Landscape Planning for Biodiversity Conservation in Agricultural ...

Landscape Planning for Biodiversity Conservation in Agricultural
Regions A Case Study from the Wheatbelt of Western Australia
Biodiversity Technical Paper, No. 2
Robert J. Lambeck, CSIRO Division of Wildlife and Ecology
© Commonwealth of Australia, 1999
ISBN 0642214239
Chapter 1 - Introduction and Background
• 1.1 Introduction
• 1.1 Background
• 1.2.1 The wheatbelt region
• 1.2.2 Wallatin Creek Sub-catchment
• 1.2.3 Clearing history
• 1.2.4 Land degradation
• 1.2.5 Changes in natural ecosystems
1.1 Introduction
Land-management strategies in Australia have generally focused on maximising economic returns through
specialisation on a single product. While such approaches have generated substantial wealth in many
regions they have also created significant problems. Land degradation, loss of biological diversity and
vulnerability to the vagaries of international markets are all manifestations of the problems that can arise
from single-objective management. Declines in the productive capacity of substantial portions of the
landscape, together with the impoverishment of Australia's unique flora and fauna (Saunders and Curry
1990; Saunders et al. 1991; Morton et al. 1995) have led to the realisation that land management strategies
based on short term profitability are not sustainable in the longer term.
Diversification of productive landscapes, combined with multiple objective management, is seen as one
means of simultaneously retaining production, conservation, social and amenity values. While the need for
integrated management of more diverse landscapes is increasingly being acknowledged (Fry 1991; Hobbs
& Saunders 1991; Hobbs et al. 1993), little progress has been made towards developing procedures for
achieving this. In fact, much of the theoretical debate has contributed little towards practical application
(Saunders et al. 1991).
If we are to successfully combine non-productive land uses, such as the conservation of biodiversity, with
other land uses, we must accomplish four essential tasks. The first of these is to set clear objectives for the
region to be managed. Secondly, it is necessary to specify what is required in a landscape to meet those
objectives. This requires definition of the composition, the amount and the configuration of essential
landscape elements. The third task is to identify how (or if) the landscape elements required for meeting
different land-use objectives can be combined in a landscape. Finally, it is necessary to assess the
economic implications of the different scenarios that may be adopted in order to meet the identified
objectives.
This report addresses the first three of these tasks with an emphasis on meeting nature conservation
objectives and integrating these with production goals. An assessment of the economic implications of the
conservation scenarios developed was beyond the scope of this project. The procedures used to address
these tasks are illustrated by a case study from the central wheatbelt of Western Australia.

A brief review of the issues affecting the region and a description of the study area are presented initially to
provide some background to the report. Chapter 2 examines a range of alternative nature conservation
objectives and explores the consequences of pursuing each of these objectives. On the basis of this
assessment, the most appropriate objective for the purpose of this study is selected and landscape designs
and management recommendations are developed to meet that objective. The extent to which the results
derived from this study can be applied to other regions is then examined. Chapter 3 presents an approach
to developing land use plans which integrate the requirements for meeting nature conservation objectives
with production and land conservation objectives. Finally, an Appendix is presented which considers the
expansion of the approaches developed in this project to a regional scale.
1.2 Background
1.2.1 The wheatbelt region
The wheatbelt of Western Australia covers an area of about 18 million ha extending in a broad band from
Geraldton in the north-west to Esperence in the south-east (Figure 1). Annual rainfall ranges from 600mm in
areas closer to the coast to 280mm on the eastern boundary. This climatic region represents a transitional
zone between the more mesic Bassean region of the south-west.
The original vegetation consisted of a mosaic of plant communities including tall, open woodland, dense
shrubland, and low heathland which reflected an underlying mosaic of landforms and soil types (Beard
1983). The region supports an important agricultural industry, based primarily on cereal cropping and sheep
grazing, which generates an estimated $4.5 billion annually (Government of Western Australia 1996).
Figure 1. The wheatbelt of Western Australia showing the location of the case-study.
1.2.2 Wallatin Creek Sub-catchment
The case study was based on the Wallatin Creek sub-catchment which occurs to the north of the central
wheatbelt town of Kellerberrin about 200 km east of Perth. The catchment covers an area of 26015 ha and
contains 19 properties. The property owners have formed an incorporated company which they have called
Wallatin Wildlife and Land Care Inc., reflecting the multiple management objectives that they have for the
catchment.

The area has an average annual rainfall of 339 mm which falls predominantly in winter. Unpredictable
rainfall may also occur in some years during the hot summer months (Hobbs 1992b).
The gently undulating topography of the region has resulted from the erosion of lateritic sand plains which
comprised part of the Great Plateau of Western Australia (McArthur 1993). The landscape has been subject
to weathering since at least the Palaeozoic era when Australia was part of Gondwana. Differential erosion
has resulted in different strata being exposed at different positions in the landscape. A typical catenary
sequence resulting from this variable weathering process is shown in Figure 2. The higher parts of the
landscape (Ulva) which form the drainage divides consist of sandy soils with laterite capping. The exposed
slopes (Booraan and Collgar), which have resulted from the dissection of the sand plain, are characterised
by outcrops of weathered granite (Danberrin) and duplex soils which vary in composition with position on the
slope. Below these weathered slopes lie broad alluvial flats of heavier soils (Belka and Merredin). The
lowest positions in the landscape are occupied by predominantly saline lakes, swamps and playas
(Nangeenan and Belka). A more detailed description of the geomorphology and topography of the region
can be found in McArthur (1993).
The variations in soil and landform types described above are associated with characteristic vegetation
types (Figure 3). The sandy uplands or Ulva, support Kwongan heath vegetation characterised by high
levels of endemism and species richness (Lamont et al. 1984). The upper slopes (Booraan) support a dense
shrub layer dominated by a variety of Melaleuca species and Allocasuarina campestris as well as eucalypt
woodland communities in some locations.
The gentle lower weathered slopes (Collgar) support diverse mallee (multi-stemmed eucalypt) communities
some of which are associated with dense thickets of Melaleuca cardiophylla.
The broad flat alluvial portions of the landscape (Belka, Nangeenan and Merredin landforms) are dominated
by woodland communities, predominantlyWheatbelt Wandoo (Eucalyptus capillosa), Salmon Gum (E.
salmonophloia) and Gimlet (E. salubris).
Areas of weathered granite rock outcrops (Danberrin) are associated with York Gum (E. loxophleba) and
Jam Wattle (Acacia acuminata) but there may also be dense stands of Casuarina (Allocasuarina
huegeliana) woodland around the margins of the rock. The lowest positions in the landscape support
Samphire communities on very saline soils and Atriplex species in less saline areas.
Figure 2. An idealised wheatbelt landscape showing the major soil landscape units. From Lantzke (1992)

1.2.3 Clearing History
Initial clearing of the native vegetation in the wheatbelt commenced in the mid 1800s but progressed
relatively slowly until the early 1900s when machinery which enabled more rapid and widespread removal of
vegetation became available. In the central wheatbelt the process of extensive clearing continued until the
1960s by which time the limits of suitability for agriculture had largely been reached (Figure 4).
Subsequent removal of vegetation has taken the form of the sporadic removal of remnants that were passed
over in the first wave of clearing. In most areas of the wheatbelt today less than 10% of the original
vegetation remains, the majority of which occurs in small remnants and roadside verges (Wallace & Moore
1987).
Figure 3. Diagrammatic cross section through a typical wheatbelt landscape displaying the relationship
between vegetation and soils (From Beard 1990).
1.2.4 Land degredation
While the extensive clearing of native vegetation for agriculture has generated significant wealth for the
nation, it has also created substantial problems. The replacement of perennial vegetation with annual crops
has significantly reduced rates of evapotran-spiration and altered patterns of water flow through the soil
(McFarlane et al. 1993). As a consequence, the naturally saline water table has risen (and is continuing to
rise) over much of the wheatbelt. Currently 1.8 million hectares of formerly productive land, or approximately
10% of the agricultural region, has been affected by salinity at an estimated cost of $1445 million
(Government of Western Australia 1996). This trend can be clearly seen in the Kellerberrin shire (Figure 5).
It is estimated that up to 30% of the wheatbelt has the potential to be salt affected if no action is taken
(Government of Western Australia 1996). Similarly the area affected by waterlogging is increasing, causing
losses of millions of dollars through lost production in the Kellerberrin area alone (McFarlane & Wheaton
1990). In addition, wind and water erosion affects 2 million ha and 0.75 million ha of the wheatbelt
respectively (Nulsen 1993). Soil structure decline and soil acidification are also significantly reducing
agricultural productivity.

Figure 4. Clearing history of the Kellerberrin study area (From Arnold and Weeldenberg 1991).
1.2.5 Changes in natural ecosystems
Changes in land cover have had a major impact on the natural ecosystems of these regions. Numerous
plant and animal species have been lost and there have been significant changes in the distribution and
abundance of those which remain. These changes to the biota have, in turn, modified the ecosystem
processes in which the various plant and animal species participated. Because it is these changes to natural
ecosystems that are the main focus of this project, they are explored in greater detail in the next chapter.
Figure 5. Area affected by secondary salinity in the Kellerberrin Shire. (From McFarlane et al. 1993)

Chapter 2 - Retaining Biodiversity in Agricultural Landscapes
• 2.1 Biodiversity: the variety of life
• 2.2 Biotic impoverishment: the impact of agriculture
• 2.3 Protecting biological diversity: the need for clear objectives
• 2.4 General enhancement
• 2.5 Strategic enhancement: using focal species to define landscape
• 2.6 Reintroductions
• 2.7 Mixed strategies in the face of partial knowledge
• 2.8 Design and management recommendations for Wallatin Creek
• 2.9 Priorities for implementation
• 2.10 Guidelines for implementation
• 2.11 Moving the goal posts: the consequences of implementation
• 2.12 Transportability of solutions
• 2.13 The role of science and data adequacy
• 2.14 Summary
2.1 Biodiversity: the variety of life
The term biodiversity was coined to describe the complex variety of life that
characterises this planet. Biodiversity encompases differences at all levels of biological
organisation. It includes differences between individuals within a species; between
species themselves; between communities, landscapes and ecosystems. Given that allencompasing
nature of the term, it is apparent that the conservation of this diversity is
not a trivial task and that it will not be possible to develop management strategies that
explicitly consider the needs of the biota at all of these levels of complexity.
Management of biological diversity in production landscapes must therefore be based on
procedures that consider particular levels of this hierarachy of biological diversity but
which also have a high probability of encompasing the other levels. In the current study,
attention was primarily directed to the consideration of species and communities at a
landscape scale. Owing to the absence of information about the range of variation within
species it is necessary to assume that strategies which will retain communities of
species will also retain the variability that occurs between the individuals of those
species.
2.2 Biotic impoverishment: the impact of agriculture
• 2.2.1 Changes in biota
• 2.2.1 Changes to ecosystem processes
Because biodiversity incorporates differences at all levels of biological organisation, from
individuals to ecosystems, and substantial change in a landscape will bring about a
change in biodiversity. Such changes are of particular concern when patterns of landuse
cause the loss of species or changes in their distribution.
In agricultural regions throughout Australia trajectories of change are consitently towards
declining natural diversity (Saunders 1989; Department of Environment, Sport and
Territories 1996). This decline in diversity results in changes to the rates and pathways
of ecosystem processes that are essential for conserving biodiversity and sustaining
agriculture (Saunders et al. 1991; Mooney et al. 1995).

2.2.1 Changes in biota
Twenty four plant species are known to have been lost from the wheatbelt as a whole
and the area now has one of the highest numbers of rare and/or endangered plant
species in Australia (Briggs & Leigh 1996). Of the 348 plant species listed as rare or
endangered in the wheatbelt, only 79 occur in designated nature reserves. Of the
remainder, 135 species occur in road verges and 53 are found on privately owned
remnants (Hopper et al. 1990).
Of the 43 mammal species that occurred in the region prior to European settlement, only
12 species are now moderately common or abundant (Kitchener et al. 1980a). Mammal
declines are continuing with several species disappearing or declining in abundance in
the Kellerberrin area over the last 15 years (Hobbs et al. 1993). While loss of habitat has
significantly reduced the population densities of these species, feral predators have also
played a major role in the process of decline and extinction (Burbidge & McKenzie
1989). It appears that bird species may be following a similar trajectory of decline, albeit
at a slower rate (Saunders 1989; Saunders & Ingram 1995). Thirty one species have
decreased in range and/or abundance in the wheatbelt over the past 90 years. Of these,
15 species no longer occur in the Kellerberrin district (Saunders & Curry 1990; Hobbs et
al. 1993). Lizards appear to have been less severely affected with no obvious
widespread loss (Kitchener et al. 1980b). While less information is available for frogs, it
would be expected that many species would be deleteriously affected by increasing
salinity both in the soil and in water courses.
While there is no evidence that changing patterns of land use have resulted in the
extinction of any invertebrate species, it is clear that they have caused substantial
changes to local invertebrate diversity (Springett 1976; Scougall et al. 1993). The
absence of evidence of invertebrate extinctions due to land use practices does not mean
that invertebrate species are secure in agricultural landscapes. Rather, it simply reflects
the absence of appropriate data for detecting such extinctions if they have occurred.
2.2.2 Changes to ecosystem processes
Changes to ecosystem functions have resulted from changes in physical processes due
to altered land-cover, as well as from changes in the diversity of species that participate
in the movement of nutrients, water and energy throughout the system.
Physical changes
Research conducted in a range of agricultural regions throughout the world indicate that
extensive land clearing can modify patterns of radiation and alter fluxes of wind, water
and nutrients across landscapes (Baudry 1989; Risser 1990; Saunders et al. 1991;
Hobbs 1992a). Radiation fluxes tend to be more variable in agricultural fields than they
are in remnant vegetation both throughout the day and between seasons, resulting in
greater local temperature extremes. Where natural bushland and agricultural fields meet,
the higher radiation from the bare ground affects the microclimate at the remnant edge
causing changes in plant composition and structure (Lovejoy et al. 1986; Palik & Murphy
1990). Altered radiation levels in fields are also likely to modify soil micro-habitats for
soil-dwelling invertebrates. Increased exposure to wind affects conditions at the edge of
remnants, with increased frequency of tree blow-downs, increased evapotranspiration

(Lovejoy et al. 1986) and increased transfer of weed seeds, nutrients and agricultural
chemicals from farmland into vegetation remnants. Similarly, increased surface water
flow brings nutrients, chemicals and weed seeds into remnants (Muir 1979; Cale &
Hobbs 1991).
Changes in vegetation cover have affected the hydrological balance in the catchment
causing a rise in saline water tables which impacts initially on vegetation remnants in the
lower valley floors but, as water tables continue to rise, vegetation located higher up the
slopes is increasingly being threatened.
Biotic changes
Considerable attention has been paid to changes in ecosystem function associated with
changes in biodiversity (Hobbs 1992b; Lambeck 1992; Main 1992; Mooney et al. 1995).
Much of the debate revolves around the question of whether greater levels of diversity
result in greater ecosystem stability or resilience in the face of environmental variability.
MacArthur (1957) argued that a greater number of species provides more pathways for
energy to reach a consumer. Consequently, the loss of any one pathway would be less
likely to have a detrimental impact on that consumer. Elton (1958) also contended that
the increasing complexity of an ecosystem that resulted from greater species diversity
should generate greater stability. However, Cody (1986) found that high diversity in
Mediterranean-climate regions tends to be associated with a greater probability of rarity
and a consequent greater risk of extinction. However, the loss of a rare species from a
diverse system is likely to have less impact on ecosystem processes than would the loss
of a more abundant species from a less diverse system. This is because a more diverse
system will have a greater probability of containing other species which play a similar
functional role and hence can compensate for the loss of a particular species. Such
compensation may not occur if a 'keystone' species is lost. This is a species which has
an impact on ecosystem processes that is disproportionate to its representation in the
community (Power & Mills 1995). In such circumstances, a small change in diversity can
have a significant impact on ecosystem function.
The impact of changes in biodiversity on function will therefore depend upon a number
of factors including the number of species in the system, the relative abundance of those
species, the functional role played by the species lost or gained and the relationships
between these and other species in the system (Mooney et al. 1995).
2.3 Protecting biological diversity: the need for clear objectives
Before any attempt is made to address conservation issues in landscapes used for
production it is essential that the objective of the exercise is clearly identified. Two broad
approaches to nature conservation can be considered:
General enhancement, which attempts to maximise the number of indigenous species
retained or, alternatively, to minimise the number lost within constraints imposed by
other land use objectives.
Strategic enhancement, which aims to ensure the persistence of particular species,
groups of species, or all species that currently occur in a landscape. This type of

approach could be extended to include the reintroduction of species that have been lost
from the landscape being managed.
The objective of the first type of approach - to maximise the number of species retained,
or to minimise the number lost - is an open-ended objective. It identifies a general
trajectory along which to proceed, but does not specify targets which can be used to
assess success or failure. This type of approach does not consider which species will be
conserved or lost but simply aims to increase the probability that any given species will
persist. It is a 'general enhancement' objective, in that it aims to 'make things better'or
'minimise the impact' in an unspecified manner within the constraints of other land uses.
Approaches based on retaining or reintroducing specified components of the biota can
be considered to be 'strategic' because they require specification of the landscape
elements and management regimes that are required to meet a specific objective. They
are more rigorous because they have quantifiable outcomes by which we can judge the
effectiveness of our actions. If our objective is to retain all of the biota or a particular
component of that biota, then the loss of any species, or of particular designated species
would constitute failure. Similarly, if our goal is to reintroduce particular species that
have been lost from an area, the failure of those species to become established is a
clear indication that the objective has not been met.
This distinction between these two broad types of objectives is an important one
because the management strategy adopted will differ considerably depending on the
objective chosen. A general enhancement strategy must rely on ecological principles,
while a strategic enhancement approach depends on having information about the
requirements of the species to be retained or restored. These alternative strategies will
result in different recommendations about the spatial and compositional characteristics
of a landscape and will have significantly different conservation outcomes. The different
types of objectives associated with these different approaches are therefore considered
below in greater detail and the design and management implications of each are
examined.
2.4 General enhancement
General enhancement strategies attempt to identify ways in which a landscape can be
improved in order to reduce the probability of species being lost. Such strategies tend to
draw on general ecological relationships between landscape or habitat characteristics
and attributes of biological communities. These attributes typically include species
richness or diversity, guild richness, biomass, or trophic structure. These types of
relationships have received considerable attention at a theoretical level. For example,
species-area models suggest that habitats of different sizes will support different
numbers of species (Munroe 1953; MacArthur & Wilson 1963); patch dynamic theories
(Cody 1975; Roth 1976; Pickett & White 1985) imply that areas having greater horizontal
diversity (patchiness) both within and between habitats will support greater numbers of
species; niche theory (Hutchinson 1958; MacArthur & MacArthur 1961; Wiens 1989)
similarly predicts that areas having greater vertical and horizontal complexity will provide
more niches and will therefore support more species; the intermediate disturbance
hypothesis (Grime 1973; Connell 1978) implies that different intensities or frequencies of
disturbance will produce communities with different levels of species richness; and
island biogeographic theory (MacArthur & Wilson 1967) and metapopulation theory

(Levins 1970; Gilpin & Hanski 1991) predict that more fragmented or isolated habitats
will support fewer species than habitats that are more connected, and that populations in
more fragmented and isolated habitats are less likely to persist through time.
Where management objectives aim simply to maximise the number of species present,
these general principles can provide a useful guide to the attributes that we should retain
in existing habitat or incorporate into revegetation strategies. If such an objective is to be
employed, emphasis should be placed on maximising the number of locally indigenous
species, rather than simply maximising species numbers per se.
Vertical heterogeneity
Much of the literature relating structural diversity to species numbers is based on studies
of bird communities. MacArthur and MacArthur (1961) and Recher (1969) identified
strong relationships between foliage height diversity and bird-species diversity in both
America and Australia. While other studies have revealed weaker relationships, or in
some cases no relationships between these variables (eg. Emlen 1977; Rice et al.
1983), they often found a correlation between bird-species diversity and some other
aspect of vegetation structure (Wiens 1989). Mammal species also respond strongly to
differences in vegetation structure. Rosenzweig and Winakur (1969) and August (1983)
found a positive correlation between habitat complexity (number of vertical strata) and
the total number of mammal species in tropical forests. In Australia, the diversity and
abundance of small mammals and arboreal marsupials are also influenced by vegetation
structure (Barnett et al. 1978; Laidlaw & Wilson 1989; Lindenmayer et al. 1994a, 1994b).
While these relationships do not hold under all circumstances (see for example, Bond et
al. 1978), the majority of observations suggest that greater structural complexity is
unlikely to have a detrimental effect on species diversity and in most cases will enhance
species numbers through the provision of a wider variety of habitats. On the basis of
these observations, attempts should be made to ensure that the appropriate number of
structural layers are maintained in vegetation remnants or are incorporated into
reconstructed habitat. Depending on the vegetation types typical of the area, these
layers may include some form of ground cover, understorey shrubs, taller middle-story
shrubs or smaller trees, and an upper canopy of taller tree species. Guidance as to
appropriate levels of structural diversity could be acquired by examining non-degraded
vegetation remnants in the area being managed.
Horizontal heterogeneity
The importance of heterogeneity in general and patchiness in particular, has received
considerable attention in the ecological literature (eg. Pickett & White 1985; Kolasa &
Pickett 1991). Patchiness in a landscape not only provides a greater range of habitats
which can be occupied by a more diverse fauna, but also provides multiple
representation of equivalent patch types. In the event of a disturbance which impacts
differently on different patches of the same habitat, the biota in the less affected patches
can provide source populations to recolonise the more affected patches (Grubb 1977).
Such patchiness must be retained in agricultural landscapes and should be incorporated
into reconstructed habitat. Guidance as to the appropriate scale of this patchiness for the
area being managed can be obtained by observing the characteristics of existing
remnants or by examining patterns in the distribution of soils and landform types.

Configuration
Two important aspects of configuration include patch proximity and connectivity between
patches. Both of these parameters affect the movement of individuals between patches
and therefore influence the probability of local extinction of populations and the likelihood
of recolonisation if local extinction does occur. This is particularly important in highly
fragmented landscapes where sub-populations are subject to extinction as a result of
natural catastrophes, human impacts, or stochastic fluctuations in population sizes
(Fahrig & Merriam 1985; Newmark 1985; Fahrig & Paloheimo 1988; Pulliam 1988). In
general, patches that are closer together and are connected by habitat which allows
movement of individuals are more likely to sustain populations of their constituent biota
than are more isolated patches (Forman & Godron 1986; Saunders & Hobbs 1991). In
general, the distance between patches should be minimised and where possible, the
width and structural complexity of vegetation linking remnants should be maximised.
Area
There has been lengthy debate over whether a single large area is better than several
small ones for maintaining biological diversity (see for example Margules et al. 1982;
Simberloff & Abele 1982; Shaffer & Samson 1985; Margules 1987). This debate centers
around the question of whether it is area per se which results in increased species
number, or whether increasing area is correlated with greater environmental variability
which in turn contributes to greater habitat diversity and hence to higher species
diversity. Whatever the cause of this increasing diversity with increasing area, the
relationship is not a linear one. The rate at which species accumulate with increasing
area tends to decline beyond a particular size. In general, larger areas are likely to
contain more species than smaller ones. However, the number of species will also
depend upon the patchiness of those areas. This suggests that larger areas of
monocultures are likely to contain fewer species than the same area with a number of
patch types. It may also be possible for smaller areas made up of a range of patch types
to contain more species than a bigger area containing fewer patches. In regions where
levels of local endemism are high, it is possible that small areas may contain unique
species that do not occur in other larger areas. When protecting or enhancing vegetation
for conservation, it is best to aim for bigger rather than smaller areas but also to
incorporate heterogeneity into the design where possible. Small areas should be
included if they are known to contain species that are poorly represented elsewhere.
Shape
The importance of remnant shape depends, to some extent on the size of the patch or
remnant being considered. While shape may be less important for large remnants, it can
be a critical factor in determining the conservation value of smaller remnants. Long, thin
remnants, such as corridors or riparian vegetation, have a much higher edge to interior
ratio and hence are more exposed to deleterious edge effects (Diamond 1975; Wilson &
Willis 1975). These include increased nest predation and brood parasitism in birds
(AndrÚn & Anglestam 1988; Johnson & Temple 1990; Paton 1994), a shift from
characteristic tree species to weedy generalists including the invasion of exotic weeds
(Noss 1987; Hobbs & Atkins 1988; Scougall et al. 1993) and increased exposure to wind
which impacts on vegetation condition (Moen 1974; Grace 1977). Linear habitat can also
present problems for species that have to return to a particular location, such as a nest,
when foraging. In linear habitat individuals may have to travel much greater distances to

meet their food requirements than they would do if those resources were distributed
within a more compact patch (Recher et al. 1987; Saunders 1990).
In general, compact areas will be better than linear areas for the provision of habitat and
for protection against external impacts from adjoining land uses. In the absence of more
precise guidelines, patches should have the least possible edge and linking vegetation
should be as wide as is practically feasible (Wilson & Lindenmayer 1996).
Buffers
An enhancement strategy can also contribute to the protection of remnant vegetation by
providing a buffer from external impacts such as strong winds and fluxes of chemicals,
fertilisers and weeds. This buffering role may be achieved by planting bands of shrubs
and trees around the edges of existing remnants or along the edges of corridors and
stream-side vegetation. Not only will these actions help to protect the vegetation but they
may provide additional habitat for plants and animals. If the buffering vegetation is part
of a commercial timber or fodder enterprise it must be recognised that the buffering role
will diminish following harvesting as will the suitability of that vegetation for habitat.
Vegetation that is not intended for harvest should be used preferentially in areas
adjacent to remnant vegetation. Where ever possible, locally indigenous species should
be used and particular care should be taken to ensure that any non-indigenous species
used are not potential weeds.
Management within remnants
General enhancement principles can be useful for providing management guidelines
where there is a simple relationship between the process to be managed and the
conservation response. For example, if increasing levels of grazing have increasing
impacts on biodiversity, then it is necessary to reduce grazing intensity in order to
increase the conservation value of a remnant. However, many relationships are not so
straightforward. Disturbances such as fire can be detrimental at both low and high
frequencies or intensities, but may be beneficial at some intermediate level. In addition,
the nature of the relationship may differ between ecosystems and even between
communities and species within an ecosystem. A precautionary approach for managing
a process such as fire, for which the outcomes are unpredictable, may be to ensure that
any management regime is highly variable. This could be achieved by varying the interfire
intervals, the intensity of sequential burns, the season in which fire is initiated, and
the area over which fires are allowed to burn. However, without a knowledge of the
responses of the biota to fire, there will be no clear basis for specifying the appropriate
values for any of these variables. Effective management will therefore require a
knowledge of how particular species or communities respond to these processes.
Species composition
Many of the principles described above relate primarily to the spatial and structural
attributes of an ecosystem and could, to some extent, be met by using any particular
combination of plant species provided that they provided spatially and structurally
diverse habitat. However, the use of non-local species will reduce the chances of
particular critical resource needs being met and will increase the risk of creating new

problems through the introduction of potential weeds. Emphasis should therefore be
placed on using locally indigenous plant species.
The limits of general enhancement
Because the above ecological principles provide only general guidelines, it is not
possible to specify the magnitude of the response required, or to predict exactly what
contribution those guidelines will make to the maintenance of biodiversity. They indicate
that remnants should be bigger and more diverse, but cannot indicate how much bigger
or how patchy. Similarly, they cannot tell us which species (or even how many species)
will be retained or lost. They simply tell us that some designs or management practices
are likely to result in the retention of more species than will others. While such an
approach clearly has serious limitations, situations will arise where there is an urgent
need to act and the requirements of the biota in the area to be managed are not known.
Under these circumstances there will be little choice but to use such guidelines.

2.5 Strategic enhancement: using focal species to define landscape requirements for nature
conservation.
• 2.5.1 Species or pattern and process as the basis for conservation planning?
• 2.5.2 Identifying the focal groups
• 2.5.3 Selecting the focal species
• 2.5.4 Applying the focal species approach to Wallatin Creek catchment
Strategic enhancement approaches generally aim to conserve locally indigenous species in their natural
habitat. In order to achieve this, it is necessary to determine the spatial, compositional and functional
characteristics required to meet the needs of the biota that we wish to conserve.
If these characteristics are currently inadequate then a strategic approach will require the reconstruction of
habitat and changes to management practices in order to provide the minimum requirements of the plants
and animals that are at risk. This section presents a procedure for addressing the needs of the species
which occur in any given landscape. This procedure is then applied to the Wallatin Creek Catchment. While
the reintroduction of species was not considered for the study area, the issue of reintroductions is
considered briefly in Section 2.6.
When assessing the capacity of a landscape to retain its biota we need to consider two aspects of that
landscape which I term 'landscape adequacy' and 'landscape viability'. An adequate landscape is one which
contains all of the necessary resources to meet the immediate needs of the species present in that
landscape. These resources, however, may not be available in sufficient quantities or appropriate
configurations to support enough individuals to ensure persistence in the face of natural or anthropogenic
catastrophes, or under conditions of demographic, environmental, stochastic, and genetic variability (Shaffer
1987). A viable landscape, on the other hand, is one that not only contains the resources required to meet
the immediate needs of the individuals present, but can also support sufficient numbers such that the
populations of the species present are able to persist for some specified period of time.
The importance of this distinction became apparent when attempting to design a landscape to protect
conservation values at a scale as small as the Wallatin sub-catchment (26000 ha). Management at the subcatchment
scale could potentially provide all of the resources required by the species that occurred there,
but could not provide them in sufficient quantities to support viable populations. Even if the whole catchment
was revegetated, it would be unlikely to support viable populations of naturally uncommon species. Under
these circumstances, failure of the biota to persist in the catchment would not be due to the quality of the
catchment, but due to the poor quality of adjoining catchments. The catchment would be 'adequate' but not
'viable'.
Clearly, if we are to maintain the biological diversity of our production landscapes, it will be necessary to
firstly define the composition, quantities and configuration of landscape elements that must be present to
meet the immediate needs of the flora and fauna (landscape adequacy) and secondly, to define the area
over which this solution needs to be implemented to support viable populations of the species present
(landscape viability). Achievement of the latter will require a regional approach to conservation planning. In
the following sections, I present a procedure for defining landscape adequacy and discuss the implications
of this approach for defining landscape viability. Appendix 1 considers the implications of applying the
approach presented here at a regional scale.
2.5.1 Species or pattern and process as the basis for conservation planning?
There has been considerable debate in the ecological literature about whether the requirements of single
species should serve as the basis for determining landscape adequacy, or whether the analysis of
landscape pattern and process should underpin conservation planning (Franklin 1993; Hansen et al. 1993;
Orians 1993; Franklin 1994; Hobbs 1994; Tracy & Brussard 1994). Species-based approaches have taken
the form of either single-species studies, often targeted at rare or vulnerable species, or the study of groups

of species which are considered to represent components of biodiversity (SoulÚ & Wilcox 1980; Simberloff
1988; Wilson & Peter 1988; Pimm & Gilpin 1989; Brussard 1991; Kohm 1991). Species-based approaches
have been criticised on the grounds that they do not provide whole landscape solutions to conservation
problems; that they cannot be conducted at a rate sufficient to deal with the urgency of the threats; and that
they consume a disproportionate amount of conservation funding (Franklin 1993; Hobbs 1994; Walker
1995). Consequently, critics of single-species studies are calling for approaches that consider higher levels
of organisation such as ecosystems and landscapes (Noss 1983; Noss & Harris 1986; Noss 1987;
Gosselink et al. 1990; Dyer & Holland 1991; Salwasser 1991; Franklin 1993; Hobbs 1994). These
alternative approaches are based on the recognition that species requirements are not independent of the
landscape in which they occur and that landscape mosaics strongly influence the long-term viability of
species within those landscapes (Janzen 1983; Newmark 1985; Saunders et al. 1991; Anglestam 1992;
Hobbs 1993, 1994).
While approaches that consider pattern and processes at a landscape scale help to identify the elements
that need to be present in a landscape, they are unable to define the appropriate quantity and distribution of
those elements. Such approaches have tended, by and large, to be descriptive. While they can identify
relationships between landscape patterns and community measures such as species diversity or species
richness, they are unable to define the composition, configuration and quantity of landscape features
required for a landscape to retain its biota. Nor are they able to specify which species will be retained and
which will be lost. In other words, they may be useful for addressing a general enhancement objective, as
discussed in the previous section, but they are inadequate for defining the characteristics necessary to
ensure the persistence of all species in a nominated location.
Ultimately, questions such as what type of pattern is required in a landscape, or at what rate a given
process should proceed, cannot be answered without reference to the needs of the species in that
landscape. Given this, it is clear that we cannot ignore the requirements of species if we wish to define the
characteristics of a landscape which will ensure their retention. The challenge then is to find an efficient
means of meeting the needs of all species without studying each one individually. In order to overcome this
dilemma, proponents of single-species studies have developed the concept of 'umbrella' species (Murphy &
Wilcox 1986; Noss 1990; Cutler 1991; Ryti 1992; Hanley 1993; Launer & Murphy 1994; Williams & Gaston
1994). These are species whose requirements for persistence are believed to encapsulate those of an array
of additional species.
The attractiveness of umbrella species to land managers is obvious. If it is indeed possible to manage a
whole community or ecosystem by focusing on the needs of one or a few species, then the seemingly
intractable problem of considering the needs of all species is resolved. Species as diverse as owls (Franklin
1994; Kavanagh 1991), desert tortoises (Tracy & Brussard 1994), black-tailed deer (Hanley 1993), gliding
possums (Kavanagh 1991) and butterflies (Launer & Murphy 1994) have been proposed to serve an
umbrella function for the ecosystems in which they occur. However, given that the majority of species within
an ecosystem have widely differing habitat requirements, it seems unlikely that any single species could
serve as an umbrella for all other species. As Franklin (1994) points out, landscapes designed and
managed around the needs of single species may fail to capture other critical elements of the ecosystems in
which they occur. It would therefore appear that if the concept of umbrella species is to be useful, it will be
necessary to search for multi-species approaches which identify a set of species whose spatial,
compositional and functional requirements encompass those of all other species in the region.
In the following sections, I present a new method for selecting, from the total pool of species in a landscape,
a subset of 'focal species' (Lambeck 1997) whose requirements for persistence define the attributes that
must be present in a landscape if it is to meet the needs of the remaining biota. The approach extends the
umbrella species concept by clearly linking the species that are declining with the threats that are causing
that decline. A suite of species that are considered most sensitive to each of the different threats are
identified and the requirements of each of these species is used to define the landscape attributes and
management regimes that are needed to ameliorate the perceived threats. Area-limited species are used to
define the minimum acceptable size of the patch types they occupy, dispersal-limited species define
configuration and connectivity, resource-limited species define habitat composition, and 'process-limited'

species define the appropriate management regimes for threatening processes. The needs of these focal
species can be used to develop explicit guidelines regarding the composition, quantity and configuration of
patch types required in the landscape and the management regimes that must be applied to the resulting
design. The procedure for defining these focal species firstly requires identification of focal groups, the
members of which are vulnerable to the same threats. The most demanding or most sensitive species from
within each group are then selected. These become the focal species whose requirements define the limits
within which the perceived threats must be managed.
2.5.2 Identifying the focal groups
In order to select the focal species, it is necessary to firstly identify the processes responsible for declining
population sizes. Species which are considered susceptible to similar threatening processes are grouped
and then, for each threat, the species that requires the most comprehensive response is identified. In the
fragmented agricultural landscapes of Western Australia the major threats have been identified as the loss
and fragmentation of habitat; the loss of critical resources; habitat degradation due to stock grazing and
weed invasion; and inappropriate rates and intensities of ecosystem processes such as fire, nutrient cycling
and predation (Hobbs et al., 1993). Grazing pressure from rabbits or the presence of invasive diseases such
as Phytophthora may also be threats in some areas. Figure 6 outlines the sequence of decisions made,
firstly to identify groups of species whose vulnerability is attributable to common threats and subsequently to
identify those species whose requirements for mitigating the threat encompass those of the other species.
The outcome of this selection process is a suite of 'focal' species whose requirements for management or
habitat reconstruction encapsulate the needs of all other species.
At risk or secure?
The first dichotomy in Figure 6 differentiates between those species considered secure in the current
landscape and those expected to be lost in the absence of action. Species considered secure are removed
from the selection process; if the status of a species is in doubt, it should remain in the analysis. Secure
species may re-enter the analysis subsequently if their presence is identified as being the cause of
vulnerability of some other species.

Figure 6. Decision tree for developing landscape designs and management guidelines based on the focal
species. The limiting processes identified in the figure are simply illustrative. These may vary from one
landscape to another. The procedure for applying this approach is described in the text. (Sections 2.5.2 - 2.5.3)
Reconstruction or within-remnant management?
The second decision differentiates between species whose persistence in the landscape depends on some
form of habitat reconstruction and species which would be able to persist in the current landscape provided
biophysical processes were managed in a different way. This dichotomy reflects a distinction between the
relative importance of pattern and process. Generally, species which could persist under the current
landscape configuration if the landscape was managed differently are those sensitive to the rates of
particular processes, or to changes in the intensity and frequency of those processes. In Australian
agricultural landscapes these processes include altered fire regimes, predation by introduced foxes and
cats, grazing of native vegetation by stock and other herbivores, and competition between native plants and
exotic weeds. Species threatened by these changes can be considered 'process-limited'. The remaining
species are those primarily constrained by landscape patterns which limit the availability of suitable habitat
or critical resources. Such species will rely on landscape reconstruction to meet their needs.
Species requiring reconstruction
Species identified as requiring landscape reconstruction are further assessed to determine whether they are
limited by (i) a shortage of critical resources, (ii) an inability to move between suitable habitat patches, or (iii)
insufficient habitat to meet their resource needs.

i) Resource-limited species
For resource-limited species, the number of individuals that a region can support is determined by the
carrying capacity at the time of lowest resource availability. Species limited by a resource bottleneck may
exhibit a significant population response to the enhancement of resources at the time of greatest shortage.
For example, many nectarivorous birds utilise a sequence of nectar sources throughout the year. Depletion
of these resources at any stage in this sequence constrains their population size (Lambeck 1995). A
rehabilitation response targeted at alleviating the bottleneck should increase the local carrying capacity for
nectarivorous species. In such circumstances, a relatively local strategic restoration action may produce a
greater population response than would a major landscape reconstruction if the latter failed to explicitly
address the resource shortage. Species that may commonly be identified as being resource limited include
birds or mammals that require tree hollows for nesting or roosting or species which depend on the
sequential availability of seasonally variable food resources such as nectar or fruit.
ii) Dispersal-limited species
Dispersal-limited species are those for which there are suitable habitat patches which could potentially
support small populations, but the patches are beyond the distance over which individuals can move or are
separated by a matrix that is too hostile to permit movement. If individual populations are too small to be
viable in their own right, the combination of stochastic and anthropogenic impacts can result in rates of local
extinction that exceed rates of recolonisation. Such species will require increased connectivity between
habitat patches either by the provision of corridors or by reducing the 'resistance' of the intervening matrix
(Knaapen et al. 1992). This would require the adoption of land uses or management practices in the matrix
which are less hostile to nature conservation. Species which are most likely to be dispersal limited will be
sedentary habitat specialists which occur in low densities and have low mobility relative to the distances
between suitable habitat patches.
iii) Area-limited species
Area-limited species are those for which the patches of appropriate habitat are simply too small to support a
breeding pair or, in the case of colonial species, a functional social group. Area-limited species are, in
reality, resource limited but are considered in this category if the limiting resource is not obvious or
quantifiable. Habitat patches are therefore used as a surrogate for resources (Hansen et al. 1993) and it is
assumed that there is a minimum patch size of a given quality that will provide sufficient resources to
support a pair or group. In general, species which occur higher in the food chain are more likely to be area
limited than will be species in lower trophic orders. Similarly, species which have a greater reliance on the
maintenance of a cohesive social structure may have greater area requirements than do solitary species as
any given habitat patch will have to support a greater number of social individuals.
Species requiring management of ecosystem processes
While the spatial and compositional characteristics of habitat may not be the primary limiting factor for many
species, they may still be vulnerable because of inappropriate rates or intensities of important ecosystem
processes. Throughout Australia, a number of threatening processes other than land clearance contribute to
a reduction in the diversity of native ecosystems. These include predation by foxes and cats, grazing by
stock and rabbits, invasion by weeds and disease, chemical and nutrient drifts from adjoining farmland, and
inappropriate fire regimes. Species which are limited by processes such as these are grouped into
categories which reflect the range of threats that occur in the region being managed.
2.5.3 Selecting the focal species
After completing the above decision-making process, all species considered at risk will be allocated to at
least one of four major categories: area-limited, resource-limited, dispersal-limited, or process-limited. The
process-limited group will be further subdivided according to the number of processes that require

management. Some species may occur in more than one category. The species in each category are then
ranked in order of their requirements for area, connectivity, resources or management, respectively.
Area requirements
Area-limited species are categorised on the basis of the dominant patch types that they utilise. For each
patch type, species are ranked according to the size of the smallest patch in which they are observed to
occur. The species with the largest minimum occupied area for a particular patch type is identified as the
focal species for that patch type. Its spatial requirements are then used to define the minimum suitable size
for that patch type. Any patch large enough to support a breeding pair, or social group, of the focal species
is assumed to be large enough to support individuals of all other species that utilise that patch type. For any
given region there will be as many focal species which define minimum patch area as there are patch types.
A difficulty which arises when attempting to determine adequate patch area is the definition of what
constitutes a patch. It is widely recognised that patchiness in a landscape is hierarchical, with patches
defined at one scale, being subdividable into smaller patches when examined at a finer scale. In agricultural
landscapes, the dominant vegetation associations will generally provide a useful basis for defining
patchiness. These associations tend to have dominant species and characteristic structural attributes which
can be easily recognised by land managers. These associations also tend to correlate with soil types and
other landform features with which many land-holders are familiar. Consequently they provide useful
'building blocks' for landscape design and reconstruction. In some circumstances vulnerable species may be
responding to a finer landscape grain than that of the vegetation association. Attempts to reconstruct
vegetation formations should therefore ensure that this finer scale variability is incorporated where
appropriate.
Connectivity requirements
For the majority of species, dispersal is one of the least understood aspects of their ecology. The approach
taken to define the characteristics necessary for dispersal would ideally follow that used in the previous
section to define area requirements. Species would be ranked according to the minimum width, length and
structural requirements of the connecting vegetation through which they are known to move. The species
with the greatest need for wide corridors or with the least inclination to move along corridors would become
the focal species for defining corridor width and length, respectively. Similarly, species with the most
demanding structural requirements would be used to define the structural attributes of the connecting
vegetation.
Because dispersal data are rarely available, presence/absence data can be used to determine the interpatch
distance beyond which seemingly suitable habitat is unoccupied. For example, Cale (unpublished
data) found for a range of bird species that seemingly suitable patches remained unoccupied if they were
too isolated. For each patch type in a landscape, the minimum acceptable distance between patches can be
defined by the species with the shortest distance beyond which an otherwise suitable patch is not occupied.
Resource requirements
Resource-limited species are those for which critical resources can be identified and be shown to limit the
carrying capacity of consumer species in the region. Where a number of species utilise the same resource
base, the resource must be increased to a level sufficient to meet the needs of the least abundant consumer
(Lambeck 1995). This species becomes the focal species for defining the appropriate level of that resource.
Resources that are commonly limiting include nest hollows for birds and mammals, nectar supplies for birds,
insects and some small mammals such as the Honey Possum (Tarsipes rostratus), or appropriate
microhabitats for invertebrates such as Mygalomorph spiders (Main 1987).

Management requirements
Having categorised species according to their needs for management of threatening processes they are
then ranked in terms of their vulnerability to those threats. Those species most vulnerable to, or most
dependent upon a given process become the focal species for defining the intensity, rate or frequency at
which that process should be managed. For example, the species most deleteriously affected by weed
invasion will define the level of weed control required and the species most vulnerable to feral predators will
define the appropriate level of predator control.
2.5.4 Applying the focal species approach to Wallatin Creek Catchment
The application of the focal species approach would ideally be based on a comprehensive survey of the
area to be studied and a knowledge of the status and requirements of all species that occurred there.
Obviously this level of knowledge will never be available for any location. Consequently, the analysis
presented here is based on a combination of survey data and 'expert opinion'. This opinion was obtained
through a series of small workshops or through consultation with individuals who had a knowledge of the
area and expertise in particular taxonomic groups which occurred in the region. The participants were asked
to identify species they considered at risk and to identify the likely threats. This process necessarily
introduces an element of subjectivity to the analysis given that the expertise of the participants is usually
derived from a combination of rigorous study, anecdotal observations and interpretation of observations
made elsewhere.
Identifying the focal groups
At risk or secure?
Members of the expert panels were asked to nominate species they considered potentially at risk in the
central wheatbelt if no management action was taken. This analysis identified 19 species of birds, 2 species
of mammals, 5 reptiles, 7 frogs, 12 invertebrates and 11 plant species that were potentially vulnerable (see
Appendix 2 for a list of vulnerable species and their perceived threats). These species were then
categorised according to the factors that limit their distribution and abundance. Factors considered relevant
for the study area were (i) insufficient habitat (ii) habitat isolation (iii) insufficient resources (iv) feral
predators (v) weeds (vi) grazing by stock and (vii) inappropriate fire regimes.
Reconstruction or management?
The categories identified above can be separated into two broad groups which reflect the need for habitat
reconstruction, on the one hand, or the management of threatening process on the other. Species of birds,
reptiles and invertebrates were identified as requiring both habitat reconstruction and management of
ecosystem processes, whereas mammals and plants primarily required more appropriate management of
the habitat that is available. For example, predation by foxes and cats was perceived to be the main factor
limiting mammal distribution and abundance, while stock grazing and weeds were the greatest potential
threats to the vulnerable plants.
Species requiring habitat reconstruction
Species requiring habitat reconstruction were partitioned according to whether they are (i) area limited, (ii)
dispersal limited or (iii) resource limited.
The species falling into each of these categories are listed in Appendix 3.
i) Area-limited species
Ten species of birds were considered potentially area limited (Table 1). These species were grouped into
three broad habitat categories (woodland, shrubland/mallee, and heathland) and then ranked in order of the
minimum patch size in which they were known to occur.

The species with the greatest area requirement for each patch type was considered the 'focal' species for
that patch type. For woodlands, both the sittella and jacky winter had similar area requirements and hence
were equally appropriate as focal species. Western yellow robins and crested bellbirds had similar minimum
area requirements for the shrubland/mallee habitat type but these were marginally less than those of the shy
hylacola which was only recorded from five patches, none of which were less than 25 ha. Very little is known
about the field wren (Calamanthus fuliginosus). It has only been recorded on the largest remnant in the
catchment (1100ha) in a heathland patch that is approximately 25 ha in size. It is not possible to ascertain
whether its presence on this remnant is attributable to the size of the remnant or to the size of the habitat
patch as there are very few other remnants with equivalent amounts of heath. In the absence of better
information this figure of 25 ha will be used to specify the minimum suitable area of heathland.
Several invertebrate species were also identified initially as being area limited. These included
Mygalomorph spiders and several scorpion species (Main 1987; Smith 1995). However, the primary spatial
limitation for these species was considered to be amount of suitable microhabitat. This suggests that the
partitioning of the landscape at the resolution of dominant vegetation types is too coarse for these types of
species. Under these circumstances it was considered that these species be viewed as being 'resourcelimited'
with microhabitat being the limiting resource. These species will therefore be considered further in
the section on resource-limited species.
While vulnerable plants were generally considered to be threatened by the impacts of stock or inappropriate
fire regimes they could also be considered to be area limited if they require populations to have sufficient
numbers of individuals to maintain genetic diversity. While this type of information is rarely available, it has
been suggested that more than 500 individual Eucalyptus regnans are required to ensure adequate genetic
diversity for that species (Ashton 1975). Two tree species which are potentially at risk in the catchment are
salmon gums (E. salmonophloia) and gimlets (E. salubris). These species occur at average densities of 96
and 102 trees per hactare respectively (Lambeck 1995; Sarre et al. 1995). This suggests that the minimum
area required to support 500 individuals would be 5.2 ha for salmon gums and 4.9 ha for gimlets. These
area requirements are considerably less than those of the most demanding woodland birds.
On the basis of the requirements of species that are area-limited, the patch sizes considered necessary to
support the most demanding species are (i)greater than 23 ha for woodland, (ii) greater than 25 ha for
shrubland/mallee and (iii) greater than 25 ha for heathland. It was assumed that any patch big enough to be
occupied by these focal species would also be large enough to support all other species that use those
patch types.
Table 1. Area-limited bird species ranked in order of the minimum patch size in which they regularly occur.
Woodland Shrubland/Mallee Shrub/Heathland
Species Area (ha) Species Area (ha) Species Area (ha)
Jacky Winter 23 Shy hylacola 25 Field wren 25
Sittella 23 Crested bellbird 16 Blue-breated fair-wren 3
White-eared honeyeater 4 Western yellow robin 22
ii) Dispersal-limited species
White-eared honeyeater 13
Southern scrub robin 22
Inland thornbill 5
Information on dispersal is available for a limited number of species in the Kellerberrin district. Most of this
information is for movements of common, highly mobile vertebrate species (eg, Arnold et al. 1991; Saunders
& de Rebeira 1991). Very little is known about the less common species which are the most vulnerable. The
bird species considered most likely to be dispersal limited include shy hylacolas and field wrens which

appear to prefer the interior of habitat patches. Both of these species occur in only a few remnants in the
study area. Consequently, virtually nothing is known about their requirements for movement.
In the absence of direct measures of movement, estimates of potential dispersal ability can be obtained by
measuring the distances between suitable habitat patches occupied by dispersal-limited species. For a
given species the inter-patch distance beyond which seemingly suitable habitat remains unoccupied can
provide a rough estimate of distances beyond which that species will not move. The best information
available for the study area comes from Cale (1994) who identified the minimum patch size occupied by
particular bird species and the inter-patch distances beyond which they are unlikely to occupy an otherwise
suitable patch (eg. Figure 7).
Estimates of inter-patch distances between occupied and unoccupied sites in the study area were available
for only a handful of bird species. Of these, the species which displayed the greatest dispersal limitation was
the western yellow robin. As can be seen from Figure 7, this species was not found in remnants more than
two kilometres from the nearest occupied patch. While this suggests that it would not be a good idea to
locate a patch more than two kilometres from an existing one, it does not guarantee that patches less than
two kilometres apart will be occupied. In fact, some patches that are large enough to be occupied and are
less than two kilometres from the next nearest patch, do not have western yellow robins on them.
Of the species examined, the yellow robin would appear to be the focal species for defining the minimum
inter-patch distance. However, because no information is available for field wrens and hylacolas, any
recommendation for a 2 km inter-patch distance should be treated with caution. It should not be assumed
that remnants that occur within 2 km are redundant and can therefore be cleared.
Similarly, little is known about the dispersal characteristics of small vertebrates or of invertebrates.
Individuals of many of these groups are unlikely to disperse between remnants in an agricultural landscape.
The maintenance of population continuity for these species will only be possible if the linear vegetation
along road verges, waterways and fence lines provides habitat which can support resident populations.
Under these circumstances, this linear vegetation should be viewed as linear habitat and should be
designed to meet the habitat needs of dispersal-limited species.
Scorpions (Smith 1995) and some spiders (Main 1987) were identified as potentially dispersal-limited
species. Of the scorpions, Cercophonius michaelseni appears to be a very slow disperser. This species has
not been recorded for a period of 5 years following a fire, in spite of the fact that adjoining populations in
unburnt habitat occur less than 20m away. This failure to disperse could also be attributable to the fact that
the vegetation has only partly regenerated over this period and may not yet constitute suitable habitat.
Some trapdoor spiders are known to have limited dispersion powers. Anidiops, Idiosoma and Teyl sp. B
tend to aggregate around the matriarchal nest if space is available (Main 1987) although their capacity to
survive and disperse in the absence of such space is not known.
Because of our limited knowledge of the dispersal requirements of most of the species in the landscape, it is
not possible to make unequivocal statements about the characteristics of connecting vegetation.
Recommendations will therefore be based on a combination of data, where it is available, and general
principles, where it is not.
For birds, it appears that remnants should be no more than 2km apart and that connecting vegetation
should be sufficiently wide for interior habitat specialists to utilise them. In the case of Hylacolas and Field
Wrens this may require corridor widths greater than 50m. It is unlikely that the implementation of such a
recommendation for all linkages between all remnants would be feasible in the short term. Hence it is
suggested that linear vegetation of this quality should be preferentially established between sites occupied
by these species and the nearest suitable habitat patches.
If linear vegetation is to act as habitat for small mammals, reptiles and invertebrates it must be ungrazed,
relatively free of weeds and have a range of habitat types. Fencing is therefore a primary consideration. In
order to reduce edge effects linear vegetation should have a minimum width of 30 m if it consists of heath

species, with the width increasing as the plant density decreases. Where woodland species are used with
little understory, linear vegetation should ideally exceed 60 m. The linking vegetation should contain the
structural and compositional characteristics of the local vegetation types. Where woodland species are
used, substantial clumps of middle canopy and understory species should be included.
These recommendations should not be taken to imply that strips of vegetation less than 30 m have no
value. Narrow strips can be used by a large number of species but they will simply have a lower probability
of meeting the needs of the more dispersal-limited species or of species that use the linear vegetation as
habitat.
Figure 7. Pattern of shrubland patch occupancy by western yellow robins (Eopsaltria griseogularis). This
species was never found in patches of scrubland less than 20 ha or in patches more than 2km from the
nearest occupied patch. Data from Cale (1994).
iii) Resource-limited species
The only species in the region for which resources have been demonstrated to be limiting are honeyeaters
(Lambeck 1995) and predatory invertebrates (Main 1987; Smith 1995). Competition for tree hollows
between various parrot species may have contributed to the demise of regent parrots and western rosellas
but, because these species no longer occur in the study area, the availability of tree hollows appears to no
longer be a threat to the remaining species.
Populations of honeyeaters are restricted by the depletion of nectar resources in late summer and autumn.
At this time, many of the nectar-dependent species leave the area, while the more generalist species
become largely insectivorous. The capacity of the remnants to support honeyeaters at this time is lower than
their carrying capacity when nectar is abundant.
There are two possible responses to this problem. One is to consider these honeyeater species nectar
limited and therefore increase the availability of nectar-producing plants at the limiting times of the year. The
alternative is to consider them to be insectivores at this time and increase the amount of habitat available on
the assumption that more or bigger patches will provide more insects and therefore support more
individuals. The former response will be the most cost-effective if it is possible to provide a rich food source
in a relatively small area at the appropriate time.
If the alternative response of increasing the habitat area is being considered, these species should be
included in the area-limited analysis described above.
In the Wallatin case study, it was decided that the most efficient response would be to enhance access to
nectar over summer by increasing the availability of summer and autumn flowering species such as mallee
eucalypts, the acorn banksia (Banksia prionotes) and the Epacrid, Astroloma serratifolium. The acorn

anksia produces copious amounts of nectar in late summer and autumn. In the Wallatin Catchment it is
confined to a single small patch and a few scattered trees (Lambeck & Saunders 1993; Lambeck 1995).
As discussed previously, some spiders and scorpions were considered to be area-limited, but the limitation
was at the level of micro-habitat rather than the dominant vegetation associations. We can therefore
consider these microhabitats to be a limiting resource and must attempt to ensure that they are
appropriately represented in any reconstructed vegetation. It is therefore assumed that an area of a given
vegetation type that is large enough for vulnerable bird species will also be of sufficient size for any
invertebrates that occupy that patch type, provided the appropriate microhabitat requirements are present.
Of the vulnerable trapdoor spiders Idiosoma nigrum requires stable litter mats for burrow sites in York
gum/Jam wattle (Eucalyptus loxophleba / Acacia acuminata) woodland; Aganippe sp. D prefers flood-prone
depressions and flats which support myrtaceous shrub heaths. Teyl sp. B and two species of an
undescribed diplurine genus prefer open patches within the litter matrix (Main 1987).
An important issue to address in future is whether these microhabitat attributes need to be created as part of
the habitat reconstruction, or whether they will develop, over time, of their own accord if degrading
processes such as stock impacts and weed invasion are excluded from the reconstructed patches.
Species requiring management
Species requiring management were categorised according to the processes that were considered
responsible for their vulnerability (Appendix 3). Undoubtably a primary threat to nature conservation in the
wheatbelt of Western Australia is that posed by an inexorably rising saline water table which threatens all
remnant vegetation in the low-lying parts of the landscape. However, it is recognised that salinity is also the
major threat to agricultural production and hence its control will be an integral part of an agricultural
management strategy. Salinity is not considered to be a threat to biodiversity in this analysis, purely
because it is assumed that it will be managed to protect agricultural production. Guidelines for managing
hydrology are therefore presented in Chapter 3. The main threatening processes in the Wallatin Creek
Catchment were therefore considered to be (i) grazing by stock, (ii) predation by foxes and cats, (iii)
inappropriate fire regimes, and (iv) invasion of remnant vegetation by weeds. For each of these processes,
species were ranked in terms of their vulnerability to or dependence upon the process.
i) Grazing
While no specific data are available to assess the relative vulnerability of the plant species threatened by
grazing, it is apparent that a number of species are sufficiently uncommon that they should be protected
from any level of grazing. The poor recruitment of woodland species, including Salmon gums, Gimlets and
Banksias, in the presence of stock suggests that there is no level of regular grazing by stock that is
acceptable in woodlands. If occasional grazing is to be allowed in woodlands, it must be at intervals that are
sufficiently long for germination to occur and for saplings to grow to a stage where they are resilient to stock.
Heath and shrubland communities have also been severely degraded by stock, resulting in marked changes
in both plant and animal species diversity (Hobbs & Atkins 1988; Abensperg-Traun 1992; Scougall et al.
1993). These changes are a consequence of direct grazing of plants, compaction of soil, and destruction of
microhabitats. Given these impacts, the precautionary approach is to exclude stock entirely from remnant
vegetation. Because remnant vegetation has been traditionally used to provide shelter for stock this strategy
will require the planting of new woody perennial species in other parts of the landscape to provide
alternative shelter.
Rabbits have also been identified as limiting the recruitment of many plant species in other parts of Australia
(Lange 1983; Foran. 1985). However, their numbers in the study area are currently not high enough to
present a major threat. Given their potential to become a threat if populations increase, it would be
advisable to maintain a program of baiting and warren ripping.

ii) Feral predators
Four species of birds, two mammal species and numerous reptile species were considered potentially
vulnerable to predation by foxes and cats (Appendix 3). These species were selected, not because of
empirical evidence of predation, but because their behaviour increases their vulnerability. The four bird
species (field wren, shy hylacola, bush stone curlew, and southern scrub robin) are either terrestrial and
nest on the ground, or occupy low shrub habitat. The western brush wallaby (Macropus irma) is currently
known from only one remnant in the study area, where it has been sighted in patches of dense
Allocasuarina adjoining open woodland. This species is endemic to the south-west of Western Australia and
anecdotal observations suggest that it may be declining in both distribution and abundance in the wheatbelt.
The ash-grey mouse (Pseudomys albocinereus) is also uncommon, having been captured in only a few
locations where deep yellow sands occur. This species has been found to be a common prey item of cats
elsewhere in its distribution (Risbey 1997). While it is clear that both cats and foxes prey upon reptiles, there
is currently no information to suggest that any particular reptile species are at risk from predation. However,
there is the potential that species which are uncommon for reasons other than predation may be vulnerable
if predators are able to drive small isolated populations to local extinction. The mountain devil (Moloch
horridus) for example, is uncommon in the study area and local populations may be susceptible to
predation.
The species in this study that appears most obviously to be at risk from fox predation is the western brush
wallaby. Studies of feral predators elsewhere have indicated that even very low predator numbers can have
significant impacts on mammal survival (Spencer 1991; Horsup & Evans 1993). This suggests that it will be
necessary to remove virtually all foxes from Durokoppin Nature Reserve in which the remaining wallabies
occur. If this reserve is unable to support viable populations of this species even in the absence of
predation, it will be necessary to extend the distribution of this species beyond the reserve. This will require
the control of feral predators over a much larger area. The brush wallaby can therefore be considered the
focal species for fox control as its requirements for predator management will encompass the needs of other
species that are also vulnerable to fox predation. Initial baiting of the reserve should be intensive. In a fox
control program at a number of granite outcrops to the south of Kellerberrin, the state conservation agency
(CALM) laid fowl eggs injected with 1080 (sodium monofluoroacetate) at 50m intervals around the perimeter
of the reserves (Kinnear et al. 1988). This study found that foxes killed by early baiting were quickly
replaced by others moving into the area and hence an intensive baiting regime was required.
It may be possible that baiting intensity could be reduced if baits are distributed more widely throughout the
landscape. However, such a strategy has not been adopted in agricultural regions and its feasibility has not
been assessed. Whatever strategy is adopted, baiting must be implemented in a coordinated manner to
ensure that all farmers are aware of the presence of baits and can take the necessary actions to protect
farm dogs.
Little is known about the impact of predation by cats in this area, however there is anecdotal evidence to
suggest that cat numbers may increase as fox numbers decline (J. Short unpublished data). Such an
increase may potentially impact on small mammals such as the ash-grey mouse. In the absence of
information about the risk that cats pose, a precautionary approach would be to adopt a strategy which
aimed for simultaneous control of cats and foxes. There is still some debate about which methods are most
effective for cat control and in some instances it appears that the use of a range of bait types and trapping
methods may be most effective (J. Short pers. comm.). If baiting or trapping for either cats or foxes is to be
undertaken it is essential to consult the appropriate management authorities to ensure that relevant
regulations are complied with.
iii) Fire
Limited information is available about the importance of fire in the maintenance of plant or animal
populations in the catchment. Animals identified as being potentially vulnerable to fire were reptiles and
some species of scorpions (Smith 1995) and spiders (Main 1987). While the majority of species in the

egion would be vulnerable to extreme fire frequencies, the assessment of risk was based on the current
regime of very infrequent fires.
Reptile species which seem to be vulnerable to single fire events include an arboreal gecko (Diplodactylus
spinigerus), an agamid lizard (Ctenophorus reticulatus) which occurs primarily in mallee communities, and a
skink (Ctenotis pantherinus) which occupies heath/shrubland. Both Ctenophorus and Ctenotis disappeared
from study sites following experimental fires and population recovery was dependent upon recruitment from
adjoining unburnt areas (G. T. Smith unpublished data). A period of approximately 4 - 5 years was required
before mature adults of these species reappeared after fire in remnants where there was adjoining unburnt
habitat. While data are not available for D. spinigerous, its arboreal habit suggests that it would also be
sensitive to fire.
The scorpion, Cercophonius michaelseni also disappeared from its highly flammable heath habitat following
an experimental fire. Individuals of this species were able to survive the fire but were unable to persist in
subsequent seasons (Smith 1995). Given the extremely slow rate of dispersal by C. michaelseni back into
the burnt habitat, it would appear that long inter-fire intervals would be required for this species to recolonise
areas from which it had become locally extinct. The trapdoor spider Anidiops villosus showed a similar
response with adults surviving the fire itself, but suffering progressive mortality and no recruitment following
the fire due to inadequate shade and litter and possibly also due to reduced prey and increased vulnerability
to predation (Main 1995). This species is also likely to be a slow disperser and hence many years may be
required for recolonisation to occur, provided there are adjacent unaffected populations.
The persistence of many plant species may also depend on the frequency and intensity of fire. An important
factor which determines the impact of fire on plant population viability is the life history attributes of species
affected. Species which recover by re-sprouting may be more resilient to frequent fires or high intensity
burns. Of the potentially vulnerable plant species in the study area, Grevillea dryandroides and Halgania
tomentosa re-sprout after fire and hence may only be at risk if the fire frequency is sufficient to deplete
below ground nutrient stores. By contrast, fire-sensitive species which rely on the germination of seeds
stored in the canopy or in the soil will be particularly vulnerable to fire frequencies which do not allow
sufficient time for post-fire recruits to reach maturity and replenish the seed bank. Verticordia hauganii is an
obligate re-seeder and may be vulnerable to a sequence of burns at intervals less than the two to three
years following germination that this species may require to set seed. Fire ephemerals which set seed within
a short time of a fire are unlikely to be at risk as they will replenish seed bank before the area has sufficient
vegetation to support a subsequent burn. Flannel flowers (Actinotus superbus) are post fire annuals which
only appear for a few years following a burn (Hester & Hobbs 1992). Their seeds presumably remain viable
in the soil seed bank for many decades as they will readily appear following fire in areas that have not been
burnt for long periods of time.
Other fire-sensitive species may take longer to reach maturity and therefore may be more vulnerable to
frequent fires. Grevillea eriostachya, for example, may be killed by an intense burn and may not flower for
up to 4 years post-fire. For this species, several more years may be required before an adequate soil seed
bank can be established. This would suggest that a minimum inter-fire interval of at least 5-6 years may be
required to ensure that local populations of this species were not threatened. Numerous Banksia species
are also killed by fire and may take several years before they produce flowers and accumulate sufficient
seeds to restore the population following fire (Gill & McMahon 1986). Where such species occur their
requirements should be used to specify the minimum appropriate inter-fire period.
Short-lived, fire-dependent species that do not have a long lasting seed bank should be used to identify the
maximum acceptable inter-fire period. For example, Banksia coccinea, a species which occurs in the
southern coastal areas of Western Australia requires fire to stimulate seed release and germination but
senesces after approximately 15 years. This species therefore requires a fire frequency of less than that
time interval. The only Banksia species in the study area is Banksia prionotes which senesces after
approximately 40 years. However, this is one of the few Banksias which readily recruits in the absence of
fire.

For the Kellerberrin area, there is no clear evidence to suggest that a frequent fire regime is necessary or
appropriate. Some plant species appear to be potentially vulnerable if fire intervals are less than 5-6 years.
Some animal species, on the other hand appear to require much longer intervals between fires to allow for
the regeneration of appropriate habitat and for their slow rates of dispersal into those regenerating habitats.
In the absence of evidence to the contrary it is recommended that fire be used as a management tool very
infrequently and only where there is evidence that vegetation communities are senescing. For most of the
vegetation types in the region, this would suggest that there is no need to burn at interval of less than 50
years. It is important to recognise however that the basis for such recommendations is somewhat limited
and that classification of species in terms of their responses to fire would be an important aid to managers.
Gill and Nicholls (1989) provide a methodology for monitoring fire sensitive plants and point out the
importance of using monitoring as a tool for learning more about the fire responses of the flora.
In areas where there is little or no information about the specific needs of the biota, land managers have
traditionally reverted to general management principles. These principles suggest that the optimal approach
is to ensure a fire regime that is as varied as possible in terms of season, intensity, frequency, patch size
and availability of adjacent unburnt areas for recolonisation. While this type of strategy may be possible in
large areas of continuous habitat where a range of options are available, it will not be possible to adopt such
an approach in small remnants of native vegetation. Many such remnants are too small to apply a spatially
and temporally variable fire regime and a hit-and-miss approach based on general principles will have a
high probability of contributing to the local extinction of some species. If these remnants are to be
considered to be a part of the conservation network, and if fire plays a critical part in the dynamics of the
plants and animals that occupy them, it will be necessary to gain a better understanding of the fire
responses of the species present and to develop fire management strategies which meet their needs.
iv) Weeds
The presence of weeds in vegetation remnants can have a significant influence on plant communities by
inhibiting the regeneration of native plant species (Hobbs & Atkins 1991), by changing micro-climatic
conditions and by altering fuel loads and hence fire dynamics. The resulting changes in plant community
composition and structure can subsequently influence animal diversity due to modified habitat
characteristics. In the study area, none of the vulnerable species were considered to be directly threatened
by weed invasion. Consequently no species have been nominated to define appropriate levels of weed
control. In spite of this, it is apparent that weeds have the potential to threaten plant species and
communities and hence should be closely monitored and controlled if they are known to be highly invasive
and where they appear to be threatening native plant communities. Control of weeds is likely to be critically
important when trying to re-establish native vegetation (Panetta & Groves 1990).
2.6 Reintroductions
The second of the strategic enhancement objectives identified in section 2.3 aims not only to retain the
current species in the landscape but also to restore species that previously occurred in the region but are no
longer present.
The reintroduction of species which have disappeared as a result of habitat removal and fragmentation
would ideally be considered in the context of a regional planning approach. Information about the habitat
and resource requirements of the species to be reintroduced will rarely be available for the site where the
reintroduction is to occur and therefore needs to be acquired from places where the target species still
exists. However, it cannot be assumed that the local site characteristics where a species currently occurs
will also be adequate for that species if duplicated elsewhere. For example, a species which has been lost
from the central wheatbelt is the rufous treecreeper, a medium sized, insectivorous, woodland bird.
Examination of its habitat requirements in a location where it does still occur reveals that it occupies a wide
range of woodland patches ranging from large pristine reserves through to small patches of degraded
woodland. Simple observations of habitat use would suggest that small patches of woodland are adequate
for this species. However, there are many such patches in the Kellerberrin area and yet the species has
disappeared.

It appears that this species is able to occupy these small patches only because of the regional context in
which it is found. The presence of large pristine reserves nearby probably provides a source of new recruits
to these smaller remnants. In fact, these small remnants may act as sinks into which individuals disperse,
but from which there is no recruitment. It is also possible that rufous tree-creepers may not be secure in the
areas where they currently occur. Their presence may simply reflect a slower rate of decline due to the less
fragmented landscape than was the case in Kellerberrin.
The important point to be drawn from this example is that it is not possible to simply duplicate, in another
location, the local habitat patterns of an area where the target species still persists and expect them to
deliver the same conservation value. Any translocation of a species between locations will require a detailed
knowledge of the ecology of that species, which takes into account not only the local habitat requirements,
but also the regional context in which that species occurs. Such an analysis was beyond the scope of the
current project but is worthy of further investigation if restoration objectives are to be considered seriously. A
list of species known to have disappeared from the Kellerberrin area is provided in Appendix 4.
2.7 Mixed strategies in the face of partial knowledge
Ideally, any strategy for conserving the biota in a region would be based on an understanding of the
requirements of the flora and fauna. In reality, such complete knowledge will never be available, resulting in
a need to develop strategies based on a combination of quantitative data, expert opinion and general
principles. The dilemma facing managers is that our major concerns are for those elements of the biota that
are at greatest risk. By and large, these tend to be less common species which are less amenable to
rigorous scientific investigation. As a consequence we are forced to rely on expert opinion. While this
opinion is often based on many years of accumulated experience and anecdotal observations by a number
of people, its reliability cannot be easily assessed. General principles on the other hand provide even less
guidance for ensuring that particular species will persist.
If we are to adopt objectives which aim to achieve more than simply improving on our current position in
some unspecified manner, we will need to develop ways of efficiently acquiring the essential knowledge
required for managing particular landscapes. However, in situations where there is an urgent need to act, or
where the resources required for effective management are unlikely to become available, it will be
necessary to increasingly rely on expert opinion and general principles. It must be recognised, however, that
the greater the emphasis on general principles, the less likely we will be to meet the needs of particular
species that are at risk in our agricultural landscapes.
2.8 Summary of design and management recommendations for Wallatin Creek
Because the focal species approach presented here is a novel one, much of the data that are required for
such an analysis were not available for the study area. The approach requires information about the most
sensitive and vulnerable species in the landscape. This information is difficult to obtain because such
species are usually uncommon. Where the required information was not available or insufficiently detailed it
was necessary to draw upon the informed opinion of professional ecologists and knowledgeable locals.
Where even this level of knowledge was unavailable it was necessary to rely on general ecological
principles.
By applying the focal-species procedure it was possible to identify area-limited and dispersal-limited focal
species whose requirements were used to develop estimates of the spatial characteristics of habitat patches
and the maximum acceptable distances between patches. The recommended minimum area for the
dominant patch types are greater than 23 ha for woodland, 25 ha for shrubland/mallee, and 25 ha for
heathland.
Information about corridor characteristics were more difficult to prescribe, but it is clear that for many small
vertebrates and less mobile invertebrates, this linear vegetation must act as habitat in which population
processes must persist, rather than as conduits through which individuals move. On this basis it was

possible to present some tentative recommendations about the preferred lengths and widths of linear
vegetation. The distance between remnants should not exceed two kilometres. Linear vegetation linking
occupied habitat patches of dispersal-limited birds to the nearest suitable patch should be approximately
50m wide. Other linear vegetation, if it is to provide habitat for smaller animals and be resilient to weed
invasion should be greater than 30m wide for heath communities and greater than 60m wide for more open
vegetation types. It was not possible, in the time available for this project, to use the focal species approach
to specify the compositional attributes of linear vegetation. Consequently, it was necessary to draw on
general principles which suggest that the type of species planted should be of local provenance and that
they should be planted on the appropriate soil type. In order to meet the habitat needs of some of the
vulnerable invertebrates and small vertebrates, areas reconstructed with open vegetation types such as
woodland should also contain clumps of denser understorey vegetation.
The requirements of resource-limited species suggest that habitat reconstruction should emphasise plant
species that produce nectar over the summer-autumn period and that reconstructed habitat patches should
contain areas of intact litter and areas in which the canopy is sufficiently closed to provide the more mesic
conditions required by some vulnerable scorpion species.
Recommendations for managing other threatening processes in the catchment include (i) that feral
predators be reduced to the lowest possible densities with an ultimate objective of eradication; (ii) that
grazing by stock of remnant vegetation and reconstructed habitat be stopped (iii) that fire be used as a
management tool only in situations where senescence of the dominant vegetation is apparent and (iv) that
inter-fire periods should exceed 50 years unless additional evidence becomes available to suggest
otherwise. It is important to recognise that these recommendations have been derived specifically for the
Wallatin Catchment and should not be transferred to other locations without a critical assessment of their
appropriateness.
2.9 Priorities for implementation
Sufficient resources will rarely be available to immediately implement the results of a planning exercise.
Consequently, it is necessary to identify those actions which would contribute most to meeting the
objectives of the exercise and hence should be preferentially implemented. A general principle for setting
priorities for habitat reconstruction is to firstly protect areas of high conservation value and then to build
outwards from these with the aim of providing adequate habitat and maximising the connectivity of the
landscape.
Beyond this very general principle, it is best to base priorities firstly, on identifying areas which best
represent the biological diversity and secondly, on ensuring that the biota is able to persist in the selected
sites. Previous sections have addressed the issue of species retention. This section will focus on attempts
to represent biological diversity in the landscape.
2.9.1 Representing biological diversity
If detailed surveys of the biota have not been conducted, priorities can simply be set on the basis of easily
assessed variables such as the size and condition of the remnants and the range of patch types that they
contain (Safstrom 1995). Such an approach will increase the probability of capturing the diversity of the
region but there will always be the risk that important elements will be overlooked. Ideally the setting of
priorities would be based on more detailed surveys of the flora and fauna in order to identify those remnants
which most efficiently represent the full range of biological diversity in the area.
Priority setting should also take into account the broader regional context in which the planning is being
conducted. Species which are common locally and hence do not appear to be a priority, may not occur
elsewhere and hence are significant from a regional perspective. Conversely species that are locally rare
may be common elsewhere. By having a regional context for planning it will be possible to better identify the

conservation priorities and determine the most effective contribution that local people can make towards
addressing these regional priorities.
The irregular distribution of conservation values in landscapes has underpinned a significant effort over the
last decade to develop systematic methods to represent the full array of biological diversity in a region.
These methods include simple scoring or ranking procedures (review by Margules & Usher 1981; Safstrom
1995), linear programming (Cocks & Baird 1989) and iterative procedures (Kirkpatrick 1983; Margules &
Nicholls 1987; Margules et al. 1988; Margules 1989; Pressey & Nicholls 1989; Pressey et al. 1990; Pressey
& Nicholls 1991; Ryti 1992; Brooker & Margules 1996). These procedures vary in their complexity and data
requirements and the reader is referred to the references provided for more information about their
capabilities.
Priority areas for the Wallatin Creek case study
Formal assessment of the representativeness of the biota has been conducted for plant species in both the
Kellerberrin district and in the Wallatin Catchment (Brooker & Margules 1996). Additional surveys have been
conducted for plant assemblages (Arnold & Weeldenburg 1991; Lambeck & Wallace 1993), birds (Cale &
Lambeck unpublished data), reptiles (Smith unpublished data) and invertebrates (Abensperg-Traun et al.
1996). The invertebrate and reptile surveys have concentrated on selected vegetation assemblages and
hence are not comprehensive. They do however enable us to assess the relative value of the sampled
vegetation types for representing the greatest numbers of locally indigenous species.
Vegetation assemblages
Arnold and Weeldenburg (1991) estimated the representation of different landform types in remnants and in
road verges. Landforms provide a rough surrogate for vegetation communities because each landform type
tends to correspond with characteristic vegetation associations (Beard 1980). They found that the
proportional representation of the different landform types in road verges was roughly equivalent to that in
the landscape as a whole, but that the vegetation remnants over-represented rock outcrops and underrepresented
the low-lying fertile land units which supported salmon gum/gimlet woodlands and York gum
woodlands. The high correspondence within road verge vegetation reflects the fact that the road network
covers the whole region in a north-south, east-west grid and provides an effective sample of the whole
range of vegetation diversity.
Further assessment of the vegetation communities using Landsat TM imagery combined with air photo
interpretation and ground survey (Lambeck & Wallace 1993) found that within the remnants that occurred in
the district, only 30% of the remnant vegetation could be considered to be healthy examples of the dominant
vegetation types in the region. Many of the remnants contained substantial areas of rock which were not
suitable for agriculture and of the remainder, most are degraded to varying degrees by stock grazing, weed
invasion, rubbish dumping, vehicle tracks, and gravel extraction. A comparison of the percentage of
vegetation types in good condition with landform types in each remnant showed that woodland and sandplain
heath communities were under-represented and rock outcrops over-represented. Banksia woodlands
were particularly poorly represented with only 1 small patch and a few scattered trees remaining in the local
area. Of the other woodland types, Salmon gum and Gimlet woodlands are less well represented than are
Wandoo and York gum. From these studies, it is clear that the remaining vegetation does not represent the
original vegetation in terms of its distribution, proportions or condition and that priority should be directed
towards protecting and enhancing the less well represented vegetation types.
Plant species
The study area contains 11 species listed as Declared Rare Flora by the Department of Conservation and
Land Management (Mollemans et al. 1993). Of these, 6 species occur on private land or in road verges. All
remnants in the catchment containing these species were nominated as priorities for protection and
enhancement.

Further prioritisation of remnants was conducted by Brooker and Margules (1996) who ranked remnants in
terms of their relative importance for representing plant diversity at both local and regional scales. In this
study, plant species, together with a suite of environmental variables, were recorded at 125 sample sites.
From these variables, a model was developed which predicted the plant species diversity in other remnants
in the catchment based on their environmental characteristics. Figures 8 and 9 show the 10 most important
remnants based on regional and local analyses. Comparison of these figures reveals that remnant priorities
are scale dependent with some remnants considered equally important at both scales while others are only
important when assessed at a local scale.
Figure 8. Map of the Kellerberrin region showing the top ten remnants for representing the plant species
diversity of the region. From Brooker & Margules (1996).
Mammals
The only mammal species considered at risk in the catchment is the brush wallaby (Macropus irma) which is
known only from Durokoppin Nature Reserve. The distribution and abundance of other mammal species in
the catchment is not known. It is therefore not possible to rank remnants according to their importance for
mammals, apart from considering Durokoppin Nature Reserve to be the highest priority.
Birds
Survey data from all remnants in the catchment (Lambeck & Cale unpublished data) were used to rank
remnants according to the number of bird species that were considered a conservation priority in the region.
Selection of priority species was based on records of species known to have declined in the region
(Saunders 1989; Saunders & Ingram 1995) and which were considered at risk of further decline or local
extinction in the absence of interventionary management. The remnants selected on the basis of
presence/absence of priority species are illustrated in Figure 10.

Reptiles
No comprehensive survey has been conducted for reptiles in the study area. However, examination of lizard
species richness in selected vegetation types indicates weak relationships with native plant species
richness, vegetation structure and remnant area in Gimlet woodland, but no relationships with vegetation
attributes or area in shrublands (Abensperg-Traun et al. 1996). On the basis of these limited data, the only
criteria for prioritising remnants for reptiles would be on woodland area. However, the data were not
considered to be sufficiently comprehensive to warrant such an assessment.
Figure 9. The top ten remnants in Wallatin Catchment ranked on the basis of their contribution towards
representing the plant species diversity of the catchment. From Brooker & Margules (1996).
Invertebrates
Invertebrate diversity has been sampled only in woodland and shrubland communities in the study area and
the study considered species richness of only a selected subset of the arthropod fauna (beetles, ants,
scorpions, isopods, cockroaches, termites, earwigs, hemipterans and butterflies (Abensperg-Traun et al.
1996)). This study indicated that vegetation characteristics such as structural diversity or plant species
richness were poor predictors of invertebrate diversity. In the absence of such predictors, it is considered
inappropriate to attempt to rank remnants in terms of their contribution to representing invertebrate diversity.
The above information indicates that the only available data appropriate for ranking remnants in terms of
their relative importance for representing the biota are those for plants and birds. Comparison of the
rankings for birds and plants indicates that the same remnants have different values from the perspective of
each group. It should be expected therefore that similar surveys for reptiles, mammals or invertebrates
would produce other combinations of sites for best representing the diversity of each group. As a
consequence, priorities for protection and reconstruction in the Wallatin Catchment are firstly, remnants
known to contain vulnerable plant and bird species and secondly, remnants that best represent various
taxonomic groups.

It is important to bear in mind that, for the purpose of this exercise, all remnants are considered worthy of
retention and the priority setting process is aimed primarily at identifying where protection or restoration
actions should preferentially occur. Because a remnant is not identified as a priority remnant does not imply
that it is not important in the landscape and hence does not need to be protected.
Ranking remnants in terms of their conservation value is clearly only the first step in a conservation plan.
There is no point in simply identifying important remnants if they are degrading and losing the values that
they currently contain. The focal species approach described previously should subsequently be applied to
ensure that the remnants are able to retain their biota.
Figure 10. The Wallatin Creek Catchment illustrating the top 10 remnants for representing the most vulnerable
bird species. Many of these remnants extend beyond the catchment boundary and remnants 4 & 8 lie outside
of the catchment but are owned by land-holders within the catchment.
2.10 Guidelines for implementation
The outcome of the focal species approach is a list of species whose requirements for key habitat and
resource variables define different attributes that must be present if a landscape is to meet the needs of its
constituent flora and fauna. Ideally, the list would include focal species to define the minimum area of each
patch type; species to define the minium width, length and structure of connecting vegetation; species to
define appropriate levels of critical limiting resources; and species to define the appropriate rates or
intensities of each potentially threatening process. Insufficient information is currently available to identify
focal species for all of these variables. While it is possible to use the methodology to identify the minimum
size of each patch type, general principles and anecdotal observations will have to be invoked to define the
appropriate configurations of patches and the attributes of connecting vegetation.
In the Wallatin Creek case study, maps of abiotic attributes, including soils and landforms, which correlate
with the relevant patch types, were used to identify positions in the landscape which were suitable for
reconstruction of the main habitat types. Existing vegetation types were mapped using satellite imagery. Airphoto
interpretation and ground surveys were used to validate the satellite-derived maps and to identify
significant vegetation associations within each formation. The resultant maps were incorporated into a

geographic information system (GIS) and interrogated to identify all vegetation patches which failed to meet
the spatial requirements identified by the focal species analysis. GIS routines were developed which
specified areas which needed to be added to these patches in order to meet the minimum area requirement
(Figure 11).
This analysis revealed that 61 of the 113 mapped habitat patches in the catchment did not meet the
minimum size required by the focal species. Expansion of these patches to meet these minimum sizes
would require the revegetation of 1,121 ha. or 4.3% of the catchment. When this area is combined with the
total area already in remnant vegetation (1,925 ha. or 7.4% of the catchment), the total area required to
produce an adequate landscape for nature conservation is 3,046 ha or 11.7% of the catchment. The
proportion of each landform type that requires revegetation is shown in Table 2.
Table 2. The amount of each landform type that is required to expand small vegetation patches to a size where
they meet the minimum area requirements defined by the focal species.
Landform Area (hectares) % of each landform required
Ulva 194.3 4.6
Booraan 270.9 4.2
Collgar 168.7 5.2
Merredin 85.1 1.2
Belka 85.8 10.6
Danberrin 316.7 7.3
This analysis does not take into account the area required for providing linear vegetation to link the remnant
vegetation. In the Wallatin Catchment, the majority of remnants are connected to some degree, although the
width of the connecting vegetation varies from a few meters to up to 15m. For this exercise, it was
considered that a general recommendation be adopted which ensures that all remnants are connected by
linear vegetation at least 30m wide with greater than 60m being the preferred width for open vegetation
types such as woodlands and a minimum of 50m for linear vegetation linking sites occupied by dispersallimited
species to nearby suitable habitat patches. Any reconstruction of connecting vegetation should
ideally use local species planted according to soil type.
From the procedure described above, the following guideline was developed to allocate land to nature
conservation:
• Commit all vegetation remnants (public or private) to remnant vegetation protection.
This commitment was considered necessary because the existing remnant vegetation appears to be
inadequate for meeting the objective of ensuring no further loss of species. The evidence for this concern
comes from the continuing decline of species that are not threatened by processes such as predation and
fire (Saunders 1989). Consequently all patches should be retained and should serve as the framework to
which new habitat should be added.
The following guidelines are for habitat expansion based on the results of the focal species analysis:
• Revegetate areas adjacent to mixed heath patches of less than 25 ha with mixed heath species to
produce a minimum area of 25 ha.
• Revegetate areas on Booraan land-units adjacent to woodland patches of less than 20 ha with
mixed woodland species to produce a minimum area of 20 ha.
• Revegetate areas adjacent to patches of mallee/shrubland less than 25h with mallee species to
produce a minimum area of 25 ha.

• Revegetate areas on Merredin land-units adjacent to woodland patches of less than 20 ha with
salmon gum/gimlet woodland species to produce a minimum area of 20 ha.
• Revegetate areas on Belka land-units adjacent to woodland patches of less than 20 ha with salmon
gum / gimlet / mallee woodland species to produce a minimum area of 20 ha.
Three additional guidelines were provided which recommend that suitable areas be allocated to
reconstruction with salmon gum/gimlet woodland and Banksia woodland. This reflects their underrepresentation
in the catchment (Section 2.8):
• Allocate landform Merredin to revegetation with salmon gum / gimlet woodland.
• Allocate landform Belka to revegetation with salmon gum / gimlet / mallee woodland.
• Allocate areas < 20 ha on gutless sands adjacent to existing remnants to revegetation with Banksia
woodland.
These guidelines formed the nature conservation input to the integrated planning exercise described in
Chapter 3. Further details about the role of these guidelines in the planning process are presented in that
chapter.
Figure 11. Map of the Wallatin Catchment indicating areas requiring habitat reconstruction to meet the needs
of area-limited focal species.
2.11 Moving the goal posts: the consequences of implementation
While focal species can be used to determine the attributes required in a landscape, it must be remembered
that the initial assessment of risk considered only immediate threats. It is possible that a species currently

limited by landscape configuration may, when the configuration is altered, change in numbers only to a level
whereby a new limit is imposed by another factor. In addition, changes in landscape pattern may result in
altered species responses to the new landscape configuration. For example, as the number of patches in a
landscape increases, it may be possible for individuals of a species to occupy smaller patches than they
could when fewer patches were available. Similarly, changes in the quality of corridors may alter the
minimum inter-patch distance over which individuals of some species can move. Not only will interactions
between species and their habitat change as a result of changing configurations, but interactions between
species may also change as a result of different species responding in different ways to altered
configurations.
These unpredictable interactions, together with the obvious risks associated with the initial assumption that
the needs of the focal species encapsulate the needs of all other species, make the establishment of a
monitoring process critically important. The monitoring program must be designed to test the underlying
assumptions and must have a capacity to detect deviations from predicted responses at the earliest
possible time. The monitoring program must focus primarily on the focal species, but must also consider the
responses of a suite of additional non-focal taxa. These additional species should be selected to represent a
range of life-history characteristics in a variety of taxonomic groups. Using this approach, a strategic
monitoring program based on a limited set of species would provide an indication of the changes occurring
in the landscape in response to the management actions.
Any monitoring strategy will obviously require the allocation of additional resources. However, the failure to
implement a monitoring strategy will ensure that we will fail to learn from our actions. The focal-species
approach provides a theoretical basis for a monitoring strategy which tests clearly stated hypotheses and
assumptions and enables an assessment of performance against nominated objectives. Because general
enhancement approaches are unable to specify the outcome expected, monitoring will simply indicate that
changes are occurring but will not be able to specify what the causes of those changes are or what remedial
actions should be taken in response to unexpected changes. The level of resources required for monitoring
will not be trivial which provides yet another reason for adopting a bioregional scale approach to
conservation planning with a carefully designed monitoring program distributed strategically throughout the
region.
2.12 Transportability of solutions
The hope of many conservation managers is that it will be possible to identify a limited set of landscape
variables which, if present, will ensure the retention of all of the biota in the landscape that they are
managing. In other words, they are seeking relatively simple templates that specify a proportion of the
landscape that should be allocated to native vegetation in order to ensure the persistence of the biota.
Expectations that general templates can be developed stem from an overly simplified interpretation of a
body of theory that developed in the 1960s. This theory is based on the species-area relationship developed
by MacArthur and Wilson (1967) which showed that the number of species on oceanic islands increased
with the size of the island and decreased with the degree of isolation from source populations. This theory
had intuitive appeal to conservation biologists who were viewing terrestrial vegetation remnants as islands in
a sea of crops.
The limitation of this theory in a modified terrestrial environment is that it does not take into account the
combination of habitat heterogeneity (patchiness) in the pre-clearing landscape and the selective clearing of
that landscape. Particular patterns of clearing in a patchy landscape can have impacts on biodiversity that
are disproportionate to the area modified. For example, the selective removal of small resource-rich areas,
or of refuges used by animals during times of environmental stress, may cause a disproportionately large
decline in species richness. Conversely, the removal of large resource-poor areas may have a smaller
impact on species numbers than would be expected on the basis of a simple relationship between species
numbers and area. These factors are particularly pertinent in the highly fragmented wheatbelt of Western
Australia which is renowned for its species richness combined with high levels of local endemism.

In order to test whether the nature conservation recommendations produced for the Wallatin Catchment
were relevant elsewhere, a preliminary survey was conducted in the Dumbleyung Shire, approximately 200
km to the south of the current study. The objective of the survey was to determine patterns of habitat use of
species that were considered vulnerable in Wallatin, and of species that have disappeared from Wallatin but
still persist in Dumbleyung. The area surveyed in Dumbleyung had a similar configuration of remnants to the
Wallatin Catchment, with approximately 10% cover of remnant vegetation. Both areas were also similar with
respect to the structural attributes of the dominant vegetation associations although there were differences
in plant species composition.
It was immediately apparent from even a superficial survey that there were significant differences in patterns
of patch use by bird species in the two areas. For example, western yellow robins and southern scrub robins
occurred in much smaller patches of shrubland in Dumbleyung than was the case in Wallatin. Rufous tree
creepers, which no longer occur in Wallatin were found in 4 ha degraded patches of woodland. Patches of
habitat which are suitable for these species in Dumbleyung are obviously not adequate in Wallatin. There
are many equivalent patches in Wallatin Catchment and the areas surrounding it and yet they are not
occupied by these species.
The reason for this discrepancy is likely to be attributable, at least in part, to differences in landscape pattern
at a regional scale. While the sites are similar at a local scale, the two locations are very different when
viewed from a regional perspective. The distribution of vegetation remnants in the Kellerberrin shire, where
Wallatin Creek is located, is roughly the same throughout the shire. In other words, the Wallatin site is
broadly typical of the rest of the region. The Dumbleyung site, on the other hand, was located within 15 km
of extensive bushland which includes Dongolocking Nature Reserve. The close proximity of this high quality
habitat is likely to be an important factor in determining the patterns of habitat use observed at Dumbleyung.
Clearing history may also influence the observed habitat patterns and any detailed comparisons between
areas should therefore take into account the time since clearing.
The fact that a solution generated in one part of the wheatbelt clearly cannot be transported to another is
likely to concern conservation managers. However, it is critical that this fact is acknowledged as any
attempts to take solutions from one location and apply them to another will produce results which will be
inadequate in some instances and excessive in others. The focal species approach does not provide a
template that can be simply transported to a new location. Rather, it provides a procedure for deriving
landscape designs and management guidelines in any landscape.
While it is apparent that the results derived from Wallatin Creek could not be applied 200 km away, it may
be reasonable to expect that the solution could be usefully applied in the catchment immediately adjacent to
Wallatin Creek, given the similarities in the flora and fauna and in landscape configuration. If it can be
applied to the next sub-catchment, could it be applied to the one beyond that? This raises a critical question
for conservation planning: To what extent can a solution be legitimately extended beyond the location where
it was generated?
A potential solution to this problem could be found if it is possible to partition a large region, such as the
wheatbelt of Western Australia, into smaller sub-regions which are internally homogeneous, not only with
respect to environmental variables, as is the case with bioregions, but also with respect to their landscape
configuration. These sub-regions, which could be termed Conservation Management Zones (CMZs) would
form a rational framework for conservation planning.
If such Conservation Management Zones can be identified, then a solution derived from one location within
a CMZ could more reliably be applied to the remainder of that zone.
Such an analysis of landscape pattern would help to define the appropriate scale for conservation planning.
In areas that are relatively uniform with respect to both biophysical and anthropogenic parameters, it may be
possible to develop solutions that can be applied over a large region. On the other hand, where patterns of
clearing are more spatially variable it may be necessary to develop plans at a smaller spatial scale.
Appendix 1 presents a framework for planning for nature conservation at a regional scale.

2.13 The role of science and data adequacy
Three major problems underpin the acquisition of knowledge for managing biological diversity in production
landscapes. The first of these is the absence of comprehensive survey data and a detailed knowledge of the
ecology of the biota we wish to protect. Consequently we have limited knowledge of what we are trying to
conserve, or of how the species that we wish to protect are using the landscape that we are managing.
Unfortunately the acquisition of comprehensive survey data and detailed ecological understanding is
extremely labor intensive and expensive (Burbidge 1991). Consequently, the information that is available to
managers is neither comprehensive, nor detailed. Where detail does exist, it invariably deals with a very
small component of the total problem.
The second difficulty stems from a perception that the acquisition of information for management should be
based on sound scientific methodology. Such scientific rigour demands replicated experiments with
appropriate controls and large sample sizes to provide sufficient statistical power. As a consequence,
practitioners of 'good' science often focus on those aspects of a landscape that can provide the required
rigour. We therefore count those things which are easily counted and describe events commonly, and hence
reliably, observed. The result is a wealth of knowledge about the common and conspicuous features in the
landscape. However, when that landscape is deteriorating, this is not necessarily the knowledge that
managers require. We do not need to manage those features in a landscape that are secure. Unfortunately,
the components that are threatened, and therefore do need to be managed, are often rare and hence less
amenable to rigorous scientific investigation.
The third fundamental problem again relates to the practice of science. The scientific method has
traditionally focussed on deriving general principles from a series of observations that appear to exhibit
common patterns. We therefore gather data from a number of specific events in particular locations and
derive a general principle which we believe should hold over a range of circumstances. Unfortunately, while
it is relatively easy to progress from a series of site-specific observations to a general principle, it is more
difficult to return from the general principle to a prescription for a new location. Ultimately, general principles
enable us to make statements about the relative merits of alternative actions, but they cannot provide the
absolute values that managers require when they are competing for limited physical and monetary
resources. For example, a general principle which states that 'more is better than less' for nature
conservation is of little use to a conservation manager confronted with a land owner who will only sacrifice
the minimum possible amount of productive land. General principles can tell us a direction in which to
proceed, but provide little guidance as to how far to go.
The resolution of these problems requires firstly, a recognition that they are problems and secondly, a
conscious effort to broaden the horizons of ecological science from the tradition of description to the
challenge of prescription. Given the complexities of ecological systems and the spatial and temporal
variability both within and between landscapes it is unlikely that efficient 'template' solutions can be
developed that will hold far beyond the source of their development. Consequently, the aim of science in
conservation management should not be to provide 'answers' that apply to all circumstances, but rather, to
provide procedures for arriving at an answer that is appropriate for the situation at hand in the most efficient
manner possible. The focal species approach presented in this section represents an attempt to provide
such a procedure. The 'answer' generated by the procedure will depend upon the objectives, the location,
and the degree of biological impoverishment in the area being considered.
This alternative view of the role of science in conservation management brings into question the type of data
required for making management decisions. If our objective is to prevent the loss of elements at risk in our
landscapes then we need to focus on those elements. This requires increasing attention to the needs of the
less common features in our landscapes. This is not to advocate a return to single species studies of rare
and endangered species for their own sake. Rather, it is an argument for focusing on those vulnerable
elements of a landscape which, if managed effectively, can deliver the greatest benefit to the widest range
of additional species.

If a focal species approach to nature conservation is to be adopted, there are a number of data sets that
must be acquired. The first of these is a map which partitions the landscape into biologically meaningful
units which can be used as the building blocks for landscape design. These units will invariably be a
compromise between the true complexity of the ecological system being managed and the capacity of
managers to incorporate that complexity into their management regime. For the current study the dominant
vegetation associations were considered appropriate units for landscape assessment and planning.
Subsequent design was based on creating different configurations of these units which meet the needs of
the focal species.
A range of methodologies are available to create such maps including satellite remote sensing, air photo
interpretation and ground survey. Often the combination of these methods can provide more reliable data. In
an attempt to improve the efficiency of the mapping process, predictive models of landscape pattern are
often generated to enable knowledge generated from one location to inform us about conditions elsewhere.
The reliability of such predictive modelling relies on the strength of the relationship between the predictive
variables and the response variable. For example, vegetation classification from satellite imagery relies on
unique and consistent spectral responses from the different vegetation types that we wish to map. The
strength of this relationship will depend on the degree of variation within patch types relative to that between
patch types. If patches are internally uniform and significantly different from other patches, such technology
can produce relatively reliable results. However, when the patches are themselves variable and the
differences between patches are subtle, the reliability of these predictive approaches will diminish. By
combining satellite imagery with models of landform pattern, it may be possible to improve the quality of the
result. For example, areas which are not separable on the basis of their spectral properties may be
distinguished on the basis of their position in the landscape. It has yet to be demonstrated that this will
improve the quality of our broad-scale mapping capacity. However, given its potential to significantly
improve that capacity this represents an area of research that warrants further investigation. Whatever
approach is adopted, ground truthing of the results to assess their accuracy is essential.
The second data set that is required is a list of species that are potentially vulnerable in the region being
managed. Species or communities should be classified as vulnerable if there is evidence of decline over
time. This should apply equally to species which are currently common but declining and those which are
rare. Rarity per se is not necessarily an indicator of vulnerability. Some species may be naturally rare and
have stable populations. However, in the absence of information on population changes over time it would
be sensible to adopt a cautious approach and consider rarity to indicate potential vulnerability.
Having identified the vulnerable species in a landscape it is necessary to have a capacity to derive basic
population parameters for key species. As described in Section 2.5, the focal species can currently only
deliver 'adequate' landscapes unless applied over an area that is large enough to have a high probability of
also being viable. Specification of what is a sufficiently large area will require an assessment of the
population viability of the focal species and hence estimates of appropriate demographic variables for these
species.
If further investigation of the focal species approach reveals that there are characteristic types of species
that are regularly identified as playing an umbrella role, then it may be possible to design efficient survey
procedures that specifically assess the status of these types of species. These species are likely to be
relatively sedentary habitat specialists. Surveys should therefore be strategically designed to target these
less common species and should have a capacity to be repeated at appropriate intervals to assess
population trajectories. The routine collection of species-based information at a catchment scale will clearly
not be possible in all cases further indicating the need to conduct such analyses at larger, regional scales.
The strategic acquisition of data at these larger scales may be cost effective in the long run if such
information can contribute to the development of strategies that require less effort per unit area to achieve
the desired conservation outcome.

2.14 Summary
Conservation objectives and strategies
Approaches to managing biological diversity in agricultural landscapes will differ depending on the
objectives that have been set. The two broad types of objectives that were identified general versus
strategic enhancement rely on different types of information and result in different landscape responses.
General enhancement strategies can be based on ecological principles whereas strategic approaches
require a knowledge of the needs of the biota.
General enhancement strategies are unable to provide clear targets for dealing with immediate conservation
problems. They aim to minimise the number of species lost or increases the abundance of species present.
They are unable to specify which species will be lost or retained or which species may increase or decrease
in numbers. Consequently there are no criteria for assessing success or failure apart from general changes
in community structure and composition. General enhancement approaches can identify directions in which
to proceed but provide no indication of the magnitude of the response required.
Strategic enhancement strategies, on the other hand, have clearly identified targets against which success
can be judged. These targets will take the form of stable or increasing distributions and population sizes of
individual species or groups of species. These types of objectives can also be used to develop explicit
guidelines about how much of a landscape is required for meeting a specified objective.
Focal species in conservation planning
The focal species approach presented in this report provides a means to address the traditional dichotomy
between single-species and landscape approaches to conservation management. While species are
employed for the assessment of landscape adequacy and for guiding management strategies, the choice of
species is based on their capacity to encapsulate the needs of other species in the landscape. These focal
species can be used to identify the appropriate spatial and functional parameters that must be present in a
landscape if it is to retain the flora and fauna that occur there. Area-limited species define the spatial
attributes of each patch type, dispersal-limited species define patch configurations and connectivity
characteristics, resource-limited species define the compositional attributes, and process-limited species
define the management regimes that have to be implemented.
A critical feature of this approach is that it does not provide a template to apply across all landscapes but it
does provide a procedure by which to determine the actions required in any given landscape. These actions
are guided by the needs of a subset of the species present with recognition that the composition of this set
will differ from one place to another due to both environmental differences and differences in the amount of
human disturbance.
In relatively undisturbed landscapes where much of the original community composition remains intact, the
focal taxa are more likely to be sedentary resource specialists that prefer patch interiors rather than edges.
They are also likely to be larger vertebrates which have the greatest demands for habitat. In landscapes that
have progressed further along the continuum of habitat reduction, fragmentation and degradation, many of
the more demanding species may have already been lost. Smaller vertebrates, plants, or even invertebrates
will have an increasing probability of being identified as the most demanding species that remain in the
landscape. By applying the focal species approach it is possible to specify the attributes that are required to
meet the needs of the species present in the landscape being managed regardless of where that landscape
sits on the degradation continuum.
Mixed strategies in the absence of perfect knowledge
Ideally, nature conservation planning would always be based on an understanding of the needs of the biota
in the landscape to be managed. However, it will probably never be the case that a complete understanding
will be acquired for any one location, let alone for a whole region. It will therefore always be necessary to

combine strategic procedures, such as the focal species approach, with general principles when the critical
information is not available and cannot be acquired within the time-frame or budget of the management
plan. It is essential to recognise, however, that the greater the reliance on general principles the less certain
will be the conservation outcome.
Spatial scales for conservation planning
The solution generated using the focal species approach was surprisingly efficient with a small but strategic
increase in the amount of habitat creating an additional 60 habitat patches that would be large enough to
support the most habitat-demanding species. It is improbable that such an efficient solution could be
developed by using general principles alone although this assertion remains to be tested. While the
approach required a relatively small increase in the total area under native vegetation to produce an
adequate landscape, it did not resolve the issue of landscape viability. The solution will meet the immediate
habitat needs of the species in the catchment but in its current form the procedure cannot identify the area
over which the solution must be extended before sufficient numbers of individuals of rarer species are
included for the landscape to also become a viable one. Landscape viability is more likely to be achieved if
assessment and planning is conducted at a bioregional scale. However, in order to ascertain the appropriate
scale for achieving particular conservation outcomes it will be necessary to determine whether a landscape
that can ensure population viability for focal species can also provide the requirements to support viable
populations of all other species.
It is also critical to determine whether a solution developed at a larger spatial scale, such as a bioregion, will
reduce the effort required per unit area to meet a specified conservation objective. For example, an attempt
to provide landscape viability as well as adequacy in the Wallatin Catchment would probably require
revegetation of the majority of the catchment and even then will not guarantee the persistence of sparsely
distributed species. Such a strategy would obviously jeopardise the economic viability of the catchment.
However, if conservation planning and action was conducted at a regional scale the land-holders would only
have to contribute to the solution rather than achieve it on their own. This would require a much smaller
proportion of their catchment to be allocated to nature conservation in order to meet the conservation
objective. These questions of landscape viability and the effect of scale on conservation effort need to be
explored in the context of a larger regional planning initiative.
Chapter 3 - Integrating Biodiversity Conservation with Other Land Uses
• 3.1 Introduction
• 3.2 Requirements for integrated land-use planning
• 3.3 Stakeholder participation
• 3.4 Management issues
• 3.5 Land uses and land suitability
• 3.6 Land allocation
• 3.7 Stakeholder responses
• 3.8 Data adequacy
3.1 Introduction
Land management in production landscapes is largely concerned with identifying the land uses and
management regimes which are best suited to different parts of the landscape. Where the sole objective is
to maximise profits in the short term this is simply a matter of determining the land capability of the various
land units and applying the most profitable land use, or sequence of land uses to those units. These
calculations are based on factors such as market prices, potential yields and management costs. Land-use
management becomes much more difficult when the manager is attempting to deal with a number of
objectives, some of which are less directly related to profits but which compete for land which is valuable for
production. The different time scales required to address some of these objectives will present additional
difficulties. Conventional production objectives have tended to be managed with a view to short-term

outcomes, whereas land conservation and nature conservation will generate benefits over a much longer
time frame. The challenge then becomes one of apportioning the landscape in a way that ensures
satisfactory performance on all objectives with minimum impact on farm profits. In order to make informed
decisions about how different portions of a landscape should be managed to meet a range of objectives it is
necessary to determine, for each objective, what land uses need to be represented, in what quantities, and
what parts of the landscape are best suited for those uses. If this can be determined, it is then necessary to
combine the different land uses in a way that maximises profits but also meets other objectives.
An important issue that needs to be addressed when allocating land among different uses is the impact on
one goal of implementing an action which addresses another. For example, what is the impact on yields and
therefore on profits, if land is allocated to nature conservation? It is essential to be able to determine the
extent to which different objectives will be met by different landscape designs and to be able to explore
alternative designs and examine their implications.
For the current planning exercise the Wallatin Catchment was treated as a single management unit in order
to assess the effect of scale on management decisions and production outcomes. In the land-allocation
scenarios developed in this case study, property boundaries were removed in order to determine how the
catchment would best be managed without the constraints of current patterns of property ownership. While
conceptually challenging for landholders, this analysis enables an assessment of whether the benefits to be
gained from whole catchment management exceed the sum of the benefits from managing each property
individually. Increasingly, calls are being made for 'integrated catchment management' but the benefits from
attempting such an exercise have yet to be demonstrated. If clear benefits can be shown, this may provide
an incentive for considering planning and management approaches which enable those benefits to be
distributed equitably throughout the catchment group.
3.2 Requirements for integrated land-use planning
Because of the complexity of addressing multiple goal planning, it is often useful to use a decision-support
system to help explore alternative strategies. Essential features of a decision-support tool are:
• An ability to reflect the interests of each of the different stakeholder groups.
• Transparency in the decision-making process so that all stakeholders can see how outcomes are
arrived at and how their interests are performing.
• An ability to quantify the extent to which different objectives are being met.
• An ability to determine the impact on all other objectives of allocating a land unit to a particular
objective.
• An ability to run 'what if' scenarios to assess the likely impact of alternative decisions.
For the Wallatin Creek case study, the land allocation package LUPIS (Land Use Planning and Information
System (Ive & Cocks 1988)) was chosen. This choice was made on the basis that (i) it is conceptually
simple (ii) it is not dependent on the quantification of all variables that are used in decision making and (iii) it
has the capacity to reflect social preferences through a transparent weighting system. Other approaches,
such as mathematical programming, have been have been used to address multi-criteria planning problems
(Field 1973; Cocklin 1989a,b), but their computational requirements and conceptual complexity can
potentially alienate stakeholder groups.
The essential elements of the planning process include;
• stakeholder participation;
• identification of the main land management issues;
• identification of the full range of potential land uses;
• assessment of land suitability/capability for those uses; and
• development of guidelines for land allocation.

Each of these issues are examined in the following sections.
3.3 Stakeholder participation
Participants in the allocation process should include all parties with an interest in the region being
considered. The list of participants will vary from region to region depending primarily upon the enterprises
in the region and land tenure arrangements. In the current study, all of the land being considered is either
freehold or is owned by the Department of Conservation and Land Management. Shire councils have
responsibility for maintaining roads and road verges in the area. Primary stakeholders are therefore the land
owners and local management authorities. Additional groups have an interest in the region as a result of
ongoing research involvement and can therefore act in an advisory capacity. These groups include
Agriculture Western Australia and CSIRO Divisions of Wildlife and Ecology and Water Resources. In other
areas, groups that may need to be involved could include aboriginal groups, mining companies, pastoralists,
tourism planners and operators, water and power authorities, other government departments such as the
Department of Defence and representatives of industry and conservation organisations.
Participation by local land-holders was achieved through liaison with key members of the local Land
Conservation District Committee. Some land-holders were understandably reluctant to embark on a
facilitated planning process involving computer-based support systems when they could not envisage the
likely outcomes of such an approach. It was therefore decided to develop a prototype analysis of the
catchment using a preliminary set of guidelines which broadly reflected the issues that may be relevant to
the group. It was then possible to demonstrate the procedure to the group and illustrate the type of results
that could be expected. The participants were then able to assess the information used and modify it to
better reflect their interests, rather than having to be involved in the developmental phase of the project.
This approach is, to some extent, contrary to the prevailing view that stakeholders should be involved from
the very beginning of a planning exercise. This study indicated the importance of establishing a clear
framework which could be taken to the group and modified and developed (or rejected) by them, rather than
attempting to seek their involvement in the development of a process which does not have immediately
apparent benefits. By developing a prototype for their particular catchment they were immediately able to
see the implications of the process for the catchment and for their particular property.
Government agency participation was dependent on the goodwill of particular individuals who provided the
best information available within the constraints of other job commitments. Because of the developmental
nature of the process, the agencies, like the land-holders, needed to be convinced of its benefits before they
committed resources to the exercise. Any subsequent application of a procedure such as this must ensure
that all participants are committed to the process from the outset and must also ensure that there are
sufficient resources available to acquire the necessary information.
Aboriginal issues were not addressed in this study because of the small area being considered. However,
Aboriginal people do have an interest in the broader region and any regional planning must include their
participation and have a capacity to reflect their interests.
Each stakeholder group was required to specify:
• the issues that they considered to be important;
• the land uses which they wished to see represented in the landscape; and
• the areas in the landscape which they considered best suited to the specified land uses.
This information provides the basis for developing guidelines for the planning exercise.
3.4 Management issues
The issues that will be relevant for any land-allocation exercise will depend on where that exercise is
located. For the case-study area, the primary issues identified by the stakeholders were:

• the need for a profitable and economically sustainable agricultural enterprise;
• the need to retain biological diversity;
• the need to control land degradation which threatens both the production and conservation values of
the catchment.
In essence, they are seeking an ecologically sustainable agricultural system. In other areas, additional
issues such as those associated with Aboriginal interests, mining or pastoralism may also need to be
considered.
3.5 Land uses and land suitability
• 3.5.1 Biodiversity land uses
• 3.5.2 Agricultural land uses
• 3.5.3 Hydrological land uses
• 3.5.4 Interactions between land uses
For each of the nominated issues the stakeholders were required to identify the range of relevant land uses
and to specify the suitability/capability of different land units for those uses.
3.5.1 Biodiversity land uses
The ongoing decline in abundance of species that require large patches of habitat indicates that the existing
remnant vegetation is inadequate for retaining the plants and animals in the catchment. It was therefore
considered necessary to not only protect the existing remnant vegetation but also to increase the amount of
available habitat. Because all stakeholders involved in the current study agreed that the existing remnant
vegetation in the catchment is inadequate there was no need to specifically deal with issues of further
clearing. Consequently all existing native vegetation was considered suitable for retention. Figure 12 shows
the current distribution of remnant vegetation in the Wallatin Catchment.
Figure 12. The distribution of remnant vegetation in the Wallatin Catchment. Data courtesy of M.G.Brooker
(CSIRO Widlife and Ecology) and the Spatial Resource Information Group, Agriculture Western Australia
(Beeston et al. 1994).

The land uses that were considered appropriate for addressing nature conservation objectives in the
Wallatin Catchment were therefore considered to be:
• retention of existing private and public remnants of native vegetation (including linear strips of
habitat, or corridors);
• reconstructed habitat patches; and
• reconstructed corridors.
Because it is yet to be demonstrated whether it is possible to reconstruct all of the complexity of natural
habitat, it was considered that habitat reconstruction should initially be based on the re-establishment of the
dominant species which make up the main vegetation types that occur in the catchment. Land suitability for
re-establishing these different habitat types was derived from a knowledge of the relationships between
vegetation assemblages and landforms (Beard 1983). Maps of these landform types (Figure 13) enable an
assessment of the suitability of different parts of the landscape for different reconstruction actions. The
landform types which occur in the catchment and their corresponding vegetation types are listed in Table 3.
Figure 13. The distribution of landforms in the Wallatin Catchment. Data from McArthur (1992)provided by the
Spatial Resource Information Group, Agriculture Western Australia.

Table 3. Land suitability for addressing biodiversity objectives. Landform units are those described in Figure 2
(Section 1.2.2).
Land unit
Existing public reserve
Existing private remnants
Landforms
Ulva
Booraan
Collgar
Merredin/Belka
Danberrin
Rock
Gutless sands
Land use
Remnant protection
Remnant protection
Heath
Wandoo woodland
Mallee
Salmon gum/gimlet woodland
York gum woodland
Jam wattle/York gum woodland
Banksia woodlands
The focal species approach described in Section 2.5 provided guidelines for the minimum area of each
vegetation type that is required to meet the needs of the most demanding species that utilise that patch
type. A Geographic Information System (GIS) was used to identify all existing patches which were less than
the desired size and to identify suitable adjoining areas that would need to be added to these patches in
order to meet the specified minimum sizes (See Figure 11 in Chapter 2).
3.5.2 Agricultural land uses
Land capability assessment has been the basis for agricultural planning since the wheatbelt was first
cleared. Lower lying soils supporting York gum and salmon gum woodlands were recognised at an early
stage as being the most productive for growing wheat (Burvill 1979). Other soil types, such as the upland
sandy soils, were originally considered less suitable for cropping until it was realised that the addition of
trace elements improved their capability.
Soil types for the study area and their relationships with landform units were described by Hawkins (1990).
The classification produced by Hawkins was used by farmers in the catchment to produce soil maps for
each farm. These maps were digitised and incorporated into the regional Agriculture Western Australia GIS
at Merredin (Figure 14). The suitability of these soil types for agriculture, as perceived by the landholders, is
shown in Table 4.
Figure 14. Soil map of the Wallatin Catchment. Modified from data provided by Agriculture Western Australia.
Soil types match those listed in Table 4.

Table 4. Suitability of different soil types for agricultural land uses. Terminology follows Hawkins (1990).
Ulva
Booraan
Collgar
Merredin/Belka
Danberrin
Rock
Land unit
Hawkins classification Land use
Deep yellow sand
Sa
Lupins/Cereals
Pasture/Cereals
Gravelly uplands Sg Lupins/Cereals
Gravelly uplands
St
Lupins/Cereals
Pasture/Cereals
Gravelly sands
Sandy loams
Sgt
Sp
Pasture/Cereals
Pasture/Pasture/Cereals
Lupins/Cereals
Pasture/Cereals
Breakaways Bd Revegetation with perennial vegetation
Shallow clays Bgc Revegetation with perennial vegetation
White gum soils Be Pasture/Cereals
Shallow gravelly duplex Dug Pulses/Cereals
Deep duplex Dud Lupins/Cereals
White gum duplex Duw Pasture/Cereals
Shallow grey duplex Dusg Pulses/Cereals
Salmon gum/gimlet
BE
Pasture/Cereals
Pulses/Cereals/Cereals
Salmon gum/gimlet soils
Msl
Pulses/Cereals/Canola/Cereals
Pasture/Cereals
Pulses/Cereals/Canola/Cereals
Pulses/Cereals/Canola/Cereals/Pasture
Pasture/Cereals
Salmon gum/gimlet soils Mgc Pulses/Cereals/Cereals
Pulses/Cereals/Canola/Cereals
York/Jam Y/J Pulses/Cereals/Pasture/Cereals
York Y Pulses/Cereals/Pasture/Pasture/Cereals
Jam/Rock J/R Revegetation with perennial vegetation
Revegetation with perennial vegetation
The land uses nominated by the farmers in this exercise tend to reflect traditional cropping rotations based
on cereals, pulses and pasture. Annual rainfall is insufficient to support a timber industry based on
Tasmanian blue gums (Eucalyptus globulus) as is the case in the wetter areas to the south-west. Other
timber species such as maritime pine (Pinus pinaster) and eucalypt mallees for producing oil may be
potential crops in the future. The recent nomination of this catchment as a 'Focus Catchment' in the State
Government Salinity Action Statement (Government of Western Australia 1996) may provide an impetus for
exploring a wider range of agricultural land-use options.

3.5.3 Hydrological land uses
The increase in land and stream salinisation throughout the south-west of Western Australia is caused by
increasing groundwater discharge as a result of the clearing of native vegetation (Salama et al. 1993, 1994).
Possible responses to this increased discharge include the re-establishment of woody perennial vegetation,
development of high water-use crops, and implementation of drainage strategies. The nature of the
response will depend on the level of understanding of the factors that influence catchment hydrology.
The Wallatin Catchment represents a relatively complex hydrological system. It is bordered in the higher
parts of the landscape by lateritic duricrust and granitic outcrops. The remainder of the catchment is
dissected by dolerite dykes with associated basement highs, quartz veins and faults (Salama et al. 1993).
Thin sedimentary deposits occur throughout the catchment, particularly in the alluvial channels. Each of
these features differs in its recharge and discharge characteristics and in its influence on salt storage.
Three different patterns of recharge can be identified within the catchment (Figure 15).
• Monotonically rising water levels: Sand plains, lateritic duricrust and basement outcrops in the
watershed zones of the catchment were identified by Salama et al. (1993) as being the most
significant recharge areas. In these areas, water levels are rising uniformly at up to 1.6 mm/day
(Figure 15a).
• Continuously rising water levels with seasonal fluctuations: In the midslopes of the catchments water
levels fluctuate in response to rainfall but display a long-term rising trend. The short-term fluctuations
result from seasonal recharge by winter rainfall accompanied by lateral groundwater movement. In
periods of high rainfall the recharge rate exceeds the rate of lateral flow resulting in a build up in
groundwater storage and a subsequent rise in water levels (Figure 15b).
• Seasonally fluctuating water levels: Water levels in the lower areas of the catchments near streams
show seasonal fluctuations with water reaching a peak following winter rains and falling to a
minimum in summer (Figure 15c).
The distribution of soluble salts varies throughout the catchment. Total soluble salt levels are highest in the
area covered by thin sediments and in the channel deposits. Salt storage increases from divide to valley
floor with higher salt storage occurring in the relict channels and palaeochannels. Salinity stores can vary
from 10 t/ha in the divides to 350 t/ha in the valleys with levels as high as 4000 t/ha upstream of geological
structures (Salama et al. 1993).
There are two revegetation responses that can be adopted in the face of this complexity. The first of these is
to acquire a detailed knowledge of recharge and discharge characteristics and strategically target
revegetation to those areas where it will have the greatest impact. In situations where this knowledge is not
available, an alternative response is to distribute revegetation throughout the catchment in order to intercept
water regardless of where it falls. On the basis of the hydrological characteristics of the Wallatin Catchment
a number of broad hydrological 'land uses' were identified. These ranged from dense planting of perennial
species, through alleys of perennial tree or shrub species interspersed with agricultural rotations and drains,
to scattered plantings of perennial species around local recharge features. The circumstances under which
each of these options is most appropriate are listed in Table 5.

Figure 15. Long-term water level trends: a) monotonically rising; b) continuously rising; c) seasonally
fluctuating. Modified from Salama et al. (1991).
Table 5. Suitability of different land units for hydrological land uses.
Ulva
Booraan
Collgar
Merredin/Belka
Danberrin
(York/Jam)
Rock
Land unit
Gravelly uplands
(Sg)
Gravelly uplands
(St)
Breakaways (Bd)
Land use options
Revegetation to deep rooted perennial vegetation
Plantation perennials
Plantation perennials
Alleys of deep-rooted perennials planted along the contour,
interspersed with deep-rooted annuals
Revegetation with perennial species
Shallow clays (Bgc) Revegetation with perennial species
Shallow gravelly
duplex (Dug)
Alleys of deep-rooted perennials planted along the contour,
interspersed with deep-rooted annuals
Alley farming with drains
Deep duplex (Dud) Alley farming with drains
White gum duplex
(Duw)
Shallow grey duplex
(Dusg)
Jam/Rock
Alley farming with drains
Cropping, interspersed with perennials around recharge features
Revegetation for hydrology
Deep-rooted perennials planted along the contour, interspersed
with deep-rooted annuals
Revegtation with perennial species
Alleys of deep-rooted perennials planted along the contour,
interspersed with deep-rooted annuals where soils are sufficiently
deep

3.5.4 Interactions between land uses
As outlined in the previous sections, the objectives of any land-allocation exercise are most likely to be
achieved where there is a good understanding of the suitability of different parts of the landscape for
different land uses. In many instances, however, a given land unit may be suitable for more than one use.
Where this is the case, it will be necessary to either (i) allocate the land unit to one or other of the competing
uses or (ii) seek ways to integrate the uses so that the same land unit can be managed for both objectives
simultaneously. Where the first option is taken it is likely that one objective will suffer at the expense of the
other with the risk that one may not be met. In the second scenario neither objective may attain maximum
performance but an adequate result may be achieved for both.
By comparing the land uses which contribute to nature conservation, agriculture and hydrology objectives, it
was possible to identify new combinations of land uses which address two or more objectives
simultaneously. For example, the requirement for dense plantings of perennial species on some recharge
areas to manage hydrology can be combined with a nature conservation objective by selecting native plants
of local provenance which may provide resources required by the native biota. Alternatively, hydrological
and production goals can be merged by planting timber producing species or oil mallees in areas where
they contribute most to managing recharge. This could be achieved either through block plantings of timber
species or by alley farming. Both of these strategies can bring further benefits including shelter for stock and
reduced wind and water erosion.
Table 6 lists composite land uses which were identified by looking for correspondence between the
recommended uses for nature conservation, hydrology and agriculture objectives. Such melding of land
uses can only be carried out where the various stakeholders acknowledge mutual benefit. If such an
agreement cannot be reached, then the individual land uses should be used in the planning exercise. By
using decision support tools such as LUPIS it is possible to explore the implications of combined versus
separate guidelines by including both the original individual guidelines as well as the combined ones and
exploring the outcomes that result from changing the relative weightings applied to each. Failure to
recognise these potential interactions could result in land uses being considered to be competing and
mutually exclusive when they are in fact compatible, if not complementary.

Table 6. Compatible land uses which can potentially be combined into new composite uses.
Land unit Compatible land uses Composite land uses
Ulva
Deep rooted perennial vegetation
Lupin/cereal rotation
Lupin/cereal rotation between alleys of deep
rooted perennials
Pasture/cereal rotation
Pasture/cereal rotation between alleys of deep
rooted perennials
Deep rooted perennial vegetation
Lupin/cereal rotation
Lupin/cereal rotation between alleys of deep
rooted perennials
Deep rooted perennial vegetation
Lupin/cereal rotation
Lupin/cereal rotation between alleys of deep
rooted perennials
Pasture/cereal rotation
Pasture/cereal rotation between alleys of deep
rooted perennials
Deep rooted perennial vegetation
Revegetation with Banksia woodland Banksia revegetation
Deep yellow
sand (Sa)
Gravelly uplands
(Sg)
Gravelly uplands
(St)
Gutless sands
(Sgt)
Sandy loams
(Sp)
Breakaways (Bd)
Shallow clays
(Bgc)
White gum soils
(Be)
Shallow gravelly
duplex (Dug)
Deep duplex
Pasture/cereal rotation
Pasture/pasture/cereal rotation
Deep rooted perennial vegetation
Lupin/cereal rotation
Pasture/cereal rotation
Revegetation with perennial
vegetation
Revegetation with Wandoo woodland
spp.
Revegetation with perennial
vegetation
Revegetation with Wandoo woodland
spp.
Alleys of deep rooted perennials
Pasture/cereal rotation
Alley farming with drains
Pulse/cereal rotation
Alley farming with drains
Pasture/cereal rotation between alleys of deep
rooted perennials
Pasture/pasture/cereal rotation between alleys of
deep rooted perennials
Lupin/cereal rotation between alleys of deep
rooted perennials
Pasture/cereal rotation between alleys of deep
rooted perennials
Booraan
Wandoo revegetation
Wandoo revegetation
Pasture/cereal rotation between alleys of deep
rooted perennials
Collgar
Pulse/cereal rotation with drains and alleys of
deep rooted perennials

(Dud)
White gum
duplex (Duw)
Shallow grey
duplex (Dusg)
Salmon
gum/gimlet soils
(Msl)
Salmon
gum/gimlet soils
(Mgc)
Salmon
gum/gimlet soils
(BE)
York/Jam
York
Lupin/cereals rotation
Alley farming with drains
Pasture/cereal rotation
Alley farming with drains
Pulses/cereal rotation
Perennials around recharge features
Pasture/cereal rotation
Pulses/cereals/canola/cereals rotation
Pulses/cereals/canola/cereals/pasture
rotation
Perennials around recharge features
Pasture/cereal rotation
Pulses/cereals/cereals rotation
Pulses/cereals/canola/cereals rotation
Perennials around recharge features
Pasture/cereals rotation
Pulses/cereals/cereals rotation
Lupin/cereals rotation with drains and alleys of
deep rooted perennials
Pasture/cereal rotation with drains and alleys of
deep rooted perennials
Pulses/cereal rotation with drains and alleys of
deep rooted perennials
Merredin
Pasture/cereal rotation with perennials around
recharge features
Pulses/cereals/canola/cereals rotation with
perennials around recharge features
Pulses/cereal/canola/cereals/pasture rotation with
perennials around recharge features
Pasture/cereal rotation with perennials around
recharge features
Pulses/cereals/cereals rotation with perennials
around recharge features
Pulses/cereal/canola/cereals rotation with
perennials around recharge features
Pasture/cereals rotation with perennials around
recharge features
Pulses/cereals/cereals rotation with perennials
around recharge features
Pulses/cereal/canola/cereals rotation with
Pulses/cereals/canola/cereals rotation
perennials around recharge features
Danberrin
Alleys of deep rooted perennials
Pulses/cereals rotation
Pasture/cereals rotation
Alleys of deep rooted perennials
Pulses/cereals rotation
Pasture/pasture/cereals rotation
Pulse/cereals rotation between alleys of deep
rooted perennials
Pasture/cereals rotation between alleys of deep
rooted perennials
Pulse/cereals rotation between alleys of deep
rooted perennials
Pasture/pasture/cereals rotation between alleys of
deep rooted perennials
Maps reflecting land suitability for all of the different land uses (Figures 1114) were overlayed, using
standard GIS procedures. This resulted in a composite map in which the different polygons represent the
land units to which different land uses can be allocated (Figure 16).

3.6 Land allocation
• 3.6.1 Guidelines for land allocation
• 3.6.2 Generating land-use plans
3.6.1 Guidelines for land allocation
When making decisions about what land uses are to be allocated to which portions of a landscape it is
important to be clear about the basis for those decisions. This is particularly the case where there are a
number of stakeholder groups involved in the planning exercise. By formalising the 'rules' whereby decisions
are made and by making those rules transparent, all parties are aware of how decisions are being made
and how outcomes are arrived at. The term 'guidelines' is used in a LUPIS exercise, rather than 'rules' to
reflect the fact that they are simply recommendations and that their relative importance can be changed.
Three types of guidelines are used in an allocation exercise. These are termed commitment, exclusion and
preference guidelines. Commitment guidelines are used where there is a consensus among all stakeholders
that a particular land unit can be used for one use only. For example, in the Wallatin Catchment, all
participants agreed that there should be no further clearing of remnant vegetation. This objective was
captured in a commitment guideline which stated that all existing remnants be committed to remain as
remnants. Exclusion guidelines, as the name implies, exclude specified land uses from particular land units
but allow other suitable uses. Preference guidelines allow any suitable land use to be applied to specified
land units with the decision as to which use is allocated being determined by the relative weighting applied
to each guideline.
The procedure adopted in a LUPIS exercise does not require the participants to agree about the content of
those rules before the exercise commences. Each stakeholder group or individual can submit their own
guidelines. The allocation process is then used to explore the implications of implementing the different
guidelines or of altering their relative importance.
The land uses that were identified by the various stakeholders are as follows:
• remnant vegetation protection
• heath revegetation
• Wandoo woodland revegetation
• Mallee revegetation
• Salmon gum/gimlet revegetation
• York gum/Jam wattle revegetation
• Jam wattle/York gum revegetation
• Banksia woodland revegetation
• alleys with lupin/cereal rotation
• alleys with pasture/cereal rotation
• alleys with pasture/pasture/cereal rotation
• alleys with pulses/cereals rotation
• pasture/cereals rotation with perennials around recharge features
• pulses/cereals/canola/cereals rotation with perennials around recharge features
• pulses/cereals/canola/cereals/pasture rotation with perennials around recharge features
• pulses/cereals/cereals rotation with perennials around recharge features.
The guidelines identified by the different stakeholders were then used to allocate each of the land uses to
different land units. The following statements illustrate the structure of the guidelines used for allocating the
land uses:

Figure 16. Composite map of the Wallatin Catchment showing the land units to which different land uses will
be allocated.
Commitment guidelines
• Commit all vegetation remnants (public or private) to remnant vegetation protection.
Only one commitment guideline was deemed necessary for the issues being considered. This is a nature
conservation guideline which reflects the consensus among stakeholders that all remaining patches should
be retained.
Preference guidelines
The following guidelines reflect the preferred land use options as perceived by different stakeholders. The
guidelines presented here are simply examples to provide a flavour of the type of guidelines used. The full
list of preference guidelines used in the Wallatin case study is presented in Appendix 5.
• Allocate landform Merredin to revegetation with Salmon gum /Gimlet woodland.
• Allocate areas < 20 ha on gutless sands to revegetation with Banksia woodland.
• Allocate Gravelly uplands to alleys of deep rooted perennials interspersed with lupin/cereal rotation.
• Allocate Salmon gum /gimlet soils to a pasture/cereal rotation with perennials around recharge
features.

The first two of these guidelines are designed to address nature conservation objectives while the second
two simultaneously address production and hydrological goals.
Exclusion guidelines
No exclusion guidelines were considered necessary for the current exercise. In some exercises a potential
exclusion guideline may take the form: 'Exclude grazing from all remnant vegetation'. This was not
necessary in the current exercise as the commitment guideline to retain remnant vegetation as remnants
effectively precluded this option.
3.6.2 Generating land-use plans
Having identified the stakeholders, determined the relevant issues, nominated the land uses appropriate for
addressing those issues, assessed the suitability of the landscape for those uses and developed guidelines
for allocating land uses to different parts of the landscape, the next phase is to embark on the allocation
process.
The guidelines identified in the previous section are converted to computer code and incorporated into
LUPIS. Weightings which reflect the social preferences of the stake-holders are then attached to each of the
guidelines. On the basis of these weightings LUPIS calculates the relative suitability of each land unit for
each land use and allocates each unit to the use with the highest suitability score (Ive 1992).
An important feature of the weighting procedure is that it reflects social preferences rather than relying on a
common 'currency' that can be compared across all objectives. If such a currency could be developed it
could be used in a LUPIS exercise. However, there is considerable scepticism among many stakeholders
about attempts to express non-monetary values, such as amenity and conservation, in dollar equivalents. In
a LUPIS exercise the performance of each guideline is assessed against the objectives that the guideline
was formulated to address. For example, the performance of conservation guidelines is assessed by the
capacity of the solution to deliver a specified conservation outcome. Similarly, the performance of
hydrological guidelines is judged according to the extent to which the allocation resolves a hydrological
problem. Valuing the different land uses in a common currency (inevitably dollars) and comparing the
relative performance of each guideline is meaningless if it is unable to assess performance against stated
objectives.
Because there is no common currency by which different objectives can be meaningfully compared, it is not
possible to specify the relative values that should be attributed to each guideline. For example it is not
possible to specify a priori whether a hydrological guideline should have a higher, lower or intermediate
weighting than a production or conservation guideline. As stated above, our interest in a LUPIS exercise is
to assess the performance of each guideline against the objectives that the guideline is designed to
address. LUPIS therefore allows the initial weightings to be set arbitrarily. A convenient starting point is to
assume that all guidelines are equally important and to give them all an equivalent weighting. An interim
land-use plan is generated and is assessed against the different objectives identified by the various
stakeholders. If the initial 'solution' is considered not to be performing adequately for any objective, the
weightings on the guidelines which address that objective can be increased and the exercise re-run. The
process is therefore an iterative one with subsequent plans converging towards a solution that is acceptable
to all stakeholders.
Figure 17 represents the distribution of different land uses in the Wallatin Creek Catchment following an
initial allocation based on equal weights for all guidelines. The proportion of the landscape allocated to each
land use when guidelines are equally weighted is displayed in Figure 18. For presentation purposes land
uses were combined into the following broad categories:
• remnant vegetation protection
• revegetation for nature conservation
• continuous cropping in alleys

• pasture/cropping rotation in alleys
• continuous cropping with strategic perennial plantings
• pasture/cropping rotation with strategic perennial plantings.
While Figures 17 and 18 indicate how much land has been allocated to each land use it does not tell us how
well any particular guideline is performing relative to its potential. In other words it doesn't tell us how much
of the land that is suitable for a particular use has actually been allocated to that use. Figure 19 represents
an extract from a LUPIS output which indicates the proportion of suitable land that has actually been
allocated to the use specified in each guideline. It provides an assessment of the extent to which the various
guidelines have been satisfied. Because different guidelines may represent the interests of different
stakeholder groups it is possible for all stakeholders to assess how well their interests are being addressed.
Figure 17. Results of a land-use allocation exercise for the Wallatin Catchment based on an equal weighting
for all guidelines.
Figure 18. Percent of the Wallatin Catchment allocated to each land use when all guidelines have equal
weightings.

Figure 19. Bar chart indicating the extent to which the different guidelines in a preliminary LUPIS exercise
have been satisfied. Each column represents a different guideline. The height of the column represents 100%
of the total area that could potentially have been allocated to the relevant land use as recommended by that
guideline. The blue portion of the column indicates the proportion of suitable land that has been allocated,
while the purple represents the area that is potentially suitable but which was allocated to another land use
competing for the same land unit. For clarity of presentation only, the first 10 of the total of 39 guidelines have
been shown. Guideline numbers correspond to those in Appendix 4.
A change in the performance of any one guideline as a result of changing its weighting will result in a
change in the performance of one or more other guidelines. Consequently it is possible to assess the impact
on all objectives of changing the weightings on guidelines pertaining to a particular objective. It is therefore
possible to quantify the likely impact on production of addressing nature conservation and hydrology
objectives. Such assessments are based on the assumption that productivity for any specified land use is
equivalent on equivalent soil types throughout the catchment.
Figure 20. Preferred land uses solution with no additional conservation commitment beyond the retention of
existing remnants.

Figure 21. Preferred land uses solution with conservation committment in which all 'non-viable' patches are
enhanced to meet minimum size requirements.

Figure 22. Preferred land uses solution with additional conservation commitment in which all 'non-viable'
patches are enhanced to meet minimum size requirements and additional areas are allocated to salmon
gum/gimlet reconstruction in recognition of the under-representation of this patch type in the catchment.
Figures 20 22 show the results of three different nature conservation scenarios. Figure 20 represents a
solution where conservation is limited to the protection of existing remnants. This was achieved by applying
a rating of zero to all conservation guidelines.
In Figure 21 the weightings were increased on guidelines for re-establishing habitat adjacent to patches that
failed to meet the minimum size requirements identified in Chapter 2. This solution increased the area
allocated to nature conservation by 4.2% which, when added to the existing remnants, accounts for 11.6%
of the catchment.
Figure 22 represents the results of an allocation exercise with an additional emphasis on conservation
enhancement aimed at increasing the area of salmon gum/gimlet woodland. This patch type is considered
to be under-represented in the area. This particular solution requires an additional 6% of the catchment.
Changes to the proportion of the catchment allocated to other land uses in response to these changes can
be seen in Figure 23.

Figure 23. Proportion of landscape allocated to primary land uses with three different levels of nature
conservation.
3.7 Stakeholder responses
Stakeholder groups were all understandably cautious about entering into an integrated management
process for which the outcomes were unclear. It is clearly unreasonable to expect stakeholder participation
in a process for which the level of commitment of all stakeholders and the type of outcomes are not clear at
the outset. It was therefore considered necessary to develop a prototype model for the catchment based on
preliminary information. This prototype was then presented to the various stakeholder groups in the
catchment to demonstrate its capabilities and to enable them to assess its appropriateness for the
management of their catchment. Having viewed the prototype, the various groups were then able to ensure
that the appropriate issues and objectives were being adequately addressed by the process.
Each stakeholder group had the opportunity to specify the issues, landuses, and guidelines that they
considered appropriate for the region from their own perspective. Unfortunately, because of the
incompleteness of the available information for both land and nature conservation, and for alternative and
more sustainable production land-uses, the recommended outcomes must still be considered preliminary.
Further development of guidelines will be required before an acceptable catchment plan can be generated
and implemented.
Responses by stakeholders to the preliminary solutions were positive. There was a general perception that
the requirements for addressing nature conservation objectives were within the scope of what they were
prepared to undertake and that these actions would also make a positive contribution to managing
hydrological issues and to enhancing the amenity value of the landscape. It must be recognised, however,
that land-owners in the Wallatin Catchment are among the most progressive in the wheatbelt when it comes
to the consideration of nature conservation as a legitimate land-use objective.
Further development of hydrological guidelines will be necessary before the landholders can assess their
acceptability. The strategy used in the current exercise was one of widespread alley planting. More strategic
solutions if available, would be viewed more favourably by the group.
Identification of the Wallatin Catchment as a Focus Catchment under the State Government Salinity Action
Strategy will ensure that further advice is made available to the catchment group. The procedures
developed in this study will provide a basis for catchment planning under this strategy. Ideally, the future
development of this approach should include consideration of a wider array of land-uses such as the
commercial use of indigenous plants and animals, including wildflower enterprises and the production of oil
from Eucalyptus and Melaleuca species. Future planning should also consider the increasing demand for
'cleaner and greener' produce. However, it will be necessary to convince land-holders that these new
activities will be commercially viable.

3.8 Data adequacy
A prerequisite for reliable land-use allocation is the availability of accurate maps which reliably reflect land
capability. The map sets available for this exercise were derived from a number of different sources.
Landform maps were produced by W. M. McArthur using air photo interpretation and ground survey. These
maps were produced with the objective of partitioning the landscape into units which reflected a combination
of attributes including vegetation type, soil type, and position in the landscape. Remnant vegetation was
mapped from aerial photographs and digitised for incorporation into a geographic information system by
Agriculture WA and CSIRO Division of Wildlife and Ecology. Vegetation communities within remnants were
mapped using Landsat MSS and Landsat TM satellite images (Hobbs et al. 1989; Lambeck & Wallace
1993) combined with air photo interpretation and ground surveys to validate and, where necessary, correct
the satellite data. Soil types outside of the remnants were mapped by individual farmers to reflect suitability
for agricultural land uses using procedures described in Hawkins (1990).
In practice, many of the above map sets should be correlated to some degree. The distribution of soil types
should reflect the landforms in the area and vegetation should show some correspondence to both soils and
landforms. However, the extent of these relationships between map sets appeared to vary throughout the
study area reflecting differences in the interpretation of the landscape by different people. This was
particularly evident where individual farmers were responsible for developing soil maps of their properties.
For some farms, the soil was mapped at a very fine resolution, reflecting small scale variation in soil types.
Other farmers appeared more pragmatic in their mapping, primarily capturing differences that best reflected
land capability for current farming practice. Neither of these approaches is more correct than the other. They
simply reflect differences between individuals in their perception of the purpose of the mapping exercise. In
addition, attempts to create discrete boundaries between soil types which grade into each other is a
subjective process. No two individuals could derive the same map and probably no one individual could
produce the same map twice. As a consequence, soil boundaries often did not match on opposite sides of a
farm boundary and soil boundaries failed to match vegetation boundaries at the edges of remnants. In
addition, if remnant vegetation as mapped by Government agencies does not match the maps generated by
landholders, further errors can be introduced into the exercise and the reliability of the outcomes will
diminish correspondingly.
Because of the discrepancies between vegetation and soil maps it was necessary, in this exercise, to use
landform maps as a surrogate for vegetation types as this was the only data set that was available both
inside and outside of the remnants. However, it is clear from satellite mapping that the match between
vegetation types and mapped landforms is also relatively poor. Consequently the results generated by the
land allocation exercises can only be considered as a general guide to what can be done rather than a
prescription that must be implemented exactly as mapped.
For this procedure to more effectively reflect the needs for agriculture and the needs of the plants and
animals in the landscape, it will be necessary to ensure greater correspondence between data sets. This
should be based on standardised data capture protocols and appropriate validation of maps.
Hydrological guidelines for this exercise were based on general landscape principles. As a consequence,
there are no criteria for assessing the likely impact of the specified actions on water table levels. If such
predictive capacities are required it will be necessary to develop hydrological models based on reliable
maps of geomorphological features and high resolution digital elevation models. Failure to develop effective
strategies for managing hydrology will not only compromise the agricultural values of the catchment, but will
also threaten much of the remnant vegetation that is essential for protecting the remaining biota.
The ramifications of these shortcoming in data quality are significant. The reliability of the results from any
planning process will only be as good as the quality of the information that is used. Even the best models
and procedures cannot convert poor information into reliable solutions. If planning seeks to provide reliable
outcomes, then it must be based on reliable inputs. The absence of standards and consistency in data
collection, the lack of agreement about the type of data that should be collected, and the lack of resources
for gathering the information even if such agreement can be reached, present major problems for any

egional planning process. Planning based on poor quality data will result in actions which do not deliver the
required outcomes and will alienate stakeholders who enter into the planning process in good faith with an
expectation that their problems will be addressed. However, the absence of quality information should not
provide an excuse for not undertaking actions that are urgently needed. Responses to data limitations for
hydrological objectives should be treated in a similar way to those recommended for nature conservation:
where there is an urgent need to act and the necessary information is not available, then general principles
should be employed. Where the situation is less urgent, attempts should be made to acquire the information
required for a more strategic approach.
Regional approaches to planning can help to reduce these data limitations. By identifying areas that are
sufficiently similar in regard to natural and human-induced pattern it will be possible to develop
recommendations based on quality data acquired for a subset of the region and legitimately apply those
results beyond the location where they were generated.