Considerations of scale in biodiversity conservation ...

Animal Conservation. Print ISSN 1367-9430

Considerations of scale in biodiversity conservation J. T. du Toit Department of Wildland Resources, Utah State University, Logan, UT, USA

Keywords multi-scale analysis; spatiotemporal scaling; triage; environmental planning; globalization. Correspondence Johan T. du Toit, Department of Wildland Resources, Utah State University, Logan, UT 84322-5230, USA. Email: [email protected] Received 30 October 2009; accepted 26 January 2010 doi:10.1111/j.1469-1795.2010.00355.x

Abstract The dilemma of conservation practice lies in weighing the urgency for action against the need for sustainable long-term solutions, with urgent responses incurring the risk of failure and long-term solutions incurring the cost of time. Wisdom of hindsight reveals that sustainable solutions are not achieved when conservation action is initiated at an inappropriate scale. Here, I review recent studies that have included considerations of scale to illustrate how conservation problems and solutions might be unapparent, or even counterintuitive, to conservation practitioners responding to issues at the scales at which they were first perceived. Case studies cover the conservation of ecosystems, ecosystem services, species and populations. These studies collectively illustrate how most biodiversity conservation efforts can be improved by considering the problem at a broader spatiotemporal scale than that at which local natural resource management has traditionally operated. Globalization is increasingly challenging conservation practitioners to search for solutions across an ever-wider range of spatiotemporal scales and institutional levels. Identifying real problems and threats at relevant scales is part of conservation triage, when opportunity costs and cost efficiencies of alternative interventions are evaluated and ranked, before action is implemented through the appropriate institutional levels.

Introduction When a population of a charismatic wildlife species is steadily declining within a large and well-managed national park, and is even projected to disappear within the next decade or two, then few conservation practitioners would disagree that something should be done soon. However, if an analysis finds it to be a declining subpopulation at the southern fringe of its large species range, which appears to be contracting northward in response to climatic changes, then pragmatists might suggest allocating conservation funds elsewhere. This situation applies to no less than three antelope species in the Kruger National Park, South Africa: roan Hippotragus equinus, sable Hippotragus niger and tsessebe Damaliscus lunatus (Ogutu & Owen-Smith, 2003). While there is every reason to continue monitoring and researching such population declines, my purpose in highlighting such examples – and in writing this review – is to illustrate the scale dependence of conservation problems and their solutions. It is now well recognized in the conservation profession that scientists, planners and field practitioners all need to do a better job of engaging science with effective on-the-ground action (du Toit, Walker & Campbell, 2004; Knight, Cowling & Campbell, 2006; Robinson, 2006; Garnett, Sayer & du Toit, 2007). To this end, two separate panels of representatives from relevant organizations have each distilled out 100 questions to be answered by researchers to achieve greater

effect in conserving biological diversity in the UK (Sutherland et al., 2006) and the world (Sutherland et al., 2009). However, there is one key question missing from both lists, perhaps because it addresses a key procedure rather than a specific research topic, and that is, before planning and implementing any conservation action, has the perceived conservation problem been examined across multiple scales? Effective action has to be applied at the spatiotemporal scale(s) at which the problem is generated, even if the immediate problem is perceived at a different scale. Finding the appropriate scale for examining a conservation problem is analogous to using a standard light microscope to examine a specimen. The investigator explores multiple magnifications by trying different objective lenses and racking up and down with coarse- and fine-adjustment knobs before choosing the optimal setting; but getting the right setting is impossible without first viewing the specimen at higher and lower magnifications. In the same way, if a conservation effort is immediately implemented without considerations of scale, it could soon end in wasted effort, resources and credibility. This is simply because the causes of the problem might occur at a different scale than was assumed when the need for conservation action was initially perceived. Furthermore, there are multiple axes to which the need for an initial cross-scale examination might apply, including space, time, institutional level, financial investment, political integrity and so on. I present below some examples to illustrate the need for considerations of scale in biodiversity

c 2010 The Authors. Journal compilation c 2010 The Zoological Society of London Animal Conservation 13 (2010) 229–236

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conservation and then suggest a robust approach for incorporating these considerations into the triage stage of conservation planning.

Thinking across spatial scales The Millenium Ecosystem Assessment (2003) found, unsurprisingly, that the analysis of social–ecological systems and their processes is scale dependent. When analyzing the resilience of a social–ecological system, its historical profile should be developed at three scales: local, regional and multi-regional (Walker & Salt, 2006). The definition of local and regional scales depends on the system of interest but the point is that the analysis should occur at scales below, at and above that at which a need for such an analysis might be perceived, allowing cross-scale effects to be identified. Because biodiversity conservation draws on both natural and social sciences, it follows that conservation planning and action should operate within the context of social–ecological systems and thus be scale sensitive. In addition, with the maturation of conservation biology in recent decades, its focus has expanded outwards from species to include multiple levels of biological organization – from genes to ecosystems – which might suggest that multi-scale thinking is also a trend (Poiani et al., 2000). But is it? It seems not, because multi-scale studies remain remarkably uncommon in ecology – the very science that underpins biodiversity conservation. A recent survey (Sandel & Smith, 2009) found that only 7% of articles in high-impact ecological journals include the term ‘spatial scale,’ and a decade ago there were a quarter as many. Of those, most simply describe the scale of the research and suggest whether the findings might or might not be applicable at other scales. Little over half of the 7% of articles describe studies in which a multi-scale component had been built in, and articles that specifically incorporate a multi-scale experimental design have consistently remained extremely rare. For clearly localized conservation issues, such as with populations that are isolated and declining for well-established and quickly reversible reasons, there might not seem to be any obvious reason for worrying about multi-scale considerations. An example is the ‘genetic rescue’ of the Florida panther Puma concolor coryi, which declined to 30 inbred individuals in an isolated remnant population that is now recovering after a bold, controversial, but so-far successful intervention (Pimm, Dollar & Bass, 2006). Eight females of the most closely related Texas subspecies Puma concolor stanleyana were transplanted into the Florida panther population in 1995 and they and their hybrid offspring have bred well. The population is now growing and expanding its range in south-east Florida. Nevertheless, it is hard to imagine the ever-present effects of human encroachment on wildlife habitat and additive mortality from vehicle collisions will not cap the genetically ‘rescued’ population too. Unless the spatial scale of the Florida panther recovery program can be increased to secure substantially more habitat soon, then the temporal scale of the success story will be short. The population will not be able to 230

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exceed the size required to prevent the recurrence of inbreeding depression and so further genomic infusions will surely be required. Steady introgression of Texas genes will inevitably result in the loss of adaptive Florida alleles through genetic drift (Maehr & Lacy, 2002). This might actually be of little ecological consequence as long as the functional properties of this large predator are maintained in what remains of the ecosystem. Yet, if the citizens of Florida wish to preserve ‘their’ panther, P. concolor coryi, then the spatial extent of the conservation action will have to be scaled up soon. Even a very large population in a very large protected area might be at risk to conservation threats if considerations of scale are not adequately incorporated into its management (Woodroffe & Ginsberg, 1998). An example is the wildebeest Connochaetes taurinus population that has maintained a size of 1.3 million animals for the past three decades in the vast Serengeti–Mara Ecosystem (25 000 km2) of East Africa. This migratory population is regulated by its dry season food supply and is thus crucially dependent on its northern dry season range, which straddles the border between Masai Mara National Reserve (Kenya) and Serengeti National Park (Tanzania). Habitat loss and poaching have been the long-held focus of concern for this migration, especially along its north-western edge (Thirgood et al., 2004), and poaching can be effectively controlled if adequate resources are allocated to antipoaching operations (Hilborn et al., 2006). That is not entirely reassuring, however, because if we ‘zoom out’ to view the migration at a wider spatial scale then a much more immediate conservation threat comes into focus. Being grazers, wildebeest are water dependent and able to utilize the northern Serengeti–Mara dry season range by having access to permanent drinking water in the Mara River. It follows, therefore, that the hydrological status of the Mara River and its tributaries that originate deep within the agricultural lands of south-western Kenya should be highly relevant to the conservation of the spectacular Serengeti wildebeest migration. Indeed, eco-hydrological modeling predicted deforestation, irrigation and diversion of water for hydro-electric generation would end the migration (Gereta et al., 2002). The Mara River was projected to become depleted to the extent that a drought of historical severity could cause its flow into the Serengeti–Mara Ecosystem to be insufficient to replace the water removed by evaporation and consumption by animals. If that were to happen the river would stop running, pools would dry up and the migration could be decimated. Intervention by UNESCO has prevented the water diversion (Sinclair & Byrom, 2006) but the Mara River Basin is still undergoing extensive anthropogenic land-cover change, causing severe soil erosion in the catchment area and sedimentation downstream (Mati et al., 2008). The flow regime of the Mara River has shifted, with the high flows being stronger and occurring earlier in the wet seasons and the low flows being reduced in the dry seasons. Efforts are now underway to develop and implement an integrated transboundary management plan for the Mara


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River Basin and hopefully the system has not been forced beyond its resilience threshold yet. Either way, the lesson in this example is that efforts to conserve the Serengeti migration have, until only very recently, missed a key threat because they have been restricted to the scale of the migration itself (Dobson & Lynes, 2008; Sinclair et al., 2008). This is symptomatic of traditional park management practices in semi-arid ecosystems where water is the life-blood of wildlife communities on a park scale but is strongly influenced by human activities on a basin-wide scale. Nevertheless, innovative multi-scale approaches are emerging and Kruger National Park in South Africa provides a role model for integrating biodiversity conservation into the order of business of catchment management agencies (O’Keeffe & Rogers, 2003). In north-eastern Australia too, multi-scale adaptive management strategies are being devised to protect the Great Barrier Reef from sediments and nutrients discharged into the ocean by rivers flowing from distant grazing lands in the interior of Queensland (Gordon, 2007). Some well-intended actions of natural resource managers can prove ineffective, or even counterproductive, to biodiversity conservation because the conventional institutional framework is adapted for those spatiotemporal scales relating to the life history of just one species in the system – Homo sapiens. It is now apparent that the scales at which humans manage livestock and cultivate land are inappropriate for managing free-ranging wildlife and conserving ecosystem services (Hobbs et al., 2008). The test-and-slaughter protocol for controlling brucellosis is appropriate for cattle at the ranch scale but inappropriate for bison Bison bison and elk Cervus elaphus at the scale of the Greater Yellowstone Ecosystem (Bienen & Tabor, 2006; Kilpatrick, Gillin & Daszak, 2009). Similarly, the areas and configurations of fields in croplands can leave an inadequate matrix of natural habitat for crucially important pollinating insects (Kremen et al., 2004). Such mismatches of scale are becoming increasingly apparent as globalization causes the inadvertent actions of individuals on a local scale to collectively translate into impacts on a global scale (Satake, Rudel & Onuma, 2008). Carbon emissions are an obvious case but there are others that verge on the bizarre – such as the global demand for organically grown fair-trade coffee causing ecosystem degradation in an Indian national park. Madhusudan (2005) found that local villagers had capitalized on their ‘subsistence’ grazing rights by herding as many cattle as they could into Bandipur National Park. They collected all the dung and transported it to the hill districts where, at the time of the study, there was a lucrative market in dung as organic fertilizer for coffee production. The profits were more than adequate to allow the villagers to switch from manure to industrial fertilizer for their own crops, while Bandipur was overgrazed and effectively ‘mined’ for nutrients. In principle, this is similar to the problem of bushmeat hunting in West Africa, which is exacerbated by the diversion to Europe of commercially harvested oceanic fish, causing the consumption of protein substitutes from forest wildlife to increase unsustainably (Brashares et al., 2004). Clearly, local conservation efforts to resolve such problems


cannot succeed within a conventional institutional framework. A new framework is required, and appears to be developing (McNeely, 2006), with strong linkages among coordinated conservation programs that bring enlightened economic instruments to bear and remove perverse incentives within a hierarchy of institutional levels, across spatial scales.

Thinking across temporal scales Biodiversity conservation is a crisis discipline, so when a steadily declining population trend is detected in a species of interest, it should trigger a call for action. But in the majority of cases, there is an absence of reliable population data before the onset of the downward trend. Unless it is an obvious case of documented habitat loss, the temporal scale over which conservation planners and practitioners are able to work is usually too short to interpret the ‘natural’ dynamics of the population and thereby set an ecologically meaningful target for recovery (Fig. 1). Innovative methods are needed to extend the time scale over which a population’s dynamics can be analyzed, and these could include the use of indirect evidence. One such method was devised by Drake & Naiman (2007) to reconstruct the population trends of four Pacific salmon (Oncorhynchus) species over 150–350 years in five mid-order rivers in the Pacific Northwest region of the USA. They did this by relating existing data on (1) migrating adult salmon counted at weirs or dams to (2) rates of nitrogenous release from salmon carcasses into riparian soils and then to (3) with-in year growth

Figure 1 When conservation practitioners see a marked decline in the size of a population (N) of special interest (inset, top right), the monitoring data are usually available from some time (t) that is relatively recent (vertical dashed line). This frequently leads to an assumption that the population decline needs to be reversed with conservation action, which might or might not be true depending on the long-term dynamics of the population. It is also commonly assumed that the population declined from some equilibrium (line B) approximating the level recorded when data collection began. However, the equilibrium level might also be considerably higher (line A), or even lower before a recent population eruption (line D), or else the recent decline might be part of a long-term cyclical pattern that is natural to the population (line C).


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Figure 2 Four decades of monitoring data show that the buffalo Syncerus caffer population (N) of Kruger National Park, here depicted for the central region of the park only, declines during droughts (shaded bars) and recovers in-between (Cross et al., 2009). A close association between a buffalo population response and rainfall variation is indicated by the plot of one-step geometric population growth (population size in year t divided by population size the previous year; Nt/Nt 1) overlain on the plot of a relative index of rainfall. R is mean rainfall (across five weather stations in the region) in year t divided by the long-term mean. Vegetation is influenced by present and past rainfall, so the index is the average of R in year t and the previous year [(Rt+Rt 1)/2]. Data were collected by the monitoring program of SANParks Scientific Services in Kruger National Park (http://www.sanparks.org/parks/kruger/ conservation/scientific/mission/).

responses in riparian trees. By modeling the relationship between the width of riparian tree rings and the annual abundance of spawning salmon, Drake and Naiman used dendroecological methods to reconstruct the chronologies of salmon abundance in each river. This enabled them to broaden the time scale of their demographic analysis from a few decades of monitoring data to centuries of reconstructed data, thereby elucidating the cyclical patterns of salmon abundance that fishery managers need to be aware of. In this example, the key to success was the existence of reliable decadal-scale monitoring data collected by diligent fisheries technicians working at remote sites. However, the value of monitoring can easily be overlooked by management agencies working with annual budgets because such value only accrues when the time scale of the monitoring data becomes decadal. The Kruger National Park in South Africa has a unique record of ecological monitoring, with annual aerial surveys having been conducted across this 20 000 km2 park since the late 1960s. Little was it known then that bovine tuberculosis (BTB) would be detected in an African buffalo Syncerus caffer in the south of Kruger in 1990 and spread rapidly northwards so that virtually all herds in the population are now carrying BTB and spreading it to other species including lions Panthera pardus. Although there is no detectable demographic impact of BTB on buffaloes in Kruger so far (Cross et al., 2009), the long-term conservation implications of this exotic bacterial disease remain worrying. Kruger’s managers have been unsuccessful in controlling the epidemic because the logistical and financial challenges of maintaining vaccination coverage and test-and-slaughter campaigns are prohibitive for a park-wide buffalo population of 25 000 animals. Nevertheless, ecological research at individual, herd and population levels, and at seasonal, annual and decadal time scales, has identified an opportunity that could not have been seen from any one-off veterinary survey. First, BTB-positive buffaloes lose body condition significantly faster and develop significantly higher endoparasite loads than BTB-negative animals in the dry season, 232

and then recover in the following wet season (Caron, Cross & du Toit, 2003). Second, four decades of monitoring data show that the buffalo population undergoes cycles of rapid decline and gradual recovery, with declines coinciding with periodic droughts (Fig. 2). Third, it follows that BTBpositive animals should be the most likely individuals to succumb during droughts and so herd- and population-level BTB prevalence should be lowest in the early post-drought recovery phase. The total population size is lowest then too. Fourth, with meteorological modeling it is possible to forecast impending droughts. This multi-scale collection of evidence provides a strategy for any future veterinary campaigns (e.g. test-and-slaughter and/or vaccination) against BTB in Kruger’s buffaloes: focus such efforts only on short windows of opportunity directly after each drought and conserve resources in-between. A feature of contemporary thinking on conservation is the recognition that change is natural; it is now widely accepted that nature preservation is a contradiction in terms because nature is in constant flux. Conservation practitioners are still in a dilemma, however, when mandated to either prevent some component of nature from changing or else restore it to some arbitrary former state. Conservation planning of any kind has to be based on an understanding of ecosystem dynamics (Sinclair & Byrom, 2006), which requires an appreciation of temporal scale. For example, there is political pressure to restore the Chobe ecosystem in northern Botswana to the state it was in during the 1960s when the Chobe River was fringed by a closed canopy forest of even aged Acacia trees. Virtually, all those trees have now been killed through debarking by elephants Loxodonta africana and any seedlings emerging from the seed bank are cropped by browsing antelopes (Moe et al., 2009). However, the forest of the 1960s emerged from a pulse of seedling establishment in the early 1900s when the browsing ungulates were decimated by rinderpest, an exotic cattle-borne virus and the elephant population had already been virtually extirpated by commercial ivory hunters over the preceding


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half-century (Skarpe et al., 2004). The forested state of the 1960s was thus an artifact of two prior anthropogenic disturbances, so forcing the system back into that unstable state would require continual replications of those profound disturbances. The Chobe story exemplifies how a ‘problem’ (tree and seedling destruction by elephants and antelopes) perceived at one time scale (decades) could change completely and even become an opportunity (tourist revenue from charismatic megafauna at spectacular densities) when examined at another scale (centuries). Some argue that a millennial time scale is appropriate for conservation biology because the Pleistocene megafaunal extinctions could have been at least partly driven by humans, and restoring a surrogate megafauna would be beneficial for conserving ecosystem processes in North America, for example (Donlan et al., 2005). There are extant plants, such as the Kentucky coffeetree Gymnocladus dioicus, with seed dispersal systems that seem adapted for megaherbivores now extinct for at least 10 000 years (Zaya & Howe, 2009). Those relict plant species beg the thorny question: should we try to conserve them or simply declare them anachronisms and let them slowly but surely follow their coevolved dispersers into extinction? We are also challenged to conserve the last few ancient migration corridors for the very reason that they have existed for millennia. Sprawling infrastructure in Wyoming could close a migration route that has been used by a pronghorn Antilocapra americana population every year, without variation, for at least 6000 years (Berger, Cain & Murray Berger, 2006). It could be argued that allowing that to happen would be as unethical and irresponsible as allowing a shopping mall to be built over Stonehenge. There is a danger, however, in assuming it is always safest to consider conservation issues in the context of the longest


possible time scales. Obviously, humans are driving change at an ever-increasing pace so serious conservation problems can catch us by surprise. Even in the absence of overharvesting, the consistent, efficient and highly selective predation by humans on wild prey populations can cause phenotypic traits to change more than three times faster than has been observed in non-exploited taxa (Darimont et al., 2009). This and previous examples demonstrate there is no ideal time scale to use as a rule of thumb for conservation considerations. My best suggestion is to revert to the analogy of the microscope with adjustable magnification: explore the issue across a range of time scales to identify problems, opportunities and solutions.

A suggested approach: incorporate considerations of scale into conservation triage I conclude with a suggested approach for answering the question I posed in the introduction: has the perceived conservation problem been examined across multiple scales? I suggest that considerations of scale should be incorporated into the triage stage of conservation planning (Bottrill et al., 2008). Money, qualified personnel, time and political capital are among the limiting resources needed for the effective conservation of biodiversity and so priority has to be given to those conservation problems that are remediable. Such an assessment requires that each conservation problem be examined at higher and lower scales of space and time, as described, but the same multi-scale exercise can also be applied to other dimensions that influence biodiversity conservation (Fig. 3). The effectiveness of conservation is

Figure 3 Probability of success can be quickly ‘eyeballed’ by placing a planned conservation project in context with several strongly interacting variables: project effectiveness is dependent on the quality of governance (i.e. corruption level) in the country concerned; opportunity costs rise where rainfall is sufficient to allow alternative land uses; the cost efficiency of a project depends on how ‘wild’ the area is in relation to existing investments in infrastructure that might obstruct conservation goals. Hypothetical relationships are depicted here in stylized plots. Axes will have different scales for different continents, regions, species of concern and conservation threats.


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et al., 2008). Cost efficiency, measured in units of conservation success achieved per unit of financial expenditure (Lindsey et al., 2005), is low in areas with much existing infrastructure (buildings, fields, fences, roads; i.e. built capital). It increases as the ratio of natural:built capital increases, with a slight tail-off where conservation investments in pristine systems yield fewer immediate conservation gains but might secure biodiversity reserves for the future (du Toit, 2010). Exploring a perceived conservation problem at larger and smaller spatiotemporal scales might reveal that the problems or its causes are different from those initially inferred (Fig. 4). If so, then the a priori probability of success can be gauged (Fig. 3) for any planned intervention before it is triaged with further rigor (sensu Bottrill et al., 2008). If the problem and its inferred causes do not change, or if they do but an intervention is still worth implementing, then the preceding multi-scale analysis will have helped identify the appropriate institutional level(s) to engage with. If the problem is caused by perverse economic incentives (global scale), then the most appropriate level would be that of international trade organizations; if it is caused by unsustainable land use (landscape scale) then interventions should be sequenced across multiple levels of governance, from national and local to village and household (Garnett et al., 2007). This process brings the conservation plan to a stage at which adaptive management of the intervention can begin, with a feedback loop for review so that the conservation problem, its causes and prescribed action, are constantly reevaluated at multiple scales (Fig. 4). In principle, this is analogous to using a microscope to observe a specimen, with the greatest attribute of that instrument being its versatility in allowing the observer to zoom in and out.

Acknowledgments Figure 4 A simple decision diagram is suggested to ensure that perceived conservation problems are examined across multiple scales as part of the triage process, and then the appropriate institutional levels are identified for engagement before any conservation action commences. Adaptive management of the action involves back-loops for review, so new evaluations and decisions can be made as more information becomes available.

positively related to the quality of governance (or inversely related to the corruption level) in the country concerned (Smith et al., 2003). Opportunity costs of conservation increase across the rainfall gradient as competing agronomic opportunities become increasingly profitable and then decline somewhat in very high rainfall areas where soils are leached. This explains why community-based conservation projects in arid regions like Namibia are very much more successful than in moister regions like Indonesia, where nebulous payments for biodiversity cannot outcompete revenues from logging and palm oil plantations (Wunder 234

Assistance with the Kruger climate and buffalo population data were kindly provided by Paul Cross and Sandra MacFadyen, while Spencer Clayton assisted with graphics.

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