Threats to Australian biodiversity

Assessment of

Australia’s Terrestrial Biodiversity

2008

Chapter 5 Threats to Australian biodiversity

These pages have been extracted from the full document which is available at: http://www.environment.gov.au/biodiversity/publications/terrestrial-assessment/index.html

© Commonwealth of Australia 2009

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Disclaimer The then National Land and Water Resources Audit’s Biodiversity Working Group had a major role in providing information and oversighting the preparation of this report. The views it contains are not necessarily those of the Commonwealth or of state and territory governments. The Commonwealth does not accept responsibility in respect of any information or advice given in relation to or as a consequence of anything contained herein.

Cover photographs: Perth sunset, aquatic ecologists Bendora Reservoir ACT, kangaroo paw: Andrew Tatnell. Ecologist at New Well SA: Mike Jensen

Editor: Biotext Pty Ltd and Department of the Environment, Water, Heritage and the Arts

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Chapter 5 Threats to Australian biodiversity

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The threats to Australia’s plants, animals and ecological systems are numerous and wide

ranging. The Australian landscape shows the legacy of past threats that are still actively

impacting on our biodiversity. New and emerging threats, particularly climate change

and water scarcity, are intensifying and will persist well into the future.

It is important to keep in mind that the direct drivers of biodiversity loss result from a

complex of interacting economic, socio-political, and scientific factors, which influence

human activities (Secretariat of the Convention of Biodiversity 2006).

Efforts to address some key threats have been scaled up. In particular, broad-scale land

clearing has been largely brought under control in the jurisdictions that accounted for

most of the clearing in 2002: New regulatory frameworks in Queensland and New South

Wales have dramatically reduced the level of approved clearing of remnant vegetation

nationally in the past five years.

Reforms to water management through recent Council of Australian Governments

(COAG) initiatives have the potential to address another key threat: altered hydrology.

The new reforms include mandatory consideration of environmental (aquatic ecosystem)

requirements for water in all new water allocation and planning.

The significance of climate change as a threat to biodiversity has become much more

widely recognised since 2002; climate change now ranks as an important threat to

Australian biodiversity overall. Knowledge of climate change scenarios and impacts is

rapidly growing as Australian scientists and science institutions increase efforts to

understand this threat.

The key findings from the Assessment of Australia’s Terrestrial Biodiversity 2008

(hereafter referred to as the ‘Assessment’) are listed below in section 5.1.

5.1 Key findings

Key threats are habitat fragmentation and the spread of invasive species.

A national analysis of the documentation and recovery plans for threatened species and communities listed under the Environment Protection and Biodiversity Conservation Act 1999 (EPBC Act) found that the most frequently cited threats are those of habitat fragmentation and the spread of invasive species

Weeds remain a threat to biodiversity but their impacts on biodiversity are not generally assessed.

Weed management strategies and policies have historically failed to address impacts on biodiversity adequately.

There are already a number of observed impacts from changes in the climate.

Observations of changes in native species and natural systems linked to climate change in Australia include: the expansion of rainforest at the expense of savanna; the encroachment by snow gums into sub-alpine grasslands at higher elevations; saltwater intrusion into freshwater swamps; and changes in bird behaviour including arrival of migratory birds, range shifts and sea-surface temperature induced reproductive changes.

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Selected case studies illustrate specific impacts of climate change.

A number of findings emerge from the selected case studies: impacts of threats are complex and variable; impacts are difficult to predict and generic modelling will require substantial localised input to be relevant; long-term monitoring is required.

Case studies of land use change show the threat posed to biodiversity by such change.

A high percentage of species are absent from cleared areas. Most species, however, can occur in regrowth (Queensland) or corridors of native vegetation (Northern Territory).

Understanding species’ requirements in terms of patch size and connectivity may allow the ‘design’ of landscapes with some clearing that retain vertebrate biodiversity.

Grazing pressure is a longstanding threat over much of the Australian landscape.

Almost 60 per cent of the Australian land mass is used for the production of livestock based on native pastures.

Strong linkages between artificial watering points and impacts on biodiversity can serve as an indicator of grazing pressure.

Information on monitoring fire distribution and frequency has improved greatly.

Fire frequency maps over the period 1997 to 2006 illustrate higher frequencies of fire in the central arid lands, northern savannas; a clear relationship with extended aridity; and frequent uncontrolled wildfires in southern Western Australia, south-east Victoria and coastal southern New South Wales.

Our knowledge of biodiversity responses to fire is still patchy.

Altered fire regimes affect biodiversity and interact with other threats in complex ways that are not yet fully understood.

5.2 Indicators

Indicators reported in this chapter are listed in Table 5.1.

Table 5.1 Indicators

Indicator Current reporting capacity rating

Trends in habitat fragmentation and decline in ecosystem function

Poor nationally

Moderate at case-study level

The range and relative importance of threats to biodiversity over time

Poor nationally

Moderate for listed species and communities

• Trends in the impacts of climate change on biodiversity

Poor nationally

Poor at case-study level

Trends in the impacts of land use change on biodiversity

• Trends in land clearing rates

Poor nationally

Good for clearing rates in Queensland


Trends in the impacts of invasive species and pathogens on biodiversity

• Extent and distribution of important invasive species

Poor nationally

Good at case-study level

• Trends in the impacts of grazing pressure on biodiversity

Poor nationally


• Trends in the impacts of altered fire regimes on biodiversity

Poor nationally

Good at case-study level

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5.3 The range and relative importance of threats to biodiversity

The key threats to biodiversity are:

• fragmentation

• climate change

• land use change

• invasive species and pathogens

• grazing pressure

• altered fire regimes, and

• changed hydrology.

These interacting threats vary in their impacts across the country and over time. A

national analysis of the documentation and recovery plans for threatened species and

ecological communities listed under the EPBC Act found that the most frequently cited

threats are those of habitat fragmentation and the spread of invasive species (Figure 5.1).

The nominations of threats for EPBC listings have changed. Before 2002, climate change

was rarely noted, but more recently it is identified as a threat for every new listed species

and ecological community.

Figure 5.1 Habitat fragmentation and invasive species pressure on EPBC listed species and communities

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5.4 The impacts of climate change on biodiversity

Evidence that climate change is causing global and regional warming is unequivocal

(Intergovernmental Panel on Climate Change 2007). Regional warming has been

associated with changes in physical and biological systems in many parts of the globe.

Australia has warmed by 0.9˚C since 1900 and is expected to warm a further 1˚C over

the next two decades (Olsen 2007).

There is significant uncertainty regarding how species and ecological systems will be

impacted by climate change. Current regional climate models suggest that impacts will

be widespread and that a ‘business as usual’ scenario over the next few decades will

result in global mass extinctions on a scale previously unseen in human history (IPCC

Working Group II 2007). There is mounting evidence that, even with concerted

mitigation effort, it may not be possible to avoid impacts of climate change such as the

loss of large components of biodiversity including freshwater systems, coral reefs and

coastal mangroves (Steffen 2008). The north Australian wetlands and the Great Barrier

Reef are among these threatened assets.

Recent studies show that impacts in Australia will be complex and highly variable

(CSIRO and Australian Bureau of Meteorology 2007a; Figure 5.2). The distribution,

diversity and abundance of species and the functioning and dynamics of ecosystems will

change, with some responding better than others. The most vulnerable species include

those with very restricted geographic and climatic range, those unlikely to migrate

successfully and/or those already highly compromised by small populations, fragmented

habitat and other threats. These include some of Australia’s most threatened and iconic

species (World Wildlife Fund 2008).

The threat of climate change include the direct impacts on habitat, ecosystem functioning

and populations of higher concentrations of carbon dioxide; altered rainfall and

temperature patterns; rising sea levels; increased sea temperatures and acidity; and more

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frequent extreme storms, floods and heatwaves. Many species are highly sensitive to

changes in climate and weather-related patterns and events. These can disrupt seasonal

food supplies and other resources, life cycle events, development, mortality, breeding

and fertility, such that entire reproductive strategies become less successful. Expected

direct impacts on species populations include:

• changes in species abundance

• changes in distribution, and

• changes in genetics over the long term as species adapt.

Climate change will compound other threats to biodiversity, including changed

hydrology, fire and invasive species. Warmer, drier conditions in southern Australia, in

particular, are predicted to lead to more frequent severe drought and wildfires. The

changing climate is also likely to favour invasive species in many areas and reduce the

competitiveness of Australian flora and fauna in their existing ranges. Migration of

native species into new and locally more favourable areas will also have implications for

extant populations.

Figure 5.2 Trends in temperature and rainfall

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The ability of species to adapt to changing conditions and recover after extreme climatic

events will be compromised by the legacy of fragmentation, habitat loss and other

pressures that have collectively reduced overall diversity, population sizes and resilience

in many species.

Less direct threats include the impacts of development shifts in response to changing

water availability. The north of Australia includes vast areas of relatively intact native

forest, woodlands and grasslands, and regions of rich biodiversity that may be at risk

from future development and changing land uses.

5.4.1 Observed impacts on natural systems and biota

The recent Intergovernmental Panel on Climate Change Working Group II assessment

collated available evidence and observations of changes in native species and natural

systems linked to climate change in Australia (Hennessy et al 2007) including:

expansion of rainforest at the expense of savanna (Bowman et al 2001, Hughes 2003);

encroachment by snow gums into sub-alpine grasslands at higher elevations (Wearne and

Morgan 2001); increased movement of feral mammals into alpine areas and prolonged

winter presence of macropods (Green and Pickering 2002); saltwater intrusion into

freshwater swamps possibly associated with sea level rise (Winn et al 2006); changes in

bird behaviour including in arrival of migratory birds, range shifts and sea-surface

temperature induced reproductive changes (Smithers et al 2003, Chambers 2005,

Chambers et al 2005, Beaumont et al 2006); change in genetic constitution of Drosophila

equivalent to a 4° latitude shift (Umina et al 2005).

5.4.2 Case studies of the impacts of climate change on biodiversity

Although many studies are in progress and numerous lines of evidence and observations

are emerging in relation to climate change impacts on biodiversity (Dunlop and Brown

2008, WWF 2008), many uncertainties remain, and it is not yet possible to provide a

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national picture. While there is no single indicator of the impact of climate change a

number of indicators already used in this Assessment can serve, e.g. trends in the

condition of native vegetation, and trends in the extent and distribution of threatened

species and communities. This is consistent with the approach taken internationally

(Secretariat of the Convention of Biodiversity, 2006). Trends in the extent and

distribution of wetlands may also be considered.

This Assessment commissioned a range of case studies to investigate specific impacts of

climate change in particularly vulnerable areas, and to illustrate some of the new

methodologies being developed to help us understand how climate change is likely to

impact on ecosystems, habitat, species and populations over time (Table 5.2).

Table 5.2 Case studies of climate change impacts on biodiversity

Case study Jurisdiction/agency

Will climate change affect Australian birds? National

Impacts of climate change on the biodiversity of the Victorian Alps Victoria

Sea level rise and biodiversity in the Coorong South Australia

Predicting climate change impacts on World Heritage rainforests in south-east Queensland

Queensland

Monitoring the impact of climate change on biodiversity in Tasmania

Tasmania

Community-level modelling of climate change impacts on biodiversity in New South Wales

NSW

Climate change and soil biodiversity CSIRO

The following findings from the seven selected case studies provide early signals about

the probable effects of climate change on biodiversity:

• Impacts of climate change on biodiversity will be complex and highly variable. For

example, changing phenology can change the competitive advantages of species

within a community and thus community composition.

• The impacts of climate change on biodiversity will be very difficult to predict, and

generic models will require substantial localised input in order to be relevant.

• Monitoring of biotic responses is critical to understanding the direction and speed of

the changes.

• Most current research deals with the direct impacts of climate conditions on plants

and animals. Very little research examines the more complex ecological interactions

likely to result from climate change.

• Responses to other variables, including CO2 concentrations, and the interactions

with other key threats are still unclear.

• The resilience of natural systems, species and populations to climate change is

largely unknown. Many taxa are already compromised by the ongoing effects of

other threats and are therefore highly vulnerable to shifts in climate-related

conditions.

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• Some species have greater inherent genetic adaptive capacity and/or inbuilt adaptive

strategies to an already variable climate that may assist them to adapt to the shifts

expected with climate change.

Case study 5.1 Will climate change affect Australian birds?

Predictions of the effects of climate change on Australian bird populations include alterations in

the abundance, breeding, migration and geographical range of many species (Chambers 2007).

Range shifts

The biophysical changes associated with warming will have a negative impact on bird species

that are highly specialised or have small populations, species that have a limited ability to adapt

and disperse (Chambers 2007) and, in particular, species whose distribution is restricted to high

altitudes and latitudes. Southwards shifts in bird populations, consistent with increasing

temperatures, are already being detected (Olsen 2007). Bird species whose distribution is

restricted to breeding islands that are low-lying, small and disconnected patches of habitat or

close to an affected food source that must shift will also suffer (Olsen 2007).

Altered hydrological patterns associated with rising temperatures will have serious consequences

for Australia’s wetland bird populations. Wetland birds that are dependent on freshwater habitats

under stress from drought conditions will be disadvantaged by climatic shifts towards lower

rainfall, as predicted for southern Australia. Wetland birds that are reliant on coastal habitats may

face a reduction in habitat range, due to potential saltwater inundations associated with rising sea

levels (Olsen 2007).

Adaptable species with broad climatic ranges or strong dispersal capabilities are likely to benefit

from the range shifts associated with warming.

Timing of migration and breeding

Strong evidence supports the relationship between climate change and changes in the timing of

migration of Australian migratory bird species (Chambers 2007). Changes in rainfall, shifts in

temperature and reduced snow cover affect the timing of migration of species such as the

nankeen kestrel (Falco cenchroides), the rainbow bee-eater (Merops ornatus) and the grey fantail

(Rhipidura fuliginosa) (Chambers 2007). These changes in migratory timing alter the duration

and timing of the reproductive period of bird populations. Changes in breeding patterns affect the

reproductive success and generational renewal of species.

Other potential impacts

Changes in agricultural production systems and coastal infrastructure will affect the habitat and

distribution of bird populations. A rise in global mean sea level—an expected consequence of

increasing global temperatures (Bennet et al 2007)—will impact on coastal and wetland habitats

through salt water inundations and coastal erosion, causing habitat loss, degradation and

fragmentation. It will also create new habitat for some coastal birds.

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Invasive species

Examples of weed species that will benefit from altered physical conditions include athel pine

(Tamarix aphylla) and gamba grass (Adropogon gayanus) (Low 2007). Such species are

advantaged by the severe weather events and increased fire frequency associated with climate

change. Athel pine outcompetes Australian river red gum (Eucalyptus camaldulensis) in periods

of severe flooding and does not provide the nesting hollows and nectar for birds that river red

gums provide (Low 2007). Gamba grass is a fire-promoting invasive species. Increased fire as a

result of climate change will have a degrading and fragmenting effect on bird habitat, which can,

in turn, promote the spread of invasive pest species. The proliferation of invasive plants causes

habitat loss, degradation and fragmentation not only for Australia’s birds, but for the species

upon which birds rely. The spread of invasive animals can cause declines in bird populations

through competition or predation.

Protecting birds from climate change

Fundamental to protecting Australia’s birds is reconnecting the natural landscape and habitats of

bird populations (Olsen 2007). Other environmental stressors, both those that result from and

those that contribute to climate change, must be reduced. These include, but are not limited to,

invasive pests, overgrazing and inadequate environmental flows (Olsen 2007). Emerging policies

on lowering greenhouse gas emissions should encourage retention of forest habitat, with

associated benefits for some bird populations, and emerging carbon markets offer opportunities

for protecting biodiversity through offsets. Ongoing monitoring, through scientific research and

volunteer contributions, will be essential in understanding the continued effect of climate change

on Australia’s birds and helping to protect them.

Australian pelican, Coorong National Park, South Australia

Photo: Paul Wainwright

Case study 5.2 Impacts of climate change on the biodiversity of the Victorian alps (DSE 2008a)

Nature of the threat

The Australian alpine region is on the mainland between the tree line and the permanent

snowline, which is now about 600 m above Mt Kosciuszko. Current alpine conditions (e.g. snow

cover) inhibit occupation by many low-altitude species (e.g. native grazers such as kangaroos)

and can change the composition of the vegetation. Declining depth and extent of snow cover

(Hennessy et al 2003) will also influence phenology (timing of ecological events). Under climate

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change, the increased days of high fire risk suggest that high-country fires, like those of 2003 and

2007, may become more frequent, again with ecosystem-modifying consequences (Hennessy et

al 2005).

The climate has been warming over the past few decades, with drought conditions over the last

several years. Biotic and abiotic responses consistent with warming have already been observed.

These include reduced snow cover, ice melt in lakes (K Green, Associate Professor, Griffith

School of Environment, Griffith University, pers. comm), eucalypt recruitment above the tree

line, altitudinal expansion of range (rabbits) and the timing of arrival of migratory species.

Higher temperature, elevated CO2 and changing rainfall patterns are a direct threat to the

biodiversity of almost all ecosystems. In alpine ecosystems, small changes, particularly to

temperature and the distribution and persistence of snow, can have large and long-lasting effects

on species and ecosystem processes (Williams et al 2006).

The current competitive advantage resulting from adaptations of the alpine species may

‘dissolve’ under new conditions that become more favourable for the surrounding lower altitude

biota. Conversely, some species may have the genetic capacity to adapt ‘on site’ and persist.

When does adaptation in one native species begin to threaten another?

Trends in biodiversity

Bioclimatic modelling has been undertaken since the early 1990s at the Arthur Rylah Institute for

Environmental Research for a range of flora and fauna of south-eastern Australia. This modelling

uses BIOCLIM/ANUCLIM to examine species’ bioclimatic envelopes under various future

climate scenarios (Bennett et al 1991, Brereton et al 1995, Newell et al 2001).

All alpine species exhibited shrinkage of their bioclimatic envelope. That of the mountain pygmy

possum (Burramys parvus) disappeared under 1°C of warming (Brereton et al 1995), whereas

that of the silky snow daisy (Celmisia serciophylla) disappeared by 2050 under all four climate

scenarios examined (Newell et al 2001). Using a spatial modelling technique, the predicted

distributions of two subalpine plant species (Pimelea alpina and Senecio pinnatifolius) were

reduced with every 1°C increase in average temperature compared with current distributions,

until local extinction of one of the species at an increase of approximately 5°C (Figure 5.3; F

Jarrad and J Elith, University of Melbourne, pers. comm).

Microhabitats may have microclimates where temperatures are different from the ambient air

temperature used in the models. However, these models provide further evidence for the

vulnerability of the alpine species. Actual future persistence of species may be determined not

purely by the abiotic variables (e.g. climate) but also by biotic interactions (e.g. competition) as

the ecosystem adapts to new conditions.

In 2003, a project was set up on the Bogong High Plains to study the likely consequences of a

warmer climate on species and plant communities of the Australian alps. The project, part of a

global investigation, consists of several integrated studies divided among experimental ecology,

genetics and ecological modelling. The main experimental design is based on the field protocol

of the International Tundra Experiment (ITEX), which has been conducted at northern

hemisphere sites for more than 15 years (http://www.geog.ubc.ca/itex/index.php). ITEX studies

and researchers make a significant contribution to the reports on global warming of the

Intergovernmental Panel on Climate Change. The experiment on the Bogong High Plains is part

of the global investigation. Small (1–2 m2) open-topped chambers passively warm ambient

temperatures within chambers by about 1.5°C to locally mimic warming. Two of the sites were

burnt by the extensive fires of 2003, and two were unburnt. These sites are the first functional

ITEX sites in the southern hemisphere, and the experimental design is the first to address the

interaction between fire and climate change in an alpine ecosystem.

Now in its fourth year, the warming experiment is beginning to yield results. Plants vary in their

response to both warming and fire. Long-term anecdotal evidence suggests that some species

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(e.g. Poa hiemata) have begun to flower earlier than usual and others later, while some have

been affected more by fire than warming. These studies complement and quantify other alpine

observations and post-fire studies. Anecdotal evidence from long-term records shows that in the

1990s usual Poa ripening was mid–late February, whereas in 2008 ripe seed was harvested two

weeks earlier than in previous years. Such changing phenology can change the competitive

advantages of species within a community, and thus community composition.

Links to bigger picture

The tundra/alpine biomes are highly vulnerable globally. Some studies (ITEX) provide key data

concerning the changing global climate (and carbon sequestration or liberation). The Australian

alpine region is a vulnerable and restricted environment of international significance.

Partnerships with international environmental research (i.e. ITEX) provide invaluable synergies

and understandings.

Figure 5.3 Predicted distribution of Pimelea alpina in the Victorian alps with increasing temperature

C lim ate + 2 C C lim ate + 1 C C ur re nt C lim ate

P robab ili ty o f occurre nce

H igh 8 5 %

Lo w : 1 %

!

!

!

!!

M t N el se

M t Hotham

M t Bogong

D inn er P l a in

M t Feathe r top

!

!

!

!!

M t N else

M t Hoth am

M t Bogong

D in n er P l a in

M t Feather top

!

!

!

!!

M t N els e

M t H otham

M t Bogo n g

Dinner P la in

M t Feather top

Source: F Jarrad and J Elith, University of Melbourne (unpublished)

Case study 5.3 Sea level rise and biodiversity in the Coorong (Seaman 2008)


The Coorong and Lower Murray Lakes—Alexandrina and Albert—form part of the Murray-

Darling Basin. They constitute a complex estuary system that empties through a relatively small

river mouth in the coastal dune barrier of the Younghusband and Sir Richard peninsulas (DEH

2000; Figure 5.4). The Coorong is an important breeding area for many species of birds.

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Figure 5.4 Location of the Coorong and Lower Lakes

In 1985, the Coorong and Lower Lakes were designated as a Wetland of International

Importance under the Ramsar Convention on Wetlands. The Coorong and Lakes Alexandrina

and Albert Ramsar Site (CLAARS) has a unique mosaic of wetlands, and provides habitat for

many waterbird species and a number of nationally threatened plant and animal species (DEH

2000).

The Coorong is about 140 km long and up to 4 km wide. It contains habitats that range from

seasonally fresh near the barrages (located in the channels linking Lake Alexandrina to the

Coorong) when large quantities of water are being released, to brackish in the Murray Mouth

area, and hypersaline in the South Lagoon (DEH 2000). Fresh water impounded in Lakes

Alexandrina and Albert by the barrages maintains a variety of permanent and ephemeral

wetlands (DEH 2000).


The overall effect of a rise in sea level on landforms and ecological systems of CLAARS is

likely to be negative. Possible consequences include:

• loss of vegetation and consequent erosion of the beach dune system of Younghusband and

Sir Richard peninsulas

• threat to the habitats of the freshwater soaks from degradation of the beach dune complex

• threat of inundation to the island habitats in the Coorong lagoons

• increased inundation, possibly benefiting Bryozoan communities, which are common on the

eastern side of the North Lagoon

• possible benefit to Samphire Flat Shrubland around the South Lagoon due to increased

inundation

• destruction of the freshwater-dependent woodland near the South Lagoon as a result of

inundation

• increased tide and wave attack on the Murray Mouth, either opening up the mouth or

depositing more sediment into the floodtide delta system

• erosion and permanent inundation of the barrages, rendering the barrage system ineffective

and resulting in a change in water chemistry of the lakes from fresh to estuarine

• reduction or even removal of mudflat habitat as a result of inundation, and

• threat to the ecosystems of the carbonate lakes from alteration of their water regime.

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Fairy Tern (Sterna nereis)

Photo: Paul Wainwright

Case study 5.4 Predicting climate change impacts on World Heritage rainforests in south-east Queensland (Butler and Accad 2008)


Bioclimatic modelling suggests that climate change might substantially alter the balance between

three rainforest regional ecosystems (Table 5.3) that form an altitudinal sequence in Lamington

National Park, one of Australia’s biodiversity icons and part of the World Heritage-listed Central

Eastern Rainforest Reserves of Australia. Rising temperatures are predicted to greatly reduce the

extent of habitat suited to both high- and low-altitude rainforest types, rather than simply driving

the current sequence toward higher altitudes.

The study uses two climate change scenarios based on CSIRO predictions for 2030 (Hennessy et

al 2006): a ‘low change scenario’, where annual mean temperature rises by 0.6°C relative to

1990 and mean annual rainfall declines by 1.5 per cent, and a ‘high change scenario’, where

annual mean temperature rises by 1.3°C and mean annual rainfall declines by 3 per cent.


The results suggest that projected climate change will substantially re-arrange the distribution of

environments suited to each of the rainforest ecosystems (Figure 5.5). Not surprisingly,

environmental conditions suited to the mid-altitude regional ecosystem (RE) RE12.8.3 are

predicted to expand uphill at the expense of the high-altitude RE12.8.5. RE12.8.3’s lower

altitudinal limit is predicted to change little. As a result, the total extent of conditions most suited

to RE12.8.3 is predicted to increase by as much as 50 per cent. The hoop pine (Araucaria

cunninghamii) rainforests that currently occupy the lowest altitudes (RE12.8.4) are predicted to

suffer the greatest reduction in extent, even though their current low altitude might suggest they

have the most room to move. Rather than moving uphill as temperature increases, the

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environmental envelope currently occupied by RE12.8.4 is predicted to diminish drastically

under the high-change scenario to as little as 2 per cent of its current extent.

Table 5.3 Predicted change in regional ecosystem extent under different climate change scenarios

Regional ecosystem

Description Predicted change %

Low scenario

High scenario

12.8.3 Complex notophyll vine forest on Cainozoic igneous rocks. Altitude <600 m.

+30 +50

12.8.4 Complex notophyll vine forest with Araucaria spp. on Cainozoic igneous rocks.

–60 –98

12.8.5 Complex notophyll vine forest on Cainozoic igneous rocks. Altitude usually >600 m.

–50 –90

Rising temperatures are likely to substantially affect the rainforests of Lamington National Park.

Although the complexity of the ecological systems involved raises questions about the RE

modelling, results indicate that increased temperatures may not simply pressure ecosystems to

move to higher altitudes. Instead, both low- and high-altitude forest types are likely to come

under serious pressure as their respective suited environments contract.

The remarkable biodiversity of Lamington National Park has received considerable attention

from biologists over many years, and their work provides an excellent baseline for ongoing

monitoring. Recently, a major international project known as IBISCA (IBISCA 2008), focusing

on the diversity and distribution of forest arthropods, established a network of plots across the

altitudinal range of Lamington National Park. Surveys have assessed a very wide range of

organisms and represent an outstanding resource to predict and monitor the impact of climate

change on biodiversity.

Figure 5.5 Mapped model results for three regional ecosystems in Lamington National Park

Source: Van Strein et al (2007)

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Case study 5.5 Monitoring the impact of climate change on biodiversity in Tasmania (Quinn and d’Arville 2008a)


Recent modelling has predicted that the effects of future climate change will include a 10 per

cent increase in annual rainfall in south-west Tasmania and a 10–20 per cent decrease in rainfall

in northeast Tasmania by 2040 (McIntosh et al 2005). The temperature is predicted to increase

by 0.5°C by 2040 and the annual potential evaporation to increase by 10 per cent in the midlands

and across the east coast. This climate scenario will have flow-on effects in terms of increased

fire risk and changes to water and nutrient availability, and may significantly alter the

distributions of many native species.

Experiments have shown clear impacts of a simulated changed climate on organisms’

physiology, anatomy and morphology. Field experiments and monitoring are vital to understand

the possible impacts of climate change, as natural processes are virtually impossible to replicate

in a laboratory or glasshouse. Establishing long-term monitoring programs will be an important

element in understanding the impact of climate change on biodiversity in Tasmania.


Tasmania is currently undertaking two major programs to investigate the impact of climate

change on biodiversity. The first is a long-term monitoring program established in 1999–2000 at

Mount Weld in south-west Tasmania as a Satellite Project for the International Biodiversity

Observation Year. In collaboration with Forestry Tasmania and the University of Tasmania, the

Department of Primary Industries and Water aims to monitor distributional change in vegetation

and faunal assemblages along an altitudinal gradient in response to climate change and other

environmental events (Doran et al 2003). The location within the Warra Long Term Ecological

Research site will allow monitoring over many years. Although the initial data have been

valuable, their true value lies in their ability to be used in future comparative studies. The site has

been developed so that it can be returned to at any time for comparisons to be made.

In the second program, the University of Tasmania has established a state-of-the-art Free-Air

CO2 Enrichment facility (TasFACE) for investigating the impact of global climate change on

native grassland in south-eastern Tasmania. The facility is located at a site that supports

approximately 60 vascular plant species, including six species listed in the Tasmanian

Threatened Species Protection Act 1995. This is a unique project for Australia and is one of very

few experiments globally that is investigating climate change impacts in natural ecosystems and

combining elevated CO2 and temperature treatments. As CO2 is the substrate for photosynthesis,

an increase in its concentration in the atmosphere will have a direct impact on many plant species

(Hovenden et al 2006). The project aims to investigate the effects of elevated CO2 and warming

on the physiology, growth, reproduction and recruitment of plant species, as well as on species

composition and interactions, and on pasture productivity and nutritional quality. The project has

been operating since 2002.

Plant response to simulated climatic conditions was varied and included some unexpected results

for growth rates under the predicted 2050 conditions. Photosynthetic theory suggests that plants

that have the ability to concentrate CO2 in their cells, and so buffer external changes in CO2, will

not show increased growth under elevated CO2 conditions, whereas plants that use an alternate

path lacking this buffering ability will (Wand et al 1999). Unexpectedly, the two dominant grass

species at the TasFACE site were found to vary in their response to the simulated predicted

climatic conditions of 2050. The Austrodanthonia caespitose (wallaby grass) showed reduced

seed production under increased CO2; when increased CO2 was combined with increased

temperature, it showed reduced seed germination and establishment (Williams et al 2007).

Themeda triandra (kangaroo grass) was not significantly affected by CO2 or warming; a slight

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increase in population growth (as defined by the matrix model approach used) was observed

(Williams et al 2007). Changes in the growth and physiological processes of these species

occurred during this trial, altering their competitiveness (Williams et al 2007). The unexpected

results highlight the need to investigate the interactive effects of CO2 and warming on individual

species’ growth and reproduction, and on community structure.

The threat of invasive species establishment is also being investigated at the TasFACE site. It is

generally thought that invasive species will be favoured in a warmer, higher-CO2 environment

(Patterson 1995). However, population growth in the two dominant herbaceous weed species

(Hypochaeris radicata, cats ear or flatweed, and Leontodon taraxacoides, hairy hawkbit) was

substantially reduced by warming and not significantly affected by elevated CO2 (Williams et al

2007). Further ecological research is investigating changes in soil microbes, soil nitrogen and

flowering times under the simulated conditions.

Links to the bigger picture

Given the immense importance of grasslands in south-eastern Australia, in terms of conservation

significance and as an agricultural resource, the outcomes of experiments conducted at the

TasFACE site will be critical in planning for the future management of grassland communities

and the threatened species within them. Potential changes in species composition and the

distribution of invasive species under a changed climatic regime will exacerbate effects on an

ecosystem already under pressure from current land practices.

Case study 5.6 Community-level modelling of climate change impacts on biodiversity in New South Wales (DECCW 2008a)


Australia’s biodiversity is already under significant stress from the impacts of 200 years of

European colonisation. Clearing, fragmentation, land use change, changes to water flows,

introductions of invasive species, and changes to fire regimes have already caused extinctions

and are continuing to threaten the persistence of biodiversity in Australia. Climate change is an

additional stressor.

Land use planners and natural resource managers throughout New South Wales are increasingly

expected to include considerations of the impacts of climate change on biodiversity in making

decisions. The scale of decision-making ranges from large-scale conservation planning processes

(eg the Great Eastern Ranges Initiative (formerly the Alps to Atherton Connectivity Conservation

Initiative)) to local processes (e.g. delivery of incentives by catchment management authorities).

To support such planning activities, the National Biodiversity and Climate Change Action Plan

2004–2007 (Natural Resource Management Ministerial Council 2004) identified the need for

ongoing action to ‘improve capacity of models to predict climate change impacts on biodiversity’

(Action 1.2.1).


The aim of the community-level modelling project is to explore the existing relationship between

biodiversity and climate by developing a continuous community compositional turnover surface

for New South Wales and relating this to available climate variables. The model is then used to

explore the potential impact on biodiversity of changes in climate.

Investigations into the potential impact of climate change on biodiversity have so far focused on

modelling distributional shifts of either individual species or community types. Both these

approaches have strengths and weaknesses. An individual species approach is computationally

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intensive, and methods are still in the development stage. Modelling with community types

whose species interrelationships are unclear can underestimate the response of biodiversity to

climate. However, this community-level approach works with spatial changes in community

composition, employing analytical techniques pioneered by Simon Ferrier (Department of

Environment and Climate Change—DECCW) and Dan Faith (Australian Museum). The

potential advantages include more rapid assessments at relatively fine resolutions, where the

assessments do not assume that species will move together, and where the results are not

sensitive to the high variability in species responses.

A trial of this approach used vegetation census data contained in the YETI database, soil

variables and key soil properties from the Bureau of Rural Sciences Australian Natural

Resources Data Library, and the existing DECCW corporate grid layers for mean annual

temperature and mean annual rainfall. Change was shown through two future climate scenarios

based on the Climate Change in Australia: Technical Report 2007 (CSIRO and Australian

Bureau of Meteorology 2007b).

Climate and soil variables used in the generalised dissimilarity model (a statistical technique for

analysing and predicting spatial patterns of turnover in community composition across large

regions) explained approximately 30.9 per cent of the variation in vegetation, only 3.4 per cent of

which was explained by the soil variables. These initial results suggest that increases in

temperature and decreases in rainfall may result in changes in biodiversity that are greatest in the

warmer and drier areas of the state. The greater distances plant species may need to disperse in

order to take advantage of new opportunities offered by changed environments could exacerbate

such pressures. This result highlights the need for impact assessments to identify the locations of

greatest pressure, so that monitoring and amelioration programs can return the greatest benefit

for investment.

5.5 Land use change

National modelling and mapping of land use indicates the overall extent of change since

European settlement (Figure 5.6). Although there have been substantial advances in

controlling large-scale modification of landscapes, particularly in highly vulnerable,

threatened ecosystems or high-value conservation areas, there are numerous instances of

smaller scale changes that cumulatively have a significant impact on biodiversity. For

example, a recent trend away from livestock to cropping across the wheat–sheep belt of

southern Australia, stretching north of the Queensland border, has resulted in small-scale

clearing of remnant vegetation and altered drainage on many properties. These remnants

are important refuges and provide linkages across highly modified landscapes for many

threatened species, including woodland birds and mammals. The cumulative impact of

their loss is unknown.

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Figure 5.6 Land use by major use categories, 2001–2002

5.5.1 Trends in land clearing rates

Outcomes of land use change for biodiversity are strongly linked to retention of healthy,

functioning native vegetation communities. Rates of land clearing are falling nationally

as broad-scale clearing regulations are tightened and implemented in key jurisdictions

(Griffin 2004, Table 5.4).

However, there is an ongoing but poorly quantified threat from approved clearing of

small remnants, clearing for approved new developments and subdivisions, and illegal

clearing.

‘…The escalating rate of clearing and other broad-scale environmental modification is

likely to increase the rate of fauna change, as dependent woodland species continue to

decline and be lost across the landscape, and be replaced by those more commensal

species favoured by landscapes sculpted for human use.’ Woinarski et al 2006

Table 5.4 Approximate clearing rates in Australian jurisdictions

Jurisdiction Annual clearing rate (ha) Trend

Qld >100 000 Falling due to implementation of the native vegetation regulations

NSW 10 000–100 000

Falling due to implementation of the native vegetation regulations

Tas 10 000–100 000

Stable

NT >10 000 Rising due to clearing for land development

WA <10 000 Stable

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Vic <10 000

Falling due to implementation of the native vegetation regulatory framework

SA <10 000

Stable

ACT <10 000

Stable

National >100 000 Falling

The figures in Table 5.4 are derived from a range of sources, including government

reports and vegetation surveys. For most jurisdictions, the estimates are determined from

data on clearing permit approvals, supplemented where possible by vegetation surveys.

Comprehensive vegetation survey data that can show trends are only available for

Queensland (a baseline and first estimate of overall change in extent and quality is

available for woody and grassy vegetation for Victoria).

The figures are approximate because not all jurisdictions provide regular data on rates of

clearing; some provide data only for certain tenures, vegetation types and periods

(Productivity Commission 2004). For instance, data on clearing in freehold tenures are

difficult to access for the Northern Territory and Tasmania. The average rates in Table

5.5 for these jurisdictions are based on figures for leasehold land plus estimates based on

available evidence of clearing rates in freehold tenures.

5.5.2 The impacts of land use change on biodiversity

The implications of land use change for biodiversity are complex. Some species have

increased their range and have benefited in other ways as a result of changes in land use,

but many others have been adversely affected. In the southern states, past land use

change has resulted in patchy, fragmented remnants and small, isolated populations

across much of the landscape.

Land uses can be ranked in terms of severity of the outcomes for biodiversity, beginning

with those that have the least severe impact:

• Biodiversity conservation at the property scale.

• Grazing based on native pastures.

• Grazing based on improved pastures.

• Dryland cropping.

• Irrigated cropping and horticulture.

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Transitions from one land use to another can impact biodiversity, such as replacing

native grasses with improved pasture, converting from dryland agriculture to irrigated

agriculture, and converting from grazing to cropping.

Other transitions can have positive outcomes for biodiversity, e.g revegetating cleared

areas and reducing the intensity of livestock or destocking.

Land management practices that have ongoing and substantial implications on

biodiversity are summarised in Table 5.5.

Table 5.5 Land management practices with ongoing substantial impacts on biodiversity

Land management practice

Impacts on biodiversity Geographic area

Removing native vegetation

Replacing native grasses with improved pasture

Habitat destruction and fragmentation, reduced population sizes, isolation of populations, reduced resilience to other threats, loss of species and diversity, depleted condition and functioning of ecosystems.

Ongoing threat to vegetation communities in south-east Queensland and northern NSW, parts of Tasmania and the Northern Territory.

Small-scale clearing threatens remnants in all settled areas of Australia.

Increasing the intensity of grazing

Habitat depletion and fragmentation, reduced population sizes, isolation of populations, reduced resilience to other threats, loss of species and diversity, depleted condition and functioning of ecosystems.

Ongoing pressure across the rangelands, the wheat–sheep zone and alpine areas, including grazing by livestock, native herbivores and invasive herbivores.

Increasing the intensity of artificial watering points

Habitat destruction, increased competition from invasive species, loss of species and diversity, depleted condition and functioning of ecosystems.

Ongoing pressure across much of the rangelands.

Extracting surface and groundwater for irrigation and other uses

Draining wetlands

Depletion of river flows and aquatic habitat, loss of species and diversity, increased vulnerability to invasive species and other threats, depletion of condition and functioning of aquatic ecosystems.

Ongoing threat in all major irrigation areas, especially in the Murray-Darling Basin.

Groundwater extraction is an ongoing threat to wildlife and natural systems in the Great Artesian basin, south-west Western Australia and many areas of southern Australia.

Converting from livestock to cropping

Habitat destruction and fragmentation, reduced population sizes, isolation of populations, reduced resilience to other threats, loss of species and diversity, depleted condition and functioning of ecosystems, depletion of the condition of aquatic and marine ecosystems.

Relatively stable but ongoing threat across the intensive land use wheat–sheep belt of south-eastern Australia.

Ongoing threat to the Great Barrier Reef.

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Altering natural fire regimes

Depleted composition and structure, reduced resilience to other threats.

Australia-wide, but impacts significant in northern savannas and fire-sensitive and fire-dependent communities (e.g. monsoon vine thickets).

Revegetating cleared areas

Increased habitat and reduced fragmentation, increased resilience to other threats.

Natural regeneration occurs throughout the country; revegetation programs are primarily implemented in southern Australia; large-scale examples aim to link remnants across landscapes.

Reducing the density of livestock

Reduced competition, increased resilience to other threats.

Occurs in response to drought throughout grazing lands and in some areas as part of sustainable grazing management programs; newly acquired conservation areas are generally destocked.

The impacts of land use change and land use practices are examined through the

following selected case studies (Table 5.6).

Table 5.6 Case studies of land use change impacts on biodiversity

Case study Jurisdiction/agency

The impacts of land clearing on vertebrate fauna in northern Australia

Northern Territory and Queensland

The impacts of vegetation clearing on Queensland’s biodiversity and changes through time

Queensland

Eucalypt regeneration in agricultural landscapes in south-eastern Australia

Victoria

Case study 5.7 The impacts of land clearing on vertebrate fauna in northern Australia (Fisher and Woinarski 2008)


Clearing of native vegetation is widely recognised as one of the primary threats to biodiversity

(Tilman et al 1994). In addition to direct habitat destruction, clearing results in fragmentation,

with subsequent impacts on biota through isolation of remnants, reduction in dispersal and

recolonisation, and habitat degradation within remnants (Andrén 1994).


Recent studies in central Queensland and the Darwin–Daly region of the Northern Territory have

investigated the effects of clearing and fragmentation on vertebrate fauna in tropical savanna

woodlands and open forest.

The Queensland study (Hannah and Thurgate 2001, Hannah et al 2007, D Hannah, QPWS, pers.

comm) focused on poplar box (Eucalyptus populnea) and silver-leaved ironbark (E.

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melanophloia) woodlands in the northern Brigalow Belt and southern Desert Uplands of central

Queensland. This area reflects Queensland’s history of extensive land clearing, with about 51 per

cent of the natural extent of these woodlands having been cleared by 1999, mostly since the mid-

1970s. Sampling was carried out at 57 sites, representing a range of clearing, fragmentation and

condition states. Of the 221 native vertebrate species recorded, 132 were sufficiently frequent for

analysis. The effects of clearing per se were assessed by comparing woodland, pasture and

regrowth sites. The effects of fragmentation were assessed by comparing extensive uncleared

tracts and remnants of various sizes, shapes and degree of connectivity. The condition of the

fragments was assessed in terms of stocking rates, density of exotic plants and fire frequency.

• The numbers of 55 per cent of species were significantly affected by one of the disturbance

factors, usually showing reduction with increasing level of disturbance.

• The most pronounced changes (‘biodiversity shortfall’) were in response to clearing,

followed by fragmentation and then to condition. A high proportion of the woodland bird

assemblage is lost with clearing; the average number of bird species in a 1-hectare site was

19.9 in uncleared woodlands, 14.6 in regrowth areas and 8.1 in cleared areas.

• Within fragments, total species richness was lowest and biodiversity shortfall was greatest in

small fragments and unconnected linear strips. Fragments of 5–30 hectares had 35–40 per

cent fewer vertebrate species than fragments of 50–300 hectares.

• The connectivity of fragments was generally not an important factor in determining their

species richness or composition.

• The total number of bird species and the individual abundance of many species within a

fragment were influenced by a complex of factors, one of which was the number of miners

(Manorina spp.). Where miners were abundant, they substantially reduced the number of co-

occurring species, particularly small arboreal insectivores such as striated pardalote

(Pardalotus striatus) and weebill (Smicrornis brevirostris). Although miners occur naturally

in these woodlands, they are likely to have benefited from clearing and other disturbances

associated with current management. Bird population composition in fragments was also

most like that in uncleared tracts where fragments had abundant fallen logs and had been

burnt relatively recently (one usual consequence of fragmentation and pastoral management

is the suppression of fire).

• The effects of site condition on biota were also complex. Of the species tested, the

abundance within fragments of 22 per cent of species was affected by buffel grass density,

and the abundance of 8 per cent of species was affected by grazing intensity. Heavy grazing

in fragments reduced the number of species that prefer a complex understorey for foraging

(such as variegated fairy wren, Malurus lamberti) and benefited those preferring sparse

groundcover (such as Australian magpie, Gymnorhina tibicen, and galah, Cacatua

roseicapilla). Five reptile, 12 bird and two mammal species were less abundant in fragments

with high buffel grass density.

The Northern Territory study (Rankmore and Price 2004, Rankmore 2006) investigated the

effects of clearing and fragmentation in Eucalyptus miniata/tetrodonta open forest, which occurs

extensively throughout the northern tropical savannas. Vertebrate fauna was sampled in a range

of fragments of varying size (0.15–100 hectares) and degree of isolation, as well as in intact

forest and completely modified land (mango plantation). Modelling was used to investigate the

relationship between fragmentation and species patterns, survival and dispersal. A total of 158

vertebrate species were recorded in the study, but only 75 were sufficiently frequent for analysis.

• Fifty-five species were significantly more common in woodlands than in modified (cleared

or partly cleared) sites, and 14 were not found in modified sites at all. Fifty species were

found to use corridors (10–15 m wide) connecting larger woodland patches.

• Twenty-four vertebrate species declined significantly in response to fragmentation. The most

important variables determining their abundance were the connectivity of fragments, the area

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of the fragment, and the proportion of bushland retained within 4 km of the site. For

example, five species were absent from fragments with less than 25 per cent of woodland

retained, while the northern quoll was absent from patches with less than 65 per cent of

vegetation retained. Some species, such as the black-footed tree rat, were absent from

patches of less than one hectare. The connectivity of fragments to intact woodland was

important for 14 species.

• There was often an interaction between fragmentation variables; for example, agile wallaby

(Macropus agilis) and black-footed tree-rat could occur in relatively small patches when

these had high connectivity to surrounding woodland, but were only found in isolated (low-

connectivity) patches if these patches were large.

• Many of the smaller bird populations showed decline in numbers, being unable to readily

move between patches, if connectivity was too low. By contrast, the abundance of the large

and mobile red-tailed black cockatoo (Calyptorhynchus banksii) was determined by habitat

features such as tree hollows and fire frequency within fragments, rather than local patterns

of fragmentation.

• Thirteen species showed a significant positive effect of fragmentation (they were more

abundant in more fragmented areas). These were mostly common and widely distributed

habitat generalists (such as torresian crow, Corvus orru, and brown honeyeater, Lichmera

indistincta), although they also included the northern brown bandicoot and pale field-rat

(Rattus tunneyi).

• Abundance of the four medium-sized mammals remained high in fragmented sites, although

their longer term future may be insecure because survival rates were lower in fragments than

in intact forest. Black-footed tree-rats had a large home range (approximately 65 hectares),

but were able to cope with low levels of habitat fragmentation by moving between smaller

patches. Habitat quality within patches was important—for example, black-footed tree-rats

required large trees and a diversity of fruiting plants, which were more frequent in long

unburnt sites. Brushtail possum appeared to be able to persist in small patches, even where

there was low connectivity and a low proportion of woodland in the surrounding landscape.

These studies demonstrated both the substantial impacts of land clearing on woodland

biodiversity and the potential for careful management of future clearing to maximise the

retention of native fauna. Extensive clearing, such as in the eucalypt woodland in central

Queensland, results in substantial change and diminution of fauna assemblages. Although broad-

scale clearing has been curtailed in Queensland since 2006, many native species may continue to

decline in remaining woodland patches, both because of a time lag inherent in the impacts of

fragmentation on much of the biota, and through reduction in the habitat suitability for many

species of fragments from grazing, exotic invasive grasses, changes in fire regimes and

competitive interactions with species such as miners.

Other longitudinal studies also show declines in the woodland bird and mammal faunas in this

region, including within uncleared sites (Woinarski and Catterall 2004, Woinarski et al 2006). In

the Northern Territory, native vegetation and species assemblages are generally intact, although

there is mounting pressure for further agricultural development. An understanding of species

requirements in terms of regional retention of vegetation, patch size and connectivity allows for

the careful design of landscapes where some clearing is permitted, in configurations that

maximise the retention of woodland fauna.

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Case study 5.8 The impacts of vegetation clearing on Queensland’s biodiversity and changes through time (Butler 2008a)

Nature of threat

The rate of land clearing increased substantially in Queensland from the mid-twentieth century

onwards. Since 1972, more forest has been cleared in Queensland than in all the rest of Australia

(DEH 2000). Land clearing for pasture has accounted for an increasing proportion of this

clearing in Queensland, rising from about 85 per cent in the 1980s to about 95 per cent this

century (DNRW 2007).

A moratorium was placed on tree clearing applications in 2003, and in 2004 major changes were

introduced that saw legal broad-scale tree clearing of remnant vegetation cease at the end of

2006. Clearing can still occur for specified purposes such as firebreaks and essential

infrastructure, and non-remnant vegetation can generally be cleared. The biodiversity impact of

recent clearing will take decades to play out. Secondary threats stemming from clearing,

including the spread of exotic grasses, are still an issue. However, there is reason to hope that

tree clearing is no longer the principal threat to biodiversity in Queensland.


The impact of land clearing on biodiversity is both profound and complex. Clearing causes

habitat loss, reducing local populations of species and increasing their risk of local extinction.

Cumulative losses of local populations deplete variety within species (a key component of

biodiversity) and increase their risk of vanishing at increasing scales—locally, regionally and

ultimately overall. Species strongly associated with ‘productive’ soil ecosystems, which are

targeted by clearing, suffer the greatest degree of habitat reduction.

Conversion of woodland to pastures has a large impact on the diversity and abundance of

species. Research on woodland animals’ response to clearing in Queensland has focused on birds

and provides very strong evidence for substantial impacts (Green and Catterall 1998, Hannah et

al 2007, Martin and McIntyre 2007). Conversion to pastures consistently favours a few generalist

species like magpies and crows, but reduces most of the more specialised woodland species of

birds. Most woodland plants can survive clearing per se, but broadscale tree clearing for pasture

in Queensland is typically accompanied by sowing introduced species (Bortolussi et al 2005),

which displace most native species (McIvor 1998, Fensham and Fairfax 2003, Jackson 2005).

Scaling up from local-scale to broader landscape and regional scales increases the complexity

and uncertainty around the impact of tree clearing on biodiversity. Cleared lands are biologically

impoverished, consisting mainly of introduced pastures supporting a few native species suited to

such systems. Predicting the effects on the rest of the biota within the uncleared parts of the

landscape is more difficult. Some theory focusing on ‘extinction debt’ (Brooks et al 1999)

suggests that around half the bird species that will be driven into extinction by clearing will still

be lingering on 50 years after clearing stops. Such long-term declines and local extinctions have

been documented for Coomooboolaroo in central-eastern Queensland by Woinarski and Catterall

(2004). Nearly half the bird species recorded in the early twentieth century had declined or

disappeared by the end of the century, and bird species that favoured ‘scrubs’ first targeted for

clearing have declined most.

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Case study 5.9 Eucalypt regeneration in agricultural landscapes in south-eastern Australia (DSE 2008b)


Isolated individual and small patches (<1 ha) of mature trees represent a large proportion of

remaining tree cover across large parts of the agricultural landscapes of south-eastern Australia

(Ozolins et al 2001, Gibbons and Boak 2002). Maintaining and increasing tree cover is needed to

prevent biodiversity losses, reverse environmental degradation and sequester carbon throughout

the grassy woodlands of southern Australia (Vesk and MacNally 2006). Scattered paddock trees,

iconic reminders of the formerly extensive woodlands, perform numerous important ecological

functions (Manning et al 2006) and have the potential to support broadscale, cost-effective

revegetation (Dorrough and Moxham 2005)

However, under current management, scattered trees are declining at a rate of 1–2 per cent per

year; complete loss could occur within the next 150 years (Gibbons et al 2008). Grazing pressure

from livestock severely limits the potential for eucalypt regeneration and recruitment. In some

areas, regeneration potential could more than halve within 30 years (Dorrough and Moxham

2005), so there is only a narrow window of opportunity for harnessing natural regeneration to

increase woodland tree cover.


Livestock grazing is a key factor preventing eucalypt recruitment. In areas with scattered mature

trees, exclusion of livestock can trigger extensive natural regeneration (Spooner et al 2002,

Dorrough and Moxham 2005). Recent analyses have demonstrated that the costs and benefits of

different revegetation strategies are likely to vary systematically across the landscape (Dorrough

et al 2008).

Natural regeneration appears to be more cost-effective than tree planting in native pastures of

low productivity (Dorrough et al 2008). In native pastures, opportunity costs of excluding

livestock are relatively low, and competition from exotic pasture is slight. Incentives for

medium- to long-term land retirement (5–15 years) to encourage natural regeneration would be a

cost-effective alternative to replanting in these areas. In high productivity pastures, the likelihood

of eucalypt regeneration is very low, even over a 15-year timeframe. Tree planting, which has

much greater chance of a good outcome in the medium term (<5years), would be more cost-

effective in these situations.

Monitoring regeneration trends is an important component of ongoing management. There is still

much uncertainty about where and when natural regeneration will occur, making investment in

regeneration risky (Vesk and Dorrough 2006). Better understanding of broad-scale patterns of

regeneration are required to help guide managers. In particular, knowledge is required on patterns

of seedfall and likelihoods of germination and seedling survival under different climatic

conditions and in different pastures.

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5.6 Impacts of invasive species and pathogens on biodiversity

Competition from invasive species is one of the most frequently noted threats in formal

documentation for national listing and recovery of threatened species and communities.

Figure 5.7 indicates where invasive species are associated with nationally listed

threatened species and communities.

Figure 5.7 Locations where invasive species are noted as threats to nationally listed threatened species and communities

Of the 27 000 alien plant species that have been introduced into Australia, approximately

2800 have naturalised (NLWRA 2007). The rate of naturalisation is estimated at about

10 species per year. Many of these species compete successfully with Australian native

plants and have become abundant and widely distributed.

Some of the animals introduced to Australia are of particular significance for

biodiversity. These include the European fox (Vulpes vulpes), domestic cat (Felis catus),

European rabbit (Oryctolagus cuniculus), feral goat (Capri hircus), feral pig (Sus scrofa)

and cane toad (Rhinella marina).

5.6.1 Extent and distribution of invasive species

The extent and distribution of key invasive plant and animal species was mapped by the

NLWRA (2008a and 2008b). The mapping included 10 vertebrate pests, 98 weeds and

many invasive plants listed as matters of national environmental significance under the

EPBC Act. These species were selected based on recommendations from the Invasive

Animals Cooperative Research Centre.

The current distributions of invasive species of particular importance for biodiversity

show that this threat occurs in most bioregions, that different species are impacting in

different areas and that some bioregions contain numerous important invasive species.

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5.6.2 The impacts of invasive species and pathogens on biodiversity

The overall impact of invasive species on Australian biodiversity has been devastating.

Predation and competition by introduced plants and animals have severely reduced the

extent and abundance of numerous native species and have impacted on entire taxonomic

groups (e.g. woodland birds and small mammals).

Although many of the most destructive invasive species were introduced some time ago

and are well established but relatively stable, there are notable examples of species (e.g.

cane toad, Rhinella marina) that continue to extend their range into new areas and

threaten relatively intact biota.

New diseases have also caused significant damage to native species and communities

(e.g. facial tumour in the Tasmanian devil, Sarcophilus harrisii, fungal infections in

frogs and Phytophthora cinnamomi infection in southern Australian forests and

woodlands).

Invasive species tend to be abundant and well established in highly modified landscapes,

but also threaten extensive unmodified areas. For example, a recent comprehensive study

of the rangelands, comprising close to 80 per cent of the Australian landmass, concluded

that 11 invasive plant species have the capacity to permanently alter these ecosystems

(Bastin and the ACRIS Management Committee 2008). Impacts on biodiversity are

complex and highly variable, and there is no consistent national monitoring of trends in

either the threats or the effects on biodiversity.

Case studies (Table 5.7) were selected to show the range of effects of invasive species

and pathogens on biodiversity and to highlight examples where the threats are expanding

and impacting on relatively common species.

Table 5.7 Case studies of the impacts of invasive species and pathogens on biodiversity

Case study title Jurisdiction/agency

Impacts on biodiversity of Mimosa pigra in the Northern Territory Northern Territory

Impacts of Phytophthora cinnamomi on biodiversity in southern Australia

Tasmania, Western Australia, New South Wales and Victoria

Impacts of Phytophthora cinnamomi in southern Australia Southern Australia

Phytophthora cinnamomi in the eastern Stirling Range of Western Australia

Western Australia

Facial tumour disease in the Tasmanian devil (Sarcophilus harrisii) Tasmania

Disease in Australian wildlife South Australia

Impacts of invasive species on soil invertebrates CSIRO

Threats from weeds in New South Wales New South Wales

The key messages from the eight selected case studies are:

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• Invasive species and pathogens represent one of the most potent, persistent and

widespread threats to Australian biodiversity. They have both a direct negative

impact on species and communities through losses and extinctions and an indirect

impact on ecosystems and biodiversity through ecological changes brought by those

losses and extinctions.

• They alter entire ecosystem compositions and have directly led to extinctions in

most bioregions of Australia.

• These losses include loss of entire species from mainland Australia and their

contraction to neighbouring islands where the particular invasive threat is not

established.

• Establishment and persistence of invasive species are promoted by a range of other

threats, including fire, all forms of disturbance and climate change.

• There are major gaps in our understanding of the impacts of invasive species and

pathogens.

• Weed management strategies and policies have historically failed to address impacts

on biodiversity adequately.

Case study 5.10 Impacts on biodiversity of Mimosa pigra in the Northern Territory (Fisher 2008a)


Mimosa pigra (mimosa or giant sensitive plant) is a declared Weed of National Significance and

one of the most serious environmental weeds in northern Australia. Mimosa is extremely

invasive and forms very dense stands in seasonally inundated wetlands and along watercourses.

Mimosa now occurs in most major Top End river systems, with the total area of infestation

estimated at about 85 000 hectares (Figure 5.8). The most extensive infestations are in the

Adelaide, Mary and Finniss floodplains and in the Daly River Aboriginal Land Trust.

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Figure 5.8 Distribution of Mimosa pigra (NLWRA 2008)

Trends in biodiversity There have been relatively few detailed studies of the impacts of mimosa on biodiversity,

although this may reflect that such impacts are largely self-evident. Infestations of mimosa lead

to a dramatic reduction in plant diversity and alteration of vegetation structure, with the dense

stands replacing almost all native vegetation. Many of the floodplains affected by mimosa are

wetlands of national and international significance, supporting a high diversity and enormous

numbers of waterbirds and migratory waders (e.g. Finlayson et al 2006), many of which rely on

the sedgelands for breeding and feeding. Mimosa shrublands can also replace diverse riparian,

aquatic, paperbark and monsoon forest communities. At least nine plant and vertebrate species

have been identified as potentially threatened by habitat loss as a result of mimosa infestation

(Walden et al 2004).

Braithwaite et al (1989) studied two areas, on the Adelaide and Finniss Rivers, that had been

infested with mimosa for five and three years. Herbaceous species richness and the density of

native tree seedlings decreased with increasing mimosa density. Numbers of many bird and

reptile species were lower in infested sites. A few birds (such as willie wagtail, Rhipidura

leucophrys) and small mammals (dusky rat, Rattus colletti, and red-cheeked dunnart, Sminthopsis

virginiae) favoured mimosa areas. However, although these species may prefer mimosa patches

for shelter, they are likely to forage in surrounding sedgelands and are unlikely to persist once

mimosa becomes ubiquitous (Lonsdale and Braithwaite 1988).

Similar results were found in a study of the effects of mimosa infestation and control in the

Oenpelli wetlands on the eastern edge of Kakadu National Park (T Griffiths, Charles Darwin

University, pers. comm). Waterbirds, particularly species that feed primarily on herbaceous

plants (such as magpie geese, Anseranas semipalmata), were most impacted by infestation.

However, waterbird populations recovered within five years of mimosa control. One factor that

may inhibit such recovery is the replacement of mimosa in control areas by invasive grasses,

such as para grass (Brachiaria mutica). Revegetation with native species is therefore preferred

(Paynter 2004).

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Case study 5.11 Impacts of Phytophthora cinnamomi on biodiversity in southern Australia (Rudman and d’Arville 2008)


Phytophthora cinnamomi (root-rot water mould, Phytophthora dieback, cinnamon fungus) is an

introduced microscopic soil-borne plant pathogen that infects susceptible plants. The pathogen

invades a plant’s roots and stems and reduces its ability to transport water and nutrients. This

process ultimately kills susceptible species.

Plant community floristics and structure in areas of infestation will be changed, and this may

result in the local extinction of populations of susceptible species. Phytophthora has the potential

to cause species extinctions of geographically rare and restricted species.

P. cinnamomi is listed as a key threatening process under the EPBC Act. It was listed originally

as a ‘fungus’, but is now regarded as being a water mould and not a true fungus. It is widely

distributed, occurring mainly across southern Australia, mostly within the 600 mm mean annual

rainfall zone but extending to around the 400 mm rainfall zone in Western Australia and south-

eastern Australia. Growth and survival of P. cinnamomi are limited by temperature, so that it is

restricted to below about 700 m in Tasmania; however, it extends as high as 1000 m in Stirling

Range National Park, Western Australia, and 1600 m on the Atherton Tablelands in far-north

Queensland. Disease is most strongly expressed on nutrient-poor soils and acidic soils.


P. cinnamomi is known to be a threat to at least 69 plant species listed under the EPBC Act , and

many other species listed as threatened under state or territory legislation are also at risk. The

scale of the threat in Western Australia is clear from the identification of approximately 40 per

cent of 5710 described plant species in the South West Botanical Province (encompassing the

south-western quarter of Western Australia from Shark bay to Esperance, including Perth) as

susceptible to P. cinnamomi and an additional 14 per cent as highly susceptible (Shearer et al

2004). Knowledge about the impact of P. cinnamomi is still limited in some states, such as South

Australia and New South Wales, where many more plant species may be identified as being

threatened as our knowledge increases and Phytophthora spreads further.

Ecological communities may be substantially altered if keystone species or a high proportion of

the species present are susceptible. A number of state-listed or nationally listed threatened

communities are at risk from P. cinnamomi. Many species within the Western Australian

Montane Heath and Thicket community are highly susceptible to P. cinnamomi, with all

occurrences of this nationally threatened ecological community (TEC) now modified by the

disease. Many of the susceptible species are key structural components of the community and

important food sources for fauna; change in vegetation communities affects dependent fauna

(Gole 2006, Shearer et al 2007).

Evidence of a link between P. cinnamomi and fire is increasing, with observed instances of

increased disease impact after fire. Climate change is also likely to impact on distribution, and

warmer conditions may allow P. cinnamomi to extend its altitudinal range.

P. cinnamomi probably already infests many hundreds of thousands of hectares in Western

Australia, Victoria and Tasmania, and tens of thousands of hectares in South Australia

(Environment Australia 2001). It has significant local impacts on native vegetation in several

widely separated areas of eastern New South Wales. P. cinnamomi may spread slowly through

root-to-root contact, and infested soil or propagules can be carried rapidly to new areas by water

movement, wildlife and human activity, making containment difficult. There is currently no

method available for eradicating the pathogen from infested areas.

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Case study 5.12 Phytophthora cinnamomi in the eastern Stirling Range of Western Australia (Rudman and d’Arville 2008)


The impact of Phytophthora dieback on biodiversity has been particularly high in areas

characterised by plant species that are endemic to the local area, such as the Stirling Range

National Park, north of Albany, Western Australia. Two-thirds of the park is infested by the

pathogen (Grant and Barrett 2003). The Montane Heath and Thicket Community, a threatened

TEC of the eastern Stirling Range, is under significant threat from this disease (see photos). P.

cinnamomi directly threatens the survival of this community, which is currently listed on EPBC

lists. In Western Australia, the community is listed as Critically Endangered.


TEC, Moongoongoonderup, before Phytophthora cinnamomi infestation in the 1980s—rich in susceptible species

Photo: Libby Sandiford

TEC, Moongoongoonderup, in 2006, after Phytophthora cinnamomi infestation—loss of susceptible species

Photo: Sarah Barrett, Department of Environment and Conservation

Phytophthora dieback results in changes in plant community floristics and structure and may

result in the local extinction of populations of susceptible species. In areas with geographically

rare and restricted species such as the Stirling Range National Park, it has the potential to cause

species extinctions.

Many species within the Montane Heath and Thicket community are highly susceptible to P.

cinnamomi, with all examples of this TEC now modified by the disease. The abundance of

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susceptible Proteaceous species has been significantly reduced in a community that was once

rich in Banksia and Dryandra species (Barrett 1996, 2000). There have been similar impacts on

susceptible members of the Papilionaceae and Epacridaceae. Many susceptible species are key

structural components of the community and important food sources for fauna.

Case study 5.13 Facial tumour disease in the Tasmanian devil (Sarcophilus harrisii) (Mooney et al 2008)


For more than a decade, devil facial tumour disease (DFTD), a unique, lethal, transmissible

cancer, has been clearly linked to local Tasmanian devil (Sarcophilus harrisii) population

declines (Hawkins et al 2006). With an overall population decline of 50 per cent and local

declines of up to 90 per cent since DFTD emergence (McCallum et al 2007), the devil is now

listed as Endangered by Tasmania.


Signs resembling DFTD were first observed in late 1996 by photographer Christo Baars in far

northeastern Tasmania (Mooney 2004, Hawkins et al 2006). The first clinical examination of

what was later recognised as DFTD was made in 1997 in the Waterhouse Point area, 70 km to

the west of Baars’s sighting (R Loh, Tasmanian Department of Primary Industries and Water,

pers. comm).

Before the disease, Tasmanian devils were distributed throughout mainland Tasmania at varying

densities. By late 2007, the disease was confirmed at 60 locations across more than 60 per cent of

Tasmania (C Hawkins, DPIW, pers. comm), and is clearly spreading. Predictions from modelling

estimate that, at the current rate of spread, the disease will be statewide within 3–20 years

(McCallum et al 2007).

The dramatic decline of devil populations has significant consequences for other Tasmanian

species, ecosystems and communities. The potential ecological impacts of increased food

availability and diminished predation from devils are profound. They include increases in prey

and competitor numbers of indigenous and invasive animal species. Monitoring suggests that

decline in devil populations has coincided with increases in species such as foxes, ferrets and

feral cats (Figure 5.9). Establishment of these invasive species would have potentially

devastating consequences for other native species, including some widely distributed on the

mainland before introduction of foxes but now restricted to Tasmania (e.g. the eastern quoll,

Dasyurus viverrinus; Tasmanian bettong, Bettongia giamardi; Tasmanian pademelon, Thylagale

billardierii).

Passive monitoring techniques used to monitor common species failed to detect the extent and

consequences of the spread of DFTD. Attitudes towards the monitoring, management and

ultimately the conservation of common, ‘secure’ species need to be revised, so that people

become more aware of the vulnerability of these species. The time available to save the world’s

largest remaining marsupial carnivores is short, the task massive and the risk to biodiversity

immeasurable.

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Figure 5.9 Trends in feral cat and Tasmanian devil counts and evidence of fox presence

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Since the founding of what is now the Fox Eradication Branch, evidence of fox presence has increased, along with the effort to find such evidence.

Case study 5.14 Disease in Australian wildlife (Bigwood 2008)


Population decline is often a result of many compounding factors. It is not always necessary to

eradicate all of the factors to reverse the declining trend. Disease is one process that can often be

cured, prevented or controlled by intervention.

Many calculated responses, from imposed spatial isolation to field vaccination, have improved

the success of recovery projects. Examples include rabies/distemper vaccination of the black-

footed ferret, genetic improvement of the golden lion tamarin, and isolation of the northern hairy-

nosed wombat from species that might introduce the mange mite (Sarcoptes scabiei).


Outbreaks of emerging and exotic diseases often deplete animal populations. Chytridiomycosis,

the fungal frog disease, for example, is spreading across the world, having catastrophic effects on

many amphibians. In Australia, it has already caused the extinction of at least one species and is

present in eight endangered and five threatened species of native frog (Bunn and Woods 2005).

Diseases introduced into a previously unexposed population can also have dramatic effects—for

example, the rapid and catastrophic spread of West Nile Virus (which affects a range of species)

across the United States. The virulence of the pathogen is largely due to the naivety of the

population’s immune system; a rapid immune response usually relies on previous exposure to the

pathogen.

Disease can also be an indicator of ecosystem health. Every living creature is a unique ecosystem

of organisms kept in balance by an efficient immune system. Stressors that reduce the immunity

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of an animal can cause a ‘normal’ organism, previously in harmony, to become a pathogen,

overwhelming the immune system and causing disease or death. Stressors may include

starvation, ambient temperature fluctuations, sudden diet changes, reproduction, suboptimal

terrain, overcrowding, genetic problems, and intraspecific aggression. These can increase the

likelihood of infections spreading to other animals, causing the population to become more

vulnerable to all the threatening processes, not just diseases.

Threats to native fauna are increasing, as systems are increasingly put under pressure, and animal

and human habitats come into closer contact. This phenomenon is global and the solution

multidisciplinary.

Measuring parasite load and blood cell indices and monitoring causes of death in a particular

population can be valuable in assessing the health of a whole ecosystem. Woylie (brush-tailed

bettong) decline in Western Australia is a good example of veterinary scientists, epidemiologists

and ecologists working as part of a team to define the organism causing fatalities and understand

why it has recently become a problem.

A range of diseases affect wildlife in South Australia (Table 5.8).

Table 5.8 Diseases affecting wildlife in South Australia

Causative organism

Hosts

Current status in South Australia

Sarcoptes scabiei is an ectoparasite that can cause a life-threatening skin disease.

Southern hairy-nosed wombats and the common wombat in the east. Also concern for critically endangered northern hairy-nosed wombat. Human, canids and other mammals.

A starter program is currently under way through Adelaide University/Adelaide Zoo/Department for Environment and Heritage, to determine distribution and prevalence of this disease in South Australia. Impact on populations is unknown. Good research and data are available for common wombat. Genetic studies are under way to try to determine parasite origin.

Toxoplasma gondii is a protozoan that can be fatal to the host organism.

All animals, including humans, are potential hosts. Fatal in many marsupial species, including eastern-barred bandicoot in Victoria. The cat is required in the lifecycle of the organism.

Number of wildlife species affected, prevalence, distribution and impact are unknown. Surveillance project is required in both vulnerable populations and feral cats. Accurate information for humans and domestic cats is available.

Wallal virus and possibly Warrego virus cause viral infections that lead to blindness in kangaroos.

Several kangaroo species. Present in South Australia; distribution and impact unknown. The virus, thought to be a mutant, emerged in NSW in 1994 and spread across Australia to the west coast by 1995–96.

Psittacine circovirus is a virus known to cause beak and feather disease of cockatoos.

Potential to affect any avian species but of particular concern in glossy black cockatoos and small populations of the yellow-tailed black cockatoo.

Listed as Threatening Process by Department of Environment. Threat abatement plan available.

Need rapid control and prevention plans protecting the endangered cockatoos.

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Batrachochytrium

dendrobatidis is a fungus that causes chytridiomycosis.

Amphibians. Listed as Threatening Process by Department of Environment.

James Cook University in Queensland is running an investigation. Present in Adelaide. Surveillance for the rest of SA is required to determine disease-free populations in susceptible environments.

Mucor amphibiorum is a fungus that causes skin and internal pathology, often resulting in death.

Platypus and amphibians. Originally known as endemic fungus of amphibians in Qld and NT but now prevalent in Tasmanian platypus. Status of the disease in SA platypus populations unknown. Lesions have not been seen.

Case study 5.15 Impacts of invasive species on soil invertebrates (Woodman et al 2008)

The invasion of exotic species is widely recognised as an important cause of decline in

biodiversity and of habitat degradation worldwide. Invasive species, introduced intentionally or

accidentally, are known for all the major soil invertebrate groups. These invasive species may

influence native soil macroinvertebrates either directly (by competing for common resources

such as food) or indirectly (invasive weeds crowd local plants and change the composition and

hence suitability of organic surface debris for local species). Despite the overall lack of long-

term data, some studies do illustrate the potential for exotic species (plant and animal) to directly

disrupt trophic organisation, introduce strong competitive pressures, alter resource networks and

influence disturbance regimes in Australian soils.

Invasive earthworms have been both accidentally and deliberately introduced into Australia.

Little is known about their impact on native earthworm populations and soil invertebrate

ecosystems. The accidental introduction of invasive termite species through infestation of illegal

fishing boats presents a threat in areas in Queensland and the Northern Territory.

Some invasive ant species, such as the yellow crazy ant (Anoplolepsis gracilipes) and the red fire

ant (Solenopsis invicta), have the potential to spread and cause serious ecological damage.

The yellow crazy ant has been widely introduced across tropical and subtropical regions of the

world, where it forms supercolonies covering up to 750 hectares. It occupies more than 2500 km2

in the Northern Territory and is present at Goodwood Island wharf in New South Wales. It can

potentially extend throughout northern and north-eastern Australia and down the eastern

seaboard into northern and inland New South Wales.

The red fire ant was first detected in Australia in 2001 at two locations in Brisbane. Like the

yellow crazy ant, it has an aggressive disposition and non-specialised diet. These ants may

displace native species and introduce predation pressures on vulnerable ground arthropods and

on small vertebrates that frequent the ground (e.g. birds, reptiles, small mammals) (Wojcik et al

2001). The species occupies widespread isolated patches in Queensland and the Northern

Territory.

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Weed invasions are also likely to influence soil communities (Slobodchickoff and Doyen 1977,

Standish 2004, Belnap et al 2005, Wolfe and Klironomos 2005, Kappes et al 2007) and, in

particular, earthworms (Heneghan 2003, Frelich et al 2006, Knight et al 2007). Many overseas

studies have demonstrated that soil communities vary with plant cover (Lee 1985, Boettcher and

Kalisz 1991, Zou 1993, Doube and Brown 1998, Brussard 1999, Peterson et al 2001, Wardle

2002, Gormsen et al, 2004, Wardle et al 2006). Differences in plant communities will influence

microclimate and the nutrient quality of decomposing organic matter either below or above

ground.

Other processes that impact soil communities include disturbance from other invasive animals

such as feral pigs, and changes in local resource composition and availability. This composition

and availability may be patchy relative to the quality and quantity of organic matter, moisture

levels and oxygen availability.

Case study 5.16 Threats from weeds in New South Wales (DECCW 2008b)


Weeds pose a major threat to biodiversity in New South Wales (Coutts-Smith and Downey

2006). There are approximately 1400 naturalised plant species in New South Wales, but not all

of them pose a threat to biodiversity.


According to a recent study, about 330 naturalised plant species pose a threat to biodiversity.

These species have been ranked in terms of the potential threat — considering current and

potential distribution, several measure of threat/impact, the types of vegetation communities that

could be invaded and an assessment of the actual impact relative to their distribution (P Downey,

NSW Department of Environment and Conservation, pers. comm).

To illustrate the impact of weeds on biodiversity, the weed threats listed under the New South

Wales Threatened Species Conservation Act 1995 (see Coutts-Smith and Downey 2006) were

assessed in the study. Of the formally listed threatened species, 45 per cent are at risk from weed

invasions. As illustrative examples, 158 species and 28 ecological communities were identified

as at risk from bitou bush (Chrysanthemoides monilifera), and more than 1300 plant and 150

animal species are at risk from lantana (Lantana camera) nationally (P Turner and P Downey,

NSW Department of Environment and Conservation, pers. comm).

As not all species are at risk in every location where they occur, and control is not always

feasible at every location, a site model is needed to ensure that control is directed to areas where

biodiversity outcomes are likely to be achieved.

5.7 Impacts of grazing pressure on biodiversity

Almost 60 per cent of the Australian land mass is used for production of livestock

(Figure 5.6). Production systems on almost 95 per cent of this area are based on native

pastures, occupying the vast rangelands of central and northern Australia. More intensive

production systems are found in the mixed farming wheat–sheep belt of southern

Australia and are based on improved pastures and fallow rotations. Native pastures were

steadily replaced with improved pastures in all states until around 1970, continuing in

Queensland until at least 1994.

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Grazing pressure is an important threat to biodiversity in both extensive and intensive

livestock production, and its management is critical to biodiversity conservation. It is

associated with direct removal of some species; changes in the relative proportions and

mixtures of species in ecosystems such as grasslands, shrublands and woodlands;

alteration to habitat in mid and lower storeys of forests and grasslands; altered fire

regimes; and impacts on soil structure and water infiltration.

In the rangelands, grazing pressure and other direct impacts of introduced stock have

reduced the extent and abundance of native pastures, and have promoted invasion of

weeds. The more palatable native species are most severely threatened as they are

selectively grazed. Reductions in native grass cover impact on a wide range of other

species and on the functioning of arid land ecosystems.

According to a comprehensive study of Australian rangelands (Woinarski et al 2001,

cited in Bastin and the ACRIS Management Committee 2008):

‘… the terrestrial fauna…has suffered catastrophic decline in many rangeland areas. This

loss has particularly affected larger dasyurids and rodents, bandicoots and smaller

macropods…The bird fauna of many rangeland regions has suffered regional extinctions

and pronounced change….Declines appear to be continuing across much of the

rangelands.’

The study found that a lack of data severely limits our understanding of the interactions

between grazing pressure and biodiversity in the rangelands, and recommended that

long-term trends in biodiversity be monitored, in a systematic and generally consistent

manner across the rangelands (Bastin and the ACRIS Management Committee 2008).

The study found a strong linkage between artificial watering points and impacts on

biodiversity (James et al 1999, cited in Bastin and the ACRIS Management Committee

2008):

‘There appears to be a consistent message of warning coming from…different authors in

different regions: widespread provision of artificial water in previously dry landscapes is

potentially threatening to many species.’

Figure 5.10 indicates that a large number of subregions are affected by networks of

artificial watering points, resulting in much of the area being relatively close to water and

therefore subject to grazing pressures (see case study 5.17—Watering points in the

Northern Territory).

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Figure 5.10 Distance from watering points within IBRA subregions

Source Bastin G and the ACRIS Management Committee (2008)

Key findings of the study relevant to biodiversity are as follows:

• Rainfall variability is one of the major drivers of change in the rangelands.

Managing short-term (seasonal and yearly) variability within the context of longer

term climate change is a key challenge to ensuring sustained production and

biodiversity conservation.

• Historically, rangeland biodiversity has substantially declined, and there is no reason

to believe that this decline had been arrested. Our ability to report change in

biodiversity continues to be limited by inadequate data.

In the intensive grazing systems of southern Australia, native grasses have been replaced

by introduced species, and the entire production system favours establishment and

persistence of these species. In such highly modified landscapes, the original native

biodiversity is severely reduced and largely confined to isolated remnants and small

populations, where it is highly vulnerable to other threats including invasive species,

pathogens, fire and climate change. Many of Australia’s threatened flora and fauna taxa

are found in these areas.

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5.7.1 Case studies of the impacts of grazing on biodiversity

Given the lack of nationally consistent data relating to the impacts on biodiversity of

livestock production systems generally and of grazing in particular, and the highly

variable nature of the impacts, this Assessment examined linkages with biodiversity

through a series of detailed case studies (Table 5.9). The case studies were selected to

highlight the linkages between grazing pressure and biodiversity and to demonstrate the

range of impacts.

Table 5.9 Case studies of the impacts of total grazing pressure on biodiversity


Watering points in the Northern Territory Northern Territory

Impacts of grazing on biodiversity in Queensland Queensland

Impacts of grazing on biodiversity in temperate grasslands and grassy woodlands

New South Wales, Victoria

The Biodiversity in Grain and Graze (BiGG) project Land and Water Australia

Key messages from the four selected case studies are:

• Grazing pressure is a long-standing and complex threat over much of the Australian

landscape.

• Historical records indicate that overstocking and land degradation have led to

depletion and loss of the original vegetation, particularly the best fodder plants, and

that substantial changes in landscapes and vegetation occurred rapidly after pastoral

settlement.

• Substantial impacts on abundance and composition of native flora have been

demonstrated under relatively well-managed grazing without any of the obvious

signs of land degradation that tend to signal heavy grazing pressure.

• Grazing pressure is strongly determined by the distribution of watering points.

• Sophisticated grazing (that is, grazing that is responsive to changes in biodiversity)

can help to conserve biodiversity together with maintaining or developing refuges

within a range of habitat types in grazed landscapes.

Case study 5.17 Watering points in the Northern Territory (Fisher 2008b)


The distribution of grazing pressure in most rangeland regions is strongly determined by the

distribution of waterpoints. Because natural surface water is scarce and mostly ephemeral, the

development of the pastoral industry in Australian rangelands has depended on the installation of

hundreds of thousands of artificial watering points, which has generally aimed to ensure that all

grazing land is sufficiently close to water to be accessible by stock. This massive proliferation of

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waterpoints in the past century has ensured that most areas of most ecosystems are accessible to

stock (Figure 5.11).


Distance from watering points is a useful indicator for the effects of pastoral land use on

rangeland biodiversity (James et al 1999). In general, impacts due to grazing and trampling on

vegetation structure, vegetation composition, ecosystem function and habitat quality become less

pronounced with increasing distance from water. The spread of permanent water across

landscapes also facilitates the spread of native species that are water dependent or that favour

disturbed areas, with a subsequent possible impact on other species through competitive

interaction. Waterpoints may also be a focus for native and introduced predators, and facilitate

the spread of feral grazers and large macropods, adding to total grazing pressure.

Species that decline under increasing grazing pressure or become less abundant closer to

waterpoints are known as ‘decreasers’. Studies along gradients of distance to water in several

rangeland ecosystems, such as mulga woodlands and chenopod shrublands (Landsberg et al

1999, 2003) and Mitchell grasslands (Fisher 2001), have demonstrated that a significant portion

(typically in the range 10–30 per cent) of native species are ‘decreasers’. These decreaser species

are found in each ecosystem and each taxonomic group (plants, birds, mammals, reptiles, ants,

spiders and some other invertebrates) that has been studied. Lightly or little grazed areas are

therefore core habitats for these species, and decrease in the area of water-remote land can result

in decline in their range and abundance and, potentially, extinction at local, regional and national

scales (Biograze 2000). Some species may be entirely dependent on areas with little or no

grazing pressure, although their identification is difficult because of their possible rarity (and

hence difficulty in adequately sampling) and because undisturbed ‘reference’ sites are difficult to

locate in most rangeland regions and ecosystems (Landsberg et al 2002).

Another part of the rangeland biota can be identified as ‘increasers’, which become more

abundant closer to water or in heavily grazed areas. Many increasers are species that are already

widespread and/or common within the rangelands (e.g. galah, crested pigeon), so that an increase

in their range and abundance generally has no biodiversity benefit.

The exact nature of the relationship between distance from water, grazing pressure and impacts

on biodiversity will depend on a large number of factors, including age of waterpoint, type of

stock, stocking history, seasonal conditions, distribution of land types within the grazing area,

and the sensitivity of different biota. Most grazing occurs within a 5 km radius from waterpoints

for sheep and an 8 km radius for cattle, although both these livestock species will walk

considerably further from water at times.

Current trends indicate an intensification of pastoral use in many rangeland regions, as producers

seek to maintain profitability in the face of steadily increasing costs. One means of achieving this

is further ‘water-spreading’ in order to achieve more even use of all available pasture. Although

this is often promoted as pastoral ‘best practice’, and may reduce severe overgrazing and

degradation around watering points, it is likely to further reduce the area of water-remote refuges

and place greater pressure on grazing-sensitive species.

Despite the potentially significant biodiversity implications, the development of new waterpoints

on pastoral leases is not regulated, or only weakly regulated, in most rangeland jurisdictions. A

combination of regulation and incentive is required to promote the retention of adequate water-

remote areas.

Reporting ‘distance from water’ in the rangelands

The Rangelands 2008 – Taking the Pulse report (Bastin and the ACRIS Management Committee

2008) used the distance from stock watering points as an indicator for pressure on biodiversity in

drier rangelands, with the assumption that a decrease over time in the total area of water-remote

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land is likely to be an indicator of negative impact on grazing-sensitive biota. This indicator was

presented as the proportion of each rangeland IBRA subregion greater than 8 km from

waterpoints (Figure 5.11).

Not surprisingly, subregions with lower pastoral productivity tended to have a greater proportion

of water-distant land; up to 79 per cent of land on pastoral leases in the Tanami P1 region and 49

per cent in the Nullarbor NUL2 subregion were mapped as >8 km from water. In contrast, some

more productive subregions had <10 per cent of their total area distant from water, and more than

50 per cent of land within 3 km of waterpoints.

Although potentially valuable, there are a number of serious limitations to using ‘area of water-

remote land’ as a robust indicator for biodiversity in rangelands. These limitations stem from the

lack of detailed and consistent mapping of both waterpoints and land types. Fensham and Fairfax

(2008) also concluded that despite considerable research effort in arid Australia there was still no

compelling evidence that water-remote areas provide refuge for grazing-sensitive species. They

identified a need for further carefully designed studies that concentrate on identifying the role of

water-remote areas as grazing-relief refuges and on their potential as havens for those elements

of biodiversity that have been suppressed in the landscape at large.

Figure 5.11 Change in waterpoint density and distance from water between approximately 1900 and 2004 for a sample area in the southern Alice Springs pastoral district, NT

Source: Adapted from Bastin and the ACRIS Management Committee 2008; data from CSIRO, Alice Springs

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Case study 5.18 Impacts of grazing on biodiversity in Queensland (Butler 2008b)


Grazing by introduced animals was part of a suite of radical changes that swept Australian

landscapes in the nineteenth century, causing far-reaching but often poorly documented

ecological alteration.


The impact of grazing on biodiversity in Queensland is intertwined with changes that began in

the nineteenth century. Identifying the effect of grazing per se on biodiversity, against the

background of drought-induced catastrophes (McKeon et al 2004) and systematic landscape

alteration, is difficult, and this difficulty is increased by the fact that more than a century has

passed since the initial impacts.

Eye witnesses reported the changes that pastoralism brought, especially to Australia’s semiarid

rangelands. In 1901, a Royal Commission into the condition of western New South Wales heard

a series of alarming accounts of land degradation and financial ruin caused by the confluence of

poorly managed pastoralism with drought, rabbits and misguided administration (NSW

Legislative Assembly 1901, Noble 1997, Griffiths 2001). In 1892, the Royal Society of South

Australia heard that throughout the ‘Riverina, and to the western and northern runs of South

Australia, the injury to the original vegetation by overstocking has assumed so great a magnitude

as to entail a national loss’ and that ‘the results of settlement appear on the whole extremely

injurious to that portion of the vegetation which comprises the best fodder producing plants’

(Dixon 1892). Similar concerns were expressed to the Royal Society of New South Wales 10

years earlier by Ross (1882): ‘over-stocking, unfortunately almost universal in these colonies,

has had much to do with the falling off of our pastures, and the consequent depreciation of the

wool’. A record of vegetation change in south-west Queensland’s Mulga Lands based on sheep

droppings begins in about 1930, but it is not until the wet years from 1947 that grass appears in

the sheep’s diet in significant quantities (Witt et al 2006). Substantial change in landscapes and

vegetation occurred quite rapidly following pastoral settlement (Mitchell 1991).

These historical accounts of land degradation accord with subsequent scientific understanding of

the effect of heavy and continuous grazing. The more palatable and especially perennial plants

(so-called ‘decreaser’ species) are disfavoured, and any remaining plants tend to be less palatable

and more ephemeral annual plants (so-called ‘increaser’ species). In many semiarid situations,

the unpalatable increasers include native shrubs, the proliferation of which is a widely recognised

symptom of heavy grazing and associated lack of fire (Noble 1997). If grazing is severe enough

to bare the soil, subsequent rains or windstorms can cause severe erosion and watercourse

degradation, removing nutrients concentrated in the soil’s upper layers (Noble and Tongway

1986).

Severe land and watercourse degradation lies at one end of a continuum of grazing impact on

landscape and biodiversity, producing more-or-less localised catastrophes, but there is also

evidence that subtle yet important changes can be caused by relatively well-managed and

sustainable grazing. In one of very few studies in which the ‘pre-pastoral’ landscape is

appropriately contrasted with contemporary grazed situations, Fensham and Skull (1999)

compared grassy woodlands naturally inaccessible to cattle but grazed by native herbivores with

similar woodlands grazed by cattle but in good condition in north Queensland. They found that

cattle grazing had eliminated one of the woodland’s dominant grasses and substantially changed

the relative abundance of a suite of plant species. These changes resulted from apparently well-

managed grazing without any obvious signs of degradation. Similarly, Martin and McIntyre

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(2007) identified substantial impacts of even light livestock grazing on bird assemblages in

southern Queensland; declines were especially common among understorey-dependent bird

species.

Ecologists are increasingly recognising that the historical impact of grazing animals has

substantially altered composition and function of much of the Australian landscape and must be

considered in ongoing management (Lunt et al 2007). Concepts such as ‘increaser’ and

‘decreaser’ species are too simple to explain the complex responses of biodiversity to grazing

pressure (Landsberg et al 1999, Vesk and Westoby 2001). Grazing is not always incompatible

with biodiversity maintenance; sophisticated use of grazing (that is, responsive to biodiversity

responses, seasonal change etc) can in fact be a tool for nature conservation. For example,

livestock grazing is being used as a management tool in several conservation reserves in southern

states (Lunt et al 2007) and has potential to control fire danger from invasive pasture grasses in

Queensland (Butler and Fairfax 2003).

Case study 5.19 Impacts of grazing on biodiversity in temperate grasslands and grassy woodlands (DECCW 2008c)


In south-east Australia, sustained heavy grazing by domestic livestock has caused major losses of

native biodiversity and ecosystems, although in some areas these depleted secondary ecosystems

may have reached a stable state under the reduced stock levels of current management regimes

(Lunt et al 2007).

In temperate grasslands and woodlands of southern Australia, livestock grazing has converted

understoreys originally dominated by tall, perennial, summer-growing, native tussock grasses

(e.g. Themeda triandra) to dominance by short, winter-growing, exotic annuals, such as Avena,

Bromus, Hordeum and Trifolium species, with losses of associated native forbs (Moore 1970,

Pettit et al 1995, Prober and Thiele 1995, Dorrough et al 2004ab). Many of these changes appear

difficult, if not impossible, to reverse, given the absence of soil seed banks of native forbs (Lunt

1997, Morgan 1998) and self-reinforcing changes to soil nutrient cycles driven by annual species

(Prober et al. 2005).

All native temperate grasslands and woody grasslands are under some threat to their long-term

viability, with degrading disturbances also occurring within permanent reserves (Kirkpatrick et al

1995). Many grassy remnants are also under severe and immediate threat due to their isolation

and continuing changes in land use and land management practices (Tremont and McIntyre

1994). Eucalypt woodlands have experienced the greatest decline of all major vegetation groups

since European settlement (DEWR 2007).

These communities have suffered broad declines in extent and condition, and remaining areas are

generally small and highly fragmented (Figure 5.12). Since they occur on some of eastern

Australia’s most fertile soils, they have been reduced to widely scattered remnants of varying

size, quality and tenure. The understorey of the remnants is largely dependent on management

history, usually under some form of grazing regime, and composition ranges from one or two

species under a patch of trees in an improved pasture to more than 100 native species in an

undisturbed remnant (Carter et al 2003).

The key threats to the survival of these ecological communities are clearing, grazing and weed

invasion. Other threats include salinity, nutrient enrichment, altered fire regimes and the effects

of fragmentation.

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The low levels of soil phosphorus availability in Australia have played an important role in the

evolution and distribution of indigenous flora (Beadle 1954). The application of phosphorus

fertilisers in temperate Australia has had many negative impacts on native vegetation (Dorrough

et al 2006).

Soil fertility clearly affects the richness of native plants, and tends to increase the abundance of

exotics, an effect that may have more of an impact on biodiversity than grazing per se (McIntyre

and Martin 2001). Although there is no clear model of exotic species and grazing in Australia,

increasing fertiliser inputs invariably lead to declines in native plant richness. In some regions,

there is good evidence of grazing-induced losses in biodiversity, independent of fertility changes

(Prober and Thiele 1995). Usually, fertiliser inputs are added in order to increase grazing

intensities, which makes it difficult to separate grazing and fertiliser impacts.

Most studies have found that exotic plant richness is little affected by grazing. There is, however,

considerable evidence throughout the temperate grazed woodlands that the cover and abundance

of a relatively small suite of exotic plants, mostly annuals, does increase in response to either

heavy grazing or enhanced soil fertility.

Small increases in available phosphorus appear to result in dramatic declines in native plant

richness in temperate grazed woodlands in Victoria (Dorrough et al 2006). Consequently,

increases in the intensity of livestock grazing may impact on native plant richness only when soil

fertility is low. At moderate to high levels of phosphorus application (>25 mg/kg), changes to

grazing regimes are likely to have little impact on native species composition, since too few

native species persist. Similar patterns have been observed in native pastures of the Northern

Tablelands of New South Wales (Reseigh et al 2003).

Very similar negative relationships between species richness and phosphorus have been observed

elsewhere in Australia (Adam et al 1989, Morgan 1998) and in grasslands of western and central

Europe (Janssens et al 1998). In white box woodlands in central New South Wales, high levels of

soil nitrate are also associated with low richness of native plant species (Prober et al 2002a).

These similarities across multiple studies in different landscapes and vegetation types suggest a

general pattern in which small increases in available phosphorus or nitrate result in rapid declines

in the richness of native plant species.

In temperate zones, trees can influence understorey grass distribution and local plant richness

(Chilcott et al 1997, Gibbs et al 1999, Prober et al 2002b, Clarke 2003, Dorrough et al 2006). In

contrast, killing or removal of overstorey trees in subtropical and tropical woodlands alters

biomass production but has no effect on the richness of understorey vegetation (Jackson and Ash,

1998, McIvor 1998, Fairfax and Fensham 2000, McIntyre et al 2003). Differences may also be

because tree coexistence in these woodlands is driven by drought and rainfall episodes, which

lead to a shifting mosaic of woodland and grassland. In contrast, woodland patches may be more

stable in temperate zones.

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Figure 5.12 Approximate pre-European extent of temperate grasslands and grassy woodlands in NSW, Victoria and the ACT

Clarke (2003) suggested that grazing may modify the effects of trees on the spatial heterogeneity

of resources and hence plant species richness. Light, intermittent grazing could increase or

maintain resource heterogeneity and thus plant diversity in forest and woodland ecosystems,

whereas intensive grazing may result in reductions (Belsky et al 1993). Grazing intensity could

also vary between areas with low and high tree cover. Research in Tasmania found that forested

areas were subjected to both lower frequency and lower intensity of livestock grazing than

adjoining grasslands and pastures (Leonard and Kirkpatrick 2004).

Case study 5.20 The Biodiversity in Grain and Graze project (LWA 2008)

The call to action

Biodiversity in Grain and Graze (BiGG) is the first research project to investigate the relationship

between biodiversity and on-farm production systems in mixed farms. Its innovative combination

of strong scientific approaches and participatory research has achieved new understanding of

biodiversity on mixed farms and the importance of social, economic and production factors that

influence the relationship between farm management and biodiversity in rural landscapes.

The vision of success

What started out as an aspiration for solid science to address important mixed farming and

biodiversity questions rapidly became a rich mix of fundamental routine research, blue-sky

investigation and farmer participation in the research process.

The project aimed to explore:

• how farm-scale measures of biodiversity are related to agricultural production

• the influence of agricultural management on native biodiversity, and

• the relative influence of site and system features on selected measures of biodiversity.

Ultimately, the project aimed to identify practices to improve natural diversity in agricultural

landscapes, including farmland outside remnant vegetation, a distinguishing feature of the

project.

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The design

Previous research has focused on the impact of farming on biodiversity. This current study also

explored the benefits biodiversity offers mixed farms. Its participatory approach broke down the

scientific–farmer divide.

The monitoring phase ran from autumn 2006 to spring 2007. BiGG field officers interviewed

farmers and collected information from four land use types (cropping, short-term pasture,

remnant vegetation, perennial pasture) on each farm, twice a year over two years. The project

measured plants, invertebrates, birds and soil microbes.

The results

The results of BiGG are useful to farmers, catchment management authorities, land managers

and researchers. As results come to hand, they are fed back to farmers through the Grain and

Graze change-on-farm network. BiGG has already significantly increased biodiversity awareness

among participating farmers and partner organisations.

BiGG researchers and farmers found more than 193 bird species, including 33 rare and

threatened species. Team members logged more than 250 000 individual invertebrates,

representing more than 780 different species of beetles, ants and spiders. A number of new

beetles were identified, and several rare native weevils were found.

5.8 Impacts of altered fire regimes on biodiversity

The relationship between fire and Australian biodiversity is highly complex. Fire is a

crucial component of the ecology and functioning of ecosystems. A long history of fire

in the landscape has helped to sculpt the composition and structure of communities.

Native biota are adapted to a sequence of burning at specific frequencies and

intensities—that is, a specific fire regime. A fire ‘regime’ refers to the fire intensity,

interval between fires, seasonality of fires, and type of fire over a period of time.

The phenology of many species is linked to fire. Some plants require fire to trigger

elements of their reproduction (e.g. seed germination). Conversely, communities in high-

rainfall areas (e.g. rainforest) are infrequently burnt and are sensitive to a more frequent

or intense fire regime. Isolated relic communities that are geographically protected from

fire (e.g. monsoon vine thickets) are particularly sensitive and vulnerable to fire.

With European settlement, the frequency and intensity of fire in large parts of the

continent changed. In southern Australia, fire is deliberately controlled and excluded for

extended periods, with the result that wildfires tend to be relatively infrequent but highly

intense and destructive. Across much of northern Australia, extensive controlled burning

and wildfires occur at higher frequency and intensity, and later in the dry season than in

the past. These altered fire regimes place pressure on finely adapted fire-

tolerant/dependent or fire-sensitive ecosystems, and are having a significant impact on

community composition and structure. A higher incidence of hotter, later fires in the

northern savannas, for example, is associated with habitat loss and spread of introduced,

fuel-rich gamba grass (Andropogon gayanus) at the expense of the native assemblage of

annual sorghum grasses. A decreased incidence of fire in wet tropical environments

(largely due to grazing) has led to loss of grasslands and extension of rainforest.

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Fire interacts in complex ways with other threats. Drought reduces biomass build-up and

therefore fuel availability, but it also aids fuel curing and therefore makes fire more

likely. There is a strong association between large wildfires and prolonged drought. Fire

is also associated with significantly reduced inflows to rivers as vegetation recovers.

Climate change is predicted to increase the incidence of intense wildfires in areas such as

southern Australia, which will experience drier conditions and hotter temperatures.

5.8.1 The extent and frequency of fires in Australia

Technology for monitoring fire distribution and frequency has improved greatly in recent

years, with the development of national satellite-based systems. Figure 5.13 shows fire

frequency over the period 1997 to 2006. It clearly illustrates the significantly higher

frequency of fires in the central arid lands and northern savannas and a clear relationship

with seasonality (extended aridity and prolonged dry seasons).

Fires also occurred frequently in extensive areas in southern Western Australia (in areas

of uncontrolled wildfires), south-east Victoria and coastal southern New South Wales.

Figure 5.13 Australian fire frequency 1997–2006

Fires tend to be highly episodic overall. An analysis of the area of forest burned from

1995 to 2004 in Australia shows that the extensive 2003 fires burned a greater area (more

than 400 000 hectares) than occurred in other years in the period (less than 100 000

hectares).

5.8.2 Case studies of the impacts of fire on biodiversity

Case studies of the biodiversity responses to fire in different parts of the country and in

different landscapes were selected to show the complexity of the responses and how they

are being monitored (Table 5.10).

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Table 5.10 Case studies of responses in biodiversity to fire


Lack of fire threatens grassy ecosystems in the Bunya Mountains Queensland

Monitoring the impacts of fire on biodiversity in tropical savannas Northern Territory

Fire history in the semiarid landscapes of the Lake Johnston area, southern Western Australia

Western Australia

Fire flora monitoring protocol in Victoria Victoria

The plight of the King Billy pine Athrotaxis selaginoides Tasmania

Mallee fowl South Australia with NSW and Victoria

Phytophthora cinnamomi (see case studies 5.12 and 5.13) Western Australia, Tasmania, South Australia and New South Wales

Key messages from the seven selected case studies are:

• Australian biodiversity is finely adapted to specific long-term fire regimes that vary

across the country. In different ecosystems, an increase or decrease in fire frequency

and/or intensity will result in significant and relatively rapid changes in composition

and structure.

• There is a large range of adaptability to altered fire regimes in healthy, functioning,

diverse Australian ecosystems, although their composition and structure may alter

significantly as a result of a new regime.

• Altered fire regimes threaten biodiversity in complex ways as fire has different

impacts on different ecosystems and communities. They are associated with loss of

diversity, changes in distribution and abundance of species and communities, and

changes in ecosystem composition and structure.

• Fire interacts with other important threats, including climate change, drought,

invasive species and pathogens.

Case study 5.21 Lack of fire threatens grassy ecosystems in the Bunya Mountains (Butler 2008c)


Lack of fire is implicated in dramatic changes to vegetation patterns in the Bunya Mountains,

south-east Queensland, threatening grassy ecosystems and their dependent biota. The mountains

support diverse vegetation that can broadly be classified into three main types: rainforest,

eucalypt forest and grassland. The grasslands occur as small patches and are called balds. There

are more than 100 distinct balds, ranging from about 0.1 hectare to 44 hectare, and occurring in

an assortment of landscape positions within eucalypt forest and rainforest. Historical aerial

photographs show that a quarter of the grassland area in 1951 had become forest by 1991

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(Fensham and Fairfax 1996), and more recent aerial photography shows that the rate of

conversion has not slowed.

The Bunya Mountains attracted Aboriginal people from across southern Queensland to feast on

the nuts of the Bunya pine (Araucaria bidwillii). Such a significant place presumably received

intense and strategic fire management, and historical accounts mention the use of fire by

Aboriginal people, although details are lacking. Fire was certainly rare in the Bunya Mountains

during the twentieth century. Could lack of fire explain the balds’ rapid disappearance?

Aerial photographs of the Bunya Mountains (1951 left, 1996 right) showing conversion of grassland (the lightest areas) to forest

Photographs © The State of Queensland (Department of Natural Resources and Water)


Subsequent work examining 77 fires has shown that a series of several fires can stop and reverse

the succession to rainforest from grassland. Fire kills only very young seedlings of Queensland

bluegum (Eucalyptus tereticornis), the main invasive eucalypt on the Bunyas. Burning trials

suggest that eucalypt seedlings develop the ability to withstand fire so fast that very frequent

fires might slow, but would be unlikely to stop, eucalypts invading the balds (Fensham and

Fairfax 2006). It is likely that balds surrounded by rainforest (about 40 per cent of the total

number) would be relatively easy to maintain with regular burning. However, just over half the

balds adjoin eucalypt forest. More than half the balds therefore seem destined to be replaced by

eucalypt forest over the long term. This replacement has possibly been eroding the grasslands

since the last ice age.

Conversion of more than half the balds into eucalypt forest would mean a severe loss of

landscape diversity and natural heritage in the Bunya Mountains, but what would it mean for

biodiversity? Most, if not all, of the plants growing in the balds also grow in grassy eucalypt

woodland. More particularly, the four threatened plants and five plants at their northern range

limits also grow in grassy woodlands. Although a similar study is yet to be conducted for

animals, the pattern for plants seems to suggest that conversion of balds to grassy eucalypt

forests will not be a catastrophe for biodiversity in the Bunya Mountains, despite the erosion of

its landscape and cultural values.

The eucalypt forests are also threatened by lack of fire, which has resulted in widespread

rainforest development in eucalypt forest in the Bunya Mountains. This observation was initially

noted as corroborating fire’s role in the case of the disappearing balds, but for plant diversity at

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least, it now seems to be a major factor. Plants in grassy ecosystems do not persist under

rainforest development.

The disappearing balds indicate the pervasive changes across the interrelated grassy ecosystems

on the Bunya Mountains. It is biodiversity in this broader suite of grassy ecosystems that is

threatened by altered fire regimes—the balds are just the most extreme and most immediately

threatened example.


Similar situations elsewhere are not uncommon. Rainforests often expand into adjoining grassy

ecosystems. In most cases, this is probably associated with fire suppression.

The situation in the Bunya Mountains serves as a reminder that imperilled ecological

communities occur in evolved and dynamic landscapes. The biodiversity, and perhaps other

values, of such communities are often shared with and supported by other landscape components.

The balds have unique landscape and cultural values but form part of a broader suite of grassy

ecosystems. Action to protect the biodiversity of threatened ecological communities should

identify and take advantage of biological connections to other, perhaps more common,

ecosystems—in this case, the grassy forests.

Case study 5.22 Monitoring the impacts of fire on biodiversity in tropical savannas (Fisher 2008c)


Fire is a natural part of the ecology of northern Australian tropical savannas. However, there has

been a significant change in fire regimes in the past century with the disruption of Aboriginal

burning practices, generally leading to an increase in the frequency and extent of relatively

intense, late dry-season fires.

Frequent fires reduce the structural complexity of vegetation. Any uniformly applied fire regime

will tend to decrease the heterogeneity of habitat. Desirable fire patterns that maximise the

retention of biodiversity appear to be patchy at quite small scales and variable over time, as well

as leaving at least some areas unburnt for extended periods.

About 20 per cent of the tropical savanna region is burned each year, although this increases to

more than 50 per cent in parts of the Top End and Kimberley. Well-developed methods for

mapping the extent of fires from satellite imagery, and detailed fire histories extending back for

nearly three decades for some areas (such as Kakadu National Park), are available.


A substantial body of research into the effects of fire on biodiversity in the tropical savannas

(Williams et al 2003, Whitehead et al 2005, Woinarski et al 2007) has shown that much of the

flora and fauna of the tropical savannas is resilient to the effects of fire and variations in fire

regime. However, some fire-sensitive species and communities of high biodiversity value, such

as sandstone heathland and monsoon rainforests, are threatened by frequent fires. Of particular

concern are plant communities that include a high proportion of ‘obligate seeders’, species that

reproduce only by seed and take five or more years to reach maturity. One prominent example is

the northern cypress pine (Callitris intratropica), a distinctive tree that has declined dramatically

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in northern Australia (see Chapter 4). Many fire-sensitive species occur in heathlands in rugged

sandstone terrain, particularly on the Arnhem Land plateau, which has a very high diversity of

endemic plant species (Woinarski et al 2006). In recent decades, most of this landform

experienced fires more frequently than once in five years, resulting in declines in obligate

seeders.

Changes in fire regimes have been implicated in the widespread decline in the tropical savannas

of granivorous birds and small mammals. A number of studies have shown small mammals to be

relatively susceptible to fire impacts. Individual fires benefit some species, such as raptors and

frill-necked lizards that forage in burnt ground, but may also disadvantage other species

dependent on dense grass or fallen logs, or grass- and ground-nesting species such as quails and

wrens.

Case study 5.23 Fire history in the semiarid landscapes of the Lake Johnston area, southern Western Australia (DEC 2008)


Extensive areas of semiarid woodland and shrubland occur in a relatively undisturbed state at the

eastern margins of the southern wheatbelt in south-west Western Australia. Although there are

some significant conservation reserves in this region—including Frank Hann and Peak Charles

National Parks—much of this land is still unallocated crown land and is subject to only minimal

intervention for fire management. Lightning storms are common during the summer months, and

extensive fires occur every few years. Fires can burn for months at a time, with their spread

largely dictated by previous burn history, natural barriers formed by salt lakes, and limited

chaining and back-burning of vegetation along the interface with farming properties. The

incidence of human-caused ignition is low, reflecting the low population density and general

absence of human activity.

The occurrence of large, intense fires appears incompatible with the development and persistence

of the mature eucalypt woodlands, a unique and ecologically important element of the landscape

in this semiarid environment. Further study is needed to provide a historical baseline for fire

regimes and to investigate whether the incidence of fire ignition from summer lightning storms

has increased in recent decades.

This case study focuses on the area defined by the Lake Johnston 1:250 000 map sheet (Figure

5.14), covering an area of 15 500 km2.

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Figure 5.14 Topographic map of Lake Johnston, Western Australia

Source: AUSLIG 1998

Visible fire patterns have been mapped from black-and-white aerial photography from 1958 and

satellite imagery available since the mid-1970s. Fire scar dates have been validated with

reference to tree-ring records from the stems of Callitris, a long-lived native conifer. Several

species of Callitris occur throughout the study area. Fire scars have been dated back to 1930, and

in a few cases to several decades earlier.


Analysis of fire-scar data indicates that the age class distribution of vegetation within the Lake

Johnston study area is highly skewed. Forty-seven per cent of the landscape has been burned

within the past 20 years, mostly by several very large fires ignited by lightning. Only 17 per cent

of the vegetation falls within the 20–70-year age class, while 36 per cent of the study area has not

been burned for more than 70 years. The spatial distribution of fire activity area is strongly

biased towards the south and west, possibly reflecting greater lightning activity in this area or a

difference in fuel conditions. Thirty-two per cent of the study area has been burned two or more

times during the past 70 years.

The consequence for biodiversity and ecosystem function of repeated burning by large, high-

intensity fires include:

• declines in the population of fire-sensitive plant species with long juvenile periods, including

Callitris, Hakea and some species of woodland eucalypt

• loss of mature eucalypt woodland and associated habitat components, such as hollows in

standing trees and coarse woody debris on the ground (see photos below)

• potential decline in vegetation cover leading to greater vulnerability to wind erosion, and

• declining carbon stocks in long-lived woody vegetation.

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Recovery to eucalypt woodlands from a wildfire in January 1991, Lake Johnston, Western Australia

Photos: Lachie McCaw, Department of Environment and Conservation

Case study 5.24 Fire flora monitoring protocol in Victoria (DSE 2008c)

Advances in monitoring

In 2005, a review of existing monitoring in Victoria found that there was limited monitoring of

the ecological outcomes of prescribed burning throughout the state. Ecological values were

rarely monitored; any monitoring that did occur was largely restricted to threatened species and

communities.

Standard monitoring procedures have now been developed to assess flora response to prescribed

burning (DSE 2008c). The initial focus is on flora, due to a lack of data on fauna (which is being

addressed through a separate research project).

A new fire flora monitoring protocol

At a statewide level, the Fire Code (DSE 2006) requires ‘fire regimes and fire management

activities to be appropriate for maintaining and enhancing the vigour and diversity in populations

of species and communities of the State’s indigenous terrestrial and aquatic flora and fauna’.

The overall objective of monitoring in the flora monitoring protocols is to test and refine the land

management framework on which current ecological fire planning is based. This framework

provides minimum and maximum tolerable fire intervals for Ecological Vegetation Classes

(EVCs), and fire is planned to achieve a mosaic of age classes across the landscape. The tolerable

fire intervals are defined by Key Fire Response Species (KFRS), which are ‘those species within

an EVC whose vital attributes indicate that they are vulnerable to either a regime of frequent fires

or to long periods of fire exclusion’ (Fire Ecology Working Group 2004).

This general monitoring objective has been further broken down into a series of smaller

monitoring objectives, which are linked to one or more of the methods in the protocols. These

objectives include:

• obtaining information on flora vital attributes for species that lack such data

1995 2006

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• predicting whether the vegetation in a proposed burn is likely to respond positively to

burning at a particular time

• detecting changes in the presence and abundance of indicator species for an EVC following

fire

• determining the extent to which KFRS can be used as indicators for all species following fire

• detecting changes in species composition in an EVC following fire.

Aspects of fire regimes (season, intensity, frequency or scale) are likely to be major factors

affecting flora species composition, but monitoring alone cannot prove cause and effect.

Table 5.11 describes the monitoring methods recommended for each objective. The ‘indicator

species’ and ‘all species’ assessments require repeated observations over time to detect change

(before-burn and after-burn observations).

Indicator species include KFRS, which are either fire sensitive (usually defining the minimum

fire interval) or fire dependent (usually defining the maximum fire interval). These species are

identified from their vital attributes, such as method of persistence, conditions for establishment

and relative longevity in response to fire. They are used as a surrogate for all species within the

EVC.

Table 5.11 Flora monitoring methods

Objective Assessment type

Description

Baseline species data collection

Vital attribute assessment

An opportunistic assessment of flora vital attributes (Noble and Slatyer 1980) in areas with different fire histories. This information will assist in the selection of indicator species for ecological burn planning and monitoring.

Site-specific burn planning

Life stage assessment for burn planning

A rapid, qualitative assessment of life stage for indicator species in potential burn areas to determine how flora species will respond (positively or negatively) to burning at a particular time.

Monitoring change in indicator species

Indicator species assessment

Measurement of indicator species in >20 plots in a monitoring area to detect change in species composition in an EVC.

Change is detected either in a particular area or across the landscape.

Testing assumptions about KFRS

All species assessment

Comparison of response of KFRS with all species to fire at a site.

Monitoring change in species composition

All species assessment

Measurement of all species in a few plots per monitoring area to assess change in species composition across the landscape in response to fire regime. Change is detected across the landscape, with data combined across a number of areas.

Initial trials to test the methods in 2006 were expanded into a larger prototype phase, which

began in spring 2007. Monitoring sites were located in the North-West, South-West, North-East,

Gippsland and Port Phillip regions and were selected to target knowledge gaps, or to identify

vegetation or species considered likely to be more sensitive to burning. GIS mapping of actual

and desired fire regimes and biodiversity requirements across the landscape assist identification

of these locations. Over time, a more comprehensive coverage of sites will meet a wider range of

monitoring objectives.

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Post-burn monitoring of these sites and a preliminary analysis will occur in spring 2010. The

annual program of monitoring sites will continue as part of ongoing fire management.


The monitoring procedures will become part of the fire management planning process, as

outlined in the Code of Practice for Fire Management on Public Land (Revision No. 1) (DSE

2006). Monitoring is a core element in the adaptive management cycle (planning,

implementation, review), where management practices can be improved in an ongoing, iterative

process.

Specifically, the monitoring program will:

• improve knowledge of the vital attributes of flora species and hence refine tolerable fire

intervals for EVCs

• assist with burn planning and selection of burn units

• assess the effectiveness of management actions

• test whether the actual outcomes of management actions matched expected outcomes, and

• incorporate factors such as fire severity or burn season into burn planning.

The program can be adapted and applied in other jurisdictions where flora vital attributes are

known, providing a monitoring and reporting framework to measure achievement of stated

protection or ecological outcomes.

Case study 5.25 The plight of the King Billy pine Athrotaxis selaginoides (Quinn & d’Arville 2008b)

Introduction

King billy pine (Athrotaxis selaginoides) is an important Tasmanian species both historically,

and scientifically. King billy pine is one of two extant species within the genus Athrotaxis; both

are endemic to Tasmania. Athrotaxis is represented in the fossil record from cretaceous deposits

in South America and New Zealand and was thus a component of the Gondwanan flora. King

billy pine is long lived (1000+ years), it is important for dendrochronology studies and has

contributed to past climate reconstructions (Allen et al, in prep).

The species is extremely fire sensitive and is limited to high rainfall habitats in western Tasmania

where it is rare in lowland rainforests but a more common component of fire protected montane

and subalpine forests, woodlands and scrubs. Jarman et al (1984) describe ten forest and

woodland communities dominated by king billy pine. The statewide vegetation map, TASVEG

1.3 (TVMMP 2008), discriminates three vegetation communities dominated or co-dominated by

king billy pine: Athrotaxis selaginoides- Nothofagus gunnii short rainforest, Athrotaxis

selaginoides rainforest and Athrotaxis selaginoides subalpine scrub (Harris and Kitchener 2005).

The impact of early commercial exploitation of king billy pine (ceasing in the 1930’s), mining

and hydro-electric development on the species’ distribution, coupled with its sensitivity to

wildfire gave rise to concern for the future of the species in the 1980’s (see Brown 1988). All

three of the vegetation communities listed above are classified as extremely fire sensitive. This

classification is based on the potential impact of a single fire on a stand of vegetation, and

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community recovery time after a fire (Pyrke & Marsden-Smedley 2005). Recently, concern for

king billy pine has arisen from the risk that wildfire poses to this extremely fire sensitive species.

Distribution

King billy pine ranges across the coldest and wettest regions in Tasmania. King billy pine has a

widespread but discontinuous distribution throughout western and southern Tasmania. The

species occurs from near sea level to 1270 metres (Cullen & Kirkpatrick 1988a); distribution is

centred around higher altitudes in subalpine forests (Brown 1988). Stands are best developed in

the wetter parts (greater that 1500mm of rainfall per annum) of this distribution, typically from

750 to 1000 metres. The species is most commonly found on well drained, nutrient poor soils

(McMurray 1982).

Ecological characteristics

Disturbance is both an opportunity and a threat for king billy pine. It is intolerant to widespread

disturbance by fire, but relies on localised canopy disturbance for recruitment. King billy pine is

long lived, individuals have been identified as over 1000 years old (McMurray 1982) and slow

growing. These factors result in the species being extremely vulnerable to frequent disturbance

events such as fire.

King billy pine displays highly episodic reproductive activity, producing large amounts of seed

every 5-7 years in what are known as ‘mast years’ (Cullen and Kirkpatrick 1988b). Mast years

are probably stimulated by a climatic signal but the precise nature of this has not been

documented. Other species also have mast years that coincide with those of King billy pine (e.g.

Huon pine Lagarostrobos franklinii and pencil pine A. cupressoides). The seed of king billy pine

regenerates readily but it is poorly dispersed, usually less than 50 metres, and is only viable for

three to four months (Cullen & Kirkpatrick 1988a).

In dense canopied forests dominated by evergreen species king billy pine relies on the formation

of large canopy gaps caused by tree death or catastrophic events such as landslides, for

successful seedling recruitment and establishment (Read 1985). In the absence of such major

disturbance, it is likely that dense stands of king billy pine will not be capable of replacing

themselves and thus the quantity of the species in such forests is likely to decline (Cullen &

Kirkpatrick 1988a). In forests dominated by deciduous beech or those with canopy openings in

which a greater amount of light reaches the forest floor, king billy pine is able to regenerate

continuously (Cullen 1987). King billy pine is not capable of vegetative reproduction (Cullen &

Kirkpatrick 1988a).

Significance

In the 100 year period between 1880 and 1980, 30% of king billy pine populations were lost to

wildfire and, those in the vicinity of Queenstown were killed along with all plant and tree

species, by the fumes from the smelters once operative there (Brown 1988). Athrotaxis

selaginoides- Nothofagus gunnii short rainforest once covered most of the east-facing slopes

between Mount Dundas and Mount Read, but nearly half of these stands have been burnt (Harris

and Kitchener 2005).

The mapped extent of the three vegetation communities in which king billy pine predominantly

occurs totals 28,600 hectares. Ninety one percent of this area is reserved on public land, 58% is

located within the Tasmanian Wilderness World Heritage Area (F Faulkner, pers. comm).

While king billy pine is not listed as a threatened species, the vegetation communities in which it

predominantly occurs are listed as threatened native vegetation communities under Schedule 3A

of the Nature Conservation Act 2002. Athrotaxis selaginoides/Nothofagus gunnii short rainforest

is considered as both rare and vulnerable; Athrotaxis selaginoides rainforest is considered as

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vulnerable; and Athrotaxis selaginoides subalpine scrub is considered as rare1. Montane king

billy pine forests have been identified as a priority community for conservation under the

Tasmanian Regional Forest Agreement (Commonwealth of Australia and State of Tasmania

1997) because of their sensitivity to fire, restricted distribution and richness in Tasmanian

endemic, primitive and relict species. As a primitive and long-lived Gondwanan genus,

populations of A. selaginoides are of international significance (Balmer et al. 2004).

Data/Information

The distribution of communities dominated or co-dominated by A. selaginoides have been

mapped at a 1:25 000 scale by the Tasmanian Vegetation Mapping Program. Brown (1988)

mapped the distribution at 1:100,000 scale of live and dead king billy pine. Cullen (1987)

documented the regeneration patterns of the species in Tasmania. Cullen and Kirkpatrick

(1988a) comprehensively documented the ecology, distribution and conservation of the genus

Athrotaxis. Read and Hill (1988) studied the autecology of the species and in particular seedling

response to light. Cullen (1991) investigated the suitability of landslides as sites for the

regeneration of king billy pine. There have been no studies of the response of the species to

different climate scenarios.

Management Issues

Fire remains the chief threat to the continued existence of the remaining stands of king billy pine.

Wild fires seem to have become less frequent, but more intense covering larger areas in recent

years. In the last decade lightening has become a prevalent ignition source for wildfires, while

arson ignited wildfires have reduced in frequency. This makes preventing wildfires more

difficult. An increase in summer wind, temperature, evaporation rates or decline in summer

rainfall, may lead to increased risk of rainforest loss due to fire (Read 1999).

The introduced pathogen Phytophthora cinnamomi is capable of affecting king billy pine when

conditions are favourable for the growth of this soil borne pathogen. In disturbed areas, where

soil temperatures are not too low the disease has the potential to cause considerable harm

(Podger et al 1990). The potential for climate change to cause soils to warm to the degree that

Phytophthora will spread into areas occupied by king billy pine is currently being investigated.

Some dieback of king billy pine individuals has been observed in some locations in Tasmania,

the cause of this is yet to be determined.

Management actions/responses

As all three of the vegetation communities king billy pine dominates or co-dominates are listed

as threatened communities under the Nature Conservation Act 2002, the species is protected

from activities resulting in clearance and conversion.

Populations of king billy pine are well reserved and predominantly managed through the

Tasmanian Wilderness World Heritage Area (TWWHA) Management Plan (Parks & Wildlife

Service 1999) which is a statutory plan implemented by the Tasmanian Parks and Wildlife

Service. There are no plans to develop a species or community specific management plan.

The Tasmanian Parks and Wildlife Service has attempted to exclude wildfire from most of the

World Heritage Area in order to protect fire sensitive communities such as A. selaginoides

rainforest (Parks &Wildlife Service 1999). The area is designated as a fuel stove only area to

1 Schedule 3A of the Nature Conservation Act 2002 lists species only as threatened. The designations of

rare, vulnerable and endangered have been derived by the Department of Primary Industries and Water, see

http://www.dpiw.tas.gov.au/inter.nsf/WebPages/AWAH-6547ZL?open

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reduce the chance of wildfire resulting from campfires. Fuel reduction burns may be necessary to

lower wildfire risk, the revised fire management plan for the TWWHA will attempt to address

this issue by increasing the area and frequency of management burns in buttongrass moorlands.

Coniferous forests and alpine areas are given priority in wildfire suppression.

Conclusion

King billy pine Athrotaxis selaginoides is a naturally uncommon endemic conifer. The species is

well reserved but the survival of this long-lived, fire sensitive species relies heavily upon careful

future management in a climate where conditions for summer wildfire may be enhanced.

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Case Study 5.26 Malleefowl (SA with NSW and Vic)


The Malleefowl, found in semi-arid and arid shrublands of southern and central Australia, is

listed as vulnerable both nationally, under the EPBC Act 1999, and by international standards

(IUCN 2001).

Over the last century the range and breeding densities of the Malleefowl have declined

significantly mainly due to habitat fragmentation and decline. The threat posed by habitat loss,

predominantly due to incremental clearing, is exacerbated by other threats such as over-grazing

by introduced (and possibly native) herbivores (Frith 1962); excessive predation by introduced

predators such as the fox (Vulpes vulpes) (Priddel 1989, Priddel et al 2007), and inappropriate

fire regimes, especially large and frequent fires from which Malleefowl usually take several

decades to recover.

The recovery effort has led to a number of significant management actions over recent years,

which are detailed in the National Malleefowl Recovery Plan 2006-2010. The plan targets a

range of management responses with a particular focus on: habitat protection through creating

new reserves on public land, covenants on private land, revegetation and connectivity projects,

improved fire management and predator control. More direct actions include programs for

captive breeding, rearing and releasing.

On a broad scale the recovery plan has had the beneficial outcomes of strengthening community

engagement (Benshemesh 2004); developing a representative and effective community based

monitoring system (Benshemesh et al 2008), and mapping the distribution of Malleefowl at a

national scale.


Despite the wealth of information on Malleefowl ecology and the considerable efforts at

reducing threats, breeding densities have declined over at least the past decade. The results of

several re-introduction/supplementation programs and the fox baiting programs have, with a few

exceptions shown disappointing results for the recovery of Malleefowl populations. There have,

however, been signs of strong recovery of the species at a few monitoring sites in SW New South

Wales, northern Victoria, and southern South Australia. The reasons for these increases are still

uncertain. In response to uncertainties, the recovery team has embarked on an ambitious and

pragmatic program of adaptive management that will examine the effectiveness of a range of

different management practices on Malleefowl abundance. Through this program, it is

anticipated that effective means will be identified to recover Malleefowl populations in a variety

of habitats and landscape patterns despite the numerous threats to the species.

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Figure 5.16 Current and historical range of Malleefowl across Australia

5.9 Impacts of changed hydrology on biodiversity

Since European settlement, there have been massive changes in the hydrology of

Australian landscapes. Clearing for livestock and crop production in the 19th

and 20th

centuries changed the rate at which rainfall infiltrated, flowed across the land and was

intercepted. Lower infiltration and faster runoff rates concentrated flow, eroding rills and

gullies. Runoff was delivered faster to streams and rivers and carried high loads of

sediment eroded from cleared and denuded areas. Much of this change occurred

relatively quickly after clearing of vegetation to change land uses.

As settlements were developed, regulation of rivers, interception of runoff in storages

and extraction of river and groundwater for irrigation, stock water and domestic supplies

gradually depleted the natural flows and radically altered the timing of floods. Cold

water released from large river storages changed the chemistry and productivity of

regulated rivers.

Many Australian rivers and wetlands are now stressed, and our largest southern rivers

and streams are over allocated. In most years, there is insufficient flow to provide secure

entitlements to all of the water users who have valid, licensed allocations, and

competition for water in many inland catchments is intense.

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Climate change and variability are compounding the pressures on our aquatic systems

through decreased runoff and increased evaporation, particularly in southern Australia.

Inflows to the Murray-Darling Basin have fallen to critical levels in the past five years.

The recent Council of Australian Governments water reforms are designed to arrest the

decline in aquatic systems and to increase the security of supply. The National Water

Initiative, the basis of these reforms, embodies two main principles:

• Environmental flow requirements are to be met first in order to protect the integrity

of aquatic systems.

• Allocations to bona fide users must be secure.

The initiative involves negotiated catchment and groundwater system water planning and

allocation based on sustainable yields. Management options are more difficult in systems

that are already severely over allocated, such as the entire southern portion of the

Murray-Darling Basin.

The recently negotiated collaborative water management arrangements for the Murray-

Darling Basin, aim to achieve greater environmental protection and higher security for

users through market-driven adjustments to allocations, and through interventions to

improve water use efficiency and management of water through the system. Government

agencies will purchase water on the market and manage it to achieve the best possible

outcomes for the environment.

The biodiversity response to changed hydrology was examined through a series of case

studies, described in Chapter 3—Aquatic ecosystems.

5.9.1 Trends in status and condition of aquatic systems

Monitoring of trends in the status and condition of aquatic systems in Australia is being

carried out by most of the states and territories and, for the Murray-Darling Basin,

through the Murray-Darling Basin Commission’s Sustainable Rivers Audit (MDBC

2008). A national river health monitoring program was partially developed under the

National River Health Program (NRHP) in phase one of the Natural Heritage Trust

(1996-2002). However, the NRHP was not continued past 2002 (see Chapter 3—

Aquatic ecosystems).

Living Murray Icon Sites

The Living Murray initiative is a major river restoration program of the Murray-Darling

Basin Commission. The initiative aims to restore the health of the River Murray system,

which has seriously declined. The Living Murray’s First Step is being implemented from

2002 to 2009 and focuses on improving the environment at six ‘icon sites’ along the

river.

These sites were chosen for their high ecological value (most are listed as internationally

significant wetlands under the Ramsar Convention) and their cultural significance to

Indigenous people and the broader community (Table 5.12). There are as yet no

monitoring data to show trends.

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Table 5.12 Living Murray icon sites

Description Living Murray First Step ecological objectives

The Barmah–Millewa Forest

The Barmah–Millewa Forest is Australia’s largest river red gum forest and the biggest ecosystem of its type in the world. Stretching for more than 66 000 hectares, it provides habitat for a huge range of species and communities. Many of these species are threatened.

• Successful breeding of thousands of colonial waterbirds in at least three years in ten.

• Healthy vegetation in at least 55% of the forest area (including virtually all of the giant rush, moira grass, river red gum forest, and some river red gum woodland).

The Gunbower–Koondrook–Perricoota Forest

The Gunbower–Koondrook–Perricoota Forest is the second largest river red gum forest in Australia. The forest’s wetlands are important breeding sites for waterbirds, including a number of migrating waterbirds that are protected under treaties with Japan and China.

• 80% of permanent and semipermanent wetlands in healthy condition.

• 30% of river red gum forest in healthy condition.

• Successful breeding of thousands of colonial waterbirds in at least three years in ten.

The Hattah Lakes

The Hattah Lakes is a unique collection of semipermanent freshwater lakes. The lakes, located within the Hattah–Kulkyne National Park, support a large variety of waterbirds as well as river red gum communities.

• Restoration of the aquatic vegetation zone in and around at least 50% of the lakes to increase fish and bird breeding and survival.

• Increase in successful breeding events of threatened colonial waterbirds to at least two in ten years (spoonbills; little, intermediate and great egrets; night herons; and bitterns).

• Increase in population size and breeding events of the endangered Murray hardy, Australian smelt, gudgeons and other wetland fish.

The Chowilla Floodplain

Due to isolation from the effects of irrigation, the Chowilla Floodplain has retained much of its original condition and remains one of the most significant floodplain ecosystems in Australia. It is nationally significant as a unique example of wetlands in the normally semidry environment. It supports a number of threatened species and native fish.

• High-value wetlands maintained.

• Current area of river red gum maintained.

• At least 20% of the original area of black box vegetation maintained.

The Lower Lakes, Coorong and Murray Mouth

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The Lower Lakes, Coorong and Murray Mouth is an extensive and diverse system of estuaries and wetlands, covering more than 140 000 hectares. The Coorong is recognized as one of the top six significant wetlands in Australia and provides habitat and breeding sites for a large number of threatened waterbirds. The health of the Coorong has recently been impacted by the lack of water flows in the area.

• Open Murray Mouth.

• More frequent estuarine fish spawning.

• Enhanced migratory wader bird habitat in the Lower Lakes.

The River Murray Channel

The River Murray Channel, the ‘main artery’ of the River Murray, links a diverse range of ecosystems, including forests, floodplains, wetlands and estuaries. These ecosystems provide habitat and breeding sites for a range of native plants, fish and animals.

• Overcoming barriers to migration of native fish species between the sea and Hume Dam.

• Maintenance of current levels of channel stability.

• Expanded ranges of many species of migratory fishes.

• Similar or lesser levels of channel erosion to those currently observed.

Documents

Threats to Australian biodiversity