When trends intersect: The challenge of protecting freshwater ecosystems under multiple land use and hydrological intensification scenarios.

STOTEN-17589; No of Pages 14 Science of the Total Environment xxx (2015) xxx–xxx

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

When trends intersect: The challenge of protecting freshwater ecosystems under multiple land use and hydrological intensification scenarios Jenny Davis a,⁎, Anthony P. O'Grady b, Allan Dale c, Angela H. Arthington d, Peter A. Gell e, Patrick D. Driver f,g, Nick Bond d, Michelle Casanova e, Max Finlayson h, Robyn J. Watts h, Samantha J. Capon d, Ivan Nagelkerken i, Reid Tingley j, Brian Fry d, Timothy J. Page d, Alison Specht k a

Institute for Applied Ecology, University of Canberra, Bruce, ACT 2617, Australia CSIRO Land and Water, Private Bag 12, Hobart TAS 7001, Australia c The Cairns Institute, James Cook University, Cairns, QLD 4871, Australia d Australian Rivers Institute, Griffith University, Nathan, QLD 4111, Australia e Federation University Australia, Water Research Network, Mt Helen, VIC 3353, Australia f Office of Water, NSW Department of Primary Industries, Orange, NSW 2800, Australia g Centre for Ecosystem Science, University of New South Wales, Kensington, NSW, Australia h Institute for Land, Water and Society, Charles Sturt University, Albury-Wodonga, NSW 2640, Australia i School of Biological Sciences and The Environment Institute, The University of Adelaide, Adelaide, SA 5005, Australia j School of BioSciences, The University of Melbourne, VIC 3010, Australia k ACEAS, Australian Centre for Ecological Analysis and Synthesis, a facility of the Terrestrial Ecosystem Research Network University of Queensland, St Lucia, QLD 4067, Australia b

H I G H L I G H T S • • • • •

This paper considers the impacts of land use and hydrological intensification on inland waters Global issues are considered through the lens of Australian examples Likely scenarios include wet regions becoming wetter, dry regions drier and storms more intense The legacies of past land use change will need to be addressed Proactive governance based on innovative science and adaptive management will be critical

a r t i c l e

i n f o

Article history: Received 26 September 2014 Received in revised form 25 March 2015 Accepted 29 March 2015 Available online xxxx Keywords: Land use intensification Hydrological intensification Climate change Freshwater biodiversity Freshwater ecosystems Extreme events Floods Droughts

a b s t r a c t Intensification of the use of natural resources is a world-wide trend driven by the increasing demand for water, food, fibre, minerals and energy. These demands are the result of a rising world population, increasing wealth and greater global focus on economic growth. Land use intensification, together with climate change, is also driving intensification of the global hydrological cycle. Both processes will have major socio-economic and ecological implications for global water availability. In this paper we focus on the implications of land use intensification for the conservation and management of freshwater ecosystems using Australia as an example. We consider this in the light of intensification of the hydrologic cycle due to climate change, and associated hydrological scenarios that include the occurrence of more intense hydrological events (extreme storms, larger floods and longer droughts). We highlight the importance of managing water quality, the value of providing environmental flows within a watershed framework and the critical role that innovative science and adaptive management must play in developing proactive and robust responses to intensification. We also suggest research priorities to support improved systemic governance, including adaptation planning and management to maximise freshwater biodiversity outcomes while supporting the socioeconomic objectives driving land use intensification. Further research priorities include: i) determining the relative contributions of surface water and groundwater in supporting freshwater ecosystems; ii) identifying and protecting freshwater biodiversity hotspots and refugia; iii) improving our capacity to model hydro-ecological relationships and predict ecological outcomes from land use intensification and climate change; iv) developing an understanding of long term ecosystem behaviour; and v) exploring systemic approaches to enhancing governance systems, including planning and management systems affecting freshwater outcomes. A major policy challenge will be the integration of land and water management, which increasingly are being considered within different policy frameworks. Crown Copyright © 2015 Published by Elsevier B.V. All rights reserved.

⁎ Corresponding author. E-mail address: [email protected] (J. Davis).

http://dx.doi.org/10.1016/j.scitotenv.2015.03.127 0048-9697/Crown Copyright © 2015 Published by Elsevier B.V. All rights reserved.

Please cite this article as: Davis, J., et al., When trends intersect: The challenge of protecting freshwater ecosystems under multiple land use and hydrological intensification ..., Sci Total Environ (2015), http://dx.doi.org/10.1016/j.scitotenv.2015.03.127

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J. Davis et al. / Science of the Total Environment xxx (2015) xxx–xxx

1. Introduction Intensification is a key characteristic of many emerging global ‘megatrends’; trends that, on a global scale, will significantly shape the ecological, social, economic and cultural landscapes of the future (Hajkowicz et al., 2012). These include urbanisation, increasing mineral extraction and energy production and the requirement to obtain more resources from a declining natural resource base. A rising world population, forecast to be 9 billion people by 2043 (UNESA, 2012), increased wealth and changing dietary preferences suggest that global food production will need to increase by 70% by 2050 (Steduto et al., 2012). Potentially constraining this production is the threat of water scarcity. Vörösmarty et al. (2010), using a global geospatial framework, showed that pandemic impacts on both human water security and freshwater biodiversity were highly coherent, although not identical. Nearly 80% of the world's population (4.5 billion in 2005) was exposed to high levels of threat to water security, while 65% of global river discharge, and the aquatic habitats supported by river flows, were classified as moderately to highly threatened (Vörösmarty et al., 2010). The most serious impacts, which include watershed changes, pollution and water resource development, coincide in regions of intensive agriculture and dense human settlement. Already, the global agricultural sector accounts for 70% of withdrawals from freshwater systems and more than 90% of consumptive water use (Steduto et al., 2012). The direct and indirect competition for water resources associated with population growth will intensify both mild and moderate droughts. Importantly, the impacts of land use intensification and increasing water demands have to be considered within the wider context of global climate change. Anthropogenically-driven climatic changes are already considered to pose a major threat to global biodiversity, including inland aquatic ecosystems and their species (Solomon, 2007; Woodward et al., 2010). Globally, freshwater ecosystems and biota are considered to be particularly vulnerable because of their physical fragmentation within terrestrial landscapes and relative isolation by catchment divides and saltwater barriers (Dudgeon et al., 2006). Many freshwater species will be unable to disperse to suitable habitats as temperatures increase and changes in precipitation disrupt migration and feeding and breeding patterns (Woodward et al., 2010). These climatically-driven changes will be accompanied by the direct and indirect effects of increasing human demands for water (Palmer et al., 2008). In regions where precipitation declines, surface and ground water resources and environmental flows will be increasingly contested. As a consequence, declining water availability will pose a significant threat to freshwater environments, as well as agriculture and human consumption. It is anticipated that by 2050, 2.3 billion people will be living in water basins experiencing severe water stress (OECD, 2012). The extent to which human communities will stay and adapt to declining conditions is difficult to predict. Most studies have focused on developing countries where mitigating circumstances, such as war and poverty, are present (Gemenne, 2011). Environmental extremes also cause hardships in developed countries, but the dynamic is likely to be different because of greater economic and political stability, and differences in agricultural technology. The implications of the interacting trends of land use intensification and hydrological intensification for the management and conservation of freshwater ecosystems are the focus of this paper. We consider the evidence for hydrological intensification, the existing legacy of land use change on the water quality and hydrological regimes of Australian river systems, the likely interacting effects of land use and hydrological intensification and the need for proactive governance and adaptive management. We also suggest a set of priority research actions to integrate land use intensification into freshwater management and conservation. 2. Evidence for intensification of the global hydrological cycle Associated with the broad trend in the expansion of the earth's population has been a major expansion of the global economy driven

largely by the exploitation of fossil fuel resources and land clearing, resulting in an associated increase in carbon emissions (Canadell et al., 2007). The rate of growth in atmospheric emissions has increased from approximately 1.3% per year during the 1990s to 3.3% per year during the period 2000–2006 (Canadell et al., 2007). This trend is likely to continue despite growing efforts to curb global emissions. Global warming is a major consequence of rising concentrations of greenhouse gases and based on current emission trajectories, temperature rises between 4 °C and 6 °C appear likely by the end of the century (Bodman et al., 2013; Peters et al., 2013). Rising temperatures are a key driver of changes in global circulation patterns and are likely linked to the global phenomenon of hydrological intensification (Durack et al., 2012b; Held and Soden, 2006; Huntington, 2006; Wild et al., 2008). A consequence of warming in the lower atmosphere is an increase in its capacity to hold water. The Classius–Claperyon expression predicts that the saturated vapour pressure of the lower troposphere increases by about 7% for each 1-K increase in temperature. This response is robust in most climate models. A key outcome of this process is the predicted intensification of the hydrological cycle such that wet areas are likely to get wetter and dry areas drier (Wentz et al., 2007; Chou et al., 2013; Held and Soden, 2006). Hydrological intensification will drive changes in the spatial and temporal distributions of water resources and an increase in the frequency and intensity of extreme events such as tropical storms, floods and droughts (Fig. 1). Disparate observational data sets generally predict that warming will likely result in increases in evaporation and precipitation, although there is little supporting evidence for predicted increases in the frequency and intensity of tropical storms and floods (Huntington, 2006). Attention is now focussed on developing an improved observational evidence base for hydrological intensification. An analysis of a network of 355 rain gauges across China for the period 1960–2000, for example, found that although there was no trend in the country-wide average rainfall, rainfall in the drier north-eastern regions of China had declined by approximately 12% since 1960, with declines mostly occurring in summer and autumn. In contrast, rainfall in southern China increased, particularly during summer and winter (Piao et al., 2010b). Associated analysis of stream flow records revealed a weak trend for increasing runoff in the Yangtze River in southern China and a significant decline in river runoff in the Yellow River in northern China. Direct attribution of changes in runoff to changes in precipitation regimes in both rivers, however, was problematic owing to the intense human pressure on water resources (Piao et al., 2010b). Examination of rainfall and discharge records for the Amazon have also revealed a marked increase in river discharge associated with an increase in the Amazon basin integrated precipitation, a trend consistent with intensification of the hydrological cycle (Gloor et al., 2013). Treydte et al. (2006) used tree ring analysis to construct a millennium scale precipitation record in northern Pakistan. They found dry conditions at the beginning of the last millennium and through the 18th and 19th centuries, with a trend for increasing precipitation in the latter half of the 19th century and throughout the 20th century, which tracks the initiation and expansion of the industrial revolution. There is also evidence for intensification of the hydrological cycle in ocean surface salinity data. Sea surface salinities in the ocean are widely measured and are used to interpret changes in the fluxes of freshwater, freshwater transport and local ocean mixing, key components of climate dynamics (Curry et al., 2003; Durack et al., 2012a). Examination of trends in ocean salinities from the 1950s to the 1990s along a transect through the western Atlantic spanning 50°S to 60°N revealed systematic freshening at both poleward ends and increasing sea salinities at low latitudes (Curry et al., 2003). Durack et al. (2012a) observed similar trends in ocean salinities and concluded that broad belts of increasing salinities through the tropics, with areas of freshening in mid and high latitudes, provided robust evidence for an intensification of hydrological cycles at a rate approximately double that predicted by climate models. Thus several lines of evidence suggest intensification of the global hydrological cycle. Global climate models also predict that this pattern



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Fig. 1. Conceptualisation of the major drivers and responses resulting in intensification of the global hydrological cycle. Population growth and associated increases in food, water and energy demands are key drivers of climate change through rising emissions of greenhouse gases. These rising emissions change the energy balance of the atmosphere, with resultant changes in the global hydrological cycle varying from an increase in water availability in some areas to a decline in water availability in others. Overall the variability and intensity of the hydrological cycle are predicted to increase with wide ranging impacts for aquatic ecosystems.

of change will occur at the seasonal scale within regions. Chou et al. (2013) used rainfall data to test this prediction in the Global Precipitation Climatology Project and found that, not only has the annual range of precipitation increased within regions, there were also trends for wet seasons to become wetter and dry seasons drier. Also, despite observed increases in rainfall and runoff in the Amazon basin (Gloor et al., 2013) during the same period, the region experienced two 1 in 100 year droughts (in 2005 and 2010) providing support for predictions that extreme conditions within regions may also increase (Lewis et al., 2011). Increasing rates of glacial melt and increases in the water use efficiency of vegetation associated with rising CO2 concentrations in the atmosphere will also affect global hydrological cycles. As temperatures rise, less winter rain falls as snow and the spring melt begins earlier, resulting in peak flows occurring earlier and declining throughout summer when demand for this water is highest (Barnett et al., 2005). Reports of glacial retreats are a widespread and well documented phenomenon (Scambos et al., 2004; Piao et al., 2010a; Chen et al., 2013). Rising CO2 increases the instantaneous water use efficiency of photosynthesis (Eamus and Jarvis, 1989) and observed increases in global runoff may be partially due to the increased water use efficiency of vegetation (Gedney et al., 2006; Betts et al., 2007). Piao et al. (2007) suggested that when the CO2 fertilisation effect is taken into account, global runoff should decrease. Considerable uncertainty exists as to how CO2 fertilisation will be manifested, however, recent empirical evidence suggests that vegetation is responding to rising CO2 concentrations. Donohue et al. (2013) demonstrated a greening of vegetation in arid and semi-arid regions and Keenan et al. (2013) found an increase in water use efficiency of vegetation in a meta-analysis of data from

eddy flux towers. Despite these observations, the hydrological impact of these changes remains difficult to quantify. Using the framework developed by Donohue et al. (2013) one of us (O'Grady) used a detailed growth model to predict that increases in runoff are more likely to occur in wet environments where the growth of vegetation and evapotranspiration are limited by incoming radiation (i.e. energylimited environments, E b P, where E = evapotranspiration and P = precipitation). 3. Australia as an exemplar study region The large areal extent of the Australian continent provides an opportunity to compare the impacts of both land use and hydrological intensification across multiple biomes because the continental landmass spans equatorial, tropical, arid and temperate climatic zones (Fig. 2). Generally, Australia is characterised by water-limited conditions with high climatic variability and soils that are low in nutrients and high in salts (Diamond, 2005). However, the Great Dividing Range in eastern Australia and almost two thirds of Tasmania contain pockets of energy-limited climates (i.e. regions where rainfall exceeds evaporation). On mainland Australia these energy-limited environments are the important water generating regions for Australia's inland river systems. For example, almost 50% of the flows into the Murray–Darling Basin are generated from 8.4% of the total basin area (Donohue et al., 2011; McVicar et al., 2012). Mean annual temperatures across Australia have increased by 0.9 °C on average since 1910. Patterns of rainfall have also shifted significantly over the instrumented record. The Australia-wide average annual rainfall has increased slightly, with large increases in annual rainfall in the


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Fig. 2. Major Australian climatic zones. Source: Bureau of Meteorology.

Fig. 3. Map of Australian resource use (2005–2006). Source: ABARES. Reprinted from Lesslie and Mewett (2013).



north-west and declines in autumn and early winter rainfall in the south-west and the south-east (CSIRO, 2014). Under future climate scenarios, the incidence of intense El Niño weather patterns is expected to double (Cai et al., 2014). The shifting rainfall distributions in Australia are consistent with observations of an expansion of the Hadley Cells (Cai et al., 2014). Recent modelling also suggests that, if these trends continue, they would result in changes to existing Koppen Geiger climate types (Fig. 2): the area of tropical climates in Australia would increase from approximately 8.8% to 9.1%, arid climates would increase from 76% to 81%, and that temperate and cool climates would decline from 14.7% to 9.1% and 0.0016% to 0.0001%, respectively (Crosbie et al., 2012). These changes are likely to result in declines in flows to systems such as the Murray–Darling Basin. Primary production (livestock grazing, dryland and irrigated agriculture) occurs over nearly 60% of the Australian continent. Areas reserved for nature conservation occupy 7% and other protected areas (mainly Indigenous Protected Areas, IPAs), cover approximately 13%. Forestry is confined to higher rainfall regions and covers approximately 2%. Intensive land use, mostly urbanisation, covers only 0.2%. Minimal land use occurs where deserts are present, an area comprising approximately 18% of the continent (Lesslie and Mewett, 2013) (Fig. 3). Over the longer term the impacts of increased climate variability and population pressure may strongly influence the location of agricultural and non-agricultural activities. Changes in agricultural activities are already emerging and there have been several significant changes in land use in recent years. For example, the area of grazing decreased by 6% between 1992–93 and 2005– 06 and over the same period the area of land used for cropping increased by 39% (Fig. 4) (Mewett et al., 2013). In Northern Australia, an estimated 20–40,000 ha of land currently used for grazing could be converted to irrigated agriculture. Nearly 80% of Australia's population (23.4 million in

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April, 2014) live in urban regions, with the greatest concentration in south-eastern Australia. Urbanisation is likely to continue, with many Australian cities displaying the highest urban growth rates (in 2013) in the developed world (www.infrastructure.gov.au). Australia's comprehensive governance of natural resources, including the role of State and Commonwealth government departments, also makes the continent an exemplar study region. Natural resource management includes integrated extension services and extensive onground land management networks. Additionally there is the capacity for listing both species and ecological communities as deserving of special protection at national (EPBC_Act, 1999) and State levels (e.g. Victoria's Flora and Fauna Guarantee Act). 4. Impacts of land cover and land use change on Australian freshwater ecosystems The changes in water quality, quantity and ecological health of Australian freshwater ecosystems resulting from changes in land use and land cover since European settlement have been profound (Boulton et al., 2014). These include sedimentation, eutrophication, salinisation, acidification, pollution (heavy metals and pesticides), altered flow regimes, degradation of urban streams, also known as the urban stream syndrome (Walsh et al., 2001), the loss and degradation of some types of aquatic habitats and ecosystems (notably shallow wetlands), species loss and increasing presence of invasive species (Fig. 5). A multi-decadal record of Australian wetland change has been extended in recent times using paleo-limnological techniques that focus on preserved indicators (e.g. diatoms, cladocerans) of wetland condition. These have been supplemented since the 1990s with other records (see Mills et al., 2013). In many instances these records, which

Fig. 4. Change in the area used for cropping between 1992–93 and 2005–06. Source: ABARES. Reprinted from Lesslie and Mewett (2013).


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Fig. 5. Land use change in Australia has resulted in: salinisation (top left: saline wetlands in the Western Australian wheatbelt); eutrophication (top right: an algal bloom in Herdsman Lake, Western Australia); sedimentation (bottom left: high turbidity downstream of a logging coupe in the Otway Ranges, Victoria); urban stream syndrome (bottom right: a channelised stream in Melbourne, Victoria).

provide a pre-European benchmark, reveal that much wetland change commenced soon after European settlement in the early 1800s. While the instrumental record shows significant impacts of climate, water resource development and contemporary primary industry, the palaeo-records suggest that these underestimate the total impact of natural resource development on Australian wetlands (Gell et al., 2013) Wetland salinity has increased in south-eastern Australia, either through decreased effective precipitation, historic water use (Gell et al., 2005), irrigated agriculture (Gell et al., 2007) or inter-basin transfers to support irrigation (Ghassemi and White, 2007; MacGregor et al., 2005). Salinity changes are mostly in the order of a ten-fold increase (Gell et al., 2009) although the Snowy River hydroelectric scheme diversion and floodplain clearance led to a 50-fold increase in the salinity of a downstream waterbody. The construction of barriers to secure freshwater for irrigated agriculture caused a shift from tidal, subsaline conditions to hypersaline conditions in a coastal lagoon (The Coorong) at the mouth of Australia's largest river system, the Murray–Darling. Widespread increases in the salinity of rivers and wetlands in the wheat belt region of south-western Australia have been created by landscape-scale clearing of deep-rooted vegetation resulting in rising saline water tables and mobilisation of salt previously stored deep within soil profiles (Davis et al., 2010). Widespread changes in the turbidity of Australian inland waters began with the increased flux of mineral matter soon after European settlement due to tree clearance, agriculture and goldmining activities. The impact of water regulation for agriculture and domestic consumption on turbidity and sedimentation has also been implicated through increased bank erosion. Some wetlands along the River Murray shifted from macrophyte dominance to phytoplankton dominance particularly after the advent of river regulation and the introduction of European carp to waterways. Carp contribute to poor water quality by uprooting vegetation and stirring up sediments during feeding (Koehn, 2004).

Sedimentation has led to the shallowing of some wetlands, allowing diverse macrophyte beds to develop (Gell et al., 2005) while other wetlands are undergoing a transition to dry land (Gell et al., 2009). Regulation of river flows, initially for navigation but ultimately to provide water for intensive irrigation, provides a refuge for carp in regulated rivers during low flow periods (Driver et al., 2005a) and has drowned seasonal, floodplain lakes, killed river red gum forests in near-channel areas while expanding suitable habitat for emergent species such as Typha (Gell, 2012) and the floating fern, Azolla. The uneven spatial distribution and extreme temporal variability of surface and groundwater supplies in Australia have led to the construction of a multitude of water regulation structures, including many thousands of weirs (3600 in the Murray–Darling Basin alone), locks and floodplain levee banks, 446 large dams (N 10 m crest height) and over 50 intra- and inter-basin water transfer schemes. The hydrological variability that typifies the natural flow regime of most Australian rivers (Finlayson and McMahon, 1988) and the water regimes of many wetlands have been altered across a range of temporal and spatial scales. Flow regulation is widely acknowledged to be a major cause of degradation in many Australian river and floodplain ecosystems (Bunn and Arthington, 2002; Cullen and Lake, 1995). Ecological changes in regulated river systems include massive loss of wetlands, decline of riparian forests, invasion of dewatered river channels and wetlands by vegetation, changes in aquatic plant community structure, population and diversity declines of invertebrates, fish and waterbirds, and several invertebrate extinctions (Arthington and Pusey, 2003; Kingsford, 2000). Additionally, a greater frequency and duration of drought in temporary wetlands (Driver et al., 2011; Casanova, 2012) along with an increase in the profitability of grain and oil-seed production (Tostovrsnik et al., 2010) has led to a change in land-use from biodiverse pasture grazing to monoculture cropping, with consequent threats to temporary wetlands as habitat for birds, frogs, plants and algae (EPBC_Act, 1999).


Impacts

Change in magnitude, duration & timing of flows Loss of groundwater dependant & shallow surface-fed ecosystems Loss of refugia Loss of floodplain productivity from reduced flows Eutrophication Acidification Pollution Loss of groundwater dependant & shallow surface-fed ecosystems Loss of groundwater-dependant ecosystems Salinization Metal pollution Loss of groundwater dependant & shallow surface-fed ecosystems Monitoring & management of water quality to ensure downstream receiving waters can support biodiverse ecosystems Restoration of riparian vegetation to minimise erosion

Impacts

Eutrophication & pollution Indirectly by alienation of floodplains by construction of levee banks to protect irrigated assets infrastructure Increasing prevalence of ‘urban stream syndrome’ Urbanisation

Increased erosion in inter-rotations Plantation forestry

Energy extraction – Coal seam gas

Erosion Sedimentation Indirectly by alienation of floodplains by construction of levee banks to protect irrigated assets Irrigated cropping, dairy and cattle feedlots

Manage silvicultural practices Increase streamside buffers Monitoring & management of water quality to ensure downstream receiving waters can support biodiverse ecosystems Use water sensitive urban design to reduce extent of impervious surfaces areas within catchments

Declining water tables may slow rate of salinisation Utilise extracted water to provide or supplement drought refugia

Solutions & opportunities

More efficient irrigation technology to reduce water use Declining water tables may slow rate of salinisation Limits placed on extraction from rivers Provision of environmental flows from water storages

Extended droughts

Solutions & opportunities Large floods

Currently agriculture accounts for approximately 70% of all consumptive water use in Australia, mostly for irrigation. Irrigated agriculture covers less than 1% of the total landmass, but produces approximately one third of Gross Farm Product (GFP) (Lake and Bond, 2007). In contrast, grazing and dryland cropping cover 53% and 6% of the land-surface, respectively (Lake and Bond, 2007). Both forms of agriculture have had profound impacts on river systems, both in terms of altering hydrologic regimes, and causing widespread erosion, sedimentation and salinisation. From a terrestrial view-point, much more attention has been given to the impacts of low-intensity farming, due to the effects of land-clearing for cropping and grazing, and the fact that such areas have a limited ability to support native biodiversity values (Cunningham and Duncan, 2012). Yet, from an aquatic perspective, it is often the perceived impacts of irrigated agriculture that receive the greatest attention because it is associated with very large volumes of water. It is thus interesting to consider whether a move towards more intensive (land-sparing) forms of irrigated production (together with

Hydrological intensification

6. Are some forms of land use intensification better than others?

Land use intensification

Land in Australia is being increasingly managed for multiple objectives, including the production of food, fibre, water supply, extraction of minerals and energy, biodiversity conservation, landscape amenity and in recent years, for carbon sequestration. Intensification of resource use is being driven by the need to increase productivity in response to growing demand from domestic and international markets (Lesslie and Mewett, 2013). Additionally, there is a declining resource base, in particular, arable land. In northern Australia, low soil fertility and large distances from domestic markets have not favoured intensive production. However, recent policy and investment documents have articulated the desire for northern Australia to become the ‘food bowl of Asia’ (northernaustralia.dpmc.gov.au). Subdivision of rural areas and the changing nature of rural areas along the edges of cities and towns (peri-urban growth) is a form of land use intensification that is becoming more common in some Australian regions. Examples of changes in rural land use include intensive animal raising, horticulture and viticulture, an increase in the number of small rural residential properties (hobby farms and lifestyle properties), and agricultural product manufacturing and processing. Potential impacts of rural land use development on freshwater ecosystems include increased competition for limited water supplies, lack of compliance with water regulations, building of dams, changes to water flows, changed riparian zone management, encroachment of rural dwellings onto areas of high conservation value, and impacts on groundwater availability and quality. In a broad-ranging analysis of potential future water use scenarios in Australia, Lake and Bond (2007) identified a suite of likely trends in terms of aquatic ecosystem condition across agricultural and urban landscapes. Their scenario of accelerated economic growth shares much in common with the intensification scenarios and associated impacts listed in Table 1, albeit examined through a much coarser lens. The most extreme hydrological scenarios are expressed as the occurrence of large floods and extended droughts. The scenarios of wet areas becoming wetter and dry areas becoming drier will result in similar but more chronic impacts on freshwater ecosystems relative to those outlined by Lake and Bond (2007). Often multiple solutions exist to address intensification impacts, and in some cases, opportunities to ameliorate existing water quality and quantity issues may become evident. The temporal differences between the impacts of large floods, which are instantaneous, and extended droughts that can play out over many years, require a paradigm shift in management approaches and systems of governance towards ones that embrace temporal variability and ecological dynamics (Bond et al., 2008).

Table 1 Predicted impacts of the combined effects of land use intensification and hydrological intensification (large floods and extended droughts) on freshwater ecosystems, potential solutions and opportunities.

5. Predicted effects of land use intensification and hydrological intensification on Australian freshwater ecosystems

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Design & manage perennial waterbodies fed by urban stormwater to maximise their value as aquatic refugia



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some agricultural abandonment and landscape restoration), would achieve greater net biodiversity outcomes compared to mixed ecoagricultural landscapes, despite heightened degradation of some landareas and associated river-corridors used to supply irrigation water. Views on this intensification debate are mixed (Benton, 2012; Phalan et al., 2011; Scherr and McNeely, 2008). For example, Benton (2012) suggested that “extensive and intensive farming may each be the best option for food and wildlife depending on the place. The design of the optimal landscape in terms of land sparing vs. land sharing will depend on the costs (in terms of change in yield) vs. benefits (in terms of biodiversity or ecosystem processes): in low production, high biodiversity landscapes, land sharing may be best, and vice versa.” An important factor is whether the effects of increased chemical use often associated with intensive agriculture can be controlled, and whether the longterm productivity of high intensity production can be maintained (e.g. in the face of crop pests). These sorts of comparative analyses in relation to aquatic biodiversity have yet to be undertaken systematically. For example, while there have been attempts to integrate changes in terrestrial biodiversity and production values with impacts on runoff (Vertessy et al., 2003), rarely have the production/biodiversity values and trade-offs from dryland vs irrigated agriculture been considered simultaneously. An important consideration for river systems, compared to terrestrial ecosystems, is their dendritic network geometry, which leads to comparatively more habitat (by river length) occurring in upland tributary streams, which are often comparatively unaffected by extensive flow regulation for irrigation, but which tend to be very strongly influenced by surrounding land use (often relatively marginal grazing). These low order streams often support high species richness and high levels of endemism and contribute to landscape-scale biodiversity patterns in important ways (Clarke et al., 2008). In contrast, the areas most heavily affected by flow regulation for irrigation are lowland floodplain habitats. It is these contrasting patterns of flow influence between uplands and lowlands, coupled with the large differences in economic returns from surrounding land-use, that creates new challenges (and possibly opportunities) for future restoration. 7. Management responses to the impacts of land use and hydrological intensification on freshwater ecosystems 7.1. Delivering environmental flows Various mechanisms are already in place in many Australian regions to protect the environmental water share or deliver environmental flows during extended dry periods or droughts. Environmental flows are defined as “the quantity, quality and timing of water flows required to sustain freshwater and estuarine ecosystems and the human livelihoods and well-being that depend on these ecosystems” (Brisbane Declaration, 2007). Opportunities include limiting the volume of water pumped from a river over the course of a year (sustainable diversion limits, SDLs), constraining the time of day of pumping (setting ‘cease to pump’ or CTP levels) and creating a list of high priority wetland and river habitats (typically pools) to guide environmental water allocations. Such rules are reviewed and adjusted to take account of the critical water requirements of riverine biota during drought. SDLs are estimates of the sustainable level of take using hydrology models that run for ~ 100 years applied to a whole valley (MDBA, 2010). The SDLs are assessed and audited annually. Many river pools across Australia are not gauged, so part of the water management challenge is to determine the ecological values of river pools and allocate gauging resources to those deemed to be most important. SDLs, if set appropriately, will protect all environmental assets and the water-dependent industries and towns (MDBA, 2010). CTP rules are designed for the protection of individual river pools. The rules to protect floodplain assets and functions are quite different and although Commence to Flow (CTF) thresholds have been determined to describe the heights at which wetlands require water, the process of describing the flow regime requirements

and then protecting these regimes within water planning arrangements is much more involved. The current emphasis under the Murray–Darling Basin Plan (the water management plan for the largest river basin in Australia) is to describe the environmental assets within each valley, including river pools and wetlands. Basin states are required to develop one- and five-year environmental watering plans within which assets and their environmental flow requirements are identified to ensure balanced use of environmental water across these assets. Most states already have tools to protect river assets. For example, water planners in New South Wales use desktop tools and field-based information, including hydrology, geomorphology, ecology and local knowledge, to identify and prioritise the delivery of water to specific assets (Driver et al., 2012, 2013). This approach offers a level of local detail above setting generic flow requirements for reaches and rivers using relationships between hydrology, geomorphology and ecology as described by Poff et al. (2010), but it is less detailed and more time efficient than a dedicated research programme. Sufficient knowledge of water distribution patterns and spatial arrangements makes it possible to modify the pattern of water abstraction so that particular streams or river reaches of high conservation value continue to receive some flow on a preferential basis (Bond et al., 2008). Such opportunities arise by taking a more flexible approach to the approval, uptake and use of water licences. This may include suspending water licences, or trading licences among properties, to provide some streams with respite during droughts. In other cases a “contingency flow” can be held in storage and used to provide carefully timed flow pulses to generate small floods similar to natural events (Driver et al., 2005b; Bond et al., 2008). A modest volume of flow at the right time of year can result in beneficial ecological outcomes (e.g. seedling and waterbird recruitment, fish spawning, water quality maintenance), sufficient to protect or sustain individual species or assemblages until the drought breaks (Bond et al., 2008). However, successful implementation of contingency flows requires knowledge of the habitat requirements, life history patterns and recruitment strategies of valued species (Bunn and Arthington, 2002). 7.2. Protecting drought refugia While careful planning and management should ensure some water remains in storage, short-term mitigation strategies that protect specific habitats are needed in cases where storage volumes fall to critically low levels, such that the desired environmental flows cannot be delivered. During drought there will be many situations where the best environmental flow strategy is one that will protect refuge habitats and their biota. Refugial habitats exist across a broad spectrum of aquatic ecosystem types (Davis et al., 2013; Robson et al., 2008). Riverine waterholes and floodplain lagoons are extremely valuable refugia in arid systems (Arthington et al., 2005; Bond et al., 2008) but other types of refuges have also been identified. They include logs, wet patches under banks, riffles, sub-surface stream sediments, yabby holes and stands of littoral and riparian vegetation (Boulton, 2003). These remnant habitats can help to sustain moisture during dry spells and drought, and may support obligate aquatic species for considerable periods of time. The identification and protection of diverse aquatic refugia should be given the highest priority at all times but especially during drought (Sheldon et al., 2010). These refugia require protective actions such as maintaining riparian vegetation through the provision of environmental flows and protection of river banks from livestock and other sources of rural and human disturbance (Davies, 2010). Seed banks also act as refugia. In many temporary wetlands plant, algal and invertebrate populations re-establish when inundated, through germination or hatching from a bank of desiccation-resistant seeds, eggs or propagules in the soil. This bank of propagules conveys a degree of resilience in temporary wetlands (Casanova, 2012), and many species can remain dormant in the soil for years or decades (Brock, 2011).



Recognition of seed/propagule banks as cryptic refugia, and an understanding of how they can be maintained, is essential for conservation of temporary wetland systems (and riparian zones) and their biota. 7.3. Other strategies, potential pitfalls and unintended consequences Management interventions such as drought refuge protection and environmental flow provisions may fail to provide sufficient protection for populations of rare and threatened species. For rare taxa or those represented by few populations, further losses of genetic diversity caused by local population loss have the potential to greatly increase the risk of extinctions (Bond et al., 2008). It may become necessary to implement targeted population management strategies such as the transfer of threatened populations into captivity for later reintroductions (e.g. Galaxias fuscus populations in Victoria), or the translocation of individuals from other areas once drought has abated. Reintroduction and translocation strategies carry significant risk: they can be costly, and there is potential to compromise natural population genetic structure or transfer diseases among isolated populations (Cunningham, 1996; Hughes et al., 2003; Olden et al., 2011). Identification of the situations where these sorts of interventions are needed may also highlight species and populations likely to be threatened during future droughts and changes in climate (Bond et al., 2008). The potential impact of inter-basin water transfers on the habitats and populations of normally isolated catchments is also of concern (Ghassemi and White, 2007). As well as moving water across the landscape, such schemes can transfer native species, thereby disrupting patterns of genetic isolation, and may also move alien species and pathogens between catchments (Hughes et al., 2003). Various strategies, such as screens to detain propagules and water treatment, have the capacity to limit biological inter-basin transfers to some degree, although they are not always economically cost-effective and feasible, and are frequently biologically ineffective (Ghassemi and White, 2007). As rainfall declines and temperatures rise during summer in southeastern Australia, there is an increasing, regular, dependence on ground water for stock, rather than a fall-back resource for use during extended droughts. The use of groundwater is incremental, but widespread, and the last decade has seen an intensification of the use of groundwater for agriculture in many regions (Tostovrsnik et al., 2010). Recent decades have seen an increase in the application of management interventions to meet freshwater ecosystem conservation and water resource objectives, especially revegetation of riparian zones. Climate change is likely to further intensify such investments and onground action. For instance, the need to protect in-stream biota from thermal stress has led to concepts of ‘over-restoration’ in which exotic fast-growing, shade providing trees are used in revegetation projects (Davies, 2010). However, the planting of exotic species may entail risks to riparian biodiversity as well as other functions, e.g. contribution of riparian materials to aquatic food webs and, where watershed reafforestation is widespread, the associated increase in evapotranspiration will result in large reductions in stream flows (Davies, 2010). Although it is unrealistic to expect human-occupied landscapes to be returned to any pre-European natural state (Hilderbrand et al., 2005), the potential for maladaptation to climate change from current management interventions is evident with recent investment in engineering works and measures for diverting water across floodplains being erroneously justified as a surrogate measure for natural flooding (Pittock and Finlayson, 2013). 8. The importance of improved governance in supporting freshwater ecosystems facing intensification Managing land use and hydrological intensification and their impacts is a human endeavour. Hence, consistent with Pahl-Wostl et al. (2013), contemporary intensification challenges facing the world will enhance scrutiny on governance systems from global to local scales.

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Consequently, building healthy governance systems that address the causes and impacts of intensification are needed to create the foundations to establish a more sustainable future for freshwater ecosystems. Dale et al. (2–13) cite Parker and Braithwaite (2003: 119) in defining a systems view of governance as the “intentional shaping of the flow of events so as to realize desired public good”. They cite several authors in further defining a systems view of governance as: • being distinguished from, but inclusive of, the processes of government (e.g. Thomas and Grindle, 1990): • representing a large set of processes of bargaining and negotiation among differing societal interests, leading to particular system outcomes (e.g. Dorcey, 1986); • being polycentric across different scales (e.g. Ostrom, 2008); and • being characterised by the use of both structural and functional concepts from the sociological and planning literature (e.g. Dale and Bellamy, 1998).

Below, we broadly apply the Governance Systems Analysis (GSA) approach developed by Dale et al. (2013a) for the interrogation of complex natural resource governance systems. Consistent with a very wide policy and planning literature, GSA considers the standard structural elements of natural resource decision-making and action at different scales needed to include vision and objective setting, analysis, strategy development, implementation, monitoring, evaluation and review (see also Potts et al., 2014). Importantly GSA also considers how well things operate within and across these structural elements. Sound strategy development, for example, can be undone by poor implementation. GSA also considers three cornerstone functional elements of healthy governance systems, consistent with many authors who have previously explored the governance through socio-ecological system concepts (e.g. consider Folke et al., 2007). These include connectivity across the system, institutional capacity within the system and the access to and use of knowledge (of various kinds) in decision-making within the system (see also Bellamy et al, 2001; Dale and Bellamy, 1998; Anderies et al., 2004). Broadly applying this analytical approach, we consider that Australia had, until recently, a relatively robust framework for the allocation of water resources through the auspices of the National Water Initiative (NWI). Important structural and functional weaknesses, however, do remain across the system, and some significant aspects of this system may weaken further with the recent decline of the NWI and the National Water Commission (cited online 2 March 2015 at http://www.nwc.gov. au/organisation/closure-in-2014). The NWI represented a nationallyfocussed: “intergovernmental agreement set out to achieve a nationally compatible market, regulatory and planning based system—one that manages surface and groundwater resources for rural and urban use, and optimises economic, social and environmental outcomes” (cited online 2 March 2015 at http://www.nwc.gov.au/nwi). The level of bilateralism in the governance of Australian freshwater systems is likely to weaken as a result of the demise of the National Water Commission. An additional systemic weakness in governance, at regional and local scales, is the often low levels of connection between water and land use planning and management processes. As a key aspect of good water governance, the concept of integrated land and water planning and management, has been supported in principle and in practice within various levels of both the Australian federal and State/Territory governments for some time. Despite commitment to this principle, in practice, institutional barriers still remain between organisations that manage riverine and non-riverine landscapes (Martin and Williams, 2013). Australia also has developed a cohesive but sometimes unstable system for setting water quality standards. Until 2007, for example,


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Australia was rapidly setting a robust vision for water quality management, negotiated bilaterally across all States and Territories through the National Action Plan for Salinity and Water Quality (NAPSQW). This framework was replaced by the Caring for Our Country Program in 2007, and the national water quality policy vision changed from an outcomes-based to a more limited outputs-based approach. The approach also became disconnected from any sense of national environment accounts (e.g. via the work undertaken by the National Land and Water Audit; Robins and Dovers, 2007; Wentworth Group, 2008). Additionally, while NWI-related delivery frameworks dealing with water allocation are well developed and quite mature in the south of the country (e.g., in New South Wales, Driver et al., 2013), much effort is still required to reverse the impacts of past periods of strong intensification. In the north, however, where new intensification pressures are emerging, there is not a particularly strong framework for the implementation of water resource plans (see JCU and CSIRO, 2013). Across the nation, water quality considerations were historically below capacity in regional areas (NWC, 2011) and often poorly integrated within water quantity-based allocation processes under the NWI framework. In broader functional terms, articulation of natural resource management problems must be directly connected to public values. This provides the essential foundation for ensuring adoption of management strategies and adaptive system management (Groffman et al., 2010; Folke et al, 2007). Water management stakeholder consultation can sometimes be too narrowly defined, often with an emphasis on the most obvious consumers, for example, irrigators (Bari et al., 2008; Hassall and Associates, 2007), and with an emphasis on the disciplines of hydrology and economics, instead of considering other aspects of human communities, especially the needs of Indigenous peoples (Jackson and Morrison, 2007; Martin and Williams, 2013). Alternatively, conservation objectives might be considered to the exclusion of social, economic and cultural values of water (Dale et al., 2013a). Hence, the full range of water managers and water users must be engaged to better understand the needs and uncertainties in natural resource decision-making, and Indigenous values and knowledge must incorporated (Fazey et al., 2006; Finlayson, 2001; Richter et al., 2006; Ryder et al., 2010). The incorporation of Indigenous values and knowledge in water planning and management will also improve resource management plans by incorporating species and season-specific knowledge (Finlayson, 2005; Woodward et al., 2012). In addition to the inclusion of broader human values, effective decision-making requires an understanding of long-term ecosystem behaviour (e.g. see Folke et al., 2007). A process to retain, maintain and synthesise management intervention knowledge across Australia's longterm monitoring programmes is urgently needed (Driver et al., 2013; Stewardson et al., 2012). At the national scale, frameworks for monitoring water quantity and quality, and water management have been strengthened, but also partly discontinued. In areas outside of the jurisdiction of the Murray Darling Basin Plan, which largely incorporates the intent of these former programmes (the NAPSQW and NWI, as discussed), there is an even greater need for an integrated approach to intervention monitoring across state and commonwealth government agencies. Periodic changes in national and State/Territory policy and associated management intervention require re-negotiation and a lack of effective bilateralism has been a key impediment to effective change in Australia since Federation. That is, each new era of governance is weakened by needing to re-establish systems across the Commonwealth and States/Territories post changes in governments at either the Commonwealth or State levels. Hence, with major changes such as intensification in Australia's north and in peri-urban zones fringing coastal cities, and also temporal shifts in ecosystem behaviour (more frequent prolonged droughts), there is a high risk of delayed adaptation as divides emerge between national and state activities relating to water monitoring and effective policy and investment. This also means that monitoring results may have less influence on key national and state policies (Wentworth Group, 2008).

A further contribution to such delayed adaptation may emerge from the recent decline in strategic national scale planning and allocation of resources for research in Australian water resources management. Research and development activities have become fragmented and more disconnected from planning and implementation as scientists seek funding opportunistically rather than working within programmes with clearly articulated national public good goals and outcomes (see Dale et al., 2015). Although Australian universities and the peak Australian government-funded research organisation, the Commonwealth Scientific and Industrial Research Organisation (CSIRO), retain a strong research and development capacity in qualitative and quantitative water resource science, governance and management, this capacity is likely to erode if national funding support continues to decline (Dale et al., 2013b). 9. Prioritising research to integrate land use intensification into freshwater conservation and management 9.1. Developing an understanding of long term ecosystem behaviour to support decision making Despite a long history of palaeoecological research across Australia, there has been very little integration of this evidence into natural resource decision-making (Mills et al., 2013). Evidence from natural archives of change, such as wetland sediments, tree rings and speleothems can provide considerable insight into systems change including, but not limited to: (i) benchmark conditions; (ii) initiation of human impact; (iii) the magnitude of human impacts relative to benchmarks; (iv) recent trajectories of change; (v) linear vs non-linear responses, and (vi) novelty of present condition. Syntheses of palaeolimnological records across Australia conclude that virtually no wetland escaped the combined effect of human impact and climate change over the last two centuries (Gell et al., 2013). So many changes have occurred to Australian ecosystems, including aquatic ecosystems, it seems likely that society will greatly underestimate the impact of past land use intensification, and so be ill-prepared for the consequences of any future intensification. The integration of long term records of change, with long term monitoring programmes, is critical for Australian researchers and managers to understand the impact of intensive resource use on the Australian continent. 9.2. Understanding the capacity of extant freshwater biodiversity to adapt to environmental change The capacity of freshwater biodiversity to adapt in situ to environmental change (e.g. through acclimation, genetic or epigenetic changes, behavioural change or plasticity), and climate change particularly, is poorly understood. Understanding the potential for adaptation through movement, including range shifts and migration is needed to inform the development of freshwater biodiversity climate change adaptation strategies. The identification of taxa with high adaptive capacities is needed for prioritisation of management efforts towards more vulnerable populations, species and communities. Understanding the mechanisms that underpin the resilience that freshwater ecosystems currently possess could allow us to support that capacity. 9.3. Determining the relative contributions of surface water and groundwater in supporting freshwater ecosystems There remains insufficient knowledge of Groundwater Dependent Ecosystems (GDEs) and their water requirements (Boulton et al., 2010). Although there has been a shift towards GDE preservation within planning, GDE protection often uses limited definitions and perceptions of what constitutes a groundwater-dependent ecosystem. Often the focus is on vegetation-dependent groundwater systems that can be readily identified using remote sensing techniques. However, these techniques remain poorly ground-truthed and empirical studies of



terrestrial groundwater use in Australia remain rare (O'Grady et al., 2010, 2011). The focus on remotely sensed metrics has reduced the scope of work pertaining to GDE identification, resulting in limited understanding of long-term GDE characteristics and viability (Eamus and Froend, 2006). This leads especially to neglect of GDEs that are difficult to detect and monitor (such as hyporheic and stygofauna communities in karsts and alluvial systems) in planning processes (Tomlinson and Boulton, 2008). The impacts of managed aquifer recharge on subterranean aquifers, to improve water security in cities located on subterranean aquifers are not known. Changes in land use, such as shifts from annual to perennial systems with climatic drying, are likely to have significant impacts on groundwater recharge (Crosbie et al., 2010). Enhanced effort and funding are required to support both research and planning to facilitate integrated management of surface and groundwater ecosystems (Eamus and Froend, 2006; Tomlinson and Boulton, 2008). 9.4. Improving capacity to model hydro-ecological relationships and predict ecological outcomes from environmental flows Effective management of environmental flows under different intensification scenarios will depend on a platform of quantitative hydro-ecological relationships, preferably for rivers and wetlands of different hydrological and ecological character (Poff et al., 2010). Hypothesis-based experimental investigations of species habitat requirements, life history strategies and ecological processes are needed to strengthen the existing science. They can be conducted in laboratories or greenhouses or as field experiments using existing flow regimes and flow facets as experimental treatments (Davies et al., 2014). A better understanding of the importance of freshwater inflows into estuaries and near-shore marine regions, is also needed. Estuaries and nearshore marine regions are important habitats, supporting the life cycles of many freshwater, estuarine and marine fishes and other species, including commercial species. 9.5. Identifying freshwater biodiversity hotspots and aquatic refugia Identifying aquatic refugia and barriers to dispersal and reestablishing connectivity within riverine networks are particularly important, especially in semi-arid and arid regions (Davis et al., 2013; Davies, 2010). Establishing priorities for protection of important refugia or connectivity pathways should involve principles of systematic conservation planning (comprehensiveness, adequacy, representativeness and efficiency) at catchment and broad jurisdictional or bioregional scales (Arthington, 2012). Priority habitats and waterbodies (e.g. evolutionary and ecological refugia) can then be designated and managed as protected areas, or for mitigation of threats, or to establish priorities for environmental flow delivery (Arthington, 2012). Although restoration of flow regimes to pre-disturbance conditions may be impossible in many areas, environmental flows can be used strategically to meet conservation obligations for endangered species, habitats and internationally important wetlands listed under the Ramsar Convention and migratory bird agreements (JAMBA, CAMBA). 9.6. Understanding the impacts of multiple stressors The disappearance of some freshwater species may be driven by a single stressor, yet for others the cumulative effects of multiple stressors will be detrimental. Some species will be more affected by climatedriven stressors and others by land-use stressors. Some stressors will create additive or synergistic effects, and others antagonistic effects. These effects are likely to differ for different types of waterbodies, e.g. systems with large water volumes and diverse food webs may have higher resilience than simple systems with little buffering capacity. Understanding the complexity, interactions, and hierarchy of stressor

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effects on communities is needed in order to manage them effectively (Ormerod et al., 2010). 9.7. Developing clear links between freshwater ecosystem functioning and human well-being As environmental policy and management become more utilitarian, conservation and natural resources management efforts are increasingly directed towards outcomes that can be clearly linked to human wellbeing and health based on the provision of ecosystem services (Mitsch and Gosselink, 2000; Acreman et al., 2014). Governments across the world have a role in maintaining healthy ecosystems and in mitigating climate effects under the Rio Convention, but have not yet delivered on key agreements (Tollefson and Gilbert, 2012). Describing the importance of healthy ecosystems to food security and as a requirement in meeting our international obligations is critical and also helps build the case for integrated land and water management. For example, IUCN (2013) describe how healthy ecosystems benefit food security by enabling availability, access, utilization and stability. Research is needed to more fully describe and quantitatively model a greater range of interactions between freshwater ecosystem structure and function and the benefits provided beyond those which support the provision of goods with clear monetary value (i.e. potable water and fishes) (Gordon et al., 2010). In this respect aquatic ecosystems can be seen as settings for human well-being and livelihoods, although paradoxical situations also need to be identified and assessed (Horwitz and Finlayson, 2011). 9.8. Trialling and evaluating improved governance systems for managing the impact of intensification on freshwater systems Using Australia as an example, it is evident that some aspects of the country's water resource governance system provide benchmarks of international significance (e.g. water allocation under the National Water Initiative) while other aspects of the system are particularly weak (e.g. water quality management from diffuse sources). Australia's cohesive system for governing water allocation is not well suited to, nor has it been applied effectively under, northern Australian conditions where new land use developments are planned. Much work is still needed to improve the national system of water governance as the impacts of intensification play out. While Australia has had some significant successes in improved governance with respect to both water quality and quantity planning and management, continuous improvement in the design and implementation of governance systems is needed. This makes the recent abolition of the nation's National Water Commission, for example, a retrograde development as it diminishes the likelihood of strong national leadership and deliberative bilateralism between the Commonwealth and the States in water management. It also makes growing integration between water quality and quantity aspects of planning less likely. Future initiatives should also be cognisant of the impact and influence of allied governance systems, for example, those for climate change mitigation and adaption. While Australia has had some significant successes in improved governance with respect to both water quality and quantity planning and management, continuous improvement in the processes of adaptive management is an ongoing challenge. Reinvigorating the national policy and coordination framework for water management may be critical to avoid some of the worst impacts of intensification. 9.9. Integrating aquatic and terrestrial ecological research and management Freshwater research and management is often isolated (and in some cases marginalised) from terrestrial ecological practice. With respect to governance, the high socio-economic value and frequently contested nature of water resources have resulted in the establishment of


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institutions that manage water and freshwater ecosystems in isolation from their broader landscape settings. Integrated catchment management is an exception, although its aims and objectives are also typically water related (e.g. end of system water quality) rather than being truly holistic at a landscape scale. One result of this disciplinary and institutional segregation is that hydrological change has been widely considered as the sole driver of ecological degradation in Australian aquatic systems, particularly during periods of drought. Addressing land use intensification and other non-hydrological intensification trends, however, requires a broader landscape perspective to be taken. The recent, changed emphasis in public funding of improved natural resource management illustrates the challenge with land-water segregation. Government programmes were largely based on a case-by-case funding assessment without properly linking expenditure to meaningful, standardised performance measures (Hajkowicz, 2009). This general problem is now compounded by a move towards funding based on calculated carbon-offset benefits for agricultural land management, which in turn are affected by a market-driven price of carbon. However, it is unlikely these measures will lead to significant carbon storage (Sanderman et al., 2010), and the consequences for aquatic systems are even less clear. Understanding aquatic ecosystem structure and function and managing these for both their intrinsic value and their human benefits in the future will depend on effective collaborative and transdisciplinary research and management across aquatic and terrestrial fields (Likens et al., 2009). Scientific disciplines and institutions in both arenas stand to benefit.

riverbank zones; provision and maintenance of refuge habits; provision of environmental flows in systems where much water is extracted; fishpasses to allow movement of fish around constructed barriers, and the identification and active conservation of valuable ecosystems and particular biota. Strategic research priorities to inform planning and management of freshwater systems include: identifying freshwater biodiversity hotspots and refugia; improving the capacity to model hydroecological and agro-ecological relationships; and developing a better understanding of surface water–groundwater interactions, long term ecosystem behaviour and aquatic-terrestrial links. Acknowledgements This work was funded by the Department of Industry, Innovation, Science, Research and Tertiary Education (DIISRTE) via the Northern Futures Collaborative Research Network Program (CRN), the Australian Government's Stream 2 Climate Adaptation Program and ACEAS, the Australian Centre for Ecological Analysis and Synthesis. ACEAS is a facility of the Australian Government-funded Terrestrial Ecosystem Research Network (www.tern.gov.au), a research infrastructure facility established under the National Collaborative Research Infrastructure Strategy and Education Infrastructure Fund Super Science Initiative. We also acknowledge co-investment from the Australian Research Council (ARC). The contribution of Ian Kidd in developing the reference library for this paper is also gratefully acknowledged. The views expressed in this paper are those of the authors, and not necessarily those of any Australian government agency or other organisation.

10. Conclusions Land has been cleared across most of the temperate regions of Australia and large amounts of fresh water have been diverted and extracted for the production of food and fibre, resulting in major land use and water use legacies. These must be addressed if the nation is to avoid the loss of multiple ecosystem services and an accelerated decline in aquatic biodiversity. Although adaptive management processes are in place, much of the inland aquatic landscape is affected by humans; often in non-quantified ways. Effective governance, across all nations, requires better understanding of the interacting impacts of climate change, hydrological intensification and land use intensification on social and ecological systems. In particular, a greater understanding is needed of the impacts on all communities: rural, indigenous and urban. The ‘hidden’ nature of groundwater means that groundwaterdependent ecosystems remain particularly vulnerable. Many trends, including biodiversity decline, wetland loss, salinisation, eutrophication and sedimentation are well recognised. However, remediation and restoration of affected systems often remain problematic. In the Australian context, land and water governance in southern Australian is well developed by global standards but evidence-based and participatory planning and management is needed to ensure that past mistakes are not repeated in the agricultural developments of Australia's north. The high likelihood of extreme events (more protracted droughts and more severe floods) affecting regions of intensive food and fibre production globally provides a compelling argument for improving the links between land and water governance. If land use intensification is to avoid serious social, economic and environmental costs, approaches that empower local people to manage their own landscapes will be needed, based on inclusiveness, negotiation, and flexibility, to ensure support for incomes and livelihoods (McCartney et al., 2014). Collaborative research enabled by institutions such as the Terrestrial Ecosystem Research Network (TERN) and Cooperative Research Centres (CRCs) is important because it creates data sharing and motivational frameworks that otherwise would not exist (e.g., for the Australian Centre of Ecological Analysis and Synthesis as part of TERN, see Specht et al., in press). Proactive strategies are needed to minimise the impacts of floods and droughts. These must include restoration in catchments and

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