Integration of biodiversity and primary production in...

_________________________________________________________________ Integrating Biodiversity and Primary Production in Northern and Western New South Wales 1

Integration of biodiversity and

primary production in northern and

western New South Wales

September 2008

Cathy Waters and Ron Hacker NSW Department of Primary Industries

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Report to the Namoi, Border Rivers-Gwydir and Western

Acknowledgments Denys Garden, Peter Orchard and Sue McIntyre are particularly thanked for efforts in reviewing sections of this document. We also acknowledge and thank the land managers interviewed for the case studies, Graham Strong, Cathy and Graham Finlayson, Bruce Maynard, Harvey Gaynor, Terry Haynes, Richard and Janet Doyle and Tim and Karen Wright. The authors would also like to thank the following individuals for providing assistance and/or contributing to the information found within the review; Trudie Atkinson Warwick Badgery Yin Chan Saul Cunningham Greg Curran Sam Davis Josh Dorrough Joern Fischer Toni George Philip Gibbons Christine Jones Jamie Kirkpatrick Rebecca Lines-Kelly Leah MacKinnon Adrian Manning Nicki Munro Ian Oliver Nick Reid Greg Rummery Jenny Stott Wal Whalley Rick Young


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Summary The Western, Namoi and Border Rivers-Gwydir CMA areas represent a large portion of NSW and cover a range of agricultural land use activities including higher rainfall permanent pastures, extensive crop production and semi-arid rangelands. Within these varied agricultural landscapes, the restoration or maintenance of ecosystem function should result in increases in sustainable production. Since many processes essential to ecosystem function are underpinned by biological diversity management should also aim to maintain or improve this component of agricultural landscapes. This review identifies a ‘portfolio of partial solutions’, or management practices required to maintain or improve biodiversity, clearly recognising that this outcome will require a plurality of approaches that are tailored to the local environment (including seasonal conditions, land-use and personal aspirations). An overarching theme is recognition of the central role of landscape heterogeneity in integrating biodiversity in agricultural production systems and the need for any change in land use should to be targeted towards increased heterogeneity. At the farm level, the review identifies numerous management practices that would positively impact on biodiversity including:

• agronomic practices – management for soil physical and chemical properties, microbial activity, weeds and crop pests;

• management of non-crop areas that are intrinsically linked to agricultural production – habitat for beneficial species or mitigation of negative impacts, and

• management directly for ecosystem services such as desirable wildlife or plant species, or carbon sequestration.

While farm–scale practices are the primary focus of the review we emphasise the need to consider these in the context of the large scale dimension of biodiversity and provide a limited number of examples of how a landscape approach to conservation of biodiversity has been or could be undertaken. At the farm scale, the general lack of empirical information on the impact of changed management practices on biodiversity means that novel management approaches will need to be examined in an adaptive management framework that allows feedback to other landholders, allowing them to apply new management within their own context. In addition, the monitoring and evaluation of on-going and past ‘restoration’ activities, as well as long-term changes in management, will be essential to provide a basic understanding of these relationships. This review identifies a vital deficiency in both the appreciation and evaluation of ecosystem services within the agricultural sector. Little attention has been paid to understanding the relationships between farming system inputs and ecosystem outputs, and the role of biodiversity in mediating these relationships. Provision of ecosystem services is likely to be particularly prone to market failure since tradeoffs are involved between the private benefits of marketed agricultural products and the predominantly public benefits of non-marketed ecosystems services such as biodiversity. In some instances, public and private interests may be compatible (e.g. the provision of cropping benefits through pollination or beneficial insects by retention of native vegetation). In these instances the market should work efficiently to achieve both public and private benefits. However, in many instances the public and private benefits are likely to be incompatible and financial or other incentives will be required if ecosystem services are to be delivered by agricultural producers in the public interest. A means of estimating the value of ecosystem services is required in order to determine the level of such incentives but development of a means of achieving this is beyond the scope of this report.


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Contents 1. Project Scope …………………………………………………...….…… 4 1.1 Background and aims……………………………………………………. 4 1.2 Definition of biodiversity and its value to agricultural production……. 5 2. Agricultural/pastoral practices that allow for the integration

of biodiversity and production…………………………………………9 2.1 Rearrangement or re-design of land use…………………………… 10 2.1.1 Importance of heterogeneity…………………………………………..... 10 2.1.2 Diversification and re-design of land-use to incorporate conservation values……………………………………………………… 12 2.1.3 The cost of re-arranging land use……………………………………….14 2.2 Remnant management………………………………………………….16 2.2.1 Remnant enhancement………………………………………………….. 16 2.2.2 Genetic integrity and local seed sources……………………………… 20 2.3 Grazing management………………………………………………….. 22 2.3.1 Domestic livestock and native vegetation……………………………... 22 2.3.2 Grazing management options...………………………………………… 23 2.3.3 Grazing impacts in the semi-arid rangelands…………………………. 28 2.3.4 Total grazing pressure management…..………………………………. 32 3.3.5 Fertility and native pastures…………………………………………….. 33 2.4 Farming practice.............................................................................. 35 2.4.1 Intensive cropping………………………………………………………...37 2.4.2 Tillage systems…………………………………………………………… 38 2.4.3 Cropping rotations……………………………………………………...... 40 2.4.4 Pasture Cropping………………………………………………………… 40 2.4.5 Alley farming……………………………………………………………… 42 2.5 Weed and pest management…………………………………………. 42 2.5.1 Weed management……………………………………………………… 42 2.5.2 Integrated pest management…………………………………………… 43 2.5.3 Invasive native scrub…………………………………………………….. 44 2.6 Water point management……………………………………………… 45 2.6.1 Riparian areas, farm dams and irrigation channels…………………... 46 2.6.2 Water point management in rangelands………………………………. 46


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3. A whole of landscape approach for integration of biodiversity and production……………………………………….. 49

4. Indicators of biodiversity health relevant

to agricultural production systems………………………………….. 52 5. Case Studies…………………………………………………………….. 53

1. Strong Family “Acacida” Narrandera…………………………... 54 2. Tim and Karen Wright “Lana” Uralla…………………………… 55 3. Auscott “Midkin” Moree………………………………………….. 56 4. Graham and Kathy Finlayson “Bokhara Plains” Brewarrina… 57 5. Bruce Maynard “Willydah” Narromine…………………………. 58 6. Richard and Janet Doyle “Malgarai” Boggabilla……………… 59

6. Recommendations………………………………………………………60 6.1 Management focused explicitly on the creation of heterogeneity 6.2 Re-design of land use for biodiversity outcomes 6.3 Options for micro-restoration of highly modified landscapes 6.4 Assessment of long-term impacts of high intensity short duration

stocking on native pastures 6.5 Biodiversity indicators of utility value for agricultural producers 6.6 Flexible farming and grazing management 7. Bibliography………………………………………………………………63


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1. Project scope

1.1 Background and aims

The International Day for Biological Diversity held in May 2008, recognised that sustainable agriculture promotes and is enhanced by biodiversity, but that agriculture has also been a major cause of biodiversity loss. To date, the responses to address these declines in biodiversity have been to set up conservation reserve schemes, and (in Australia) legislation to conserve vegetation and threatened species. However, both local and global biodiversity continues to decline at alarming rates (Millennium Ecosystem Assessment 2005). Conservation of biodiversity within agricultural landscapes is essential but, on its own, reserving land for conservation does not provide an effective means of doing this. Ultimately, the sustainability of our agricultural systems will be judged not only by their success in sustaining world food production, agricultural livelihoods and enhanced human wellbeing, but also by the extent to which they conserve biodiversity (The Convention on Biodiversity, 2008). In Australia, ‘Caring for our Country’, due to commence in July 2008, has as a major goal to provide, “an environment that is healthy, better protected, well managed, resilient and provides essential ecosystem services in a changing climate” (Caring for Our Country, 2008). While some synergy between on-farm management of biodiversity and long-term financial viability has generally been assumed, concerns have been raised by both landholders and Catchment Management Authority (CMA) staff regarding the specific nature of this interaction. Documentation of management practices that are most likely to achieve benefits for both farm financial viability and biodiversity conservation is lacking. Responding to these concerns, a Joint CMA Biodiversity and Production Working Group (B&PWG) was formed between Western (WCMA), Namoi (NCMA) and Border Rivers-Gwydir (BR-GCMA) CMAs to identify opportunities for more effective integration of production and biodiversity conservation. Specifically, the aim of the B&PWG was to review the available scientific evidence relating agricultural management practices to biodiversity outcomes at farm and regional scales. This report represents a synthesis of knowledge which can provide practical recommendations and identify knowledge gaps. Case studies that illustrate the practical application of some of these recommendations are also provided. Within this document we clearly recognise that four approaches are available to achieve biodiversity conservation on private land. These include regulation, voluntary measures, economic incentives and motivation approaches (education and community-based programs). The relative merits and short-comings of each approach have been reviewed by Curran (2000). The value of education and community-based programs is that they allow individuals to respond to a range of measures, whether voluntary, financial or regulatory. Importantly, this report also provides a basis for the development of training workshops and development packages in the second stage of the project.


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1.2 Definition of biodiversity and its value to agricultural production Definitions of biodiversity generally cover all life forms, are vague, and only of limited utility value to the agricultural sector. For example, a classic definition describes biodiversity as, ‘the variety and variability among living organisms and the ecological complexes in which they occur’ (Office of Technology Assessment 1987). Here, biodiversity includes not only biotic1 but also abiotic2 ecological processes. A more recent, simpler definition, closely following that of Noss (1990), describes biodiversity as, “the variety of life, its composition, structure and function, at a range of scales” (Freudenberger and Harvey 2003). This definition incorporates the important notion that biodiversity is related to scale, a major issue (as will be shown below) for the conservation of biodiversity in agricultural landscapes. This notion of ‘scale-dependency’ includes not only the understanding that individual organisms vary enormously in their requirements for space but also that conservation of the ‘variety of life’ implies a requirement for maintenance of species populations and their inherent variability across their natural ranges. Conservation of biodiversity must thus be considered at both the small (e.g. paddock or farm) and large (e.g. sub-catchment or catchment) scales. Because the definitions of biodiversity are broad, different people will inevitably attach different interpretations to biodiversity. For example, naturalists may be more interested in describing an unambiguous name for a species, a scientist may be interested in the genetic diversity of a species and an environmentalist may be more motivated by protection or preservation of a species (Mayer 2006). These definitions provide no clear application to real life and a definition of biodiversity (simple, comprehensive or operational) that is responsive to real-life management is unlikely to be found. Mayer (2006) suggests a more useful definition would be to describe biodiversity as an intrinsic character, with each of the attributes of structure, function and composition operating at a range of temporal and bio-geographical scales (Table 1.1). Ecosystems are generally considered as the largest units of biodiversity and comprise, “an amalgam of habitats, the species within them and importantly the processes occurring within them” (Doherty et al. 2000). While an ecosystem function may be characterised by, for example, the ability of soils to break down plant matter and liberate carbon dioxide, nutrients and water, ecosystem processes that support this function may be nitrogen mineralisation and decomposition (Doherty et al. 2000). Recognition that conservation of biodiversity involves ‘bio-geographic’ scales from genes to landscapes provides the major rationale for attempts to conserve biodiversity within agricultural production systems, since all of these ‘scales’ cannot be encompassed within discrete areas set aside for 1 Biotic factors include the living component of the ecosystem e.g. plants, animals and fungi 2 Abiotic factors include non-living or the physical component of an ecosystem e.g. soils, water, climate and plant litter


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conservation. This is particularly true at the ‘genetic’ scale, since conservation of biodiversity at this level requires maintenance of viable species populations throughout their natural range. Indeed, the importance for biodiversity conservation of beneficial integration with agricultural production increases as we move down the hierarchy from landscape to genetic scale – representative conservation areas may well hope to preserve examples of landscapes, but can never expect to conserve the genetic variation within species. While biodiversity in the sense discussed above - subsequently called biodiversity per se -, is considered important in its own right (e.g. the potential for future benefits or simply the right to exist), it also underpins the functioning of ecosystem services (ES), the conditions and process through which ecosystems sustain human life (Daily 1997). Landscapes are primarily managed to optimise ‘provisioning’ ES of food and fibre production from agriculture, but depend on ‘regulating’ services (e.g. pollination, drought and flood control) and, more importantly, ‘supporting’ services (e.g. soil fertility and nutrient cycling) that underpin all other ecosystem services (Zhang et al. 2007) (Table 1.2). Critical provisioning services to agriculture (which are often forgotten) are management practices that maintain, restore or regenerate ecological processes that depend in some way on biodiversity (Zhang et al. 2007). In agriculture, recognising the link between ecological functions and biodiversity is the first step towards understanding the utility value of biodiversity. However, acknowledging the potential for agriculture to adversely impact on biodiversity per se at a range of scales is also essential to recognising the dynamic relationship between them. Broadly, agroecosystems comprise managed productive areas (intensive agriculture) and the semi-natural or natural areas surrounding these and areas of human settlement. Semi-natural areas may be seen as a threat to the productive areas (e.g. source of pests) and productive areas seen to have a negative impact on biodiversity per se, with little regard for their actual or potential interdependence. Agricultural management practices should aim to maximise the positive benefits of biodiversity in the form of ecosystem services, while minimising the negative impacts on biodiversity per se. In the following section, major agricultural practises are examined in the context of their positive and negative impacts. However, these impacts should be considered within the context of the condition of the landscape and its capacity to respond to changes in management. For example, Figure1 shows that highly modified areas (low biodiversity) tend to the most productive agricultural areas and management should aim to minimise off-site impacts; in areas of high biodiversity, agricultural production is generally low and these areas may be most suited to conservation; where agricultural production losses are associated with biodiversity benefits, financial incentives may be required.


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Table 1.1 Factors which are important in describing the character of the three major

attributes of structure, function and composition of biodiversity. Each attribute occurs at different scales highlighting the hierarchical nature of biodiversity (Adapted from Noss 1990)

Attribute

Composition

Structure

Function

Landscape scale

Landscape types or ecosystems Species distribution and richness Proportions of vegetation patches

Landscape pattern Heterogeneity and connectivity

Landscape processes Disturbance processes (crop and pasture rotations, seasonality of crop/pasture growth, intensity of grazing) Nutrient cycling rates Persistence of vegetation patches Rates of erosion Disruption of hydrological processes (e.g. salinity)

Community ecosystem

Community ecosystems Relative abundance, richness, evenness and diversity of species Proportion of endemics, exotic, endangered, threatened species Proportion of C4:C3 plant species

Habitat structure Edaphic characteristics, aspect, slope; vegetation biomass, horizontal patchiness Abundance of features (cliffs, outcrops, gilgais, logs)

Inter-specific interactions Biomass and resource productivity Parasitism and predation rates Colonisation and local extinction rates Fine scale disturbance processes Nutrient cycling rates

Population/species scale

Species populations Absolute or relative abundance Frequency Importance of cover Biomass Density

Population structure Population structure (sex and age ratios) Habitat variables (see community ecosystem above) Morphological variability

Demographic processes Fertility, recruitment rates, mortality rates Flowering characteristics

Genetic scale

Genes Allelic diversity Presence of particular rare alleles Deleterious recessives Karyotypic variation

Genetic structure Population size Heterozygosity Chromosomal phenotypic variation

Genetic processes Inbreeding/outbreeding depression Rate of genetic drift Gene flow Mutation rates Selection intensity


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Table 1.2 Major ecosystem services and disservices supporting agricultural ecosystems, the scales over which they are provided and the groups of species that underpin the service. (Adapted from Zhang et al. 2007)

Ecosystem service

Paddock

Farm

Landscape

Supporting Services

• Soil formation and fertility, nutrient cycling

Microbes, invertebrate communities; legumes

Vegetation cover

• genetic diversity Crop diversity for pest and disease resistance

Genetic resource of wild populations

• Water cycling (infiltration and runoff)

Plant residues, cover crops, rotations, soil microbes (soil structure)

Plant residues, cover crops, rotations, soil microbes (soil structure)

Large scale remnant vegetation patches

Regulating Services

• Soil retention

Perennial ground cover Cover crops

Perennial ground cover Cover crops

Riparian and flood plain vegetation, vegetation cover in catchment

• Pollination

Ground nesting bees (majority native species) and other pollinating animals; native vegetation patches or paddock borders

Bees (largely exotic species) and other pollinating animals; native vegetation patched or paddock borders.

Remnant vegetation

• Dung burial

Native and exotic dung beetles



• Natural control of plant pests

Predators and parasitoids Predators and parasitoids

• Food sources and habitat for beneficial insects

Remnant vegetation Remnant vegetation Remnant vegetation

• Water purification

Remnant vegetation Vegetation around drainage ditches and dams

Vegetation cover in catchment

• Atmosphere/climate

Vegetation influencing microclimate (e.g. agroforestry, shelter belts)

Vegetation influencing microclimate

Vegetation influencing stability of local climate; amount of rainfall, temperature Vegetation and soils for carbon sequestration and storage

Ecosystem disservices

• pest damage

Insects, snails, birds mammals, fungi, bacteria, viruses, weeds

Remnant vegetation. Birds, mammals, weeds

Remnant vegetation, birds, mammals, weeds, insects (e.g. locusts)

• Competition for water from other ecosystems

Weeds Vegetation cover Vegetation cover in catchment

• Competition for pollination

Flowering weeds Flowering weeds Flowering plants in catchment

Agricultural production

Bio

dive

rsity

Conservation agreements, land sparing

Incentives required Duty of care

(biodiversity improvement without loss in production)

Production only (duty of care, minimise off-site impacts)

Conservation Native vegetation

Native pastures-fertilised/unfertilised

Cleared areas, highly productive intensive agriculture

Figure 1. Anticipated changes in agricultural production with increasing in biodiversity

(native and exotic) and associated requirements for duty of care and financial incentives.

2. Agricultural/pastoral practices that allow for the integration of

biodiversity and production At the farm scale, if the objective is to return a site to some historical condition, the context of contemporary land use is largely ignored. In this case conservation and primary production will always be set up as conflicting interests. However, setting goals that are compatible with current or proposed land use will ensure that biodiversity and production goals can be integrated. In this section, we promote the concept that the benefits of biodiversity to primary production are largely dependent on the maintenance of ecosystem processes that can ensure both long-term landscape viability and long-term agricultural production. Goals for integrating biodiversity and production should therefore focus on restoring or maintaining the viability of native species and plant communities that support these ecosystem processes, rather than on restoring plant community composition in some historical sense. Thus, agricultural practices should seek to maintain, for example, vital ecological processes such as nutrient cycling. Since the relationship between biodiversity and ecosystem processes assumes some capacity for self-design landscapes in which these processes are maintained should have greater resilience and capacity to respond to predicted climate changes. The following sections evaluate examples of key management practices that may achieve these goals.


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2.1 Rearrangement or re-design of land use ‘Where the landscape is sufficiently heterogeneous at all spatial scales, different taxa will find their own habitats’ ‘Land use should be tailored toward or targeted to increase heterogeneity’ 2.1.1 Importance of heterogeneity Norton and Miller (2000) argue that different parts of the landscape yield different production and conservation values and that land use can be rearranged to reflect these differences. To illustrate how this can occur, the simplest rearrangement may involve agricultural land managed specifically for commodity production (typically intensive enterprises using high inputs), and biodiversity restricted to areas set aside from agriculture (land sparing). The most complex rearrangement can incorporate heterogeneity (‘wild-life friendly farming’). In this latter example, patches of vegetation are retained, a diversity of crops may be planted, paddock sizes may be smaller, scattered trees or vegetation at paddock margins retained and, typically, conservation and production areas become less well defined (Fischer et al. 2008). A major contrast between these two extremes is that ‘wild-life friendly’ farming characteristically involves greater variability (heterogeneity) at smaller spatial scales than land sparing. However, the notion of heterogeneity may not simply be confined to a description of habitat heterogeneity (Oliver et al. 2007) but extended to include ecological heterogeneity, where the landscape supports many ecological processes (Benton et al. 2003). Thus, where land capability varies and land use is to be matched to this variation, it also implies variation in land use or management. Where landscapes are homogeneous (biodiversity is low), common land use may exist. Thus, a central idea of matching land use to land capability is recognising that heterogeneity is the key to maintaining biodiversity. There is a lot of discussion about the importance of maintaining or building ‘connectivity’ in the landscape and its relationship to the preservation of biodiversity. However, different species have different requirements in terms of connectivity (e.g. differing spatial arrangements and composition of remnant vegetation). For example, the common bird species Cacatua galerita (Sulphur Crested Cockatoo) requires open grasslands while Platycercus elegans (Crimson Rosella) requires woodland habitat (Dorrough et al. 2008a). Thus, maintaining the connectivity of, say, grasslands and woodlands may be of equal importance for the preservation of habitat to support a suite of species. Although the specific habitat requirements of plants and animals may differ, if the environment is sufficiently heterogeneous at all spatial scales, different taxa will find their own habitats (Part and Soderstrom 1999). Using birds as an example again, a recent British study found that complex landscapes with diverse vegetation structure and composition resulted in higher species richness with populations that were more stable or persistent (Devictor and


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Jiguet 2007). In Australia, over past decades a good deal of time has been spent tree planting, largely as strip plantings along fence lines. While the biodiversity benefits of these activities may be debated, (see section on remnant management), this has essentially resulted in a similar landscape ‘design’, or management practice across whole catchments. Despite the benefits of tree planting in highly modified landscapes, the common land management practice of pastures with paddock perimeter tree planting is still working to homogenise landscapes (and minimise further heterogeneity). However, ecological heterogeneity is also lost as agricultural intensification occurs. For example, inappropriate grazing management in the WCMA has resulted in widespread native woody shrub invasion and an associated decline in farm profitability and loss of floristic biodiversity. Here, the competitive interaction between woody shrubs and perennial grasses is shifted to favour unpalatable woody weeds as grazing pressures are increased. In the NCMA and BR-GCMA, intensification of cropping activities in some areas has led to large contiguous areas dominated by tillage under a common management, increasing the vulnerability of cropping enterprises to variable climatic conditions. The global evidence that links farm biodiversity decline to agricultural intensification and landscape homogenisation is clearly recognised (Benton et al. 2003; Millennium Ecosystem Assessment 2005). Thus, the re-establishment of heterogeneity in agricultural areas may provide an improvement in biodiversity. Theoretically, all agricultural practices can be tailored to increase rather than eliminate heterogeneity. Currently, CMA’s are not focusing explicitly on the creation and management of heterogeneity on agricultural lands. Where land use is to be rearranged, the capability of different land management units must be recognised. This can be facilitated by identifying differences in land capability either through training within courses that specifically match land use to land capability3 or as part of whole property planning. It is important to note that for the most productive agricultural areas (highest productive capability) there is likely to be a conflict between achieving biodiversity gains and losing agricultural production. The capacity to fence off areas to meet the management requirements for different enterprises e.g. for saltbush lamb production [Case Study 1] or for conservation [Case Study 2] is fundamental to the matching of land use and land capability on specific land units. There are other examples where fencing has been used with the specific aim of excluding additional grazing pressure from kangaroos or the exclusion of livestock for water point management [Case Studies 3 and 4]. While there has been some past criticism over the allocation of funds for fencing for biodiversity enhancement, in most case studies reported here part funding for fencing costs had been crucial in achieving both production and conservation/regeneration outcomes. ’Wildlife-friendly fencing’ essentially involves the removal/replacement of barbed wire or making it more visible (Wild-life Friendly Fencing 2008). While there has essentially been no research evaluating the effectiveness of ‘wild-life friendly fencing’, many

3 A course that explicitly does this is Landscan® as part of the NSW DPI PROfarm course, Landscan (2008)


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landholders we spoke to were using portable electric fencing without barbed wire. 2.1.2 Diversification and re-design of land use to incorporate

conservation values In Europe, the concept of ‘multifunctional’ landscape is seen as a mechanism for delivering both ecological and economic outcomes through the incorporation of agricultural subsidies that target ecosystem goods and services (Otte and Simmering 2007). Increasingly, incentive and market-based programs are being piloted in Australia as a mechanism for providing payment for conservation outcomes. Stewardship schemes provide payment to a landholder for voluntarily undertaking a “stewardship service” to either maintain or to improve the current natural resource condition of their land. Importantly, payment is offered for services to a part or whole of a property above the expected minimum acceptable practice. In the United States, the Conservation Reserve Program, introduced in 1985 and running for over 10 years, included payments for the retirement of land from cropping into permanent perennial pastures. The subsequent US Conservation Security Program ran for less time but rewarded landholders for demonstrating environmental leadership with the motto, “reward the best, motivate the rest”. European stewardship schemes have been operating since the 1980s but became part of the Common Agricultural Policy Reform of the EU in 1992 (Hanrahan and Zinn 2005). The equivalent of almost four billion euros is reported to be paid annually through agro-environmental schemes in Europe and America and has proven successful in reversing declines in wildlife populations (Donald and Evans 2006). These authors suggest that additional theoretical benefits of such schemes include the restoration of areas that surrounding agricultural farm land, allowing species to adapt to climate change and slowing the spread of invasive species. The Environmental Stewardship Program run in the United Kingdom provides payment for effective environmental management. For example, entry into the Rural Stewardship Scheme in Scotland occurs through an inspection or environmental audit that uses a ranking system to calculate a total score reflecting the conservation benefit of a management proposal (The Rural Stewardship Scheme 2005). In this way, the procedure is not dissimilar to that of the project assessment process run by various CMA’s throughout Australia. However, the European situation differs from that in Australia in that European ecosystems have evolved under centuries of farming so that certain farming activities are seen as integral to the maintenance of biodiversity (e.g. coppicing and pollarding of trees - see Dobbs and Pretty 2004 for more detail). In this European context, the abandonment of agricultural land use, as much as agricultural intensification, is seen as resulting in declines in biodiversity. As a result, farmers may receive income support, but also be eligible to participate in stewardship arrangements (Comerford et al. 2006). In Australia, our ecosystems have not evolved with long-term European agricultural practices, so stewardship payments are being delivered to effect the abandonment of some inappropriate European practices. However, there


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has been little long-term empirical data collected to support the adoption of alternative practices, highlighting the importance of monitoring and evaluation of changed management practices wherever possible. Enterprise Based Conservation (EBC) is an on-going Australia program run by the WCMA aimed at developing conservation as a long-term, competitive alternative land use (Enterprise Based Conservation 2008). Here, payments are made to landholders that match returns from local land uses such as grazing. As with the European stewardship programs, specific management is undertaken beyond what is considered as common practice. These payments are linked to an assessment of landscape condition (e.g maintaining ground cover or maintaining control of total grazing pressure) and, importantly, scaled to local seasonal conditions [Case Study 4]. The development of a national, public funded stewardship scheme such as EBC would not only result in biodiversity benefits but may also represent a more effective use of public funds than Exceptional Circumstance assistance. Comparable programs for maintaining ground cover are being implemented within BR-GCMA under a Sustainable Grazing program where landholders are involved in monitoring ground cover. However, while the environmental benefits of maintaining ground cover may be many, the relationship between ground cover and biodiversity has not been clearly established. The European and USA examples of stewardship schemes have resulted in substantial environmental benefits, and studies have indicated that such schemes can be designed to maintain essential ecosystem services in Australia (Comerford et al. 2006). The soon-to-commence Australian Government initiative, Environmental Stewardship Programme, supports a market-based approach to environmental management where targeted environmental assets (largely extensive, high quality assets) can be enhanced through incentive payments (Caring for our Country, 2008). There appears to be no plan for national schemes such as EBC that specifically target good agricultural or grazing practices. It is also unlikely that bodies such as CMAs will be able to financially support such payments. However a possible mechanism to maintain these schemes may be the formation of alliances between local councils and CMAs. For example, the Hume City Council (Western Australia) commenced such an alliance in early 2007 to support management activities associated with the protection of ecologically significant vegetation. While additional financial support for this alliance has been provided by Federal funding through NHT, and some management activities are to be undertaken by landholders, works such as fencing, weed and vertebrate pest control are also planned to be carried out by contractors (Ecotracks, 2007). As the development of national stewardship schemes specifically for primary industries may be unlikely in the near to medium term, identifying management practices that integrate production and biodiversity becomes more important in providing motivational tools for landholders. Biodiversity Banking (BioBanking) is an Australian off-set scheme that is due to commence in mid 2008. Here, development impacts can be off-set by negotiating conservation activities on privately owned land. Offsets are measured in terms of credits created by the landholder for enhancing and


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protecting biodiversity values and can be sold for income and bought by developers. An agreement between the developer and landholder is supported by the BioBanking Trust Fund, administered and monitored through the Department of Environment and Climate Change (BioBanking, 2007). A key feature of this scheme is that credits are matched to the impacts on ecosystem or species. 2.1.3 The cost of re-arranging landuse Changes in land use will often, at least in the short term, incur costs. Land use can be re-arranged so that the dual purposes of agricultural production and biodiversity conservation can be maintained to minimise the impact of these costs. For example, the construction of windbreaks to benefit livestock or tree planting for erosion control may be located to also provide linkages between remnant vegetation. Sometimes specific parts of the landscape will have the potential for dual purposes, e.g. the regeneration of native vegetation in riparian areas can provide habitat for wildlife as well as buffers for the management of spray drift in intensive cropping areas [Case Study 6]. A number of case studies in this report demonstrate the effectiveness of alternative grazing regimes in increasing the native component of pastures or increasing ground cover while also increasing profitability. In the WCMA there were also some reports of replacing breeding livestock with long-term agistment options as a means of gaining flexibility to allow resting of native pastures (G. Curran pers. comm.). From the case studies used in this report, changes in management did not result in a negative economic impact on enterprise profitability. In some cases, additional income from farm-stay enterprises was important in providing income, particularly through recent drought. It is unlikely that ecotourism or farm-stay opportunities represent a successful regional scale alternative land use option, and re-arrangement of land use may prove more broadly applicable. Bathgate et al. (2008) evaluated a range of land use scenarios in terms of economic and biodiversity outcomes at a catchment scale in central western NSW. While shifts in farming systems (varying levels of adoption of alley farming with forage shrubs or pasture cropping on suitable land units) did achieve small improvements in catchment-scale biodiversity and increase farm profit, scenarios that resulted in substantial improvements in biodiversity resulted in major reductions in farm profit. Similarly, a study of three mixed farms in southern Queensland and northern NSW showed that re-arrangement of farm activities to meet conservation outcomes did adversely affect income and that there was little opportunity for offsetting these losses within the farm enterprise (House et al. 2008). In this study, scenario modelling was used to evaluate the economic impacts of a number of options to improve protection and connectivity of the landscape. Landscape configuration was enhanced through the conservation of high quality habitat areas, increased connectivity of existing vegetation and increasing the core size of vegetation patches. Opportunities to increase agricultural production were made through either the addition of pasture, or the removal of trees in cropping areas as a means of off-setting the cost of


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applying enhancement scenarios. Under these constraints, improvements in conservation were linked to a reduction in agricultural production, and the authors suggest that until national schemes to subsidise this lost production are in place, regional or catchment scale improvement in natural resource is unlikely. Modelling different options for multifunction agricultural landscapes in Europe to include increases in plant diversity, spatial arrangement of vegetation and reduced nutrient losses at the farm scale provide additional evidence for the negative impacts of altered landscape design on gross margins (Groot et al. 2007). A number of important issues should be considered in evaluating the applicability of the results of the studies and literature reports reviewed here. Firstly, where multiple ‘farms’ were studied, investment returns tended to vary between farms - in some cases no or small positive impacts on farm profitability were achieved. Secondly, it was generally assumed that the area under agricultural production related to profitability, so increases in gross margins were associated with increases in area under primary production. Also, no account was made of increased efficiencies (in particular, reduced input costs), and no allowance for off-set payments (e.g. carbon accreditation schemes). More importantly, no costing of natural capital or ecosystem service value was included in these analyses. This appears to be a feature of economic modelling generally, despite global acknowledgement of its importance (Millennium Ecosystem Assessment 2005) and an ability to value these services (Aronson et al. 2007). Activities to improve landscape configuration beyond removing agricultural land from production (setting aside areas for conservation) are largely opportunistic or based on best-bet scenarios with little information available to assess improvements to profits and biodiversity either in the short or long terms. While the case studies seem to reveal individual farm profitability may not be adversely affected by small changes in land use to enhance biodiversity, whether these changes equate to significant large-scale (catchment scale) biodiversity gains remains unknown.


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Landscape heterogeneity

Impact on Biodiversity per se.

Impact on production Negative

Positive

Positive

Agricultural intensification with larger farm size and large contiguous areas dominated by common land use (e.g. simplified cropping rotations, higher input farming systems, higher off-site negative impacts on biodiversity; pasture improvement reducing species diversity and habitat diversity).

Mixed farm enterprises where land use is matched to land capability; the incorporation of multiple purpose land use and ecological or environmental indicators incorporated in management decisions (e.g. ground cover or species utilisation level).

Negative

Intense agricultural production operating where key ecological processes have collapsed (e.g. patch dynamics and hydrological functions lost, fire excluded from fire-dependent systems; woody weed infested areas).

Land sparing (e.g. land taken out of agricultural production and set aside for conservation purposes only).

2.2 Remnant management ‘Remnant vegetation represents an important component of native biodiversity on private land and if managed correctly, provides a vital contribution towards enabling the integrating of biodiversity and production’ Native vegetation performs direct benefits to agriculture such as maintaining hydrological processes or the provision of shelter for livestock. Indirectly, native vegetation benefits agriculture by providing vital ecosystem functions such as nutrient cycling, landscape stability, fodder production for livestock, habitat for native species as well as carbon storage. It is also important to highlight that the private value of land with native vegetation is dependent on the context of farm management and also dependant on changes in management (Crosthwaite and MacLeod 2000). In general, practical guidelines for the on-going management of vegetation that forms remnants (natural or enhanced) is lacking and this deficiency is presenting particular problems in some situations [Case Study 3]. 2.2.1 Remnant enhancement

In Australia, the management of ‘remnant enhancement’ is largely confined to the consideration of woody species, essentially involving tree planting, perhaps with the occasional addition of shrubby species. It is less common for understorey species to be included in direct seeding activities. However, grazing management is often aimed at increasing the number and abundance of native perennial grasses and, therefore, may also be viewed as remnant


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enhancement (see section 2.3). In Western Europe, perimeter plantings of grassy strips in arable fields have been shown to provide enhancement of ecosystem services such as soil fertility and pest control (Smith et al. 2008). This study provides evidence that these grass strips may increase the permeability of the agricultural matrix for species (common and rare), but these effects are confined to isolated patches, We found no comparable Australian studies. Enhancement of remnants (through the creation of tree buffers/corridors or additional habitats) can potentially offer a dual role of providing habitat and maintaining ecological processes such as the control of water table height and recharge to combat salinity (Zhang et al. 2007). Eldridge and Freudenberger (2005) describe trees as ‘ecosystem wicks’ which moderate large run-off events by providing a greater proportion of macropores, allowing reduction of regional groundwater recharge. These authors provide evidence that soil surface litter and foliage cover associated with remnant vegetation also play a role in maintaining the integrity of these macropores. Remnant vegetation also provides important associations with ectomycorrhizal fungi which can aid in the uptake of nutrients and the recovery of native vegetation following disturbances (Tommerup and Bougher 1999). In each case, these benefits are realised at a local scale, in close proximity to the remnant areas. The widely promoted use of retaining perennial grasses on hill tops for the control of dry-land salinity is also seen as a management option that has beneficial off-site impacts. Paddock trees directly benefit agricultural production by providing shelter for pastures, crops and livestock from wind, heat and cold. For example, Walpole (1999) in a simulation study, found that maximum gross margins for pasture production were achieved where tree cover across farms was around 34 and that any increase in woody cover beyond this threshold resulted in no additional productive benefit. A similar threshold of around 30% tree cover for fragmented landscapes has been reported by McIntyre et al. (2002) and Andren (1994). It is suggested that trees provide a stimulatory impact on soil nutrients (augmenting the addition of organic matter and reducing losses of nutrients) and microclimates (reduction in wind velocity), resulting in increased pasture growth. Jackson and Ash (2001) have also reported enhanced production of the tropical woodland native grasses Heteropogon and Chrysophogon under trees due to favourable soil nutrient conditions, and suggested that these effects are enhanced in low fertility situations. The use of trees and shrubs as wind breaks and associated benefits for increased animal production has long been established (Reid and Landsberg 1999; Reid and Thompson 1999). Nationally, lambing mortalities of around 20% in unsheltered paddocks represent a significant impact on farm profitability. The provision of shelter can reduce wind velocity across lambing paddocks, reducing heat loss from lambs and significantly improving their survival. The national project EverGraze is currently investigating the use of ’biological sheds‘ or living shelters constructed from pastures or crops with high standing dry matter, tussocks, pasture hedgerows, shrubs and trees (MLAPrograzier 2006). Some shrub species have been reported to eliminate


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lambing mortalities due to exposure. While this project is examining the potential role of tall exotic species, native species such as Paspalidium spp. and Themeda spp. may equally well serve the same purpose. Because trees influence the availability of resources (soil nutrients, light and moisture) they can also enhance spatial and temporal landscape heterogeneity resulting in increased plant diversity within tree-grass mosaics (Prober et al. 2002; Clarke 2003). However, there is some evidence from overseas and elsewhere within Australia (see Invasive native shrub) that high density trees can reduce diversity (Scoles and Archer 1997), and that small tree patches may provide a focus for grazing (high grazing pressure) reducing spatial heterogeneity and diversity associated with these patches (Belsky et al 1993). Nevertheless, recent research in the New England area of northern NSW has shown that even small isolated paddock trees provide benefits to native invertebrates, especially in close proximity to trees where increased litter and changes in soil fertility are greatest (Oliver et al. 2006). The value of scattered trees for nutrient conservation has also been reported for the south-east slopes of NSW (Eldridge and Wong 2005). Apart from these advantages, a single paddock tree may also provide valuable habitat in hollows and fallen branches which takes significant time to form. Manning et al. (2006a) point out that in highly modified agricultural landscapes, scattered trees should be viewed as ‘keystone structures’, performing important ecological functions at both local and landscape scales. Apart from the local microclimate effects described above, scattered trees also serve as ‘stepping stones’ to increase connectivity for some faunal species, may allow genetic connectivity for some tree species, and provide central points for the recovery of ecosystems (Manning et al. 2006a). As single or small patches of trees (<1 ha) are a common feature in agricultural landscapes, in some areas representing the only significant portion of the tree cover, it is important that they be retained (Freudenberger and Ozolins 2000; Gibbons and Boak 2002). Some predictions suggest that these small patches could be lost within 40 years (Gibbons and Boak 2002) and, given the benefits outlined here, single trees or small patches therefore warrant special management and should not be excluded from remnant enhancement management activities. While most revegetation programs have multiple goals (e.g. perennial grasses for salinity control and tree species for the provision of shelter), conservation of native biodiversity is also an important priority. Often lost in revegetation planning is a consideration of the habitat resource trees provide for fauna. In a recent review, Munro et al. (2007) outline the response of fauna to revegetation activities on agricultural land. While revegetation (tree corridors) has probably had a positive effect on bird species, these authors highlight the need for studies of its effects on other faunal groups, and of the impact of greater structural complexity arising either from planned revegetation or the aging of existing stands. Unfortunately, revegetation or re-seeding activities tend to be treated as one-off events, with little follow-up activity that may allow for the creation of greater diversity in age structure and species composition in the long-term.


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There is evidence that some habitat resources accumulate more rapidly than others. For example, dense shrub canopies can accumulate within 10 years to provide nectar sources, but tree bark resources may take up to 20 years (Vesk et al. 2008). This implies that when calculating offsets for habitat destruction it is wrong to assume an immediate equivalence of remnant and restored vegetation as there is a lag time in resource provision. This will be particularly important when the viability of faunal populations depends on the continuation of resource supply. Management choices may also influence the trajectory of resource provision. A major benefit of direct seeding is that higher stem densities can be produced. While high density plantings may be capable of persisting in the long term, and have the benefit of providing shelter for small birds, high density plantings they may also retard provision of desirable attributes such as spreading canopy, large boughs, tree hollows, fallen timber and tree and shrub recruitment (Vesk et al. 2008). This highlights the importance of retaining logs and dead trees as they can tide over the time lags expected within existing revegetation areas. Lower density plantings may offer a greater chance of designing self-regenerating landscapes, as a greater number of recruitment sites are likely to become available (Spooner and Briggs 2008). Given these contrasting attributes, a mix of planting densities may prove to be a practice that results in more desirable biodiversity outcomes generally. The majority of recent tree planting activities have focused on planting in strips or clumps with little allowance for enhancement of paddock trees or the incorporation of structural diversity. Management activities that can either retain or enhance scattered paddock trees are lacking. Manning et al. (2006) suggest agroforestry and green tree retention (retaining a certain number of trees post-harvest) can provide the multiple benefits of enhancing landscape connectivity, enrichment of established stands with a structural element, as well as ’life-boating‘ of species and ecological processes until tree-cover is re-established. There are no reports of micro-restoration techniques such as erection of temporary fences around mature trees or groups of trees, direct seeding, soil scarifying or burning to facilitate tree regeneration within grazing landscapes. These techniques may also be useful in complementing existing revegetation programs to consolidate patches of woody cover but, as yet, appear unexplored. While natural regeneration provides a cost-effective alternative to direct seeding, studies undertaken to examine the influence of grazing patterns on regeneration capacity suggest tree recruitment will involve some destocking. Replanting woody vegetation through seed and/or tube stock, although widely used, can be expensive and results can be variable. Dorrough et al. (2008b) used farm economic modelling to examine the costs of revegetation activities on farms. This study showed direct seeding and planting (active regeneration) was more cost efficient in highly productive landscapes but, in areas of low productivity, natural regeneration was less costly and would require a more long-term approach to land retirement. These authors suggest that, to achieve catchment targets, a mix of both long- and short-term investment is required.


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2.2.2 Genetic integrity and local seed sources In part, the variable results in direct seeding and tree planting may be due to the use of seed poorly adapted to local site conditions. There has been considerable debate over issues of genetic integrity and the use of local provenance seed sources. The use of local seed sources in revegetation is presumed to avoid problems of genetic pollution of resident populations. Potts et al. (2003) highlight the importance of using local seed in farm tree planting and forestry in order to maintain the genetic resources of valuable Australian forestry species. However, there is good evidence to suggest that mixing of seed sources, particularly in fragmented landscapes may result in long-term viability of re-seeded populations (Waters et al. 2007). The arguments in favour of using local provenance seed sources revolve around a presumed ‘home-site’ advantage of local populations and, globally, there are numerous examples of adaptive advantages of local populations (see Lenssen et al. 2004 and Capelle and Neema 2005 for recent examples), leaving little doubt that local adaptation can occur. However, the difficulty is in understanding how to define ‘local’. Generally, adaptation to a site decreases with increasing geographic distance. However, in Australia there are numerous examples of local populations which display superior growth and survival characteristics over broad geographic ranges for both forestry species (Potts et al. 2003) and native grasses (Garden et al. 2005; Waters et al. 2005). There are also a number of examples of the genetic impoverishment of local seed sources due to inbreeding (e.g. the native pea Swainsona recta - Buza et al. 2000), or even the complete loss of genes which are important for reproductive effort (e.g. the native daisy Rutidosis leptorrhychoides - Young and Murray 2000). For both these rare and endangered species, the use of non-local seed sources, or mixing of seed sources to increase genetic diversity, may prove to be the best option. However, wide geographic range may not necessarily preclude the need to mix seed sources. For example, populations of the widespread native grass Austrodanthonia caespitosa (Wallaby Grass) showed reduced reproductive effort when collected from small isolated fragments within intensive cropping areas of central western NSW (Waters 2007). Sourcing seed from multiple populations may therefore be a better option in agricultural landscapes where vegetation is restricted to small patches or isolated trees. Compositing this genetic diversity may also have the added advantage of ensuring remnant populations have sufficient genetic diversity and adaptive capacity to survive predicted climate change scenarios. Carr (2005) suggests that the implicit genetic integrity of local provenance seed sources is irrelevant to revegetation where the only goal is site amelioration. Waters et al. (2007) extend this view, suggesting that clear recognition of the restoration/revegetation objectives be considered within a risk assessment framework. These authors outline a decision-making tool which identifies an acceptable seed collection range by matching the quality of the seed collection site with a consideration of the natural capital benefits of


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the revegetation activity. For example, in high value conservation areas, local seed sources may provide the most appropriate seed sources when local populations are of high genetic quality. Alternatively, when local populations are of poor genetic quality, seed may need to be sourced from more distant locations. It appears that the conservation of genetic diversity in natural ecosystems is generally not considered an issue in large-scale revegetation, and that population demography is more important in determining population viability than population genetics. The evidence that genetic diversity is important for ecosystem resilience (resistance to disturbance and climate change) is generally derived from studying species-poor communities. However, in agricultural areas, understanding whether species richness or intra-specific variation among populations (usually associated with increased richness) is more important in determining the relationship between biodiversity and ecosystem functioning is largely unknown.

Remnant vegetation


Impact on production

Negative

Positive

Positive

Planting narrow tree corridors with little structural composition usually in the same locations (paddock perimeter); one off revegetation activity with little accumulation of food and habitat resources.

Remnants with structural, compositional and spatial integrity (providing quality habitat for birds and insects; potential for IPM; animal, crop and pasture protection and retention of hydrological function).

Negative

Inappropriate management of individual paddock trees (e.g. protection without habitat modification.

Planting tubestock in areas of low productive capability.


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2.3 Grazing management ‘Grazing can be managed in a way that does not make it a threatening process’ 2.3.1 Domestic livestock and native vegetation

While domestic livestock are seen as a major threatening process for conservation of biodiversity, there is considerable evidence that grazing can also play an important role in enhancing biodiversity in agricultural landscapes. The complete absence of grazing in some New Zealand grasslands has led to a replacement of native plant species by exotic species (Meurk et al. 1989), and has been shown to reduce species diversity in native pastures in central western New South Wales, at least in the short-term (Alemseged et al. unpublished). Grazing management has been used as a conservation tool in the Terrick Terrick National Park (N. Bruce pers. comm.) and in the Tom Gibson Nature Reserve (Fensham 1991) and has been recognised as providing positive conservation and production outcomes (Kirkpatrick et al. 2005). The critical point here is that grazing is managed in a way that does not allow it to become a threatening process (Whalley 2005). The influence of domestic livestock gazing on native plant species diversity (largely species richness) has been the focus of numerous recent Australian research programs (for recent examples see Allcok and Hik 2003; Clarke 2003; Dorrough et al. 2006a; Spooner and Briggs 2008). However, consensus on these effects is not easily reached. In part, this may be due to the interaction of grazing with particular pasture development practices including tree clearing, cultivation, introduction of exotic species, fertiliser application, and increased stocking rate, as well its interaction with climate (e.g. drought). There is evidence that grazing has a direct benefit for species richness in the temperate areas of Australia (Gibson and Kirkpatrick 1989; Tremont and McIntyre 1994; Morgan and Lunt 1999) and elsewhere in the world (Fynn et al. 2004; Frank 2005). A 12 year study in an Australian riparian ecosystem showed exclusion of grazing to have no effect on species richness and only a minor impact on plant composition (Lunt et al. 2007b), but that the quality and productivity of remnant areas had some influence on the response of vegetation to grazing (Lunt et al. 2007a). Where grass densities are too high (i.e. grasslands are highly productive) there are limited opportunities for inter-tussock species to establish. This has led to a general notion that highly productive grasslands competitively exclude species and that grazing will allow more species to co-exist by ‘opening up’ the grassland. Data collected as part of the national Sustainable Grazing Systems (SGS) project found that pasture productivity at a range of locations in temperate Australia reached a maximum at around 10-20 species, after which production declined, irrespective of grazing management (Kemp et al. 2003). As this threshold is reached, the number of forbs increased and these authors suggest that forbs achieved a competitive advantage due to decline in fertility which disadvantaged more productive, desirable grazing species.


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Support for this suggestion is found in many other publications that demonstrate greater pasture production from exotic species with the addition of fertiliser (Lodge et al. 2003). In drier areas, differences in species richness between grazed and un-grazed sites may be less obvious than in higher rainfall environments because of greater temporal variation in species mortality and recruitment rates. For example, native grasslands of Dichanthium sericeum (Queensland Blue Grass) and Astrebla spp (Mitchell Grass) were shown to be largely unaffected by the influences of past grazing (Lewis et al. 2008). In this long-term study (8 years), the floristic composition of the Kirramingly Nature Reserve located on the Moree Plains was compared to adjacent stock reserves. While areas that had been previously cultivated (13-22 years prior) had significantly lower native species richness, no significant effects of grazing were found. In particular, no decrease in forb species richness was indicated in the absence of grazing..This suggests that either the 8 year period of grazing exclusion was not long enough to reveal any of the competitive effects between perennial grasses and forbs, or the 10-20% bare ground reported in this study provided insufficient recruitment sites. Lewis et al. (2008) suggest that additional grazing pressure from kangaroos may provide enough disturbance to prevent grass densities from becoming too high to limit the growth of forb species. Alternatively, these grasslands may simply be more tolerant of grazing than might be expected (Orr and Holmes 1984). A similar innate stability of native grassland has been reported in higher rainfall areas of south-eastern Australia dominated by Microlaena stipoides (Weeping grass) and Austrodanthonia spp (Wallaby grass) (Garden et al. 2000). Another rare study of the long term effects of grazing exclusion is that of Spooner and Briggs (2008) who compared 7-9 year fenced and unfenced areas of remnant woodland in the Murray Catchment. Removal of grazing resulted in a greater occurrence of tree regeneration (eucalypts), consistent with the findings of Dorrough et al. (2006), greater dominance of native perennial grasses and fewer exotic annual grasses and weeds. Soil compaction was also reduced compared with unfenced sites. However, the pattern of grazing appeared to have a greater influence on compositional, structural and functional characteristics than its presence or absence. 2.3.2 Grazing management options There is a continuum of grazing management options from set stocking, rotational grazing to cell-type grazing and tactical grazing. Differences among these grazing regimes are largely based on their attempts to satisfy the requirements of animals or pastures species (Table 2.1). In a survey undertaken in the temperate grassy woodlands of the New England tablelands, most of the variation in species composition was attributed to different grazing patterns (Clarke 2003). Continuously grazed sites had a greater abundance of native grasses than other plant groups such as forbs, but fewer species were present. Generally, Clarke (2003) also found increasing grazing pressure led to a shift from a dominance of native warm-


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season perennial grasses toward short-lived exotic cool season grasses, consistent with the observations of McIntyre et al. (2003) and McIntyre and Tongway (2005). McIntyre et al. (2003) describe shifts in grassland sward structure from tall palatable species to shorter-growing tussock swards under increasing grazing pressure. Additional research has indicated that these structural and compositional shifts are associated with declines in soil surface condition, reflecting reduced infiltration and nutrient cycling rates (McIntyre and Tongway 2005). McIntyre et al. (2003) found tall patches of grasses provided nesting habitat/seed sources for birds as well as shelter for ground dwelling mammals. However, heavily grazed short patches were also important for providing high quality forage for native fauna, in much the same way as domestic livestock derive nutritional benefits from annual pasture species. When both the connectivity of tall and intermediate grass patches and soil surface condition are maintained, habitat benefits for organisms of low mobility have been demonstrated (McIntyre 2005). In this way, a range of sward structures may provide better for the needs of all species. This conclusion is in keeping with those of Perkins et al. (2000) who suggest a heterogeneous sward structure is important for the provision of habitat for British bird species found throughout agricultural areas as well as for invertebrate diversity (see McIntyre 2005 for additional detail). While the importance of understorey heterogeneity to biodiversity is in line with the general notion that management practices should target increased heterogeneity, the patchiness may have different implications in different environments. For example, when the grass patches in tropical tall grasslands described by McIntyre et al. (2003) are continuously grazed, they can become degraded and eroded. Fire is then required to even out the grass sward and prevent the continuous patch grazing. While patches or gaps between trees and perennial tussocks (if free from competition from exotic annual grasses) may provide suitable conditions for tree seedling and native species regeneration (Spooner and Briggs 2008), the optimal patch size (or percentage of bare ground) will be different for different climatic areas. In tropical environments, McIvor et al. (2005) suggest than grazing should be managed to allow 60-70% dominance of large to medium tussock grasses in native pastures. In semi-arid landscapes, resources are concentrated into patches that harvest limited nutrients and water and are associated with greater productivity (Ludwig et al. 1997). Overgrazing of these patches will lead to landscape dysfunction and unsustainable pastoral enterprises. Overgrazing may also lead to undesirable shifts in competitive interactions between species. In the semi-arid woodlands, woody weed encroachment into rangelands can occur when grazing is intense (Scholes and Archer 1997). In these drier environments the herbaceous layer tends to be discontinuous, and higher grazing pressures will provide additional opportunity for shrub seedling recruitment by reducing grass density (Harrington 1991). In this way the nature of shifts in species composition and structure are dependent on environment. For example native woody weeds may proliferate in overgrazed semi-arid environments whereas exotic annuals tend to dominate in


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overgrazed winter rainfall areas as a result of the inability of cool season native perennials species to survive summer drought. Although generally small in size (20-200 ha), travelling stock routes also provide some insight into the effects of high density/short duration (HDSD) grazing on native species diversity. It is argued that these areas provide important refugia for remnant vegetation, retaining examples of high quality grasslands and woodlands, and are therefore important in maintaining or improving grassland of high conservation value (McIntyre and Lavorel 1994; Davidson et al. 2005). Table 2.1 Contrast between different grazing management options (Adapted from Hacker

2008b unpublished and McCosker 2000)

Grazing management

options

Examples

Paddocks per herd

Other features

Stock

density relative to

continuous grazing

Continuous

Continuous grazing Set stocking

1

Stocking rate varies but no movement of animals in relation to either plant or animal requirements

_

Rotational Resting Deferred rotation grazing Merrill system

Generally 2 or less

Calendar based movements Moderate

Low intensity rotational grazing

High Intensity Low Frequency grazing (HILF)

3-7 Calendar based movements Moderate - High

1. High intensity rotational grazing (High utilisation grazing)

Non-Selective Grazing Crash grazing Short duration grazing

>7 Each paddock severely grazed before moving; calendar based movements

High

2. High intensity rotational grazing (High performance grazing)

Controlled selective grazing Short duration grazing

>7 Each paddock lightly grazed before moving; calendar based movements

High

1. Time control grazing methods (Production focus)

Block grazing Strip grazing Rational grazing High density, short duration grazing

20-40 Moves based on pasture growth rate and physiological requirement for rest; requires high stock density Focus on maximising plant and animal production

Very High

2.Time control grazing methods - (Holistic focus)

Savory grazing method Cell grazing Controlled grazing Management intensive grazing Ultra high density grazing Planned grazing

20-40 Moves as above; focus on ecosystem sustainability and optimising profits

Very High

Tactical grazing n/a Variable Movements determined by defined objectives and management strategies for individual paddocks.

Variable


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Over the last decade, there has been an increase in the number of livestock producers undertaking a form of cell grazing in temperate areas of NSW (McCosker 2000). There are also numerous examples of landholders in low rainfall areas adopting this type of grazing regime [e.g. Case studies 1 and 4]. Generally, it appears that a shift from set stocking or continuous grazing to some form of rotational grazing results in biodiversity or other natural resource benefits although the evidence if far from conclusive, and landholders often report an increase in profitability. It remains unclear, however, whether profitability is being driven by the change in grazing regime associated with an increase in carrying capacity (more efficient pasture utilisation in smaller paddocks), changes in the enterprise mix, or improved animal or financial management. Shifts to cell grazing, for example, are usually accompanied by training in other aspects of business and livestock management and often appear to be accompanied by the introduction of a livestock trading enterprise which may increase flexibility in allowing rest of pastures compared with traditional breeding enterprises under set stocking regimes. This may be particularly important in low rainfall areas, where there is greater temporal variation in pasture production and rates of mortality/recruitment. Flexible grazing management techniques, while directly applicable in low rainfall areas, are likely to have a broader application under predicted climate change scenarios. Tactical grazing management is a principle that allows exploitation of production opportunities under good seasonal conditions and avoids risks associated with drought through resting pastures. In this way, management can be changed to match seasonal conditions (Campbell and Hacker 2000). The anecdotal evidence provided by Earl and Kahn (2006) also clearly highlights the importance of tactical rest periods for perennial grass survival and ecosystem function, and other studies refer to deferred grazing to enhance native pasture composition and increase profitability (Moll et al. 2006). The Cicerone Project, recently conducted on three 54 ha farms in the New England area, provides some evidence that in high-input situations, high density, short duration (HDSD) stocking may not result in higher economic returns (Scott et al. 2006; Scott 2007). In this project, high input/high stocking and moderate input/moderate stocking systems, both using Prograze™ principles, were compared with moderate input/HDSD rotational grazing. Under HDSD higher levels of desirable pasture species were achieved but with lower legume composition. Added to this, the increased costs of labour and capital associated with fencing resulted in lower gross margins in the HDSD farm. However, HDSD stocking did result in reduced nematode faecal egg counts and was associated with superior worm control (Healey et al. 2004). In a four year study on the NW Slopes of New South Wales, Lodge et al. (2003) provided empirical evidence that shifting from continuous to rotational grazing increased total herbage mass, ground cover and litter of native pastures dominated by Bothriochloa macra (Redgrass). Rotational grazing


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also had the advantage of higher levels of soil microbial activity, soil OC, higher number of earthworms and reduced runoff rates, all resulting in a higher economic return than under continuous grazing treatments. Importantly, this study also demonstrated that these effects could be enhanced by periods of rest following grazing. In this study, rest periods (either 4 or 12 weeks) were not timed to coincide with key phenological times of seeding and seedling recruitment, which is a strategy that may also be important for lower rainfall (<500 mm) areas. A three year study in the mid-north of South Australia compared the response of native pastures to continuous grazing, tactical grazing aimed at avoiding key phenological periods occurring in autumn and/or spring, HDSD and de-stocking of domestic livestock (Earl and Khan 2006). The highest species diversity and greatest number of perennial grasses were associated with paddocks continuously or tactically grazed, but continuous grazing also resulted in a 10-fold increase the amount of bare ground. Increased pasture biomass resulted in the carrying capacity of HDSD paddocks being almost double that of continuously grazed paddocks although the number of perennial grass plants showed a small decline. The EverGraze project (Central Slopes NSW), is evaluating a number of grazing treatments on native pastures (low intensity - set stocking at 5 DSE/ha; medium intensity rotational grazing (paddocks grazed 25% of the time at 20 DSE/ha) and high intensity rotational grazing (paddocks grazed 5% of the time at 100 DSE/ha). While located on the central tablelands (>550 mm rainfall), this project will provide additional information on the effects of grazing regime on livestock and pasture composition (W. Badgery pers. comm.). Popularly promoted benefits of cell grazing have been the asserted increases in soil organic matter (OM) (which provides a higher fertility status for pasture species), increased soil water holding capacity and provision of protection from erosion. However, more recently, its potential for sequestration atmospheric carbon has been promoted. In their South Australian study, Earl and Khan (2006) reported small improvements in soil surface condition (subjective assessment) under HDSD stocking, but the first known empirical data supporting this notion was published only recently (Sanjari et al. 2008). In this study, soil physical and chemical properties under continuous and cell grazing management were compared on a property in south-eastern Queensland in the process of converting from long-term continuous grazing to cell grazing. Under cell grazing total stocking rate (DSE days/ha), litter cover, and pasture standing crop (kg/ha) all increased. No changes in bulk density occurred under cell grazing, but continuous grazing resulted in a greater degree of soil compaction. Although not statistically significant, organic carbon (OC), nitrogen (N) and C:N ratio all increased slightly under cell grazing, while under continuous grazing these variables remained unchanged. Patchy accumulation of NO3 and phosphorus (P) occurred under continuous grazing, with the potential for these nutrients to move down slope and contaminate water bodies, whereas cell grazing resulted in not only a more even distribution of nutrients, but also greater immobilisation of N and C by plants and micro-organisms. Extractable P fell under cell grazing, possibly the result


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of increased utilisation by pasture species associated with increased biomass production. However, differences in C:N, OC and N were small, and there appeared to be more variation within treatments than between the grazing regimes. Generally, the balance of the empirical literature suggests that within the agricultural context, if managed correctly, grazing by domestic livestock is a valuable management practice that can lead to biodiversity benefits. In particular, benefits appear to occur when changing from set stocking or continuous grazing to a rotational or non-continuous grazing regime. Interpreting species response to grazing has been further complicated by both the varying scales at which studies have been conducted and the interaction of grazing with tree clearing, fertiliser application and, undoubtedly, the initial condition of the pasture. Nevertheless, some general conclusions can be drawn from the studies reviewed above:

• Continuous grazing results in lower native species richness - structural and compositional changes occur under increasing grazing pressure that are linked to declines in biodiversity

• A move from set stocking/continuous grazing to a non-continuous grazing regime is beneficial to both biodiversity and production

• Strategically managed, high intensity, short duration stocking rates appear to have both production and biodiversity benefits, particularly in low-input situations

2.3.3 Grazing impacts in the semi-arid rangelands Drawing principles for grazing impacts on the semi-arid rangelands is not a trivial task, as this complex situation involves not only consideration of sustainable and profitable livestock production from diverse native plant communities but also the distribution of these communities across variable landscapes and the highly variable climate of these areas (Hacker 2008b). Comparisons of continuous and rotational grazing systems, at least at experimental scales, rarely demonstrate any advantage in favour of rotational grazing in terms of individual animal performance at comparable stocking rates, or pasture production (Briske et al. 2008). However, in large paddocks, particularly in the semi-arid zone of the WCMA, the inclusion of a large degree of landscape variability results in non-uniform grazing or patch grazing (over-utilisation of some species or areas and under-utilisation of others). Subdivision of paddocks or, to a lesser extent, establishment of additional water points, can result in more uniform utilisation of forage across the landscape (Hart et al. 1993, Hunt et al. 2007). Such improvements in the efficiency of forage utilisation may result in increases in carrying capacity at the property level which arise purely from infrastructure design and are independent of the grazing management system. Norton (1998) listed nine studies from the global literature in which grazing trials on research stations, lasting from 5-35 years, had carried stocking rates under continuous grazing from 40-90% higher than recommended levels for the


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district. This situation is not unlike that found in higher rainfall areas but clearly involves an ability to recognise similar land management units (land capability) and fencing to achieve uniform utilisation. Two relatively newly adopted grazing systems in western NSW are Time Controlled Grazing (TCG) and Tactical Grazing (TG). A distinctive feature of TCG is an emphasis on the use of large mob (herd) sizes and utilisation of a high stocking density for a short duration. The requirement for high stock density and large mob size arises from the particular importance attached to the role of ‘animal impact’ and ‘herd effect’ (Savory 1988) in facilitating nutrient cycling and ecological succession in ‘brittle’ environments (laying and incorporation of litter, breaking soil surface crusts, creation of niches suitable for seed lodgement and germination, etc). However, there is little evidence to support the importance of these processes for ecosystem function in semi-arid rangelands, particularly in Australia where rangelands have not co-evolved with large herds of hoofed grazers. Beukes and Cowling (2003) have recently demonstrated increased biological activity in the interplant spaces of areas receiving a high degree of animal impact under Non-Selective Grazing4 in the Nama Karoo (South Africa). They describe this as an example of positive impact on soil quality resulting from the herding of large herbivores at high densities. However, they also note that, ”the literature is replete with contradictory results on the effects of grazing on soils”, and that soil type may have influenced the outcome of their study. Research in semi-arid rangelands of the US generally indicates adverse rather than beneficial effects of heavy trampling under short duration grazing on infiltration rate, sediment yield and soil compaction (Wood and Blackburn 1984, Thurlow et al. 1986, Phular et al. 1987; Weltz et al. 1989). Effects of trampling on seedling establishment have ranged from severely adverse (Salihi and Norton 1987), to negligible (Bryant et al., 1989), to variable depending on rainfall, species and soil type (Winkel and Roundy 1991). Eckert et al. (1986) demonstrated that heavy trampling of big sagebrush communities in Nevada could favour undesirable botanical change, although their study did not relate specifically to TCG. In western NSW, Greene et al. (1994) measured a significant increase in soil loss following surface disturbance by sheep at Lake Mere, near Louth. Of the Atriplex (saltbush) seedling cohorts studied by Eldridge et al. (1991) near Broken Hill, only 5 per cent occurred in the ’depression‘ microsite, commonly formed by hooves of sheep or cattle. While anecdotal evidence has shown benefits resulting from animal disturbance in specific situations (e.g. scald regeneration), the general importance of animal impact/herd effect for landscape function in western NSW remains unknown. Current models of the functioning of semi-arid landscapes emphasise the importance of local concentration of water and nutrients into fertile patches through run-on/run-off mosaics (Ludwig et al. 1997). It appears that it is this patch concentration of resources which sustains production of perennial plants in semi-arid and arid landscapes. As rainfall is usually much less than 4 Essentially high intensity/short stocking of large mob size in small paddocks (Hoffman 2003; Beukes and Cowling 2003).


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evaporative demand, the uniform distribution of rainfall may severely disadvantage species required to survive long dry periods. Disruption of sealed surfaces, or cryptogamic crusts in chenopod shrub communities may improve in-situ water infiltration but any disruptions to local redistribution are likely to be counterproductive from the perspective of the perennial vegetation components. While this apparent conflict is a significant theoretical issue for rangeland ecology, the extent to which surface disruption could be achieved in practice in semi-arid rangelands may reduce its practical importance. McCosker (2000) maintains that adequate animal impact is unlikely at densities of less than 5 cattle/ha or 10 sheep/ha. At carrying capacities typical of the Western Division (say 0.2-0.4 DSE/ha) this implies a requirement for 25-50 paddocks per mob for sheep and 100-200 for cattle, assuming one beast is equivalent to 8 DSE (Hacker 1993). While the importance of animal impact and herd effect are questionable in semi-arid areas, the use of larger mob sizes may confer some advantages independently of these processes. Norton and Miller (2000) have argued that animals in large mobs tend to disperse more extensively over the landscape. Amalgamation of mobs under rotational grazing may thus achieve more efficient forage utilisation than the same number of animals grazing continuously in smaller mobs over the same paddocks. However, objective data are lacking, and the effect of mob size may well vary between sheep and cattle. Another distinctive feature of TCG is the use of periods of rest (and therefore also of grazing) determined only by plant growth rates. Since the aim is to ensure that regrowth is not prematurely grazed, rest periods are shortened in times of rapid growth and lengthened when growth is slow. For any given number of paddocks in the rotation, the grazing period must follow an identical pattern. However, in the semi arid zone of western NSW rainfall is largely aseasonal and growth occurs in pulses at any time of the year. The application of this principle in these environments may be more difficult than in environments with a distinct growing season. Furthermore, the requirement to extend grazing in the current paddock until recovery is ‘complete’ in the next paddock to be grazed ignores the need to keep utilisation at levels that will minimise the risk of plant mortality in the critical summer period (Hacker et al. 2006). The adjustment of stocking rate to match carrying capacity may reconcile these potentially conflicting requirements. However, rapid adjustment in stocking rate is not always possible, and there would seem to be a case for adjusting grazing times to account for the level of utilisation achieved as well as the degree of recovery. On balance, the use of TCG in semi-arid areas appears to ignore the complexities of environmental/landscape variability by providing an over-simplistic ‘recipe’ for grazing management. Time control grazing may have the potential to lift carrying capacity, through increased efficiency of forage utilisation resulting from paddock subdivision and the amalgamation of mobs, but a consequence of this will be the need for very responsive management in terms of variation in stocking rate, as the drought buffering capacity achieved by non-uniformity of grazing will be removed. Strict attention to the timing of stock


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movements will also be required if the high stock densities are not to impose a penalty on animal production through reduced diet quality, or lead to heavy grazing pressure on desirable species. The extent to which TCG will achieve range regeneration is difficult to assess. The long rest period that it provides should be beneficial compared with continuous grazing, but strong evidence is currently lacking. Hacker et al. (2005) found no evidence of any pasture response to TCG compared to continuous grazing at comparable stocking rates on two properties over 4 years in the Western Division, but the study was limited by poor seasonal conditions. Dowling et al. (2005) found no apparent benefits of TCG over continuous grazing for encouraging or maintaining a favourable botanical composition over five study sites and six years in the high rainfall zone of south east Australia. Failure to direct grazing and rest periods to the specific needs of the component species may be a contributing factor to this lack of response. Tactical Grazing is more a framework for developing a grazing management regime for individual properties than a grazing system or method. It does not require prescribed livestock movements or the strict application of universal principles. Key principles of TG involve the establishment of an (ecological) management objective for each management unit (generally a paddock), determination of a strategy to achieve that objective that is implemented on a day-to-day basis, and monitoring to guide tactical decisions. An example of a strategy for maintenance of semi-arid grasslands might be to:

• Maintain ground cover above 40 per cent • Maintain average utilisation of key species at approximately 30% • Spell for seeding of key species if no seed set for 3 years • Burn April or October if fuel is sufficient and undesirable shrub seedlings

are present or shrub cover is increasing (Campbell and Hacker 2000).

Because strategies for each paddock will be developed in relation to the specific needs of that paddock, maintenance or regeneration of the pasture base should be more assured than if these needs are not specifically targeted. Under TCG, for example, succession is assumed to occur because rest is provided and effective operation of the nutrient and water cycles is encouraged. While these no doubt are preconditions for pasture regeneration, the outcome is unpredictable in the absence of targeted strategies, particularly in an environment with unpredictable rainfall. Regeneration of many perennial species will only occur when there is a, “fortuitous co-occurrence of events when each event has a low probability of occurrence” (Noble 1986). Management should be flexible enough to capitalise on these events. A potential disadvantage of the TG approach is that distribution of animals around paddocks in accordance with individual paddock strategies could be a complex task. A number of paddocks may require rest at the same time, for example. Paddocks should thus be assigned a priority, based on their productivity or potential to respond to management, to help resolve these conflicts when they arise. Where strategies are defined in terms of targets for ground cover and utilisation rather than for grazing of specific species at specific


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times (as in the Aristida spp. control strategy of Lodge and Whalley, 1985), implementation may well be easier if stock are amalgamated into larger herds or flocks. In this case TG would have similarities to TCG, but without the emphasis on high stock densities and subdivision. 2.3.4 Total grazing pressure management The Western TGP (total grazing pressure) project consists of long-term, large scale demonstrations in western NSW where different grazing management options are being evaluated. The project is to be completed within the next six months but, by the end of the first phase, after five years, it was not possible to detect consistent or appreciable differences in native pasture composition or species frequency due to exclusion of grazing (goats, kangaroos). In part, this was due to a protracted drought during the course of the project, and to the failure of fencing to achieve effective grazing control. However while differences in the frequency of native grasses between grazed and ungrazed (kangaroos excluded) areas were not significant, visual differences did occur (T Atkinson pers. comm.), reflected by increased total ground cover under kangaroo exclusion (Hacker et al. 2005).

Grazing


Impact on production

Negative

Positive

Positive

Set or continuous stocking with no pasture recovery period, low-moderate stocking rate (beneficial for production per head but not for production per ha if stocking rate is too low). Paddock subdivision and/or additional water (if this impacts water- remote refuges).

High intensity grazing with long periods of rest and short grazing periods; moderate stocking rate. Pasture rest periods are tactical in response to ongoing climate conditions. Paddock subdivision and/or additional water (if this retains or expands water- remote refuges).

Negative

Set or continuous stocking with no pasture rest/recovery period, high stocking rate. All grazing removed in highly productive pasture.

High intensity grazing with long periods of rest and long grazing periods. Long periods of pasture rest where native pastures are degraded.


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2.3.5 Fertility and native pastures There is strong evidence that increased fertility (nitrogen and phosphorus) is associated with reduced native floristic biodiversity. In addition, there is a growing body of literature that suggests the effects of grazing on biodiversity (total species richness) vary according to fertility (phosphorus and nitrogen). It is argued that the innately low levels of phosphorus in Australian soils have been an important element in the evolutionary development and distribution of Australian flora and, therefore, phosphorus application, particularly throughout temperate Australia, has had a negative impact on native vegetation (Kirkpatrick et al. 2005). Generally, research on the response of specific species to phosphorus has shown variable results (Garden et al. 2000). Soil phosphorus concentrations are partly a function of paddock fertiliser history and fertiliser requirements are generally greatest under set-stocking regimes. In larger paddocks, nutrients can accumulate in dung or in sheep camps, rather than being re-distributed evenly across paddocks, and cell grazing can actively aim for ‘re-distribution’ of sheep camps (high fertility patches) (T Wright pers. comm). Tim Wright suggests that by redistributing fertility, the cost of fertiliser is reduced. This observation raises some fundamental questions about cell grazing in these environments:

• is fertility redistributed across the paddock? • are the benefits in terms of productivity based on better utilisation of

feed within a minimum fertility threshold? • can pasture productivity be expected to decline in the long-term if

fertiliser is withheld? Further research is required to investigate these issues. Dorrough et al. (2006) suggests a threshold of around 30 mg/kg of plant available (Colwell) phosphorus, above which understorey species (exotic and native) richness declines with increasing phosphorus (Figure 2.1).

0 50 100 150 200

Phosphorus (mg/kg)

Spec

ies

richn

ess/

0.9

ha

1020

3040

Threshold of around 30 mg/kg phosphorus

Figure 2.1 Response of species diversity to available soil phosphorus (Colwell) (Adapted from Dorrough et al. 2006b).


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Oelmann et al. (2007), in Germany, demonstrated that plant diversity (species or functional groups) controls soil and plant nitrogen pools and that plant communities with higher species diversity are more likely to utilise limited nitrogen resources more efficiently. Hacker et al. (2007) reduced the availability of nitrogen by applying sugar to promote the immobilisation of nitrogen on degraded native pasture in central western NSW. This technique appeared to facilitate both the establishment and growth of native perennial grasses, Enteropogon acicularis (Curly Windmill Grass) and Austrodanthonia caespitosa (Wallaby Grass) by limiting the competitive ability of annual weedy species. In this experiment application of120 kg/ha of nitrogen and 30kg/ha phosphorus increased the growth of annuals and suppressed the establishment of native perennial grasses, a result in agreement with Smallbone et al. (2007).


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2.4 Farming practice Intensive agricultural land uses such as cultivation and cropping are viewed as ‘leaky’ because of their impacts beyond the farm gate and their effects on landscape function. Expansion of this land use beyond around 30% of the landscape has also been reported to be associated with a rapid decline in some critical ecosystem functions (McIntyre et al. 2002). Below, we discuss further some of the adverse off-site impacts associated with intensive cropping/pasture activities (see sections on Weed and Pest Management and Water-point Management for additional discussion). There is considerable evidence for the beneficial role of soil biodiversity in sustaining ecosystem functioning in cropping landscapes. The main options to manage soil biodiversity include tillage, cropping rotation (and sequence) and organic matter management (Figure 2.2). While the mechanisms are unclear, the literature provides evidence that soil biodiversity underpins sustainable farming systems (see Brussaard et al. 2007 for a recent review). Specifically, plant and soil biodiversity affect two major ecological processes - nutrient and water cycling (Figure 2.3). There are currently a number of studies examining the potential capacity of soils to sequester carbon (at depth) under cropping systems, and this is being promoted as a ‘win-win’ situation in which excess CO2 can be removed from the air and soil improved by enhancing organic matter which, in turn, provides energy and nutrients for soil biota (C. Jones pers. comm.). Certainly, given the extensive areas under cropping, sequestering even small quantities of carbon will have a positive environmental benefit. However, the evidence published to date suggests that differences in soil carbon between various land uses (cropping and remnant areas) are insignificant at a depth of 0.8 m and that effects are largely restricted to surface soils (Young et al. 2005). As organic matter is most useful in cropping situations when it decays, a major focus of future research will be to develop farming systems in which trade-offs between carbon storage and utilisation can be made. ‘Sustainable farming revolves around being able to facilitate the re-cycling of organic matter’


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Decreasing population density and diversity

Increasing population density and diversity

Pesticides and soil contamination

Erosion

Intensive tillage, use of heavy machinery (soil compaction)

Monoculture

Burning of crop stubble

Acidification, uncovered (bare) soil)

Liming

Irrigation in more arid areas

High plant diversity

Crop rotationNo tillage

Green manures and cover crops

Organic farming

Agricultural m

anagement

Biodiversity

Figure 2.2 The impact of different agricultural cropping practices on soil and animal

density and biodiversity (Adapted from Brussaard et al. 2007).


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Climate Management

Crop rotation OM quality Tillage method

Wat

er

Nut

rient

S

uppl

y

Nutrient Use Efficiency Water Use Efficiency

Carbon Sequestration

W

ater

/Nut

rient

Use

Sustainable Agroecosystems

Carbon and nutrient cycles

Water cycle

Soil structure and physical/chemical processes

Plant and soil biodiversity

Figure 2.3 The effects of plant and soil biodiversity on functioning of agricultural

ecosystems. (Adapted from Brussaard et al. 2007). 2.4.1 Intensive cropping Studies in south-western Queensland have shown clearing of native vegetation results in a decline in soil organic carbon (OC) of around 30% in the top few centimetres of soil (Harms et al. 2004). Greater losses in nitrogen (N) can be expected (Harms et al. 2004), and are associated with decreased grain protein and yields (Dalal et al. 2003; Harms et al. 2004). Additional research undertaken in the Brigalow region of southern Queensland found that organic matter (OM) also declines, but at an exponential rate as the period for cropping increases (Dalal et al. 2003). Continuous cropping (> 20years) in northern NSW has resulted in declines in OC to a depth of 60cm when compared to grassy woodlands (Young et al. 2005). As higher levels of OC and OM are associated with higher crop yields, it appears that either would be a useful indicator of cropping sustainability. However, as rainfall decreases, the percentage OC also declines (Figure 2.4).


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The study of Young et al. (2005) also suggests that the maximum total carbon (soil and biomass) is maintained in areas where some tree cover is retained, providing a case for maintaining paddock trees. While the value of remnant vegetation to livestock and crop production in temperate areas, and the role of native vegetation in harbouring beneficial insects/birds, have been outlined in section 2.2, we were unable to find any literature reports which demonstrated the value of scattered trees or margins of native vegetation to intensive, broad-acre crop production in low rainfall areas beyond that of carbon storage.

450 500 550 600 650

Mean annual rainfall (mm)

Org

anic

C (%

)

0.5

1.0

1.5

2.0

Figure 2.4 Relationship between annual rainfall and soil organic carbon for uncultivated

(solid line) and cultivated soil (Adapted from Dalal and Mayer 1986) The impact of mineral and organic fertilisers, pesticides and soil fumigants on soils organisms in Australian agriculture has been reviewed by Bunemann et al. (2006). These authors noted that mineral fertilisers (e.g. urea, ammonia nitrate, sulphates and phosphates) have limited direct effects, but indirectly enhance biological activity as a result of increases in crop yield, incorporated crop residue and soil OM. Nitrogen fertilisation will also indirectly affect soil acidification which has considerable effects on soil organisms. Bunemann et al. (2006) also report the impacts of commonly used herbicides, insecticides and fungicides on non-target soil organisms in various soil types. Not all chemicals adversely affect biodiversity. For example, zero tillage relies heavily on Glyphosate which can reduce soil bacteria but also increase fungi and microbes by 9 -19% (Busse et al. 2001). However, there is little evidence for long-term changes in the structure of soil biotic communities. 2.4.2 Tillage systems Soil OC is accumulated above ground by plant biomass and below ground as root mass, with losses occurring as a result of mineralisation (release of


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carbon as CO2), largely driven by environmental condition, but also management. Tillage systems will influence the amount of soil OC. Under annual cropping systems, OC declines as carbon is removed as grain and fallow periods expose OM directly to the atmosphere. Zero-tillage increases soil biomass above that of conventional tillage systems and has been recently been reviewed by Kladivko (2001). Tillage systems also affect the soil physical and chemical environment of soil organisms and their response to these changes will in turn influence soil structure, nutrient cycling and OM decomposition. In Australia, Chan et al. (2003) found that zero-tillage (3-19 years) on light textured soils (<35% clay) resulted in significantly higher soil carbon levels than conventional tillage, but only in wetter areas (>500 mm). They also found that these differences were restricted to the top 2.5–10 cm of soil. Other unpublished information suggests carbon sequestration to depths of up to 120 cm, particularly when total carbon (organic/inorganic and phytolith carbon) is measured, can be enhanced under perennial grass ‘leys’ (C. Jones pers. comm.). In addition, the role of cropping systems and the influence of grazing management are currently being studied (Jones 2007). The potential of zero-tillage or ‘pasture cropping’ (see 2.4.4 below) for sequestering carbon needs to be critically assessed. Baker et al. (2007) examined soil OC over the whole profile and found that below 30cm depth, soil OC was lower under zero tillage than with conventional tillage methods. These authors suggest that any detection of greater soil OC under zero tillage may be due to the sampling protocol (sampling only to top few cm). Currently, the literature therefore indicates that zero tillage has the ability to increase OC in the top of the soil profile (<30 cm) (Blanco-Canqui and Lal 2008), but provides no evidence for sequestration deeper within the profile. Chan et al. (2003) concluded that reductions in soil OC under conventional tillage were most likely a result of low crop yield (low rainfall), partial removal of stubble by grazing and high decomposition rates (resulting from higher temperatures). This suggests that where continuous cropping is practised in low rainfall areas, soil OC will continue to decline even under conservation tillage practices. The practice of conventional tillage and stubble burning also results in lower levels of OC, but these effects are only apparent in the long-term (greater than 19 years) (Chan et al. 2002). The widespread adoption of zero tillage methods has shifted reliance towards herbicide use. While currently 5-10% of farmers in the Namoi CMA may utilise weed seeking herbicide technology in a bid to reduce herbicide costs the application this technology is likely to be common practice within the next 2-3 years (G. Rummery pers.comm.). On balance, these changes may result in less tillage and less herbicide use across these areas. The international literature provides clear evidence that zero tillage is beneficial to soil and water conservation as well as providing a reduction in production costs (Puget and Lal 2005). However, despite the evidence that a


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reduction in tillage results in increases in OM, microbial activity, soil structural stability and infiltration rates, Australian farmers appear to be producing variable results in terms of crop yields. Explanations for this may include associated increases in pests and diseases, toxic chemicals associated with stubble retention, greater residual effects of herbicides, inhibited root growth and surface nutrient concentration. The central process in explaining these effects involves the rhizosphere and has been reviewed by Watt et al. (2006). In this review, the authors outline three case studies (largely from higher rainfall areas in southern Australia), that detail the effects of conservation tillage on crop productivity. 2.4.3 Cropping rotations The effects of fallow management and crop/pasture rotation on soil properties and crop production were reported in a recent study in south west Queensland (Thomas et al. 2008). This study found that short-term pasture leys did not improve infiltration, but long-term leys that provided greater surface cover did increase infiltration. They also reduced the risk of run-off and soil erosion and resulted in faster movement of water through the soil profile. These authors suggest that this may result in leaching of nutrients beyond the root zone and a reduction in nutrient availability to crops and pastures. The composition of the pasture ley may also be important, as there is some evidence that mixed legume/grass pastures are more effective in reducing weed seed banks than legume-only pastures (Central Queensland Sustainable Farming Systems, 2008).

2.4.4 Pasture Cropping ’Pasture cropping’ (PC) is a novel concept that is generating widespread interest throughout eastern Australia. It is a cropping technique developed and promoted by Colin Seis, an innovative farmer from central western NSW. Pasture cropping is not intercropping (similar approach to PC but pastures generally lucerne in a degraded/weedy form), which has been a management practice with significant history. Essentially, PC is promoted as a low cost technique for sowing an annual cereal or winter crop (oats, cereal rye, wheat or lupins) directly into a summer growing (C4) native pasture (in Col’s case, Bothriochloa macra; Red Grass). The system seeks to exploit the differential growth patterns of the crop and the pasture while minimising damage to the pasture itself. The system is well suited to environments such as the central west where rainfall is aseasonal and particularly to soils with limited water holding capacity. In this situation it may provide a flexible farming option [see Case Study 5) and anecdotal evidence indicates that additional environmental benefits may result including enhanced native perennial grass recruitment and production, soil structural improvements and increases in soil organic carbon (C. Seis pers. comm.). However, its suitability for summer dominant rainfall areas, where subsoil moisture is traditionally stored using fallows for subsequent cereal crops, has been doubted. As part of the Central West/Lachlan, Grain & Graze Regional Initiative, the effects of PC and zero-tillage on production, soil moisture and soil physical


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and chemical properties are currently being evaluated (Millar and Badgery 2008). The results reported here refer only to pasture cropping under a Red Grass pasture. In this experiment, at Wellington, PC resulted in an initial reduction in native grass biomass but total biomass (cereal and native grass) was enhanced and the biomass of native pasture soon returned to a level comparable with undisturbed native pasture. This suggests that PC is capable of providing additional feed for livestock, particularly over the cooler months and that it may result in no long term damage to the native pasture. While some reduction in the number of adult native grass plants occurred initially, recruitment was generally enhanced under PC, so total native grass plant numbers recovered, despite drought conditions experienced over the experimental period. Grain yield declined by 30 to 50% compared with the no till treatment but it is unclear if the reduction in input costs of almost 50% reported by some landholders (Barton 2004) is able to offset the reduced grain production. Preliminary gross margin analyses in the Grain & Graze project suggest PC provided, at best, half the gross margin of zero tillage (though it should be noted that the red grass pasture was relatively dense and a better result might be expected in a more degraded situation). These losses in grain yield appear to be linked to declines in nitrogen. However, more importantly, the run-down of nitrate levels may limit annual weed growth in winter, encouraging the survival of newly recruited perennial species. This appears to be supported by the results of studies that have applied sugar to promote the immobilisation of N (Hacker et al. 2007; Smallbone et al. 2007). Combined, these studies suggest that PC may play a potential role in the rejuvenation of degraded native pastures by competitively excluding annual weedy species. Crucially, information on the sequence of PC within farming or livestock enterprises is currently being evaluated under this Grain & Graze initiative. Ground cover and litter appear to continually decline under successive PC to around 30 to 40% of native pasture. However, there were indications that, given adequate summer rainfall, ground cover may return to values comparable with pasture alone. During the experimental period, levels of soil moisture were not significantly different between treatments but, given the run of dry years over the experimental period, this is not surprising. The on-going project should provide additional information. It is important to note that the experimental sites for the PC Grain & Graze initiative were located in the aseasonal rainfall zone with mean annual rainfall in excess of 600 mm. The suitability of this cropping technique for summer-dominant rainfall areas remains unresolved. Given the reliance on stored soil moisture from summer rain for subsequent cereal crops in northern NSW, it may be expected that significant soil moisture would need to be stored under PC in this situation. The exception to this might be soil types that did not hold soil moisture efficiently. Pasture cropping does appear to provide a flexible farming system that has potential benefits for rejuvenation of native pastures. However, it appears that PC may be most profitable when there is a capacity to utilise the additional forage, (with livestock), made available either through the absence of summer


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fallow or by the presence of failed cereal crops. Again, the on-going project should provide this information. No-kill cropping (previously referred to as Advanced Sowing) is a modification of PC that involves no herbicide/no fertiliser and is assumed to impact least on biodiversity and on the simplification of the native grassland [Refer Case Study 5]. There are as yet no published reports comparing the effects of PC and No-kill cropping on biodiversity or profitability. 2.4.5 Alley farming The Grain & Graze Regional Initiative: Central West/Lachlan, is currently evaluating an alley farming system in which 20% of the paddock area is sown to alleys of Old Man Saltbush (OMSB), Atriplex nummularia (Hacker 2008). Conventional crop rotation is carried on between the alleys. The objective, from a production perspective, is to allow the reliable production of prime lamb, rather than wool, by ensuring a supply of relatively high quality feed over summer. Preliminary results show that after three years, the incorporation of OMSB alleys resulted in improvements in native plant species richness and native mid-storey cover compared with conventional cropping. This indicates that incorporation of OMSB alleys may provide habitat for a wide range of native plant species. This improved habitat condition did not, however, lead to an associated increase in bird diversity or abundance, although there was some evidence that over time more woodland bird species and decliner bird species may use OMSB sites compared to conventional paddocks. Hydrological studies on the experimental site indicate that OMSB is progressively ‘dewatering’ the soil profile below the crop root zone (1.0 - 2.8m). Data for several more years are required before the production benefits of the system will be clearly established. To date there is some evidence that sheep production may benefit from the incorporation of OMSB, particularly during dry periods when feed is of limited availability. There is also anecdotal evidence of benefits to native understorey species (particular grasses) through the provision of forage shrubs [Refer Case Studies 1 and 5]. 2.5 Weed and pest management 2.5.1 Weed management The use of chemicals for weed control remains a major component of agricultural activities. Considerable recent research has resulted in guidelines for reducing chemical use and retaining soil and nutrients on-farm, thus avoiding off-site contamination through runoff into riparian areas and percolation into ground water systems. Some widely promoted techniques to manage or minimise these off-site effects include erosion control to ensure chemicals are not removed with soil, matching chemical use with crop/pasture requirements and, more recently, integrated pest management. Adoption of voluntary Environmental Management Systems (EMS) in agriculture has been slow but offers a potential tool to minimise the risks of off-farm chemical impacts. The most widely used EMS is ISO 14001, but the prospect of


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positive biodiversity outcomes will be limited by the lack of appropriate biodiversity indicators or surrogates. The adoption of components of EMS without full accreditation may prove a useful approach for primary producers to reduce and better manage chemical use. Cunningham (2007) evaluated the potential of EMS to improve farm management and suggests that they are most useful when combined with other complimentary policy instruments. 2.5.2 Integrated pest management

A major function of native vegetation is the provision of the ecosystem service of pest control. Retention/enhancement/revegetation of native vegetation will allow the adoption of integrated pest management (IPM). While IPM takes into account all pest management options, including the use of insecticides, a central concept is that pests are managed to acceptable levels by utilising, as much as possible, naturally occurring beneficial insects and mites to control pest species. Native vegetation thus provides habitat for beneficial insects and birds and it offers potential benefits through IPM which can not only off-set input costs associated with insecticides, but also increase crop yields (Landis et al. 2000). The distribution of pest and beneficial insect species between shelterbelts and adjacent pastures was studied in central Victoria (Tsitsilas et al. 2006). Both pest mites and Lucerne Flea populations were lower within shelter belts than in pastures, and these differences were enhanced by greater diversity in understorey vegetation (more grass species, > 30 cm high). This study also showed that predator populations from within shelter belts were more effective at suppressing pest mite numbers than those collected within adjacent pastures. Schellhorn and Bianchi (2008) are undertaking research on the ecosystem service benefits of pest control from native vegetation. In particular, this research may shed light on the coverage of native vegetation required to significantly suppress pests. There has been some suggestion that farming practices such as stubble retention and minimum tillage have led to an increase in pest species. However, recent research conducted in Western Victoria as part of the national Grain & Graze program has shown that the environment created within crop and pastures allows significant increases in certain beneficial species (Nicholson and Horne 2006). This study also reports remnant native grassland to contain a greater diversity of beneficial insects than crop/pasture areas. The native vegetation acts as a species ‘reservoir’ for adjacent crops and pastures, but that numbers were insufficient within native vegetation to provide any immediate, direct biological control benefit in these adjacent paddocks. Importantly, these results also suggest that where native woody remnants are dominated by an understorey of exotic species such as phalaris, the diversity of some beneficial species (e.g. beetles) declines, although populations increase in size. Poor quality native vegetation sites may thus not provide the same population of beneficial species as higher quality remnants, raising questions regarding the extent to which recently planted vegetation corridors may be able to support significant populations of beneficial insects.


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Managing farming to maximise insect populations may also have implications for bird populations. A recent study in the United Kingdom linked changes in insect abundance to bird density decline and correlated with agricultural intensification (Benton et al. 2002). These authors also suggest that increasing the area of non-cropped habitat (such as wider, unsprayed, headlands or paddock margins) may provide forage habitat for bird species but, should, ideally, be adjacent to nesting sites to offer the greatest benefits. The role of native vegetation in supporting the key ecosystem service of pollination is often cited as being important for crop productivity, with over one third of world food crops depending on pollinators (Klein et al. 2007; Cunningham and Blanche 2007). Ricketts et al. (2008) provide a synthesis of the international literature that describes the relationship between animal pollination services and distance from native vegetation. They report a decrease in both pollinator richness and native species visitation rate with increasing distance from remnant vegetation. While this relationship varied with climatic region, broadly, visitation rates of pollinators tended to decrease significantly greater than 600 m from native vegetation. However, pollinator richness declined at greater distances (1.5 km). The relationship between distance to native vegetation and fruit/seed set did not necessarily follow the same relationship. While the benefit of remnant vegetation within broadacre cereal crops such as canola (Morandin and Winston 2005) and sunflower (Greenleaf and Kreme 2006) has been demonstrated in the United States and Canada, there is a need for comparable studies in Australia. 2.5.3 Invasive native shrub This review has previously reported that native species diversity is generally greater in uncleared areas compared to cleared areas, but it is difficult to separate the effects of livestock grazing and other activities (associated with clearing) from the effects of subsequently introduced exotic species, many of which are considered weeds. Generally, weeds have a negative impact on native flora and fauna. In the semi-arid rangelands, these effects have been reviewed by Grice (2004). Invasive species are seen as a major threat to biodiversity (SoE, 2006; Environment Defenders Office, 2008). Invasive native scrub (INS) has received much attention (particularly in the development and evaluation of control techniques) over past decades in the semi-arid rangelands within the WCMA, and research is currently examining management options for control in the Endangered Ecological Community of Coolibah–Black Box Woodlands of North western NSW (Good and Reid 2008). The catchment-scale impacts of INS on biodiversity have received less attention, but have been incorporated into INS tools where INS patches can be seen to help sustain greater landscape biodiversity but to have significant negative impacts on agricultural productivity. Climate change is likely to result in changes in INS distribution patterns, and thus on the heterogeneity created by INS patches, but the likely impact on regional biodiversity is largely unknown. At the paddock scale, INS areas are floristically more homogeneous compared with relatively open areas. In a unique study undertaken as part of


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the WEST2000 program, changes in biodiversity under different INS densities was examined (Ayer et al. 2001). In this study, floristic species diversity appeared to be lower in INS areas, but varied. Biodiversity, and the relative abundance of species, was similar for a range of taxa (vertebrates and invertebrates) irrespective of INS species, shrub density and location. However, the habitat value of shrubby areas varied between native fauna and regions. Shrubby areas were less suitable for ground-feeding, seed-eating birds and supported fewer understorey and groundcover species (e.g. grasses, forbs and chenopods). Where there is justification to remove INS, recent studies (Alemseged et al. 2008) suggest that short term cropping could be an effective means of promoting the restoration of native grasslands if post-cropping grazing management is satisfactory. There appears to be no intrinsic reason why healthy native grassland should not be able to re-establish following shrub removal and short term cropping. 2.6 Water-point management 2.6.1 Riparian areas, farm dams and irrigation channels Riparian areas provide the link between aquatic and terrestrial ecosystems and, therefore, are referred to as ‘keystone’ ecosystems (Gregory et al. 1991). Despite this, these areas remain generally in poor condition across agricultural landscapes (Rutherfurd et al. 2000). Key management activities to rehabilitate these areas include the re-establishment of buffer zones and the restriction of livestock access. The size of these buffer zones has been debated, but will depend on the objectives, which are usually multiple (e.g wildlife habitat, bank stability, ensuring water quality). A draft ‘Guidelines for Management of Wetlands On-farm’ is currently being reviewed to be finalised towards the end of 2008 (S. Davis pers. comm.). This document reports eight state-wide case studies to illustrate management options for wetlands and will therefore provide valuable examples of management options for the areas of interest to the B&PWG The Aquatic Habitat Rehabilitation program – ‘Fish Friendly Farms’ - provides seven management activities to enhance fish habitats on farm. These are:

• Retain large woody debris (snags) in your streams • Grow native vegetation on stream banks • Install fish friendly crossings • Control or treat agricultural run-off • Provide water for stock off-stream • Control the opening of floodgates • Protect wetlands

A major component of these activities is riverbank rehabilitation measures that may include fencing around waterways, revegetation (for stabilisation, provision of continual woody debris source) and weed control (Fish Friendly Farms 2008). In addition, fish passage restoration activities (e.g. removal of


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barriers such as road crossings or weirs to allow the migration of fish up and down stream, provision of shelter for fish by retaining fallen logs) are considered to be beneficial. Management of run-off carrying chemicals and nutrients to waterways and modification of floodgates are also important management activities in intensive cropping enterprises. 2.6.2 Water point management in rangelands Development of artificial watering points in the NSW rangelands has been so extensive that very few areas now lie outside the grazing range of domestic livestock (Figure 2.5). This massive introduction of artificial water into a previously dry landscape is generally thought to have facilitated an increase in the density of some kangaroo species, particularly those which are commercially utilised, contributing to the total grazing pressure on native vegetation. Populations of feral goats have also no doubt been favoured by this development. In a major study of biodiversity trends away from watering points in semi-arid and arid rangelands, Landsberg et al. (1997) showed that, for the several taxa studied, between 15 and 38% of species appeared to be ‘decreasers’ (i.e. their abundance declined with proximity to water), 10-33% appeared to be ‘increasers’ (whose abundance increased with proximity to water), while the remaining species exhibited no appreciable response. Transects were selected in relatively uniform country and extended for 10-12 km, the farthest sampling sites being beyond the grazing range of domestic livestock. No suitable site was identified for this study in western NSW due to the extensive water development noted above. These trends in species abundance do not necessarily reflect a response to grazing pressure alone but rather to the total influence of pastoral development. Nevertheless, it is clear that a substantial number of species are disadvantaged, though not necessarily threatened with extinction, by extensive water development. Thus, part of any program for improvement of biodiversity at landscape scale in western NSW will require the cessation of further water development, and attempts to de-water areas where feasible. Part of this de-watering process could involve the replacement of permanent watering points, or open bore drains, with temporary or portable waters. This innovation can be seen as part of a process aimed at improved management of total grazing pressure and the regeneration of native pastures. It should therefore have positive long-term production benefits. Its benefits for biodiversity are uncertain but it might be expected that biodiversity per se would benefit with the general regeneration of native pastures. Biograze (2008) is an approach to property planning that incorporates conservation of biodiversity within rangelands. Placement of water points and fencing is designed to allow a percentage of a paddock (water remote areas) to be retained for ‘decreaser’ species, disadvantaged by grazing without compromising profitability. For example, in a sheep grazing enterprise 10% of the total area is retained, but only 5% for cattle enterprises. These results were derived for Mitchell grasslands on the Barkly Tablelands for cattle


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enterprises, and from the Kingoonya region of South Australia for sheep enterprises. It is not clear if retention of the same proportions for grazing sensitive species would be sufficient to achieve a comparable level of conservation in western NSW, where chenopod shrublands and Mitchell grasslands vary in composition, or where unpalatable woody weed species occur. Nevertheless, the positioning of watering points will be of major importance in determining the distribution of grazing within paddocks (Hunt et al. 2007) and biodiversity outcomes (James et al. 1999).


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Figure 2.5. Distribution of permanent and semi-permanent waters on the White Cliffs 1:250 000 topographic map sheet before settlement (above) and at the present time (below). The solid and dashed circles in the top figure represent 10 km radii from permanent waters (including mound springs) and semi-permanent water holes, respectively. Only areas within the solid circles would have been permanently habitable for kangaroos. In the bottom figure red and green circles represent sheep grazing radii (5 km) from earth tanks (407) and bores and wells (155), respectively. Virtually all of the landscape is permanently accessible to both sheep and kangaroos.(After Hacker and McLeod 2003).


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3. A whole of landscape approach The previous sections have dealt with management targeted at biological aspects of biodiversity at the small, farm-scale. As we have previously emphasised, different species operate at different spatial scales and genetic variation within species can occur across their natural range. This is particularly important in rangeland areas where ecological and landscape processes are highly variable (spatially and temporally) and, consequently, the patterning of biodiversity is dynamic (Stafford-Smith and Ash 2005). It is therefore difficult to find a set of principles to guide the management of biodiversity at the landscape scale or that are applicable across each of the CMA areas. Because we have primarily discussed management that targets small-scale, paddock or whole-farm activities, it could be argued that such approaches are essentially only patching up the landscape. Local vegetation enhancement or revegetation activities, for example, could be seen simply as defensive strategies which do not consider the broader landscape. We acknowledged early in this review that while National Parks represent an important component of species conservation, they fail to adequately address conservation of fragmented landscapes such as those found in agricultural areas or the conservation of genetic diversity across non-fragmented landscapes such as the rangelands. Where the conservation of biodiversity is integrated within agricultural landscapes (as described in previous sections) through changes in management (‘well-managed paddock model‘), it may slow the rate of biodiversity loss, but there is little guarantee that ‘best management practices’ will be maintained with changes in land ownership. Thiele and Prober (1999) propose a Conservation Management Network (CMN) to provide a mechanism that takes advantages of both the National Park model and the ‘Well Managed paddock’ approach. A CMN, is a network of remnants, their managers and other interested parties co-ordinated through a single administrative point (e.g. the Grassy-Box Woodland CMN is housed within the Department of Environment and Conservation). The advantages of such an approach are numerous:

• Biodiversity is conserved at an ecosystem level and should capture ecosystem, species and population biodiversity

• Biodiversity is conserved at both the small and larger, landscape scale • Multiple instruments for preservation can be employed • Management can be co-ordinated across remnants, and management

of one remnant can inform the management of another • A single point for brokerage of information

The CMN model provides a mechanism to share information and responsibility for conservation of biodiversity at the large scale, both of which are vital ingredients for successful community-based projects (see Marais et al. 2007 for examples of project failure due to a lack of shared knowledge among stakeholders). Despite the shared knowledge created by the CMN model, major barriers still exist to large-scale restoration activities. Apart from funding, these include the ‘shifting baseline syndrome’, the scale and


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complexity of restoration and the long-term/open-ended nature of restoration. Manning et al. (2006b) suggest ambitious long-term ‘stretch goals’ and ‘backcasting’ or visualisation of the restoration endpoints as two possible techniques for approaching large-scale restoration activities. Currently, the potential use of software packages that allows the loss or restoration of vegetation to be visualised is being examined, and potentially offer a powerful method for the prevention of continued biodiversity loss and for restoration planning (A. Manning pers. comm.) Recently, Lindenmayer et al. (2008) documented principles for the integration of biodiversity at a landscape level but acknowledged these are strongly context-dependent. That is, they cannot necessarily be applied in all landscapes and serve more as a check list (Table 3.1). Table 3.1 Guiding principles for landscape conservation (Adapted from Lindenmayer et

al. 2008) Setting goals

• Develop a long-term shared vision with quantifiable objectives

Spatial issues

• Manage the entire mosaic (not just small bits) • Consider amount and configuration of habitat and land cover type • Identify disproportionately important species, processes and landscape

elements • Integrate aquatic and terrestrial environments • Use a landscape classification and conceptual models approach to objectives

Temporal issues

• Maintain capacity for landscape recovery • Manage for change • Recognise lag times occur and manage to reduce these or plan for them

Management approaches

• Manage in an experimental framework • Manage both species and ecosystems • Manage at multiple scales • Allow for contingency

The importance of classifying the landscape to develop large-scale management guidelines has been highlighted in a number of recent publications (McIntyre and Hobbs 1999; Hobbs and McIntyre 2005). These authors suggest the use of a large-scale agro-climatic classification aligned to IBRA bioregions to identify landscapes with common characteristics, sharing similar threats and approaches to management (Hobbs and McIntyre 2005). At a smaller regional scale, patterns of landscape modification can be identified beyond a simple habitat vs. non habitat classification to ensure consideration of the degree of disturbance is incorporated into management objectives (McIntyre and Hobbs 1999). For example, landscapes may be classified as intact, variegated, fragments or relictual based on the degree of habitat destruction, connectivity modification and pattern of modification of the remaining habitat.


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How regional natural resource management bodies ensure that management at these larger scales is linked to management on farms remains problematic. The CMN model is one possibility but it probably requires administration at a higher level than the CMA (eg by a government Department) if all of the relevant landscapes are to be included. Briggs (2001) proposes local councils are the best bodies to provide the link between the two scales. Here, it is argued that councils have a statutory ability to undertake landscape and biodiversity management. While this has occurred in some situations (e.g. Ecotracks 2007), this author also acknowledges that most council’s are under-resourced to meet such an obligation. It is however possible that alliance between CMAs, agencies and councils may support such a mechanism. Environment Management Systems (EMS) many provide another model but their adoption within the agricultural sector has been slow, probably due to their complexity and the level of detail required for planning and monitoring. In some sectors they have been replaced with Best Management Practices (BMP). However, there a number of aspects of such approaches that may facilitate large-scale integration of biodiversity and production. For example, an EMS or BMP may incorporate an audit component that serves as a reference point against which changes in management may be monitored. EMS and BMP approaches may also embed catchment biodiversity plans allowing on-farm management activities to be couched within the broader regional planning. It is doubtful if the costs and benefits of biodiversity are shared equitably among agricultural sectors let alone between urban and rural communities, given the difficulties in definition of biodiversity and the differences in its value to different sectors outlined at the beginning of this review. Within the agricultural sector there are difficulties in grasping the utility value of biodiversity. This is largely because the values of natural capital and ecosystem services supported by biodiversity are not costed and remain unappreciated, and the interdependence of agriculture and biodiversity unrecognised.


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4. Indicators of biodiversity health relevant to agricultural

production We have described biodiversity as a multiple scale character that underpins vital ecosystem processes supporting agricultural land use (see section 1.2) and, as such, the use of simple indicators to monitor the impacts of changes in land management is problematic. There has been a great deal of effort devoted to the search for biodiversity indicators that have broad applicability and can be used internationally (Teder et al. 2006), incorporate plants alone (Gibbons et al. 2008; Gibbons and Freudenberger 2006; Gibbons et al. 2005; Landsberg and Crowley 2004), incorporate a range of taxonomic groups (Kati et al. 2004) or mulit-scale attributes (Smyth and James 2004; Duelli and Obrist 2003; Noss 1990), or which can be used as benchmarks (Oliver et al. 2007; McCarthy et al. 2004). Despite these efforts, and initiatives to provide integrated software packages such as SCaRPA to predict outcomes of land use change (TOOLS2, 2005), biodiversity indicators that are both appropriate and of utility value in agricultural systems are lacking. Moonen and Barberi (2008) provide a structured framework that recognises ‘agroecosystem functional groups’ as key indicators relevant to agricultural landscapes and argue that a clear distinction needs to be made between biological indicators that reflect environmental change/status and agro-ecosystem functional groups that reflect interactions and regulation of agro-ecosystem processes. In this way, clusters of biota may provide the same agro-ecosystem service and could be grouped into the one functional group. This recognises that one biota can deliver a number of ecosystem functions. For example, for productive ecosystem services, an agro-ecological functional group may include nutrient cycling, decomposition rates, aggregate stability and organic matter formation. While ‘agro-ecosystem functional groups’ may be appropriate indicators for agricultural areas, they require considerable technical expertise to measure. The ‘Landscape Function Analysis’ (LFA) tool has been widely adopted (Tongway and Hindley 2004). This method provides a quick but rigorous monitoring procedure using simple indicators of landscape function (soil surface stability, infiltration and nutrient cycling). These indicators are measured by monitoring cryptogam, soil, litter and perennial grass cover; soil surface crusting, erosion type and severity, deposition materials, microtopography, surface resistance to disturbance, slake test and soil texture, and have been broadly applicable across a range of land uses (rangelands, mining, horticultural industries, habitat conservation) and as an indicator of soil carbon. Although this technique requires limited training, it has been widely used as part of regional, state and national resource monitoring and as an adaptive management tool for stewardship schemes (Ampt et al. 2007). The emphasis on using ground cover in the LFA technique and elsewhere highlights the importance of this indicator in landscape function.


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5. Case Studies In June 2007, a national conference in Tasmania organised by Tamar Natural Resource Management examined the integration of biodiversity and production. This conference developed The Tamar Principles5 for the integration of biodiversity (Lloyd et. al.2008): Respect: The respect for nature and natural process starts

with respect for ourselves and for others. Consider the Future: There is a duty of care to those around us as well

as to future generations. Set Goals: Clear unambiguous goals that allow management

to be matched to the scale of the goal should be set.

Be Open: To be open to novel management techniques and

ways of thinking. Learn: Personal anecdotes, the experiences of our

neighbours, scientific and technical writing: we listen and learn from them all.

Demonstrate: To record our success and our failures and

reorganise, celebrate and promote all who progress.

Share: To seek ways to share or spread across the whole

community the costs of biodiversity. A characteristic of all case studies identified in this review was that each manager was actively using some or most of these principles. Another common theme was that individuals expressed a desire to work with rather than against nature. In some cases, this required using less inputs and not trying to ‘control’ the environment (e.g. by reducing high levels of inputs, farming systems became more flexible and able to respond to different seasonal conditions). This provided additional lifestyle benefits.

5 Lloyd (2007) Tamar Principles http://tamar-nrm.org.au/BIO-Nov/conference.html


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Case Study 1. Graham, Jan and Garth Strong – Regenerative agriculture

“Arcadia”, Narrandera

“Management is about keeping the energy cycling, biodiversity is the battery that enables this.” Enterprise at a glance: Long-term mean annual rainfall is 425mm but the area has received considerable less in recent years; Dryland farming for fodder production; adopts many natural farming methods commonly used in many ‘Organic’ QA programs; 1700 self-replacing Merino ewes; first- cross lambs incorporating, development of ‘Arcadia Saltbush Lamb’ product sold directly to suppliers. Key elements: Graham sees that creating a flexible enterprise by not being locked into specific farming activities can allow the spread of seasonal risks, resulting in a less stressful working environment. In 1996 Graham commenced a succession of developments

• Increased shrub cover from 2% to 12 %, planted and direct seeded saltbush, over 250,000 trees and 120 ha of old man saltbush, managed remnant vegetation and improved the ground cover. It is planned to continue to increase these areas

• Trading carbon credits to add value to wool clip • Ceased insecticide and fungicide use on crops and pastures • Stopped burning crop stubble, commenced pasture cropping • Purchased and restored Birrego Church and developed it as a B&B • Commenced a farm stay as part of WWOOF (Willing Workers on Organic

Farms) • Regularly hosts University study tours and has an open door policy to those

wanting to learn • Derived income from bush foods and native seeds from trees • Maintained wool production and improved wool quality • Commenced a “Paddock to Plate” lamb enterprise where produce is sold

directly to markets under “Arcadia Saltbush Lamb” in Canberra, Wodonga and Sydney.

• Is currently employing a number of water harvesting techniques to redirect water and trap nutrient runoff to aid revegetation

• Also planning to incorporate ‘ponds’ within this water redirection for frog habitat


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Case Study 2. Tim and Karen Wright

“Lana”, Uralla

Enterprise at a glance: Mean annual rainfall 769mm; Historically has run both Merino ewes and cattle grazing enterprises; practise Holistic Management; Key elements: Tim and Karen’s example has been used in the past as case studies and collaborated with research projects. “Lana” represents a valuable example of long-term time controlled grazing enterprise. http://www.landcareheroes.com/profile/tim-and-karen-wright/67/23/ http://www.holisticmanagement.org.au/PDF/The%20Wrights%20LWW.pdf .


http://www.landcareheroes.com/profile/tim-and-karen-wright/67/23/

http://www.holisticmanagement.org.au/PDF/The%20Wrights%20LWW.pdf

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Case Study 3. Ripariain vegetation management and vegetation corridors

Terry Haynes and Harvey Gaynor, “Midkin Farm”, Auscott Ltd, Moree.

At a glance: This enterprise represents a major irrigation development and farming enterprise in the Gwydir valley and is a working example of biodiversity management in a highly modified landscape using vegetation enhancement and reduction of off-site impacts. About 25 % of the farm is riparian zone. Over ten years ago this company ceased grazing and farming of these areas to allow native regeneration as well as undertaking some replanting. A major focus of these activities was the prevention of erosion. Since commencement of these activities, these areas have been largely stabilised in terms of maintaining woody cover. In addition to the control of erosion, these areas are also seen as important in the provision of buffers for the control of spray drift between fields and between crops and the riparian areas. This enterprise has collaborated with numerous university studies, Landcare and CMA projects in riparian areas and elsewhere on the farm. Key elements:

• Approximately 100 ha of “Green belt” tree plantings with some native understorey natural regeneration in and around a Cotton Gin complex and utility areas; 20 ha field boundary tree planting with approximately 50% survival rate.

• This enterprise has also managed the control of noxious weeds and feral animals within these areas. Significant expenditure on selective spraying (noxious weed control) occurred within the first 24 months of tree planting with additional ongoing treatment for the following three years.

• Additional ongoing problems are occurring or the control of noxious weeds (Mimosa, Box Thorn, Tree and Tiger Pear and Galvanised Burr) within tree plantings.

• An increasing problem in the riparian remnants is the proliferation of high stem densities of Coolibah (Eucalyptus microtheca) and River Red Gun (E. camaldulensis) at the expense of herbaceous and grassy understorey species. Some thinning activities have been undertaken, with little or no success. Whether natural thinning will occur is unknown. High tree density remnants may not perform the same functional role of mitigating flood impacts. Therefore despite incurring significant past costs for revegetation programs and their on-going management the function role of these wooded areas in mitigation of flood impacts is unknown.

• Whilst the introduction of Genetically Modified Ingard® cotton has resulted in reduced insecticide use, it is believed that remnant native vegetation areas may still provide some benefits in terms of provision of habitat for beneficial insects.


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Case Study 4. Revegetating Rangelands . Graham and Kathy Finlayson

“Bokhara Plains”, Brewarrina

Enterprise at a glance: Mean annual rainfall 400 mm: in the past, predominately sheep with some cattle but shifting toward adjustment of livestock only. A 7,000 ha property under cell grazing with an additional 1,000 ha leased. Participant in the Western Enterprise Based Conservation Program (EBC) – payments based on maintaining percentage ground cover; inclusion of farm stay/conference facility enterprise “Bokhara Hutz”. Key elements: The Finlayson’s purchased “Bokhara Plains” in 1999. Since 2001 the district has experienced a seven year drought resulting in complete destocking from 2005-2006 although destocking has also occurred for shorter periods. During these periods the income from the farm stay enterprise has provided the only income.

• The whole 7,000 ha (in the district this is considered too small for profitability) is run under cell grazing and the movement of stock is based on pasture utilisation levels. Aim to maintain 40% ground cover to meet the Finlayson’s personal expectations as well as stewardship payment requirements.

• Ultimately aim to achieve 70% ground cover. • Central to grazing management is finding a balance between preserving the

vegetation resource (assessing how much grazing pressure can be applied to different vegetation communities) and using livestock to disturb soil and distribute fertility.

• Plant population densities have been maintained, or improved in some areas despite drought conditions.

• Scalded, claypan areas have been disturbed using a blade to create furrows that successfully capture moisture providing a favourable micro-site for seedling establishment. This is being used as an initial treatment to kick start plant succession.

• In the future it is planned to run adjustment livestock herd only giving the flexibility to remove stock as required.

• Cost of fencing and water has partially been met through involvement in the EBC, but also represents considerable financial outlay for the Finlayson’s. Continued subdivision of paddock is being planned.

• Additional income sources to complement primary industry activities are being actively sought.

http://www.bokharaplains.com/functions.htm


http://www.bokharaplains.com/functions.htm

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Case Study 5. No-Kill cropping and forage shrubs. Bruce Maynard

“Willydah”, Narromine

Enterprise at a glance: Mean annual rainfall 500mm, largely trading livestock, currently mostly cattle; total of 1500 ha; 350 ha old man saltbush; although this area alters between years; approximately 1/3 property under No-kill Cropping; 15% of the property under tree regeneration and an additional 4% planted for Carbon Credits; cell grazing used across whole property. Key elements: Bruce commenced cell grazing in 1994 and No Kill Cropping was developed in 1996. The first Saltbush and Tagasaste (another forage shrub) plantings were undertaken in 1988. The primary objective is to increase biodiversity; Bruce believes that livestock are an integral tool to achieving this objective.

• Aims to maintain 100% ground cover, decisions to move stock based upon an appraisal of available forage. The amount judged as available is what can be harvested without putting perennial plants at a disadvantage. This then dictates the stocking rate which is adjusted constantly.

• No-kill cropping of oats used to enhance biodiversity of native pastures and provide additional feed resource. Achieving a cereal crop is a secondary objective. The area ‘cropped’ varies according to year with a greater area sown during dry periods. When an early season break occurs, less No-kill cropping occurs but a greater likelihood of harvestable crop.

• Bruce believes the tree regeneration areas will serve dual purpose of provision of ‘deep, cool shade’ (requirement for which is likely to increase with climate change) as well as payment for carbon sequestration.

• Recently, enterprise is to include breeding flock of ewes. Some past difficulties with the introduction of adjustment or trading stock onto unfamiliar forage shrub diet may be overcome by running a small number of ‘experienced’ livestock alongside introduced flocks.

• Bruce has actively pursued information on animal behavioural science, travelling to the US to attend the latest course presented by Fred Provenza. The Provenza research has shown that animals can utilise a greater diversity of plants (including weeds) and these observations are consistent with the grazing effects that are occurring on Willydah. These grazing improvements are built upon good grazing management and stress free stockmanship.

• Bruce believes the introduction of multiple grazing species (cattle, sheep, goats and alpacas) may provide not only economic benefits, but also benefits in terms of differential grazing patterns to enhance biodiversity.

• Aims to include direct seeding of trees and additional shrub species within saltbush plantings and en.

• Different spatial arrangements of Saltbush planting has been tested, alleys, concentric circles and spirals.

• Recently adopted the use of polywire to provide more flexibility in subdivision of grazing cells

• Welcomes visitors and is actively involved in teaching No Kill Cropping and Stress Free Stockmanship methods.


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Case Study 6. Holistic grazing and pasture cropping in a summer dominant rainfall area. Richard and Janet Doyle

“Malgarai”, Boggabilla

Enterprise at a glance: A total of 8,000 acres, with 1200 acres alluvial soils and 6800 acres black clay soils. Historically about half this area was under conventional tillage (dry-land cereal crops) and about 60-70% depleted native grasses. A major objective over the past 8 years has been to ensure farm management outcomes are in line with life-style expectations; this revolves around management is targeted toward ecological benefits as well as farm profitability. This enterprise represents an early stage use of Pasture Cropping for gain production in a summer dominant rainfall area and incorporation of cell grazing over the past eight years. Key elements:

• Native pasture regeneration has been achieved through the use of cell grazing and/or pasture cropping, although the latter is in its early stages of implementation. Richard and Janet have seen increased proportions of short-lived (e.g. Austrostipa spp. and Panicum sp.) and long-lived perennial grasses (e.g. Astrebla spp) in paddocks under cell grazing

• The incorporation of Pasture Cropping for grain production is seen as integral to farm profitability

• Pasture Cropping is also viewed as a mechanism for removing chemical fallow from the enterprise

• Stock numbers have varied from 850 to 1200 live stock units, depending on seasonal conditions and animal. Importantly, their enterprise has moved away from breeding herds toward

• Richard and Janet have also undertaken extensive riparian revegetation activities


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6. Recommendations

Recommendation

Strategy (page reference)

6.1 Management focused explicitly on the creation of heterogeneity - ‘Mix it up/keep it messy’

6.1.1. Establish strategic alliances between CMA’s, local councils, agencies and business that enable the conservation of biodiversity at the regional scale using models similar to the Conservation Management Network. (p.48) 6.1.2. Promote the establishment of structurally and floristically diverse plant communities within revegetation planning. (pp18-19) 6.1.3. Promote the use of both high-density tree planting and wider spaced plantings to ensure structurally diverse revegetation (broad range of habitats). (pp18-19) 6.1.4 Promote the retention of habitat resources that accumulate over long time periods to develop greater diversity of habitat. For example, the retention of fallen timber and log hollows.(p 19) 6.1.5. Ensure that existing water-remote areas are retained for biodiversity conservation in the semi-arid zone. (p45-47) 6.1.6. Explore the potential for closure of watering points, especially in the semi-arid zone, where this will produce sizeable tracts of land that are remote from water and domestic livestock grazing. (p45-47)

6.2 Re-design of land use for biodiversity outcomes.

6.2.1. Promote and demonstrate the interdependence of agricultural production and biodiversity, based on the findings of this review. 6.2.2. Provide information and assistance to landholders for paddock subdivision where land capability can be matched to land management units. (p11) 6.2.3. Promote and demonstrate the rearrangement of land use for dual outcomes (production and biodiversity) e.g. fencing of riparian areas; retention of native vegetation for IMP, livestock shelter and crop protection; tree planting for shelter and carbon credit. 6.2.4. Promote management that offers increased flexibility in response to changes in environmental conditions to minimise degradation. e.g. no-kill cropping and livestock trading.


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6.2.5. Develop incentive schemes by exploring local council and industry alliances to seek long-term financial incentives for agricultural activities that result in biodiversity gains at the expense of private profit. 6.2.6. Promote current and developing financial off-set/stewardship schemes within each CMA area e.g. tree planting for carbon trading/offset; Enterprise Based Conservation. (p13) 6.2.7. Employ multiple policy instruments (regulation, motivation, and economic incentives) to integrate biodiversity and production. (pp12-13) 6.2.8. Investigate the extent to which stewardship schemes may provide a better basis for public expenditure, in terms of landholder support and biodiversity benefits, than Exceptional Circumstances assistance. (p13)

6.3 Demonstrate and evaluate options for micro-restoration of highly modified landscapes.

6.3.1. Promote the ongoing restoration and enhancement of native vegetation (paddock perimeter and corridor plantings) to ensure continued creation of habitat resources and more diverse age structure in vegetation. 6.3.2. Promote the benefits of retaining paddock trees. (p 19) 6.3.3. Evaluate methods for paddock tree enhancement (e.g. enhancing connectivity, consolidation and enrichment of existing trees). (p19) 6.3.4. Evaluate the biodiversity value of past revegetation activities by setting up monitoring for recent and past plantings (comparing perimeter plants, narrow spaced rows and single species tree plantings with tree plantings of varied species and structural components. 6.3.5. Evaluate the biodiversity value of planting perimeter paddock tree in extensive croplands of northern-western NSW.

6.4 Assessment of long-term impacts of high intensity short duration stocking on native pastures.

6.4.1. Support research aimed at evaluating the impact of high intensity short duration grazing regimes on landscape function, especially in the semi-arid zone. (pp28-30) 6.4.2. Continue monitoring of properties where high intensity short duration grazing is being practiced and seek additional co-operators. (pp26-28)


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6.5 Support research to address critical knowledge gaps.

6.5.1. Adopt an adaptive management approach that treats new management options as ‘experiments’ where changes in biodiversity and other natural resource outcomes can be monitored and evaluated. 6.6.2. Is it more important to manage biodiversity within the existing land use or to change the land use mix? 6.6.3. What threshold of habitat fragmentation is beyond landscape self-repair? Is the 30% intensification/woody cover applicable for all agricultural landscapes (rangelands, extensive croplands and grasslands)? 6.6.4. Support research projects that seek to quantify the economic and biodiversity benefits of changes land use patterns using production economic and biodiversity modelling techniques now available. 6.6.5. Evaluate alternative biodiversity indicators of utility value to agricultural producers.


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