √ 

Australian Magpie Cracticus tibicen √ √ √ √ √ √ √ 

Grey Butcherbird Cracticus torquatus √ √ √ √ √ √ √ 

Pied Currawong Strepera graculina √ √ √ √ √ √ √ 

Grey Currawong Strepera versicolor √ √ √ √ √ √ √ 

Rhipiduridae Grey Fantail Rhipidura albiscarpa √ √ √ √ √ √ √ 

Willie Wagtail Rhipidura leucophrys √ √ √ √ √ √ √ 

Rufous Fantail Rhipidura rufifrons √ √ √ √ √ √ √ 

Corvidae Australian Raven Corvus coronoides √ √ √ √ √ √ √ 

Little Raven Corvus mellori √ √ √ √ √ √ √ 

Forest Raven Corvus tasmanicus √ 

Monarchidae Magpie-lark Grallina cyanoleuca √ √ √ √ √ √ √ 

Black-faced Monarch Monarcha melanopsis √ √ √ √ √ 

Satin Flycatcher Myiagra cyanoleuca √ √ √ √ √ √ √ 

Restless Flycatcher Myiagra inquieta √ √ √ √ √ √ √ 

Leaden Flycatcher Myiagra rubecula √ √ √ √ √ √ √ 

Corcoracidae White-winged Chough Corcorax melanorhamphos √ √ √ √ √ √ √ 

Apostlebird Struthidea cinerea √ √ L 

Petroicidae Southern Scrub-robin Drymodes brunneopygia √



Eastern Yellow Robin Eopsaltria australis √ √ √ √ √ √ F 

Hooded Robin Melanodryas cucullata √ √ √ √ √ √ √ NT L F 

Jacky Winter Microeca fascinans √ √ √ √ √ √ √ F 

Scarlet Robin Petroica boodang √ √ √ √ √ √ √ F 

Red-capped Robin Petroica goodenovii √ √ √ √ √ √ √ F 

Flame Robin Petroica phoenicea √ √ √ √ √ √ √ 

Pink Robin Petroica rodinogaster √ √ √ √ √ √ √ F 

Rose Robin Petroica rosea √ √ √ √ √ √ 

Alaudidae Horsfield's Bushlark Mirafra javanica √ √ √ √ √ √ √ 

Cisticolidae Golden-headed Cisticola Cisticola exilis √ √ √ √ √ √ √ 

Acrocephalidae Australian Reed Warbler Acrocephalus australis √ √ √ √ √ √ √ 

Megaluridae Brown Songlark Cincloramphus cruralis √ √ √ √ √ √ √ 

Rufous Songlark Cincloramphus mathewsi √ √ √ √ √ √ √ 

Little Grassbird Megalurus gramineus √ √ √ √ √ √ √ 

Timaliidae Silvereye Zosterops lateralis √ √ √ √ √ √ √ 

Hirundinidae White-backed Swallow Cheramoeca leucosterna √ √ √ √ √ 

Fairy Martin Hirundo ariel √ √ √ √ √ √ √ 

Welcome Swallow Hirundo neoxena √ √ √ √ √ √ √ N L 

Tree Martin Hirundo nigricans √ √ √ √ √ √ √ N H 

Turnidae Bassian Thrush Zoothera lunulata √ √ √ √ √ √ √ 

Nectariniidae Mistletoebird Dicaeum hirundinaceum √ √ √ √ √ √ √ 

Estrildidae Red-browed Finch Neochmia temporalis √ √ √ √ √ √ √ 

Beautiful firetail Stagonopleura bella √ √ √ 

Diamond Firetail Stagonopleura guttata √ √ √ √ √ √ √ VU L 

Double-barred Finch Taeniopygia bichenovii √ √ √ 

Zebra Finch Taeniopygia guttata √ √ √ √ √ √ 

Motacillidae Australasian Pipit Anthus novaeseelandiae √ √ √ √ √ √ √ 

REPTILES 

Cheluidae Common Long-necked Turtle Chelodina longicollis √ √ √ √ √ √ √ S 

Murray River Turtle Emydura macquarii √ √ √ √ √ √ DD L 

Broad-shelled Turtle Macrochelodina expansa √ √ √ √ √ EN L 

Agamidae Tree Dragon Amphibolurus muricatus √ √ √ √ √ √ BFNS S HL 

Eastern Water Dragon Physignathus lesueurii √ BS S HL 

Gippsland Water Dragon Physignathus lesueurii howittii √ √ BS S HL 

Bearded Dragon Pogona barbata √ √ √ √ √ √ DD BFS 

Mountain Dragon Rankinia diemensis √ √ √ BS 

Gekkonidae Marbled Gecko Christinus marmoratus √ √ √ √ √ √ √ FNS FS L 

Southern Spiny-tailed Gecko Diplodactylus intermedius √ √ NS FS L 

Tessellated Gecko Diplodactylus tessellatus √ √ √ NS 

Wood Gecko Diplodactylus vittatus √ √ √ √ √ √ NS 

Thick-tailed Gecko Nephrurus milii √ √ √ NS 

Pygopodidae Pink-tailed Worm-lizard Aprasia parapulchella √ VU EN L S 

Aprasia Aprasia sp. √ √ S 

Southern Legless Lizard Delma australis √ S 

Striped Legless Lizard Delma impar √ √ √ VU EN L S 

Olive Legless Lizard Delma inornata √ √ √ √ √ SN 

Burton's Snake-lizard Lialis burtonis √ √ √ S 

Common Scaly-foot Pygopus lepidopodus √ √ √ S 

Hooded Scaly-foot Pygopus schraderi √ CR L 

Scincidae Eastern Three-lined Skink Bassiana duperreyi √ √ √ √ √ BFNS 

Red-throated Skink Bassiana platynotum √ √ BFNS 

Bassiana Bassiana sp. √ √ √ √ BFNS 

Southern Rainbow Skink Carlia tetradactyla √ √ √ √ √ √ BFNS



Carnaby's Wall Skink Cryptoblepharus carnabyi √ √ √ √ BFNS FS L 

Eastern Striped Skink Ctenotus orientalis √ √ √ BFS 

Large Striped Skink Ctenotus robustus √ √ √ √ √ √ √ BFS S L 

Copper-tailed Skink Ctenotus taeniolatus √ √ √ √ √ BFS 

Swamp Skink Egernia coventryi √ √ VU L BF 

Cunningham's Skink Egernia cunninghami √ √ √ √ √ √ BF S HL 

Black Rock Skink Egernia saxatilis intermedia √ √ √ √ √ √ BFNS S HL 

Tree Skink Egernia striolata √ √ √ √ √ BFNS S HL 

White's Skink Egernia whitii (group) √ √ √ √ √ √ BFS 

White's Skink (plain back morph) Egernia whitii (plain back morph) √ BFS 

White's Skink (spotted back morph) Egernia whitii (spotted back morph) √ √ √ BFS 

Yellow-bellied Water Skink Eulamprus heatwolei √ √ √ √ √ √ BFNS 

Alpine Water Skink Eulamprus kosciuskoi √ CR L BFS 

Southern Water Skink Eulamprus tympanum tympanum √ √ √ √ BFNS 

Unidentified Water Skink Eulamprus sp. √ √ √ √ √ √ √ BFNS 

Three-toed Skink Hemiergis decresiensis √ √ √ √ √ √ FNS 

Delicate Skink Lampropholis delicata √ √ √ BFNS 

Garden Skink Lampropholis guichenoti √ √ √ √ √ √ √ BFNS 

Bougainville's Skink Lerista bougainvillii √ √ √ √ √ √ √ FNS 

Spotted Burrowing Skink Lerista punctatovittata √ FS 

Grey's Skink Menetia greyii √ √ √ √ √ BFNS 

Samphire Skink Morethia adelaidensis √ EN L BFNS 

Boulenger's Skink Morethia boulengeri √ √ √ √ √ √ BFNS 

McCoy's Skink Nannoscincus maccoyi √ √ √ √ √ FNS 

Coventry's Skink Niveoscincus coventryi √ √ √ √ √ BFNS 

Metallic Skink Niveoscincus metallicus √ √ BFNS 

Glossy Grass Skink Pseudechis rawlinsoni √ √ √ NT BFS 

Southern Grass Skink Pseudemoia entrecasteauxii √ √ √ BFNS 

Tussock Skink Pseudemoia pagenstecheri √ √ √ BFS 

Tussock Skink/Alpine Bog Skink Pseudemoia pagenstecheri/cryodroma √ √ BFS 

Spencer's Skink Pseudemoia spenceri √ √ √ BFNS BFS L 

Unidentified Grass Skink Pseudemoia sp. √ √ √ √ √ BF 

Weasel Skink Saproscincus mustelinus √ √ FNS 

Blotched Blue-tongued Lizard Tiliqua nigrolutea √ √ √ √ √ S 

Stumpy-tailed Lizard Tiliqua rugosa √ √ √ √ √ √ S 

Common Blue-tongued Lizard Tiliqua scincoides √ √ √ √ √ √ S 

Varanidae Sand Goanna Varanus gouldii √ √ √ √ √ BFNS BFNS HL 

Lace Goanna Varanus varius √ √ √ √ √ √ √ VU BFS 

Boidae Carpet Python Morelia spilota metcalfei √ √ √ EN L BFNS BFNS HL 

Typhlopidae Peters's Blind Snake Ramphotyphlops bituberculatus √ √ √ √ S 

Gray's Blind Snake Ramphotyphlops nigrescens √ √ √ √ √ √ S 

Woodland Blind Snake Ramphotyphlops proximus √ √ √ √ √ √ NT S 

Elapidae Highland Copperhead Austrelaps ramsayi √ BNS 

Lowland Copperhead Austrelaps superbus √ √ √ BNS 

White-lipped Snake Drysdalia coronoides √ √ √ √ NS 

Tiger Snake Notechis scutatus √ √ √ √ √ √ √ BNS FS HL 

Red-bellied Black Snake Pseudechis porphyriacus √ √ √ √ √ √ √ NS 

Eastern Brown Snake Pseudonaja textilis √ √ √ √ √ √ √ BNS 

Eastern Small-eyed Snake Rhinoplocephalus nigrescens √ √ √ √ √ √ NS 

Coral Snake Simoselaps australis √ NS 

Dwyer's Snake Suta dwyeri √ √ √ √ √ NS 

Little Whip Snake Suta flagellum √ √ √ √ √ NS 

Mitchell's Short-tailed Snake Suta nigriceps √ √ √ NS



Curl Snake Suta suta √ √ √ NS 

Bandy Bandy Vermicella annulata √ √ √ √ √ NT L NS 

FROGS 

Hylidae Booroolong Tree Frog Litoria booroolongensis √ CR L S 

Blue Mountains Tree Frog Litoria citropa √ S 

Southern Brown Tree Frog Litoria ewingii √ √ √ √ √ √ S 

Southern Brown Tree Frog Group Litoria ewingii (group) √ √ √ √ S 

Southern/Plains Brown Tree Frog Litoria ewingii/paraewingii √ √ √ √ √ S 

Lesueur's Frog Litoria lesueuri √ √ √ √ √ S 

Large Brown Tree Frog Litoria littlejohni √ VU NT L S 

Leaf Green Tree Frog Litoria nudidigita √ √ S 

Plains Brown Tree Frog Litoria paraewingi √ √ √ √ √ √ √ S 

Peron's Tree Frog Litoria peronii √ √ √ √ √ √ √ FS S L 

Growling Grass Frog Litoria raniformis √ √ √ √ √ √ √ VU EN L BS 

Spotted Tree Frog Litoria spenceri √ √ EN CR L S 

Verreaux's Tree Frog Litoria verreauxii √ √ √ S 

Alpine Tree Frog Litoria verreauxii alpina √ √ VU CR L S 

Whistling Tree Frog Litoria verreauxii verrreauxii √ √ √ S 

Myobatrachidae Plains Froglet Crinia parinsignifera √ √ √ √ √ √ √ S 

Common Froglet Crinia signifera √ √ √ √ √ √ √ S 

Sloane's Froglet Crinia sloanei √ √ √ √ S 

Victorian Smooth Froglet Geocrinia victoriana √ √ √ √ S 

Giant Burrowing Frog Heleioporus australiacus √ VU VU L S 

Southern Bullfrog Limnodynastes dumerilii √ √ √ √ √ √ √ S 

Southern Bullfrog (northern form) Limnodynastes dumerilii dumerilii √ √ √ √ √ S 

Southern Bullfrog (south-eastern form) Limnodynastes dumerilii insularis √ √ √ S 

Barking Marsh Frog Limnodynastes fletcheri √ √ √ √ S 

Giant Bullfrog Limnodynastes interioris √ √ √ CR L S 

Striped Marsh Frog Limnodynastes peronii √ √ √ √ √ S 

Spotted Marsh Frog Limnodynastes tasmaniensis √ √ √ √ √ √ √ S 

Spotted Marsh Frog (northern call 

race) Limnodynastes tasmaniensis NCR √ √ √ √ √ S 

Spotted Marsh Frog (southern call 

race) Limnodynastes tasmaniensis SCR √ S 

Mallee Spadefoot Toad Neobatrachus pictus √ √ S 

Common Spadefoot Toad Neobatrachus sudelli √ √ √ √ √ √ √ S 

Haswell's Froglet Paracrinia haswelli √ S 

Baw Baw Frog Philoria frosti √ EN CR L S 

Brown Toadlet Pseudophryne bibronii √ √ √ √ √ √ √ EN L S 

Dendy's Toadlet Pseudophryne dendyi √ √ √ DD S 

Southern Toadlet Pseudophryne semimarmorata √ √ √ S 

Smooth Toadlet Uperoleia laevigata √ √ √ √ DD S 

Rugose Toadlet Uperoleia rugosa √ √ VU L S

Appendix 2 


All vascular plant species (from forests and woodlands), in three bioregions subject to firewood 

harvesting, that have a rare and threatened status. 

EPBC (Australian Threatened) status: E Endangered; V Vulnerable. FFG status: f = listed. Victorian (Rare 

or Threatened) status: e endangered; v vulnerable; r rare; k poorly known but suspected to be r, v or e. 

DICOTYLEDONS Common name EPBC FFG Vic 

Acacia aspera subsp. parviceps Rough Wattle r 

Acacia ausfeldii Ausfeld's Wattle v 

Acacia deanei Deane's Wattle r 

Acacia deanei subsp. deanei Deane's wattle f e 

Acacia decora Western Silver Wattle v 

Acacia doratoxylon Currawang r 

Acacia euthycarpa subsp. oblanceolata Wedderburn Wattle v 

Acacia flexifolia Bent-leaf Wattle r 

Acacia omalophylla Yarran Wattle f e 

Acacia penninervis var. penninervis Hickory Wattle r 

Acacia sporadica Pale Hickory-wattle v 

Acacia verniciflua (southern variant) Southern Varnish Wattle k 

Allocasuarina luehmannii Buloke f 

Alternanthera sp. 1 (Plains) Plains Joyweed k 

Amaranthus macrocarpus var. macrocarpus Dwarf Amaranth v 

Amyema linophylla subsp. orientale Buloke Mistletoe v 

Asperula gemella Twin-leaf Bedstraw r 

Atriplex lindleyi subsp. lindleyi Flat-top Saltbush k 

Atriplex spinibractea Spiny-fruit Saltbush e 

Boronia anemonifolia subsp. aurifodina Goldfield Boronia r 

Boronia nana var. pubescens Dwarf Boronia r 

Bossiaea cordigera Wiry Bossiaea r 

Bossiaea riparia River Leafless Bossiaea r 

Brachyscome chrysoglossa Yellow-tongue Daisy f v 

Brachyscome cuneifolia Wedge-leaf Daisy k 

Brachyscome debilis s.s. Weak Daisy v 

Brachyscome gracilis Dookie Daisy f v 

Brachyscome muelleroides Mueller Daisy V f e 

Brachyscome readeri Reader's Daisy r 

Calotis cuneifolia Blue Burr-daisy r 

Calotis lappulacea Yellow Burr-daisy r 

Cardamine moirensis Riverina Bitter-cress r 

Cardamine papillata Forest Bitter-cress r 

Cassinia diminuta Dwarf Cassinia r 

Cassinia ozothamnoides Cottony Cassinia v 

Cassinia scabrida Rough Cassinia r 

Centipeda crateriformis subsp. compacta Compact Sneezeweed r 



Centipeda nidiformis Cotton Sneezeweed r 

Centipeda pleiocephala Tall Sneezeweed e 

Centipeda thespidioides s.l. Desert Sneezeweed r 

Chenopodium desertorum subsp. virosum Frosted Goosefoot k 

Choretrum glomeratum Common Sour-bush r 

Choretrum glomeratum var. chrysanthum Golden Sour-bush r 

Convolvulus angustissimus subsp. omnigracilis Slender Bindweed k 

Cullen tenax Tough Scurf-pea f e 

Cymbonotus lawsonianus Bear's-ear r 

Desmodium varians Slender Tick-trefoil k 

Discaria pubescens Australian Anchor Plant f r 

Dodonaea boroniifolia Hairy Hop-bush r 

Dodonaea heteromorpha Maple-fruited Hop-bush x 

Dodonaea procumbens Trailing Hop-bush V v 

Eremophila debilis Winter Apple e 

Eremophila divaricata subsp. divaricata Spreading Emu-bush r 

Eremophila gibbifolia Coccid Emu-bush r 

Eremophila maculata var. maculata Spotted Emu-bush r 

Eriochlamys sp. 1 Lesser Mantle v 

Eucalyptus aff. aromaphloia (Castlemaine) Fryers Range Scentbark e 

Eucalyptus aff. porosa (Quambatook) Quambatook Mallee-box e 

Eucalyptus aggregata Black Gum f e 

Eucalyptus alligatrix subsp. limaensis Lima Stringybark V f e 

Eucalyptus froggattii Kamarooka Mallee f r 

Eucalyptus polybractea Blue Mallee r 

Eucalyptus pyrenea Pyrenees Gum r 

Eucalyptus tricarpa subsp. decora Bealiba Ironbark v 

Euphrasia collina subsp. muelleri Purple Eyebright E f e 

Euphrasia collina subsp. speciosa Purple Eyebright x 

Euphrasia scabra Rough Eyebright f e 

Geijera parviflora Wilga f e 

Glycine canescens Silky Glycine f e 

Glycine latrobeana Clover Glycine V f v 

Goodenia benthamiana Small-leaf Goodenia r 

Goodenia lunata Stiff Goodenia v 

Goodenia macbarronii Narrow Goodenia f v 

Goodia medicaginea Western Golden-tip r 

Grevillea dimorpha Flame Grevillea r 

Grevillea dryophylla Goldfields Grevillea r 

Grevillea floripendula Ben Major Grevillea V f v 

Grevillea micrantha Small-flower Grevillea r 

Grevillea obtecta Fryerstown Grevillea r 

Grevillea polybractea Crimson Grevillea r 

Grevillea repens Creeping Grevillea r 

Haloragis glauca f. glauca Bluish Raspwort k 

Hibbertia humifusa Rising Star Guineaflower 

Hibbertia humifusa subsp. erigens Euroa Guinea-flower V f v 


r


Hibbertia humifusa subsp. humifusa Rising Star Guinea- 

r 

Hovea asperifolia subsp. spinosissima flower Rough Hovea r 

Indigofera adesmiifolia Tick Indigo v 

Lepidium pseudohyssopifolium Native Peppercress k 

Lepidium pseudopapillosum Erect Peppercress V f e 

Leptorhynchos elongatus Lanky Buttons e 

Leucochrysum molle Soft Sunray v 

Lotus australis var. australis Austral Trefoil k 

Myoporum montanum Waterbush r 

Olearia pannosa subsp. cardiophylla Velvet Daisy-bush f v 

Olearia tubuliflora Rayless Daisy-bush r 

Philotheca difformis subsp. difformis Small-leaf Wax-flower f e 

Pomaderris paniculosa subsp. paniculosa Inland Pomaderris v 

Prostanthera saxicola var. bracteolata Slender Mint-bush r 

Pseudanthus ovalifolius Oval-leaf Pseudanthus r 

Ptilotus erubescens Hairy Tails f 

Ptilotus sessilifolius var. sessilifolius Crimson Tails k 

Pultenaea foliolosa Small-leaf Bush-pea r 

Pultenaea graveolens Scented Bush-pea f v 

Pultenaea juniperina s.s. Prickly Beauty r 

Pultenaea lapidosa Stony Bush-pea f v 

Pultenaea platyphylla Flat-leaf Bush-pea r 

Pultenaea reflexifolia Wombat Bush-pea r 

Pultenaea vrolandii Cupped Bush-pea r 

Quinetia urvillei Quinetia r 

Rumex stenoglottis Tongue Dock k 

Santalum lanceolatum Northern Sandalwood f e 

Sida intricata Twiggy Sida v 

Stylidium calcaratum var. ecorne Foot Triggerplant k 

Swainsona adenophylla Violet Swainson-pea f e 

Swainsona behriana Southern Swainson-pea r 

Swainsona galegifolia Smooth Darling-pea f e 

Swainsona recta Mountain Swainson-pea E f e 

Swainsona sericea Silky Swainson-pea f v 

Swainsona swainsonioides Downy Swainson-pea f e 

Templetonia egena Round Templetonia v 

Tetragonia eremaea s.s. Desert Spinach k 

Teucrium albicaule Scurfy Germander k 

Thesium australe Austral Toad-flax V f v 

Vittadinia condyloides Club-hair New Holland Daisy r 

Vittadinia cuneata var. hirsuta Fuzzy New Holland 

r 

Vittadinia cuneata var. morrisii Daisy Fuzzy New Holland 

r 

Vittadinia pterochaeta Daisy Winged New Holland Daisy v 

Westringia crassifolia Whipstick Westringia E f e 

Zieria aspalathoides subsp. aspalathoides Whorled Zieria f v 



MONOCOTYLEDONS Common name EPBC FFG Vic 

Acianthus collinus Hooded Mosquito-orchid f v 

Aristida calycina var. calycina Dark Wire-grass r 

Austrodanthonia monticola Small-flower Wallaby- 

r 

Austrodanthonia setacea var. breviseta grass Short-bristle Wallaby-grass r 

Austrostipa breviglumis Cane Spear-grass r 

Austrostipa tenuifolia Long-awn Spear-grass v 

Austrostipa trichophylla Spear-grass r 

Caladenia audasii McIvor Spider-orchid E f e 

Caladenia clavescens Midlands Spider-orchid v 

Caladenia cruciformis Red-cross Spider-orchid f e 

Caladenia fulva Tawny Spider-orchid E f e 

Caladenia oenochila Wine-lipped Spider- 

v 

Caladenia ornata orchid Ornate Pink-fingers V v 

Caladenia reticulata s.s. Veined Spider-orchid v 

Caladenia rosella Little Pink Spider-orchid E f e 

Caladenia sp. aff. fragrantissima (Central Victoria) Bendigo Spider-orchid f e 

Caladenia toxochila Bow-lip Spider-orchid f v 

Caladenia versicolor Candy Spider-orchid V f e 

Caladenia xanthochila Yellow-lip Spider-orchid E f e 

Calochilus richiae Bald-tip Beard-orchid E f e 

Corunastylis ciliata Fringed Midge-orchid k 

Deyeuxia imbricata Bent-grass v 

Dianella amoena Matted Flax-lily E e 

Dianella sp. aff. longifolia (Riverina) Pale Flax-lily v 

Dianella tarda Late-flower Flax-lily v 

Dipodium pardalinum Spotted Hyacinth-orchid r 

Diuris behrii Golden Cowslips v 

Diuris dendrobioides Wedge Diuris f e 

Diuris palustris Swamp Diuris f v 

Diuris punctata var. punctata Purple Diuris f v 

Diuris tricolor Painted Diuris f e 

Diuris X palachila Broad-lip Diuris r 

Eragrostis alveiformis Granite Love-grass k 

Hypoxis vaginata var. brevistigmata Yellow Star k 

Juncus psammophilus Sand Rush r 

Prasophyllum aff. fitzgeraldii B Elfin Leek-orchid e 

Prasophyllum aff. pyriforme (Inglewood) Trim Leek-orchid e 

Prasophyllum hygrophilum Swamp Leek-orchid f e 

Prasophyllum lindleyanum Green Leek-orchid v 

Prasophyllum pyriforme s.s. Silurian Leek-orchid e 

Prasophyllum sp. aff. fitzgeraldii A Pink-lip Leek-orchid f e 

Prasophyllum sp. aff. validum A Woodland Leek-orchid e 

Prasophyllum subbisectum Pomonal Leek-orchid E f e 

Pterostylis aciculiformis Slender Ruddyhood k 



Pterostylis boormanii Sikh's Whiskers r 

Pterostylis despectans Lowly Greenhood E f e 

Pterostylis diminuta Crowded Greenhood k 

Pterostylis hamata Scaly Greenhood r 

Pterostylis maxima Large Rustyhood v 

Pterostylis smaragdyna Emerald-lip Greenhood r 

Pterostylis sp. aff. plumosa (Woodland) Woodland Plume-orchid r 

Pterostylis woollsii Long-tail Greenhood f e 

Thelymitra epipactoides Metallic Sun-orchid E f e 

Thelymitra mackibbinii Brilliant Sun-orchid V f e 

Thelymitra X chasmogama Globe-hood Sun-orchid v 

Thelymitra X macmillanii Crimson Sun-orchid v

Ecological impacts of firewood collection - Department of ...

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