Wildlife Restoration

Biological Conservation 191 (2015) 830–838

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Discussion

Wildlife restoration: Mainstreaming translocations to keep common species common David M. Watson ⁎, Maggie J. Watson Institute for Land, Water and Society, Charles Sturt University, Australia

a r t i c l e

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Article history: Received 5 June 2015 Received in revised form 13 August 2015 Accepted 24 August 2015 Available online 14 September 2015 Keywords: Reintroduction Translocation Novel ecosystems Landscape ecology Environmental planning Climate change adaptation

a b s t r a c t In most urban and agricultural landscapes, remnants of native vegetation are surrounded by an inhospitable matrix. Although vagile species come and go, many reptiles, amphibians and small mammals are effectively stranded and declining towards local extinction. In the same landscapes, other areas where these species are absent are improving in habitat quality, both through natural regeneration and active restoration efforts. So, for many species in many domesticated landscapes, there are too many individuals in some patches of decreasing quality and no individuals in patches of increasing quality. One solution to this situation is to move animals from those areas where there are plenty to suitable areas where there are none. These targeted translocations apply lessons learned from revegetation to dispersal-limited animals to in-fill distributional ranges, increase population size and improve both demographic and genetic connectivity, pushing nonequilibrial metapopulations away from extinction via an imposed mass effect. In contrast to conventional reintroduction schemes—expensive, reactive interventions involving highly-trained specialists and captive-raised endangered species—these inexpensive, proactive, community-driven initiatives aim to avert future declines by keeping common species common. Having introduced the wildlife restoration vision, we use two scenarios to illustrate the benefits of the approach—to species, ecosystem function, ecological understanding, restoration practise and public engagement. As well as adhering to best-practise reintroduction techniques to ensure animal welfare is not compromised and avoid detrimental effects to source populations or release sites, we emphasize community participation, data quality and long-term accessibility as paramount to maximize learning opportunities. © 2015 Elsevier Ltd. All rights reserved.

1. Introduction Habitat loss, fragmentation and land-use intensification have taken a heavy toll on most ecosystems—species lost, processes disrupted and functions diminished. As global climate systems are further disrupted, the human population continues to grow and production agriculture accounts for an ever increasing proportion of primary productivity (Watson et al., 2013), these pressures will be exacerbated and ecosystems pushed progressively further into no-analogue conditions (Williams and Jackson, 2007). Despite the swelling literature on range-shifts (see Patiño and Vanderpoorten, 2015), microrefugia (Sublette Mosblech et al., 2011) and assisted colonization (Gallagher et al., 2015 and references therein), our basic understanding of the microclimatic tolerances, microhabitat requirements and other fine-scale distributional determinants remain incomplete for most species in most systems (Lunt et al., 2013), precluding predictions (Clark et al., 2003) and severely constraining on-ground interventions to avert further losses. Despite these knowledge gaps, extensive areas of many ecosystems have been revegetated, these actions ameliorating many threatening processes while improving our understanding of how to minimize ⁎ Corresponding author. E-mail address: [email protected] (D.M. Watson).

http://dx.doi.org/10.1016/j.biocon.2015.08.035 0006-3207/© 2015 Elsevier Ltd. All rights reserved.

further disruption from accelerating climate change (Maschinski and Haskins, 2012; Fischman et al., 2014). To date, however, restoration efforts and practitioners focus almost exclusively on plants (Clewel and Aronson, 2013) and, when animals are considered, only rare or endangered species are targeted. Consequently, restored ecosystems typically comprise revegetated rural and urban landscapes that are either depauperate, or overrun with invasive or weedy animal species (Munyenyembe et al., 1989; McKinney, 2008). Although supporting an increasingly complex set of plants and recreated habitats, many urban and agricultural areas contain increasingly simplified animal communities (Werner, 2011). In addition to suboptimal ecological outcomes, these practises largely ignore people living and working in these areas (Miller and Hobbs, 2002) —communities considered to be increasingly disconnected from nature (Colding, 2011). Here, we advocate incorporating translocations of dispersal-limited animals into mainstream environmental management, focusing on two classes of domesticated landscapes: urban and agricultural areas. Rather than being the sole prevail of rare species or protected areas, we see animal translocations to be philosophically and practically comparable to revegetation—useful interventions when natural recolonization is too slow and/or too patchy. By seeding highly modified landscapes with recruits of locally-extirpated animals, wildlife restoration is a proactive scheme to maintain ecosystem function and avert

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future declines while providing opportunities for local people to engage with nature, thereby building environmental literacy in the wider community. Having articulated our vision and considered how it complements existing theoretical and practical approaches, we consider candidate species and landscapes best suited to wildlife restoration, and population genetic, implementation and policy factors. We use two illustrative scenarios to showcase how wildlife restoration could be applied to landscape-scale conservation and close with five principles to guide further discussion. Although relevant to policy formulation, environmental planning and restoration practise, this perspective is intended primarily for conservation biologists, challenging them to re-imagine the domesticated landscapes where we live and work as vanguards for ecological restoration. 2. Wildlife restoration Originally involving the planting of non-native species for erosion control and aesthetic improvement (Burton and Burton, 2002), revegetation has become the main tool for restoration and remediation. This shift in motivation is reflected in advanced techniques to harvest seed, grow seedlings and maximize establishment success (of rare and common species; ruderal and climax communities alike), all guided by clearly-articulated objectives to effect lasting improvements to ecological processes and function (Burton and Burton, 2002; Maschinski and Haskins, 2012). By striving to emulate site-specific successional dynamics, current restoration efforts and practitioners rarely consider animals (Clewel and Aronson, 2013), operating under the tacit hypothesis that as habitat structure and resource availability increase, animals will progressively return (Palmer et al., 1997). Empirical tests of this hypothesis have found more mobile species are more likely to recolonise, but many

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animals are either unable or unwilling to navigate less hospitable intervening areas (Watson, 2010; Craig et al., 2012; Vergnes et al., 2013). To illustrate how metapopulation dynamics can affect animals in highly-modified landscapes, imagine a checkerboard, shaded squares denoting patches where a species is present, white where it is absent (Fig. 1). Rather than being static through time, occupied and unoccupied habitats may change, especially for vagile species for which patches blink on and off across a landscape. For sedentary species or taxa of limited vagility, patterns of occupancy may be more stable, but the trajectory of individual populations will vary—some populations growing, others declining. For those patches surrounded by an inhospitable matrix, recruits will be unable to leave the patch, often leading to crowding and lower resource availability, diminishing habitat quality over time. So, rather than uniform, occupied cells may change shade as population trajectories rise and fall (Fig. 1). But many white cells—those unoccupied areas—also vary. Through time, an increasing proportion of unoccupied patches may contain all the resources required by a species—all that is missing are recruits. Where habitats are less fragmented and functional connectivity is high, animals eventually locate these unoccupied areas (Huxel and Hastings, 1999). But, for highly-fragmented systems and dispersal-limited animals, the likelihood of recruits colonizing unoccupied patches is low and, for many woodland- and forest-dependent species, that probability is very close to zero (Craig et al., 2012 and references therein). So, in many landscapes, land-use change, improved agricultural practises and the passage of time (e.g., industrial areas rezoned as residential estates, broad-acre ploughing replaced with no-till cropping; street trees maturing and forming hollows) have resulted in large areas switching from unsuitable to suitable, yet a suite of limited-mobility animals are unable to colonize and inhabit them (Watson, 2010). Faunal relaxation through both

Fig. 1. Simplified landscape illustrating how targeted translocations of dispersal-limited animals can be used to maximize occupancy of suitable habitat. Panel 1 (top) depicts the original landscape, presumed habitat suitability for a lizard reflecting overall productivity, ranging from green (productive, high habitat suitability) to red (low productivity, low habitat suitability). Panel 2 depicts the same landscape after initial clearing—habitat in the most productive landforms was disproportionately cleared and the lizard persists only in the largest remnant (coloured lizard denotes occupied patch; white lizard = unoccupied). Panel 3 depicts the fragmented landscape ~100 years later—habitat suitability has decreased in the occupied remnant but has increased in unoccupied remnants (through both deterministic and stochastic processes) and a new patch of potential habitat has become available through natural regeneration. Panel 4 (bottom) depicts the outcome of translocating lizards into unoccupied habitat patches (two high suitability remnants and one low suitability regenerated site). In addition to increasing patch occupancy, habitat suitability in the source area increases as density-dependent effects are alleviated by reducing crowding. Successful establishment of lizards in the regenerated site refines habitat suitability criteria for future releases, the colour of this habitat patch changing from red to green.

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deterministic and stochastic processes represents a one-way ratchet towards progressively emptier habitats performing fewer ecosystem services becoming more vulnerable to invasion by exotic species. A solution to this situation is to move individual animals from those areas where there are plenty into areas where there are none (Fig. 1). Rather than occasional interventions undertaken only when regional populations become scarce, we envision community-driven environmental management involving large numbers of translocations of abundant and widespread species motivated by the axiom “keeping common animals common”. By taking excess individuals from remnants where a species persists and moving them to unoccupied but otherwise suitable habitat patches, a range of both pure and applied benefits can be realized. Considered in population terms, these actions equate to giving nonequilibrial metapopulations a push away from extinction via an imposed mass effect (after Leibold et al., 2004), while temporarily relieving density-dependent effects in source populations. From an ecological perspective, these actions collectively comprise a coordinated distributed experiment, testing current ideas about habitat preferences and resource requirements for candidate species. In ecosystem management terms, wildlife restoration increases the functional diversity of remnant habitats, making them more resistant to invasion. From a conservation perspective, this intervention spreads risk, increasing both demographic and genetic connectivity, thereby safeguarding against regional loss. From an education and engagement perspective, this community-driven initiative involves the wider public in active conservation management in their local area, empowering landholders and providing a new range of incentives to reward best-practise management. If this vision were to be realized under steady-state conditions, bestcase scenario outcomes would be the founding of a few subpopulations of already common and widespread species. Many introductions would fail, while most reintroduced subpopulations would likely blink off in a generation or two. But steady-state conditions are unlikely for many landscapes—species distributions are shifting under novel climates and no analogue communities are already developing (Williams and Jackson, 2007). Land-use patterns are changing, with entire districts previously used for production agriculture becoming better suited to carbon farming and other non-extractive uses (Watson et al., 2013). By introducing locally-extirpated species to pockets of suitable vegetation in the areas where we live and farm, we are augmenting the number of candidate native species able to persist in these altered landscapes over the long term. In addition to in-filling distributional ranges, seeding novel ecosystems and safeguarding regional populations, wildlife restoration embraces the essential human element of conservation biology. Identifying and conserving species of local significance is an effective way to promote environmental advocacy and increase linkages between people and nature (MacDonald and Johnson, 2003). Returning salamanders to an isolated wetland, skinks to a golf course, or shrews to a cemetery gives communities their own local champions while increasing their environmental literacy (Cairns, 2002) and contextualizing national and continental-scale visions within which these actions are embedded (Foreman, 2004). Wildlife restoration complements the existing portfolio of concepts and associated tools available to conservation researchers and restoration practitioners. It relies on many of the lessons learned from rare, captive-raised animals in reintroduction biology, but applies them to free living individuals of common species. It shares the same goal as assisted colonization/managed relocation (moving individuals to avoid an unacceptably high number of extinctions), but addresses this objective by infilling gaps in current distributions rather than moving species beyond current ranges. Our vision responds directly to the growing recognition that natural recolonization is too slow, too patchy, or both (Gallagher et al., 2015). Our vision embraces Manning et al. (2009) concept of landscape fluidity, with anticipatory restoration actions needed in many systems to maximize retention of species and maintenance of ecosystem function. Our approach complements the emerging concept

of rewilding (Foreman, 2004), downscaling from continents, tigers and wolves to parks, lizards and frogs. Wildlife restoration aligns with the growing understanding of the importance of climaticallydetermined microrefugia and their key role as sources of recruits under rapidly changing climates (Sublette Mosblech et al., 2011; Weeks et al., 2011). Finally, our approach borrows from aquatic ecology, specifically the emphasis of recruitment limitation as a key factor driving community assembly (Roughgarden et al., 1987). By treating translocations of common animals as a coordinated distributed experiment (after Fischman et al., 2014), we can expand our understanding of habitat requirements and distributional constraints which we can then progressively apply to restored, rehabilitated, and possibly recreated systems. By taking a proactive approach and managing species while they are still common, this approach emulates preventative medicine, achieving better outcomes more efficiently and cost effectively than waiting until problems arise (Huxel and Hastings, 1999). Finally and most importantly, rather than being just another tool for restoration ecologists to mend broken ecosystems, wildlife restoration is driven by the wider community, involving the general public in initial selection of candidate species and sites, eventual monitoring of establishment success and all intermediate stages. 3. Lessons from reintroduction biology: what we know and what we don't know Humans have been moving wild animals around for millennia, motivated by a variety of nutritional, recreational and aesthetic objectives (Griffith et al., 1989; Seddon, 2010). Although the majority of recent interventions involve restocking and enlarging the distribution of game species for hunting and fishing, they are increasingly being conducted for conservation management (Griffith et al., 1989). Translocation for conservation purposes relates almost exclusively to rare or endangered species, with many programmes using captive-raised animals (Armstrong and Seddon, 2008). Originally conducted on an ad hoc basis (often as last ditch attempts to avert extinction) the field of reintroduction biology has matured and developed into its own discipline (Serena and Williams, 1995), complete with synthetic reviews and comparisons of successful and unsuccessful initiatives (Fischer and Lindenmayer, 2000; Sheean et al., 2012). Several factors consistently emerge as predictors of successful reintroduction programmes, directly informing how to design wildlife restoration programmes for maximum benefit. Griffith et al. (1989) found habitat quality to be critical to successful establishment, and that reintroductions to core areas within prior distributional ranges are more successful than to peripheral areas. Veitch (1994) noted that the most important prerequisite to successful reintroduction was that “the original cause of their extirpation from that locality must be reversed or other repair activity undertaken to a point where the reintroduced species can survive” (see also IUCN, 1996). But identifying the driver of extirpation is rarely straightforward, and often relies on intuition and hearsay (Ewen and Armstrong, 2007). In addition to particular threats or resource requirements, this difficulty also relates to habitat suitability—where a species presently occurs need not be a reliable basis for determining the full range of habitats where that species once occurred. Moreover, given climatic changes and dynamic shifts in the gradients defining distributional ranges, prior occurrence may paint a misleading picture of future suitability. The difficulties of assessing habitat suitability underscore the need for experimental approaches, treating introductions as trials to test existing ideas about habitat preferences and refine our understanding of both bottom-up and top-down factors constraining realized distributions (Morton, 1987). Although researchers have long advocated using this approach with common species (Serena and Williams, 1995), reintroduction programmes remain focused on rare and endangered taxa, with neither the numbers of animals nor willingness to risk losing animals to improve our understanding of their needs.

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Plants, however, are regarded very differently by researchers and restoration practitioners alike. Large-scale reintroductions of common species are conducted on a routine basis, often using a variety of methods to determine which factors determine establishment success (Clewel and Aronson, 2013). As well as revegetation, transplant experiments are a standard tool to determine distributional constraints and discern genotypic from environmental effects on plant growth (Lesica and Allendorf, 1999). This traditional approach has undergone a renaissance as interest in plant distribution has switched from understanding distributional patterns to predicting the effects of climate change (Savolainen et al., 2007), with protocols refined and standardized to maximize comparability (see Kremer et al., 2012). The only group of animals where reintroductions of common species are routinely conducted is fish. Although the primary goal of most programmes is to maintain stocks for commercial and/or recreational fisheries, conservation management and maintaining ecosystem function are increasingly driving restoration efforts (Downs et al., 2002). Aside from habitat restoration for spawning substrate, fish introductions typically entail releasing large-numbers of hatchery-bred juveniles. Translocation of adults is occasionally conducted (e.g., moving adults of anadromous adults upstream of dams and other barriers) but connectivity with headwaters is typically achieved using fish ladders and other engineering solutions. Interestingly, the way fish are treated and the language used to describe these actions have more in common with plants than terrestrial animals—adults are “transplanted” (Olsson et al., 2006); hatchlings are “grown” in “nurseries”. Many of the ethical and philosophical aspects of translocation have been discussed and resolved in the fisheries management literature, and moving animals within and between watersheds is now widely practised. Although fisheries management may have been the original objective for most translocations, conservation management typically follows, with rivers unifying diverse stakeholder groups and catalysing whole-ofcatchment restoration efforts (Perrow et al., 2002). Whether through fear of upsetting the ‘natural balance’ or simply because this approach is foreign to most terrestrial ecologists, systematic and replicated translocation of common animals is rarely countenanced, let alone conducted (Serena and Williams, 1995). Rather than being relevant only for threatened species management or necessarily involving costly programmes centred on captive individuals and extensive reserves, we see translocation as a valuable tool in routine natural resource management. The greatest factor constraining assisted colonization and integrating animals explicitly into restoration is a fundamental lack of understanding of the factors defining habitat suitability (Lunt et al., 2013; Gallagher et al., 2015). In addition to testing our fragmentary understanding of habitat requirements and distributional constraints, mainstreaming translocations with common species will arrest future population declines while addressing all ten of Armstrong and Seddon's (2008) key questions in reintroduction biology. By building capacity and incorporating organizational learning opportunities for management agencies, our vision will generate the expertise and social licence required to realize the potential of assisted colonization and other bold interventions presently deemed prohibitively risky. 4. Practicalities 4.1. Candidates Instead of rare, charismatic or highly-interactive taxa, wildlife restoration focuses on common species that are dispersal limited, patchily distributed and locally abundant. Although traditionally disregarded in conservation biology, common species are disproportionately important in ecosystem structure and function (Gaston, 2010). Thus, subtle declines in their local abundance and distribution can drive pervasive cascades, leading to loss of less common species and diminished provision of ecosystem services. Dispersal ability can be difficult to quantify

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in absolute terms, but identifying those species within a local biota that are unable to move between remnant habitats is more straightforward (e.g., using expert opinion; Watson, 2003). Many amphibians, reptiles and small mammals are ideal candidates, especially those species dependent on vernal pools/wetlands (i.e., home ranges constrained by reproductive strategies) or those species reliant on burrows/hollows (i.e., central-place foragers). Rather than necessarily vertebrates, various arthropods and molluscs would also be candidates for translocation (e.g., Sherley, 1995). For long-lived animals, or those groups with low reproductive rates and high site fidelity, comprehensive surveys of potential source areas are required to ensure that removal of individuals will not have lasting detrimental impacts. In terms of candidate areas for release sites, selection criteria will be defined by the species being moved and by the motivations of the communities involved. In addition to habitat extent and quality, careful consideration of threats is needed, even though the reasons underlying local absences may be unclear. Baseline surveys to confirm absences and establish site suitability will become progressively more refined as outcomes of earlier translocations are integrated and determinants of habitat suitability are more finely resolved. For sites deemed unsuitable, these surveys may identify management interventions required to maximize suitability for future releases (e.g., installation of nest boxes or removal of incompatible exotic species; see Groenewald, 1992). Although source populations closest to release sites might seem most appropriate, peripheral populations (relative to range boundaries and microclimatic tolerances) should be considered, both to maximize retention of unique haplotypes but also recognizing that the habitats presently occupied by the species need not reflect the full range of suitable habitats (Ewen and Armstrong, 2007). Finally, although we envisage wildlife restoration becoming part of revegetation and remediation initiatives, initial interventions should relate to remnant habitat, both to maximize the likelihood of establishment and to build up a picture of the factors defining habitat suitability.

4.2. Population genetic considerations Most of the issues surrounding provenance, out-breeding depression and how best to select founders to avoid inbreeding depression have been thoroughly explored in the revegetation and reintroduction literature, but several matters warrant exploration here. Although revegetation practitioners have long championed the need to use locally-sourced seed, there is little evidence for superior growth of local ecotypes or “home site advantage” (Galloway and Fenster, 2000, see also Lesica and Allendorf, 1999). Referring explicitly to outbreeding depression in reintroductions, Weeks et al. (2011) wrote: “We suggest that perceived, but generally unsubstantiated, risks place too much constraint on current management options, commonly leading to inaction”, while Burton and Burton (2002) noted: “Indeed advocacy for the use of ‘local only’ plant materials can take on near religious proportions even when scientific imperatives are lacking”. In contrast, inbreeding avoidance requires explicit consideration, relating to the source populations used, as well as the individuals selected for translocation. Current approaches to growing native plants for revegetation use a “multi-lineal approach”—using mixtures of seed stock from multiple sources to ensure genetic breadth and population robustness, under the working assumption that if local genetic adaptation really matters, local environmental filters will select from this broader mix, ensuring a close fit (Burton and Burton, 2002). Hence, by drawing animal recruits from multiple source populations, both the amount of genetic variation represented and the adaptive potential increases, reducing the likelihood of both genetic and demographic losses across the wider region. Just as with habitat suitability, successive reintroductions represent valuable opportunities to improve our understanding of how genetic factors influence population persistence (Fenster and Dudash, 1994, see also Pääbo, 2000).

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D.M. Watson, M.J. Watson / Biological Conservation 191 (2015) 830–838 Fig. 2. Boston is a large city with both high-density urban development and a series of interconnected greenways known as the Emerald Necklace. These remnants and restored sites represent valuable habitat for many birds and other vagile taxa, but a suite of dispersal-limited taxa is missing. Many species of bumble bee are common but declining (Colla et al., 2011): translocations of red-belted bumble bee (Bombus rufocinctus) would bolster populations, resultant information then applied to less common species. Urban areas could also provide refuges for bumble bees from agricultural pesticides. Red-backed salamanders (Plethodon cinereus) thrive in urban environments, lower predation pressures in urban wetlands leading to increased survivorship of aquatic larvae. Finally, northern short-tailed shrews (Blarina brevicauda) are ideal candidates for urban translocation, more hospitable climatic conditions increasing survival over winter enabling rapid expansion from initial release sites into established gardens. Photographs by Larry Connolly (background), Sean McCann (red-belted bumble bee), Gilles Gonthier (short-tailed shrew) and Dave Huth (red-backed salamander).

4.3. Implementation Instead of being conducted solely by researchers, wildlife restoration relies on the participation of contributors with a much broader combined skill-set. Baseline surveys of potential release sites and source areas would be designed and conducted by natural resource management agency staff with at least a Masters-level education in wildlife ecology (or equivalent). While researchers may need to be involved in assisting with initial development of protocols and advice on species and site selection, these officers would be responsible for making decisions, conducting the translocations and monitoring their success. Rather than imposed, this work would necessarily be conducted in close consultation and direct participation of local stakeholders. Whether involving university students to help capture recruits, landholders contributing to baseline surveys, or citizen scientists assisting with subsequent monitoring, opportunities for community involvement will enable greater efficiency but also greater longevity of programmes by building and maintaining the necessary social licence. Careful consideration of the perceptions regarding translocation outcomes would be needed; e.g., how best to communicate the scientific value of failed efforts? In many jurisdictions, all of these proposed actions can be conducted within current legislative and management frameworks, but policies and regulations may need adjustment. These changes to regulatory frameworks need not be problematic, but could catalyse discussions with policy makers and enforcement officers at local and regional levels, adding their expertise and integrating these local actions with catchment and regional scale conservation planning.

5. Scenarios Initially, we envisage wildlife restoration to be most applicable in two classes of domesticated landscapes—those dominated by urban and agricultural development. In both settings, remnants of native vegetation are surrounded by high-contrast matrices (sensu Watson, 2002), making inter-patch movements possible only by vagile taxa (e.g., birds, bats, butterflies). With time since initial habitat loss, the matrix (i.e., suburbia and farmland) will become progressively more hospitable as

vegetation (parks and gardens; roadsides and on-farm plantings) matures and becomes more complex (e.g., Munyenyembe et al., 1989; Manning et al., 2009). These similarities notwithstanding, the emphasis and medium- to long-term objectives of wildlife restoration in these two classes of landscape differ. For urban areas, the high density of people coupled with the small size of most potential release sites means that establishing large populations is a secondary goal. Rather, engaging the wider community with biodiversity in their local area is the aim, and this will need to be considered in selecting potential species and candidate areas for release. By integrating wildlife restoration with existing programmes centred around particular parks or reserves and buildingin opportunities for volunteers, the overall benefits will extend well beyond the particular species and places involved. Given that cities characteristically develop in places with high fertility soils, reliable rainfall and more hospitable climates than surrounding areas (Werner, 2011), we anticipate establishment of new populations to be most successful for species associated with high productivity landforms. By contrast, the lower density of people living in agricultural landscapes and larger remnants of vegetation associated with low productivity landforms means that establishing new populations and filling in distributional gaps will be the main priority. Rather than being conducted in isolation, we envisage wildlife restoration to become part of existing natural resource management approaches, assisting in efforts to restore functional connectivity to minimize local extinctions as regional climates change. Engagement will be a secondary aim and may best be achieved via targeted programmes through local schools, allowing students to reconcile their understanding of apex predators and rewilding initiatives with tangible examples of local, familiar animals (MacDonald and Johnson, 2003). 5.1. Scenario 1—American city Most cities, especially those in the Northern hemisphere that have been urbanized for centuries, contain networks of small habitat patches (i.e., parks and reserves, Fig. 2) initially set aside or specifically planted to provide aesthetic amenity and recreation (see McKinney, 2008; Colding, 2011) linked by smaller habitat patches (i.e. street trees and

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Fig. 3. The south-west slopes region of New South Wales (southeastern Australia) is representative of many agricultural landscapes—habitats on the more productive soils disproportionately cleared and most remaining habitat restricted to low fertility landforms and linear strips beside creek-lines and roads. As landuse intensifies, the area used for cultivated crops and livestock production decreases, with both extent and quality of native habitats increasing in many areas. Of the dispersal-limited, patchilydistributed and locally abundant candidates, two reptiles and one mammal are ideal candidates for translocation within this landscape. Blue-tongue lizards (Tiliqua scincoides) are a widespread skink and, although common in gardens and large remnants, they are frequently killed by vehicles on roads and thus unable to recolonize otherwise suitable remnants (Koenig et al., 2001). Squirrel gliders depend on large hollow-bearing trees as denning sites but, although they readily use nest boxes, they do not cross open ground so are unable to access isolated woodlands. Lace monitors are large-bodied predatory lizards that persist in disturbed landscapes but were inadvertently extirpated from many farming areas through secondary poisoning (from historic rabbit control efforts). Photographs by David M. Watson (background), Roger Smith (lace monitor), Alex Clarke (blue-tongue lizard) and Michael Todd (squirrel glider).

backyards). As suburbs age, gardens mature and structural connections develop (e.g., transport corridors beside roads and railways), habitat extent and quality increase for many species, with birds, butterflies and bats typically increasing in diversity (McKinney, 2008; Werner, 2011). Only the more vagile subset of the original biota can recolonise these improving habitats, however with under-exploited resources potentially facilitating invasive species and subsidizing native mesopredators and other opportunistic generalists (e.g., raccoons Kluza et al., 2000). Those dispersal-limited taxa are represented by relictual populations along city margins and riparian buffers, threatened both by densitydependent factors and habitat loss as cities expand and new developments encroach. Of the dispersal-limited species distributed along the fringes of urban areas, bumble bees are one of the more speciose groups that tend to thrive in urban habitat, are not aggressive, are easily translocated using artificial nest boxes (Hobbs et al., 1960) and can benefit both native vegetation and gardens through pollination services. As an initial trial, ten schools add and monitor ten bee nest boxes in nearby green spaces or school grounds (e.g., Golick et al., 2003). Each nest box location is assessed (plant diversity, temperature, aspect) prior to translocation, and continually monitored by students for two years. This trial provides a large scale, high sample-size translocation experiment where all aspects—from nest box building to release and subsequent monitoring—are conducted by the community. South-facing nest boxes within 50 m of native flowering shrubs and within 100 m of a reliable water source may prove more successful, refining criteria for homeowners who wish to join into the city-wide bumble bee network. After this proof of concept trial, school and community groups can be recruited to participate in a larger-scale release of salamander eggs across the whole city. Relocating the eggs of common salamander species into urban environments spreads the risk of regional extinction and teaches students about nature while potentially founding additional breeding populations. Releases can be timed to coincide with water quality assessments or rainfall events, thus providing experimental data to determine the exact requirements for egg translocations, while providing recreational opportunities for local groups and reinforcing ties to their local natural spaces. No matter the exact species chosen, the improvements to local diversity will increase regional population stability and seed new

microrefugia and source populations from which these species can spread. Most of all, projects such as these create incentives for increased public engagement in ecosystem management, forging partnerships between disparate community groups and increasing awareness and literacy in local environmental issues (Cairns, 2002; Colding, 2011). As a final candidate for translocations, moving shrews into parks and cemeteries will facilitate expansion from relocation sites to adjoining backyards (Vergnes et al., 2013), and pave the way for future translocations of other predatory species. As with many small mammals, shrews are morphologically conservative (Fig. 2), and are frequently confused with rodents. These identification issues will encourage educational opportunities at local schools and community groups to learn about the diversity of small mammals found in the area. Indeed, such a programme might lead to the discovery of previously undocumented relictual populations of small mammals persisting in the area. By selecting a small, non-threatening predator, discussions of other carnivorous species could be initiated, potentially catalysing the social change necessary to develop future translocation projects that contribute to larger-scale rewilding programmes. 5.2. Scenario 2: Australian farmland Most southern Australian agricultural regions are managed for wheat and sheep production, but large-scale revegetation has become part of many properties' ongoing land management. Initially undertaken to combat gulley erosion around drainage lines and salinization in floodplains, these plantings are augmented by shelterbelts and farm forestry plantations. These activities, coupled with natural regeneration and regrowth on abandoned farmland and adjacent roadsides has resulted in the proportion of native vegetation increasing in many farming districts (Geddes et al., 2011; Watson et al., 2013). Although most uncleared vegetation is restricted to ridges and other low productivity landforms (Watson, 2011; Fig. 3), revegetation and regeneration has increased tree cover in riparian areas, valley floors and other high productivity sites, improving habitat quality and extent for many woodland dependent animals (Geddes et al., 2011). Reintroductions are increasingly being used as management tools for rare and threatened species (e.g., Manning et al., 2011), often in concert with feral animal control and other threat abatement measures.

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Of the dispersal-limited, patchily-distributed and locally abundant candidates, skinks are one of the more diverse groups, most species reliant on coarse woody debris, litter and outcrops (Koenig et al., 2001; Michael et al., 2011; see also Craig et al., 2012). Preliminary communications with potential landholders explain the need for trials with multiple species to improve understanding of habitat requirements, thereby progressively increasing the likelihood of success. As an initial trial, five species of skink are translocated to five properties, each containing five patches of remnant woodland. Having conducted surveys to determine suitability and locate nearby source populations, all translocations are carried out in a weekend. Although a relatively minor exercise for an environmental manager, to a quantitative animal ecologist, this represents 125 separate experimental trials to determine habitat preferences, 25 replicates for each species (providing the basis for two PhD projects). Two years after release, repeat surveys find skinks were more likely to establish in those woodlands with standing dead trees, crash grazing (rather than set stocking or no livestock grazing) and if recruits were drawn from more than one source population. In addition to refined protocols for future releases, set stocking and removal of dead trees for firewood are prioritized by the regional land management agency as issues to emphasize when considering applications for small grants to maximize biodiversity values and carbon sequestration on farms. After this proof-of-concept trial, another set of translocations is undertaken, moving two groups of eight squirrel gliders (Petaurus norfolcensis; sourced from four adjacent populations) to unoccupied roadside remnants where ongoing fox control and retention of large hollow-bearing trees meet the habitat suitability criteria. The animals only establish in one of the two release sites, which differs from the unsuccessful site in having twice as many available hollows and greater density of scattered trees in adjacent fields. In addition to refining the habitat suitability criteria, this gives landholders an additional incentive to install nest boxes on their properties and maintain fox control, thereby increasing habitat quality for a wide range of other woodlanddependent species. By scaling up strategic translocations to encompass entire catchments, these local actions rapidly build momentum, continually becoming more successful as lessons learned from previous trials are applied. The final candidate for translocation in this system is the Lace Monitor (Varanus varius), a large predatory lizard found throughout eastern Australia in woodland and forests. Eight gravid females are moved into two farming landscapes where monitors hadn't been seen for decades: an established grazing property raising sheep with extensive woodland along the ridgelines and creeks; and a group of adjoining lifestyle properties and gardens with remnant roadside woodland. Pre-release surveys confirm the presence of termite mounds (used as egg-laying sites) and high densities of European rabbits (an invasive species and frequent prey item). Follow up surveys conducted by undergraduate students confirm the presence of at least five individual monitors in both landscapes three years after release, with dietary analysis revealing rabbits, carrion, cats and locusts are important dietary items. Although landholders in both landscapes report monitors occasionally taking poultry, they also note a decrease in rabbit activity, both in woodland and adjoining pastures used for livestock grazing (Fig. 3). By selecting a charismatic predator, the pros and cons of larger future translocation projects and rewilding schemes become part of the lexicon of these land owners, making wildlife restoration less of an imposition and more of an element of best-practise property management.

6. Principles Building on established principles underpinning successful reintroductions of rare and threatened species (e.g., the five principles of Fyfe, 1978; the 10 principles of the IUCN, 1996), we propose the following five guiding principles to maximize the lasting benefits of wildlife restoration, ranked in order of importance.

1. Ensure source populations are not adversely impacted Given that one of primary aims of wildlife restoration is to ensure the long-term viability of common species by founding new subpopulations, adverse impacts to existing subpopulations must be avoided. In addition to determining the minimum size of populations to be used as source areas and maximum proportions of populations to be removed, additive impacts need to be given careful consideration. Thus, minimizing trapping for monitoring purposes, reducing handling of individuals to a minimum, and instituting hygienic protocols to ensure no pathogens or potential invasive species are introduced to source areas during fieldwork. Inadvertent impacts could be assessed by comparing source areas used with potential source populations not used, both to identify potential adverse impacts before they become threats, but also to determine whether populations benefit from reduced density-dependent factors by removing individuals. Finally, potential on-ground actions to improve habitats in source areas should be actively investigated, applying the same approach used in revegetation to ‘seed orchards’ to maximize future recruits. 2. Ensure release sites are not adversely impacted The same precautionary approach needs to be applied to all areas receiving individuals, ensuring best practise animal welfare principles are upheld and no pathogens or invasive species are accidentally introduced. As well as the initial release of animals, this relates to subsequent surveys to evaluate success of the intervention. As part of baseline surveys to determine habitat suitability for the target species, potential interactions should be given careful consideration, especially if potential release sites contain relictual populations of plants or animals poorly represented elsewhere. By comparing release sites visited frequently for post-release monitoring with potential release sites only visited for initial surveys of habitat suitability, any unforeseen consequences (positive and negative) of translocations will be detected early, informing management of affected sites and refining protocols for future interventions. 3. Monitor, monitor, monitor The importance of monitoring every stage of every translocation cannot be overstated. Rather than necessitating annual visits to all sites, monitoring must balance flexibility with consistency, working with landholders and other community members to ensure adherence to the same set of practises. Minimally, this monitoring should follow basic BACI principles, comparing before and after interventions in sites where translocations were conducted and sites where no animals were released. In addition to measuring the successes central to wider uptake, this monitoring will maximize learning opportunities from inevitable failures. Rather than simply relating to sites and species, parallel monitoring of ecological functions and aspects of the wider community should be undertaken concurrently, enabling evaluation of how wildlife restoration alters community perceptions and values. 4. Wildlife restoration is by the people, for the people In stark contrast to interventions involving rare or endangered species, these translocations cannot be driven by biologists. To realize maximum benefits to species, sites and communities, wildlife restoration actions must be driven by the same kinds of people who have rallied behind revegetation, restoration and rewilding. Rather than giving carte blanche for anyone to move anything anywhere, this principle draws on sociological research on the determinants of behavioural change, maximizing the likelihood of meaningful change by ensuring inclusion at all stages (Cairns, 2002). Rather than supplanting other approaches, we envisage wildlife restoration to complement traditional revegetation efforts, combining translocations with indigenous plantings to restore connectivity and enhance the flora and fauna of existing remnants. By adapting existing regulatory frameworks, adjusting relevant policy instruments where needed and incorporating training opportunities for key stakeholders, consistency, efficiency and broader uptake will be entrained.

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5. All outcomes must be shared The learning opportunities from these coordinated distributed experiments will only be realized if outcomes of all translocations are reported and widely accessible (Fischman et al., 2014). Rather than constraining individual groups or enforcing conformation to rigid protocols developed elsewhere, these reporting requirements will enable early detection of determinants of successful initiatives, thereby ensuring iterative improvement of techniques and onground outcomes. Web-accessible distributed databases can be readily constructed to facilitate both information storage and retrieval—of both data related to outcomes and associated metadata relating to actions (who, what, where and when). By establishing these overarching frameworks from the outset, large-scale conservation management can become progressively more evidence-based.

production-dominated properties with the plants that used to grow there—why not give them the opportunity to put some of the animals back as well?

7. Prospect

Armstrong, D.P., Seddon, P., 2008. Directions in reintroduction biology. Trends Ecol. Evol. 23, 20–25. Burton, P.J., Burton, C.M., 2002. Promoting genetic diversity in the production of large quantities of native plant seed. Ecol. Restor. 20, 117–123. Cairns Jr., J., 2002. Rationale for restoration. In: Perrow, M.R., Davy, A.J. (Eds.), Handbook of Ecological RestorationPrinciples of Restoration 1. Cambridge University Press, Cambridge, pp. 10–23. Clark, J.S., Lewis, M., McLachlan, J.S., HilleRisLambers, J., 2003. Estimating population spread: what can we forecast and how well? Ecology 84, 1979–1988. Clewel, A.F., Aronson, J., 2013. Ecological Restoration: Principles, Values, and Structure of an Emerging Profession. 2nd ed. Island Press, Washington DC. Colding, J., 2011. Creating incentives for increased public engagement in ecosystem management through urban commons. In: Boyd, E., Folke, C. (Eds.), Adapting Institutions: Governance, Complexity and Social-Ecological Resilience. Cambridge University Press, Cambridge, pp. 101–124. Colla, S., Richardson, L., Williams, P., 2011. Bumble Bees of the Eastern United States. USDA Forest Service. Craig, M.D., Hardy, G.E.S.J., Fontaine, J.B., Garkakalis, M.J., Grigg, A.H., Grant, C.D., Fleming, P.A., Hobbs, R.J., 2012. Identifying unidirectional and dynamic habitat filters to faunal recolonisation in restored mine-pits. J. Appl. Ecol. 49, 919–928. Downs, P.W., Skinner, K.S., Kondolf, G.M., 2002. Rivers and streams. In: Perrow, M.R., Davy, A.J. (Eds.), Handbook of Ecological RestorationRestoration in Practice 2. Cambridge University Press, Cambridge, pp. 267–296. Ewen, J.G., Armstrong, D.P., 2007. Strategic monitoring of reintroductions in ecological restoration programmes. Ecoscience 14, 401–409. Fenster, C.B., Dudash, M.R., 1994. Genetic considerations for plant population restoration and conservation. In: Bowles, M.L., Whelan, C.J. (Eds.), Restoration of Endangered Species: Conceptual Issues, Planning and Implementation. Cambridge University Press, Cambridge, pp. 34–62. Fischer, J., Lindenmayer, D.B., 2000. An assessment of the published results of animal relocations. Biol. Conserv. 96, 1–11. Fischman, R.L., Meretsky, V.J., Babko, A., Kennedy, M., Liu, L., Robinson, M., Wambugu, S., 2014. Planning for adaptation to climate change: lessons from the US national wildlife refuge system. Bioscience 64, 993–1005. Foreman, D., 2004. Rewilding North America: a Vision for Conservation in the 21st century. Island Press, Washington. Fyfe, R.W., 1978. Reintroducing endangered birds to the wild: a review. In: Temple, S.A. (Ed.), Endangered Birds: Management Techniques for Preserving Threatened Species. Croom Helm, London, pp. 323–329. Gallagher, R.V., Makinson, R.O., Hogbin, P.M., Hancock, N., 2015. Assisted colonization as a climate change adaptation tool. Austral Ecol. 40, 12–20. Galloway, L.F., Fenster, C.B., 2000. Population differentiation in an annual legume: local adaptation. Evolution 54, 1173–1181. Gaston, K.J., 2010. Valuing common species. Science 327, 154–155. Geddes, L.S., Lunt, I.D., Smallbone, L.T., Morgan, J.W., 2011. Old field colonization by native trees and shrubs following land use change: could this be Victoria's largest example of landscape recovery? Ecol. Manag. Restor. 12, 31–36. Golick, D.A., Schlesselman, D.M., Ellis, M.D., Brooks, D.W., 2003. Bumble boosters: students doing real science. J. Sci. Educ. Technol. 12, 149–152. Griffith, B., Johnston, C.A., Scott, J.M., Carpenter, J.W., Reed, C., 1989. Translocation as a species conservation tool: status and strategy. Science 245, 477–480. Groenewald, G.H., 1992. The relocation of Cordylus giganteus in the Orange Free State, South Africa: pitfalls and their possible prevention. J. Herp. Ass. Afr. 40, 72–77. Hobbs, G.A., Virostek, J.F., Nummi, W.O., 1960. Establishment of Bombus spp. (hymenopetera: apidae) in artificial domiciles in Southern Alberta. Can. Entomol. 92, 868–872. Hoegh-Guldberg, O., Hughes, L., McIntyre, S., Lindenmayer, D.B., Parmesan, C., Possingham, H.P., Thomas, C.D., 2008. Assisted colonization and rapid climate change. Science 321, 345–346. Huxel, G.R., Hastings, A., 1999. Habitat loss, fragmentation and restoration. Restor. Ecol. 7, 309–315. IUCN, 1996. IUCN/SSC guidelines for re-introductions. 41st Meeting of the IUCN Council, Gland Switzerland, May 1995. http://iucn.org/themes/ssc/pubs/policy/reinte.htm. Kluza, D.A., Griffin, C.R., Degraaf, R.M., 2000. Housing developments in rural New England: effects on forest birds. Anim. Conserv. 3, 15–26. Koenig, J., Shine, R., Shea, G., 2001. The ecology of an Australian reptile icon: how do bluetongued lizards (Tiliqua scincoides) survive in suburbia? Wildl. Res. 28, 214–227.

Conservation Biology has moved well past discussions of what is ‘natural’, preservationist ideals (see Smith and Winslow, 2001) replaced with the more pragmatic ‘novel ecosystems’ paradigm, with functionality and purpose paramount. Although we share the views of HoeghGuldberg et al. (2008) that assisted colonization and other bold initiatives must be considered if we are to minimize the loss of entire lineages from extinction, we also recognize the caution with which these ideas have been received (Lunt et al., 2013; Gallagher et al., 2015). We simply do not know enough about the resource needs and habitat requirements of species to complete risk management procedures for assisted colonization with due diligence—we cannot be sure that the treatment will end up being more destructive than the underlying disease. If we adopt interventionist management with common species in the landscapes where we live, work and grow our food, valuable lessons can be learned to foster this much-needed confidence. By building capacity and gaining the trust of the wider community, we can generate the requisite momentum to start pushing the population extinction ratchet in the opposite direction, one local good news story at a time. 8. Conclusions Restoration ecology has undergone a rapid transformation from reactive, conservation-driven practise to a theoretically-grounded scientific discipline, achieving improved on-ground outcomes while addressing fundamental questions about ecological process. Here, we identify the next key challenge for restoration ecology to embrace: to explicitly include animals in restoration efforts. Many animals are dispersal-limited, many sites cannot be reconnected and many suitable habitats are missing key biotic elements. Applying lessons learned from revegetation and principles of reintroduction biology developed with rare and endangered species, we outline a novel approach we call Wildlife Restoration. By moving dispersal-limited animals stranded in isolated remnants to unoccupied but otherwise suitable habitat patches, distributional ranges can be in-filled, local extinctions avoided and regional populations of common species safeguarded. As well as being far more cost-effective than reactive management to reverse declines, this proactive approach has several other benefits. By designing each set of translocations using consistent protocols, these individual conservation actions collectively represent a coordinated distributed experiment, revealing the mechanistic basis of observed habitat preferences and identifying the climate and resource-based determinants of realized distributions. By returning native species, functional diversity and ecosystem service provision are augmented, thereby reducing the likelihood of colonization by invasive species. Finally, all aspects of Wildlife Restoration involve members of the local community, building in meaningful opportunities for engagement and increasing connectedness between people and the places where they live and work. The ongoing success of revegetation initiatives has demonstrated the great willingness of landholders to repopulate their gardens, neighbourhoods and

Acknowledgments Discussions with Amos Bouskila, Larry Connolly, Michael Craig, Damian Michael, Dale Nimmo, Euan Ritchie and Rodney van der Ree clarified the ideas expressed herein. MJW recognizes the support of the Institute for Land Water and Society, Conservation Evidence and the Technion University. References

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Kremer, A., Ronce, O., Robledo-Arnuncio, J.J., Guillaume, F., Bohrer, G., Nathan, R., Bridle, J.R., Gomulkiewicz, R., Klein, E.K., Ritland, K., Kuparinen, A., Gerber, S., Schueler, S., 2012. Long-distance gene flow and adaptation of forest trees to rapid climate change. Ecol. Lett. 15, 365–377. Leibold, M.A., Holyoak, M., Mouquet, N., Amarasekare, P., Chase, J.M., Hoopes, M.F., Holt, R.D., Shurin, J.B., Law, R., Tilman, D., Loreau, M., Gonzalez, A., 2004. The metacommunity concept: a framework for multi-scale community ecology. Ecol. Lett. 7, 601–613. Lesica, P., Allendorf, F.W., 1999. Ecological genetics and the restoration of plant communities: mix or match? Restor. Ecol. 7, 42–50. Lunt, I.D., Byrne, M., Hellmann, J.J., Mitchell, N.J., Garnett, S.T., Hayward, M.W., Martin, T.G., Mcdonald-Madden, E., Williams, S.E., Zander, K.K., 2013. Using assisted colonisation to conserve biodiversity and restore ecosystem function under climate change. Biol. Conserv. 157, 172–177. MacDonald, D.W., Johnson, P.J., 2003. Farmers as custodians of the countryside: links between perception and practice. In: Tattersall, F.H., Manley, W. (Eds.), Farming and Mammals. Occasional Publication of the Linnean Society, pp. 2–16. Manning, A.D., Fischer, J., Felton, A., Newell, B., Steffen, W., Lindenmayer, D.B., 2009. Landscape fluidity—a unifying perspective for understanding and adapting to global change. J. Biogeogr. 35, 193–199. Manning, A.D., Wood, J.T., Cunningham, R.B., McIntyre, S., Shorthouse, D.J., Gordon, I.J., Lindenmayer, D.B., 2011. Integrating research and restoration: the establishment of a long term woodland experiment in south-eastern Australia. Aust. Zool. 35, 633–648. Maschinski, J., Haskins, K.E. (Eds.), 2012. Plant Reintroduction in a Changing Climate. Island Press, Washington D.C., U.S.A. McKinney, M.L., 2008. Effects of urbanization on species richness: a review of plants and animals. Urban Ecosyst. 11, 161–176. Michael, D.R., Cunningham, R.B., Lindenmayer, D.B., 2011. Regrowth and revegetation in temperate Australia presents a conservation challenge for reptile fauna in agricultural landscapes. Biol. Conserv. 144, 407–415. Miller, J.R., Hobbs, R.J., 2002. Conservation where people live and work. Conserv. Biol. 16, 330–337. Morton, E.S., 1987. Reintroduction as a method of studying bird behavior and ecology. In: Jordan III, W.R., Gilpin, M.E., Aber, J.D. (Eds.), Restoration Ecology: A Synthetic Approach to Ecological Research. Cambridge University Press, pp. 166–172. Munyenyembe, F., Harris, J., Hone, J., Nix, H., 1989. Determinants of bird populations in an urban area. Austral Ecol. 14, 549–557. Olsson, I.C., Greenberg, L.A., Bergman, E., Wysujack, K., 2006. Environmentally induced migration: the importance of food. Ecol. Lett. 9, 645–651. Pääbo, S., 2000. Of bears, conservation genetics and the value of time travel. Proc. Natl. Acad. Sci. U. S. A. 97, 1320–1321. Palmer, M.A., Ambrose, R.F., Poff, N.L., 1997. Ecological theory and community restoration ecology. Restor. Ecol. 5, 291–300. Patiño, J., Vanderpoorten, A., 2015. How to define nativeness in organisms with high dispersal capacities. A response to Essl et al. J. Biogeogr. 7, 1360–1362. Perrow, M.R., Tomlinson, M.L., Zambrano, L., 2002. Fish. In: Perrow, M.R., Davy, A.J. (Eds.), Handbook of Ecological RestorationPrinciples of Restoration 1. Cambridge University Press, Cambridge, pp. 324–354.

Roughgarden, J., Gaines, S.D., Pacala, S.W., 1987. Supply-side ecology. The role of physical transport processes. In: Gee, J.H.R., Giller, P.S. (Eds.), Organization of Communities: Past and Present. Blackwell Scientific Press, London, pp. 491–518. Savolainen, O., Pyhäjärvi, T., Knürr, T., 2007. Gene flow and local adaptation in trees. Annu. Rev. Ecol. Evol. Syst. 38, 595–619. Seddon, P.J., 2010. Moving along the conservation translocation spectrum. Restor. Ecol. 18, 796–802. Serena, M., Williams, G.A., 1995. Wildlife conservation and reintroduction: an Australian perspective. In: Serena, M. (Ed.), Reintroduction Biology of Australian and New Zealand Fauna. Surrey Beatty & Sons, Chipping Norton, pp. 247–252. Sheean, V.A., Manning, A.D., Lindenmayer, D.B., 2012. An assessment of scientific approaches towards species relocations in Australia. Austral Ecol. 37, 204–215. Sherley, G., 1995. Translocations of the mahoenui giant weta Deinacrida n. sp. and Placostylus land snails in New Zealand: what have we learned? In: Serena, M. (Ed.), Reintroduction Biology of Australian and New Zealand Fauna. Surrey Beatty & Sons, Chipping Norton, pp. 57–63 Smith, S.E., Winslow, S.R., 2001. Comparing perceptions of native status. Nativ. Plants J. 2, 5–11. Sublette Mosblech, N.A., Bush, M.B., van Woesik, R., 2011. On metapopulations and microrefugia: palaeoecological insights. J. Biogeogr. 38, 419–429. Veitch, C.R., 1994. Habitat repair: a necessary prerequisite to translocation of threatened birds. In: Serena, M. (Ed.), Reintroduction Biology of Australian and New Zealand Fauna. Surrey Beatty & Sons, Chipping Norton, pp. 97–104. Vergnes, A., Kerbirious, C., Clergeau, P., 2013. Ecological corridors also operate in an urban matrix: a test case with garden shrews. Urban Ecosyst. 16, 511–525. Watson, D.M., 2002. A conceptual framework for the study of species composition in islands, fragments and other patchy habitats. J. Biogeogr. 29, 823–834. Watson, D.M., 2003. Long-term consequences of habitat fragmentation: highland birds in Oaxaca, Mexico. Biol. Conserv. 111, 283–303. Watson, D.M., 2010. From pattern to process: towards understanding drivers of diversity in temperate woodlands. In: Lindenmayer, D., Bennett, A., Hobbs, R. (Eds.), Temperate Woodland Conservation and Management. CSIRO Publishing, Collingwood, pp. 159–166. Watson, D.M., 2011. A productivity-based explanation for woodland bird declines: poorer soils yield less food. Emu 111, 10–18. Watson, S., Luck, G., Spooner, P., Watson, D.M., 2013. Human-induced land-cover change: incorporating the interacting effects of frequency, sequence, time-span and magnitude of changes on biota. Front. Ecol. Environ. 12, 241–249. Weeks, A.R., Sgro, C.M., Young, A.G., Frankham, R., Mitchell, N.J., Miller, K.A., Byrne, M., Coates, D.J., Eldridge, M.D.B., Sunnucks, P., Breed, M.F., James, E.A., Hoffmann, A.A., 2011. Assessing the benefits and risks of translocations in changing environments: a genetic perspective. Evol. Appl. 4, 709–725. Werner, P., 2011. The ecology of urban areas and their functions for species diversity. Landsc. Ecol. Eng. 7, 231–240. Williams, J.W., Jackson, S.T., 2007. Novel climates, no-analog communities, and ecological surprises. Front. Ecol. Environ. 5, 475–482.