Scale Mismatches, Conservation Planning, and the Value of Social-Network Analyses
Published: Jul 17, 2015
ANGELA M. GUERRERO, RYAN R. J. MCALLISTER, JONATHAN CORCORAN, AND KERRIE A. WILSON
Article first published online: 10 JAN 2013
Conservation Biology, Volume 27, No. 1, 35–44
© 2013 Society for Conservation Biology
Many of the challenges conservation professionals face can be framed as scale mismatches. The problem of scale mismatch occurs when the planning for and implementation of conservation actions is at a scale that does not reflect the scale of the conservation problem. The challenges in conservation planning related to scale mismatch include ecosystem or ecological process transcendence of governance boundaries; limited availability of fine-resolution data; lack of operational capacity for implementation; lack of understanding of social-ecological system components; threats to ecological diversity that operate at diverse spatial and temporal scales; mismatch between funding and the long-term nature of ecological processes; rate of action implementation that does not reflect the rate of change of the ecological system; lack of appropriate indicators for monitoring activities; and occurrence of ecological change at scales smaller or larger than the scale of implementation or monitoring. Not recognizing and accounting for these challenges when planning for conservation can result in actions that do not address the multiscale nature of conservation problems and that do not achieve conservation objectives. Social networks link organizations and individuals across space and time and determine the scale of conservation actions; thus, an understanding of the social networks associated with conservation planning will help determine the potential for implementing conservation actions at the required scales. Social-network analyses can be used to explore whether these networks constrain or enable key social processes and how multiple scales of action are linked. Results of network analyses can be used to mitigate scale mismatches in assessing, planning, implementing, and monitoring conservation projects.
Discordancia de Escalas, Planificación de la Conservación y el Valor del Análisis de Redes Sociales
Muchos de los retos que enfrentan los profesionales de la conservación pueden ser catalogados como discordancia de escalas. El problema de discordancia de escalas ocurre cuando la planificación e implementación de acciones de conservación se llevan a cabo en una escala que no refleja la escala del problema de conservación. Los retos de la planificación de la conservación relacionados con la discordancia de escala incluyen el rebase de límites de gobernanza por los procesos ecológicos; la disponibilidad limitada de datos de resolución fina; la carencia de capacidad operativa para la implementación; la falta de entendimiento de los componentes socio-ecológicos del sistema; amenazas a la diversidad ecológica que operan en escalas espaciales y temporales diversas; discordancia entre el financiamiento y la naturaleza a largo plazo de los procesos ecológicos; tasa de implementación de acciones que no refleja la tasa de cambio del sistema ecológico; ausencia de indicadores apropiados de las actividades de monitoreo y ocurrencia de cambio ecológico en escalas menores o mayores que la escala de implementación o monitoreo. El no reconocimiento y consideración de estos retos al planificar la conservación puede resultar en acciones que no abordan la naturaleza multiescala de los problemas de conservación y que no se alcancen los objetivos de conservación. Las redes sociales enlazan organizaciones e individuos en el tiempo y espacio y determinan la escala de las acciones de conservación; por lo tanto, el entendimiento de las redes sociales asociadas con la planificación de la conservación ayudará a determinar el potencial para la implementación de acciones de conservación en las escalas requeridas. El análisis de redes sociales puede ser utilizado para explorar si esas redes constriñen o facilitan procesos sociales claves y como se relacionan las múltiples escalas de acción. Los resultados del análisis de redes pueden ser utilizados para mitigar la discordancia de escalas en la evaluación, planificación, implementación y monitoreo de proyectos de conservación.
Scale mismatch, also referred to as the “problem of fit,” has emerged in the literature of natural resource management and refers to a mismatch between the extent and resolution of management actions and the ecological system of interest (Lee 1993; Young 2002; Cumming et al. 2006). The problem of scale mismatch in conservation settings occurs when conservation actions are undertaken at a scale that does not reflect the scale(s) required to solve a particular conservation problem. For example, scale mismatches are a common problem in the management of migratory species (e.g., Berkes 2006) and when the relatively short time horizons of planners and politicians conflict with longer-term ecological and social changes (Folke et al. 1998b). Cumming et al. (2006) explored scale mismatch in the management of natural resources and explained its causes and consequences. They highlight that scale mismatches are generated by a wide range of social, ecological, and linked social-ecological processes and conclude that how to best resolve scale mismatches remains an open question. An understanding of how scale mismatches transpire and their likely consequences would be of value to conservation professionals because it would further the development of strategies to address problems of scale.
Conservation planning is evolving from being primarily concerned with the systematic identification of protected areas (Margules & Pressey 2000) to a process of prioritizing, implementing, and managing actions for the conservation of biological diversity and other natural resources, inside and outside protected areas (Wilson et al. 2009). Effectiveness of conservation planning is hindered by a lack of funding, support for only short-term projects, lack of consideration of ecological processes and dynamic threats that determine the persistence of biological diversity (Pressey et al. 2007), limited extent to which science and research results inform on-the-ground action (Balmford & Cowling 2006; Pressey & Bottrill 2009), unacknowledged diversity of human value systems (Wondolleck & Yafee 2000; Van Houtan 2006), and unrecognized, opposing or conflicting goals that obstruct objective decision making (Biggs et al. 2011). Many of these challenges emerge as a result of scale mismatches, primarily because conservation problems often require multiple actions that are each associated with different ecological and management scales (Sarkar et al. 2006). The problem of scale mismatch lies not in fitting conservation action to match a particular scale. Rather, the multiscale nature of conservation problems needs to be understood and negotiated so that strategies and actions are developed and applied at appropriate temporal and spatial scales. Governance and management arrangements that have the capacity to alleviate mismatches across the range of actions are therefore required. However, there is often insufficient institutional structures or mechanisms to adapt to the multiscale nature of conservation problems and effectively manage across scales (Folke et al. 1998a; Young 2002; Wyborn 2011).
Conservation planning needs to include stages dedicated to understanding the social-ecological system in which conservation actions are to be implemented, including cultural, economic, and institutional contexts (Polasky 2008; Pressey & Bottrill 2008), and the norms, values, and other human factors that underpin opportunities for and constraints on effective conservation (e.g., Cowling & Wilhelm-Rechmann 2007; Guerrero et al. 2010; Knight et al. 2010). The identification and involvement of stakeholders is key to effective conservation planning. It can facilitate the identification of new knowledge and opportunities for and barriers to implementation and engender trust and support for implementation (Pierce et al. 2005; Knight et al. 2006a; Pressey & Bottrill 2009).
Network theory has been useful for explaining social phenomena across a diversity of disciplines (Borgatti et al. 2009). Social networks link organizations and individuals across space and time and hence are critical in determining the collective scale of conservation actions, which in turn underpins the magnitude of mismatch in scale. We sought to understand challenges mismatches of scale pose to the conservation-planning process. We explored this issue across scales associated with the different stages of conservation planning. We considered emerging conservation-planning approaches that may be useful in addressing scale mismatches and how social-network analyses (SNA) can be applied to the management of scale-mismatch problems.
Scale Mismatches and Conservation Planning
Planning and implementing conservation actions (Fig. 1) involves definition of the conservation problem, formulation of actions, and determination of how action will be implemented. Conservation problems are often complex and involve competing objectives, multiple actors, and a diversity of possible conservation actions. Decisions can be made at spatial and temporal scales that may not match the scale of the ecological patterns or processes relevant to the conservation problem, a situation that creates a scale mismatch. For example, actions and strategies may be formulated at a regional scale but the conservation problem also requires action at a finer scale (Briggs 2001; Sarkar et al. 2006), or a plan may be formulated at an appropriate scale for action, but the operational capacity for implementation may be lacking.
We applied a modified version of Cumming et al.'s (2006) classification of scale mismatches (spatial, temporal, and functional scale mismatches) to show how scale mismatch manifests itself in diverse ways and at each stage in the conservation-planning process (i.e., problem assessment, strategy and action formulation, and plan implementation, evaluation, and adaptation) (Table 1). Spatial scale mismatches occur when the geographic extent of the solution differs greatly from the geographic extent of the problem. For example, when conservation action is applied at a fine scale, such as vegetation patches, but the problem prevails at a broader scale, such as the landscape scale (Cash et al. 2006). Temporal scale mismatches relate to processes that occur over different time scales (Cash et al. 2006). Both temporal and spatial scales also have grain, which is the resolution at which observations are made (i.e., data resolution). Functional scale mismatches occur when the scope of processes considered for solving the conservation problem differs greatly from the scope of processes within the system associated with, or that affect, the conservation problem (Lee 1993; Folke et al. 1998b). For example, a very narrow focus on a few ecological features compared with a broad focus on many ecosystem processes.
*Stages of project development and implementation (see Fig. 1).
In conservation planning, one of the first tasks is to define the extent of the planning region. In some instances, regions are defined solely on the basis of institutional boundaries without accounting for ecological boundaries (Table 1). This can result in a plan that addresses only part or none of the conservation problem. For over 100 years the Murray-Darling Basin, one of the most important river systems in Australia, has provided water for irrigation, livestock, and industry, and domestic use across 4 Australian states. Increased water diversion fueled by the expansion of irrigation in the basin resulted in a 40% reduction in water flow (Cosler et al. 2010). As a result, ecosystems collapsed and native fishes, riparian vegetation, and wetlands of national significance have been negatively affected. Diverse but unconnected institutions (e.g., governments of different states) have attempted to repair water flow, and these efforts have lead to a lack of effective governance of the basin as a whole. This is an example of a mismatch of spatial scale; the planning region did not reflect the boundaries of the ecological systems of the basin and instead encompassed areas of the basin occurring in each state. Linked to this spatial mismatch was a functional mismatch in which the full scope of features and ecological processes (e.g., patterns of river flow, condition of wetlands) occurring across the basin were not accounted for (Murray–Darling Basin Authority 2012). More recently, attempts to manage these scale mismatches include creation of institutions operating at a federal level (e.g., Commonwealth Water Act of 2007) and formation of the Murray-Darling Basin Authority. The authority is responsible for the formulation of an integrated management plan to set water-diversion limits for the entire basin (Commonwealth Water Act of 2007) and for the development of specific conservation programs in conjunction with state governments (e.g., Rivers Environmental Restoration program). When identifying areas for conservation action, decisions about data resolution affect which and how many areas are selected (Pressey & Logan 1995; Rouget 2003). A spatial scale mismatch can occur when the resolution of the data used to understand the ecological and social setting fails to reflect the heterogeneity of the area (Table 1), which can limit the effectiveness of planning decisions (e.g., Rouget 2003). The limited availability of fine-resolution data across a planning region and limited resources for acquiring new data (Margules et al. 2002) result in the use of coarse-resolution data (Mills et al. 2010).
Most spatial conservation-planning exercises involve representation of species diversity patterns, but relatively few consider ecological processes or dynamic threats to biological diversity (Pressey et al. 2007; Pressey & Bottrill 2009). Lack of consideration of key ecological processes that sustain biological diversity at the assessment stage can lead to functional mismatches (i.e., failure of conservation actions prevent disruption of these processes) (Pressey et al. 2007).
Formulation of Actions and Strategies
When conservation actions are not formulated at appropriate scales, the social-ecological components of the system that affect the success of conservation actions (e.g., institutional barriers, cultural practices, livelihood activities) may not be accounted for. An example of scale mismatch is when actions are formulated at a particular governance level, such as a state or county, but are applied to an ecosystem or ecological process that transcends governance boundaries. For instance, actions may be developed for species that migrate across countries but may not be developed for species that migrate within a country (e.g., Gilmore et al. 2007). In the United Kingdom, regulations on recreational use of inland waters are based on short-term behavioral responses of birds to disturbance that are averaged across sites and habitats (O'Connell et al. 2007). This generalized approach to planning does not account for site- and time-specific human disturbances and results in spatial and temporal mismatches. For example, human activity may only occur at particular times of the year or in specific locations and birds may use different lakes for different purposes (O'Connell et al. 2007).
Threats to biological diversity operate at diverse spatial and temporal scales. Therefore effective conservation planning requires the scheduling of multiple actions that can operate at these diverse scales. Some actions may need to be threat specific (Salafsky et al. 2002; Pressey et al. 2007) to address relevant ecological processes such as those associated with connectivity, population dynamics in fragments, and maintenance of patch dynamics (Carwardine et al. 2008) and thereby ameliorate the potential for mismatches of functional scale.
Implementation and Management
The need for more effective implementation of conservation actions is recognized as a key challenge in conservation planning (Balmford & Cowling 2006; Knight et al. 2008; Pressey & Bottrill 2009). Many of the challenges of implementation stem from a disjointed planning process in which early stages are not integrated into a broader planning framework that focuses on implementation. This occurs, for example, when spatial prioritization analyses do not account for implementation constraints and opportunities (Pierce et al. 2005; Knight et al. 2008) or when planning units used in the prioritization of areas are dissimilar to areas where management will be implemented, which makes it difficult to translate plans into actions (Pierce et al. 2005).
Spatial scale mismatches in implementation lead to actions occurring at scales that do not resolve the conservation issue (Table 1). Spatial scale mismatch is sometimes driven by a lack of resources for implementation or occurs because key organizations or individuals were not engaged in the planning process (e.g., Waudby et al. 2007). Conservation of Australia's endangered bridled nailtail wallaby (Onychogalea fraenata) (Commonwealth Environment Protection and Biodiversity Conservation Act of 1999) consisted of a centralized state program that was not effectively implemented at a local scale or over a long period; thus, subpopulations could not be maintained and the program failed to stop the decline of the species (Kearney et al. 2012).
Temporal scale mismatches at the implementation stage occur, for example, when funding does not match the long-term nature of ecological processes relevant to the conservation problem. This mismatch results in partially attained conservation objectives (e.g., Waudby et al. 2007). Temporal scale mismatches can also occur when actions are implemented at a rate that does not reflect the rate of change of the ecological system of interest, for example when actions are delayed due to political timeframes or in the pursuit of scientific certainty (e.g., Grantham et al. 2009).
Another temporal scale mismatch occurs when the same stakeholders cannot be involved throughout the planning and implementation processes (Pierce et al. 2005; Walters 2007; Pressey & Bottrill 2009). Implementation is an incremental and often lengthy process that requires the long-term participation of stakeholders so that plans can be adapted to reflect changes in ecological and social systems (Pierce et al. 2005; Pressey & Bottrill 2009; Grantham et al. 2010) (e.g., changes in areas of interest, new data on threats and species diversity, changes in funding, and changes in interests of local communities) (Pressey & Bottrill 2009). Consistent participation of stakeholders facilitates the adoption of plans into the regular activities of organizations responsible for planning and development (Pressey & Bottrill 2009). Some conservation plans account for this temporal mismatch by ensuring long-term involvement of stakeholders (e.g., Green et al. 2009; Henson et al. 2009).
Plan Monitoring and Adaptation
Monitoring is key to evaluating outcomes, and it facilitates learning and adaptive management (Stem et al. 2005; Ferraro & Pattanayak 2006; Field et al. 2007; Lindenmayer & Likens 2010). Scale mismatches at the monitoring and adaptation stage of a plan manifest themselves when ecological changes occur at scales smaller or larger (or longer or shorter) than the scale of monitoring operations and thus go undetected (Table 1). Such mismatches limit ones’ ability to respond to change, which can limit effectiveness of adaptive management.
To monitor conservation outcomes one must decide which ecological metrics to use, where to conduct monitoring, and the duration and frequency of monitoring (Spellerberg 1994; Lindenmayer & Likens 2010). These decisions can result in spatial, temporal, or functional scale mismatches. For example, choosing appropriate indicators (Lambeck 1997; Carignan & Villard 2002; Tulloch et al. 2011) is an uncertain process that may result in indicators that do not provide a whole-system view of the problem (Simberloff 1998) and may not account for multiscale requirements of the species or ecological features for which the indicator is assumed to be a surrogate (Lindenmayer et al. 2002). Insufficient data, cost of monitoring, and the potential difficulties of applying the most appropriate indicator (Tulloch et al. 2011) are obstacles related to the problem of scale mismatch (Lindenmayer et al. 2002; Lindenmayer & Likens 2010).
Addressing Scale Mismatch in Conservation Planning Emerging Approaches
For conservation planning to operate at diverse spatial, functional, and temporal scales, conservation practitioners need to apply tools that account for the multiscalar nature of conservation problems. Planning approaches that account for functional scale mismatches in the problem-assessment and plan-formulation stages are emerging. For example, Pressey et al. (2007) discuss approaches for planning for physical and biological processes that require management over large areas or of areas with unique topography. Such approaches include conservation areas that may change in extent and location over time, variable representation targets, and the use of specific design criteria (e.g., Briers 2002; Nicholson et al. 2006; Leroux et al. 2007). Threats are also being considered in, for example, the scheduling of actions so that threatened areas or species are given priority and areas with nonabatable threats are avoided (e.g., Burgman et al. 2001; Game et al. 2008) and through the explicit consideration of the effects of multiple threats (e.g., Evans et al. 2011). New developments in conservation planning may address spatial and temporal mismatches inherent in more traditional planning methods, which account only for static views of the ecological, human, and social characteristics of an area. New methods balance divergent priorities at multiple spatial scales (Moilanen & Arponen 2011) and prioritize actions over time in the face of dynamic threats, uncertainty, and changing costs of activities (Costello & Polasky 2004; Meir et al. 2004; Wilson et al. 2006).
These new quantitative planning methods are useful for addressing scale mismatches that arise during problem assessment and action formulation stages of the planning process (Fig. 1). During these stages species diversity and other biological data are compiled, conservation targets are set, and priority conservation areas or actions are identified (Margules & Pressey 2000). However, scale mismatches at the implementation and monitoring and adaptation stages can still transpire. In addition, the need to embed quantitative planning methods in a social process that facilitates effective implementation is increasingly being recognized (Knight et al. 2006a; Pressey & Bottrill 2009; Reyers et al. 2010) and acted on (e.g., Pierce et al. 2005; Knight et al. 2006b; Game et al. 2010). It is therefore timely to explore tools and approaches that can help deal with scale mismatches that impede effective implementation.
SNA may provide guidance on how implementation might be approached in the management of problems of scale mismatch. Some authors suggest integrating ecological assessments with social assessments of a region (Cowling & Wilhelm-Rechmann 2007) to facilitate an understanding of the social-ecological effects on valued nature and of the opportunities for and constraints to implementation. Such social assessments could include an examination of the social networks that exist to determine key people affecting conservation outcomes (either through their involvement with conservation activities, or with economic, subsistent and other types of activities that have a direct effect on conservation outcomes); how the people involved are connected to each other through partnerships for action or other types of collaborations (e.g., Prell et al. 2009; Vance-Borland & Holley 2011); and what spatial, temporal, or functional scales of operation or influence these partnerships have. Social-network theory can be used to characterize networks of collaborations and social relations and to facilitate multiscalar conservation. For example, SNA can be used to determine the links between actors (individuals, groups, or organizations) that could be used to promote cooperation and coordination of key activities at particular and required scales of action (e.g., Gass et al. 2009).
We define conservation social networks as the networks of relationships that link actors involved in conservation activities across space. These networks are the basis of social norms and community learning; hence, they also link actors across time. Networks can be formal or informal. Informal networks are present where conservation action is to occur (e.g., a group of citizens concerned about specific issues) (e.g., Newman & Dale 2007; Vance-Borland & Holley 2011) and take many forms, for example, farmer advice networks (e.g., Isaac et al. 2007). Formal networks (e.g., Carlsson & Sandstrom 2008) are formed during the conservation-planning process through the establishment of formal agreements or partnerships, for example between nongovernmental organizations or government agencies, around a particular conservation objective (e.g., Bode et al. 2010). The different patterns of interactions among actors in a network give rise to different network structures (Borgatti & Foster 2003) that can inhibit or enable social processes that are often needed in conservation planning, such as cooperation, knowledge generation, learning, and conflict resolution (e.g., Hahn et al. 2006; Olsson et al. 2007; Bodin & Crona 2009). SNA are used to analyze the behavior of actors in a network on the basis of its structure (i.e., pattern of relations) (Emirbayer & Goodwin 1994). For example, one can study the density of ties within a network (extent to which all actors are connected) to understand the capacity of integration and sharing of knowledge within that network (Bodin & Crona 2009), whereas the level of fragmentation of a network (presence or lack of presence of distinct subgroups) can be useful for understanding capacity for collaboration within the network (Granovetter 1973) and access to new knowledge (Newman & Dale 2007; Bodin & Crona 2009). Structural analyses of conservation social networks can help inform implementation strategies. For example, a network that is connected through a few key actors (Fig. 2a) may indicate the best strategy is to engage with these few key actors so that they can then coordinate action through their own networks. Alternatively, a network that is highly fragmented (Fig. 2b) may require engagement with many different actors and thus a greater financial investment at the implementation stage.
Analyzing network structures can help in the understanding of the degree to which multiple scales of action are linked or being coordinated. For example, network analyses can be used to identify bridging actors (e.g., Olsson et al. 2007), or scale-crossing brokers, who link those operating at different scales who would otherwise be disconnected (Bodin et al. 2006). They can also help identify different subgroups of actors in the network that are related to particular required scales of action and thus could drive implementation at those particular scales. In work to recover the endangered Australian Glossy Black-Cockatoo (Calyptorhynchus lathami) (Environment Protection and Biodiversity Conservation Act 1999) a variety of agencies, community groups, landowners, and volunteers operating at different scales cooperated effectively to implement actions required for the persistence of this species (Waudby et al. 2007). Although, to our knowledge, a social-network analysis was not performed as part of this recovery plan, this example shows how identification and engagement of key groups associated with different scales of action could play a key role in the success of the project. The use of SNA to identify stakeholders allows for a targeted approach to stakeholder selection (Prell et al. 2009).
Results of SNA may be most useful when they are combined with other information about the social-ecological system. It is useful to understand not only how each actor relates to others, but also how they relate to the ecological features of interest (Fig. 2c) (Janssen et al. 2006). For example, different fishers harvest different fish species at different fishing locations, and some of those species and locations will be of greater importance for achieving conservation outcomes. As well as identifying key actors who can connect to all other relevant actors—and other scales, it is also important to identify those actors connected to the most important ecological features. Such connections enable the targeting of actions to the most appropriate spatial scales.
There are other benefits of applying SNA in conservation planning. Engagement of stakeholders is an expensive process and SNA can help minimize costs by identifying either well-connected actors or actors linked to others who could prove difficult or costly to engage with directly (e.g., Prell et al. 2009). It can also be used to identify actors who could help maximize understanding of the system's complexity because of their connections to actors who hold different types of knowledge. It may also help uncover particular collaboration gaps that if addressed might connect key groups or actors who could collectively enhance conservation success (Vance-Borland & Holley 2011).
Structural analyses of networks can provide insights into how social networks affect planned outcomes by enabling or constraining key social processes needed in the planning and implementation of conservation actions. However, acquiring a deep appreciation of the role of social networks likely requires not only an understanding of structural aspects, such as the presence or absence of links between two or more key actors or groups, but also information on the value or effectiveness of such links. For example, engaging an actor that is well connected to many other actors operating at different scales (a structural characteristic) may not be beneficial if that actor is not trusted by other actors (e.g., Gass et al. 2009), if the actor lacks legitimacy (Tyler 2006), if the actor's presence in the network over time is uncertain (McAllister et al. 2008), or if cultural, institutional, and other contextual aspects affect the actor's willingness to participate (e.g., Bodin & Crona 2008).
Strategic decisions made at the onset of a conservation project can be informed by an understanding of some of the challenges that can arise during the process of development and implementation of conservation actions, which include potential mismatches in spatial, temporal, and functional scales. We discussed how scale mismatches can manifest at each stage of the conservation-planning process and can lead to a plan that does not account for the threats, risks, constraints, and opportunities posed by the complexities and dynamics of the social-ecological system or a plan that cannot be implemented fully or partially. In addition, scale mismatches can also affect the adaptive capacity of conservation institutions during project development and implementation because the institution's ability to detect, and therefore learn from, ecological changes occurring at scales other than the scale of operation is impeded.
An understanding of how these scale mismatches manifest themselves at various stages of project development and implementation can be used to predict the likelihood of success of conservation initiatives. Anticipating the potential for scale mismatches can inform the conservation-planning process so that mismatch problems can be dealt with effectively. Strategies to avoid scale mismatches may involve a spectrum of alternatives, where at one extreme the mismatch is addressed and strategies and actions are developed and applied at temporal and spatial scales that are appropriate for the problem, and at the other extreme the mismatch is not addressed and the likelihood (however reduced) that some positive conservation outcomes may transpire is hoped for. Whether scale mismatches are addressed may depend on the resources available, on competing considerations that shape decisions about scale (Mills et al. 2010), and on the viability of strategies and actions that could address the mismatch.
The importance of social networks to solving conservation problems is associated with the nature of environmental problems. Environmental problems are multiscaled and constantly evolving. Thus, solutions to these problems require local actors to have connections to broad levels of society (and vice versa) and require a flexible and open process. Solving environmental problems also requires transdisciplinary processes involving experts, government, and local stakeholders (Newman & Dale 2007). We have shown how SNA can be applied to conservation planning to improve the effectiveness of conservation action. Specifically we showed how it can be used to help conservation actions be applied at the required spatial, temporal, and functional scales.
We are thankful for feedback on the manuscript from R. Loyola and 3 anonymous reviewers. We also thank the Australian Government's National Environmental Research Program, the Australian Research Council Centre of Excellence for Environmental Decisions, and CSIRO's Climate Adaptation Flagship for funding and support.
Balmford, A., and R. Cowling. 2006. Fusion or failure? The future of
conservation biology. Conservation Biology 20:692–695.
Berkes, F. 2006. From community-based resource management to complex
systems: the scale issue and marine commons. Ecology and Society
(accessed October 2012).
Biggs, D., N. Abel, A. T. Knight, A. Leitch, A. Langston, and N. C. Ban.
2011. The implementation crisis in conservation planning: Could
“mental models” help? Conservation Letters 4 DOI: 10.1111/j.1755-
Bode, M., W. Probert, W. R. Turner, K. A. Wilson, and O. Venter. 2010.
Conservation planning with multiple organizations and objectives.
Conservation Biology 25:295–304.
Bodin, O., B. Crona, and H. Ernstson. 2006. Social networks in
natural resource management: What is there to learn from a
structural perspective? Ecology and Society 11(2): http://www.
ecologyandsociety.org/vol11/iss2/resp2/ (accessed October 2012).
Bodin, O., and B. I. Crona. 2008. Management of natural resources at the
community level: Exploring the role of social capital and leadership
in a rural fishing community. World Development 36:2763–2779.
Bodin, O., and B. I. Crona. 2009. The role of social networks in natural
resource governance: What relational patterns make a difference?
Global Environmental Change-Human and Policy Dimensions
Borgatti, S. P., and P. C. Foster. 2003. The network paradigm in organizational
research: a review and typology. Journal of Management
Borgatti, S. P., A. Mehra, D. J. Brass, and G. Labianca. 2009. Network
analysis in the social sciences. Science 323:892–895.
Briers, R. A. 2002. Incorporating connectivity into reserve selection
procedures. Biological Conservation 103:77–83.
Briggs, B. S. V. 2001. Linking ecological scales and institutional
frameworks for landscape rehabilitation. Ecological Management
& Restoration 2:28–35.
Burgman, M. A., H. P. Possingham, A. J. J. Lynch, D. A. Keith, M. A.
McCarthy, S. D. Hopper, W. L. Drury, J. A. Passioura, and R. J.
Devries. 2001. A method for setting the size of plant conservation
target areas. Conservation Biology 15:603–616.
Carignan, V., and M. A. Villard. 2002. Selecting indicator species to
monitor ecological integrity: a review. Environmental Monitoring
and Assessment 78:45–61.
Carlsson, L., and A. Sandstrom. 2008. Network governance of the commons.
International Journal of the Commons 2:33–54.
Carwardine, J., C. J. Klein, K. A. Wilson, R. L. Pressey, and H. P.
Possingham. 2008. Hitting the target and missing the point: targetbased
conservation planning in context. Conservation Letters 2 DOI:
Cash, D. W., W. N. Adger, F. Berkes, P. Garden, L. Lebel, P. Olsson,
L. Pritchard, and O. Young. 2006. Scale and cross-scale dynamics:
governance and information in a multilevel world. Ecology
and Society 11(2): http://www.ecologyandsociety.org/
vol11/iss2/art8/ (accessed October 2012).
Cosler, P., T. Flannery, R. Harding, D. Karoly, H. Possingham, R.
Purves, D. Saunders, B. Thom, J. Williams, and M. Young. 2010.
Sustainable diversions in the Murray-Darling Basin. Wentworth
Group of Concerned Scientists, Sydney. Available from http://www.
darling-basin (accessed August 2012).
Costello, C., and S. Polasky. 2004. Dynamic reserve site selection. Resource
and Energy Economics 26:157–174.
Cowling, R. M., and A. Wilhelm-Rechmann. 2007. Social assessment as
a key to conservation success. Oryx 41:135–136.
Cumming, G. S., D. H. M. Cumming, and C. L. Redman. 2006. Scale mismatches
in social-ecological systems: causes, consequences, and solutions.
Ecology and Society 11(1): http://www.ecologyandsociety.
org/vol11/iss1/art14/ (accessed October 2012).
Emirbayer, M., and J. Goodwin. 1994. Network Analysis, culture, and
the problem of agency. American Journal of Sociology 99:1411–
Evans, M. C. E. M. C., H. P. Possingham, and K. A. Wilson. 2011. What
to do in the face of multiple threats? Incorporating dependencies
within a return on investment framework for conservation. Diversity
and Distributions 17:437–450.
Ferraro, P. J., and S. K. Pattanayak. 2006. Money for nothing? A call for
empirical evaluation of biodiversity conservation investments. Public
Library of Science Biology 4 DOI: 10.1371/journal.pbio.0040105.
Field, S. A., P. J. O’Connor, A. J. Tyre, and H. P. Possingham. 2007.
Making monitoring meaningful. Austral Ecology 32:485–491.
Folke, C., F. Berkes, and J. Colding. 1998a. Ecological practices and
social mechanisms for building resilience and sustainability. Pages
414–436 in F. Berkes and C. Folke, editors. Linking social and ecological
systems. Cambridge University Press, London.
Folke, C., L. Pritchard, F. Berkes, J. Colding, and U. Svedin. 1998b. The
problem of fit between ecosystems and institutions. IHDP working
paper 2. International Human Dimensions Programme on Global
Environmental Change (IHDP), Bonn, Germany.
Game, E. T., G. Lipsett-Moore, R. Hamilton, N. Peterson, J. Kereseka,W.
Atu, M. Watts, and H. Possingham. 2010. Informed opportunism for
conservation planning in the Solomon Islands (letter). Conservation
Letters DOI: 10.1111/j.1755-263X.2010.00140.x.
Game, E. T., M. E. Watts, S. Wooldridge, and H. P. Possingham. 2008.
Planning for persistence in marine reserves: a question of catastrophic
importance. Ecological Applications 18:670–680.
Gass, R., M. Rickenbach, L. Schulte, and K. Zeuli. 2009. Cross-boundary
coordination on forested landscapes: investigating alternatives for
implementation. Environmental Management 43:107–117.
Gilmore, S., B. Mackey, and S. Berry. 2007. The extent of dispersive
movement behaviour in australian vertebrate animals, possible
causes, and some implications for conservation. Pacific Conservation
Granovetter, M. 1973. Strength of weak ties. American Journal of Sociology
Grantham, H. S., M. Bode, E. McDonald-Madden, E. T. Game, A. T.
Knight, and H. P. Possingham. 2010. Effective conservation planning
requires learning and adaptation. Frontiers in Ecology and the
Grantham, H. S., K. A.Wilson, A. Moilanen, T. Rebelo, and H. P. Possingham.
2009. Delaying conservation actions for improved knowledge:
How long should we wait? Ecology Letters 12:293–301.
Green, A., et al. 2009. Designing a resilient network of marine protected
areas for Kimbe Bay, Papua New Guinea. Oryx 43:488–498.
Guerrero, A. M., A. T. Knight, H. S. Grantham, R. M. Cowling,
and K. A. Wilson. 2010. Predicting willingness-to-sell and its utility
for assessing conservation opportunity for expanding protected
area networks. Conservation Letters 3 DOI: 10.1111/j.1755-
Hahn, T., P. Olsson, C. Folke, and K. Johansson. 2006. Trust-building,
knowledge generation and organizational innovations: the role of
a bridging organization for adaptive comanagement of a wetland
landscape around Kristianstad, Sweden. Human Ecology 34:573–
Henson, A., D. Williams, J. Dupain, H. Gichohi, and P. Muruthi. 2009.
The Heartland conservation process: enhancing biodiversity conservation
and livelihoods through landscape-scale conservation planning
in Africa. Oryx 43:508–519.
Isaac, M. E., B. H. Erickson, S. J. Quashie-Sam, and V. R. Timmer. 2007.
Transfer of knowledge on agroforestry management practices: the
structure of farmer advice networks. Ecology and Society 12(2):
Janssen, M. A., O. Bodin, J. M. Anderies, T. Elmqvist, H. Ernstson,
R. R. J. McAllister, P. Olsson, and P. Ryan. 2006. Toward a network
perspective of the study of resilience in social-ecological systems.
Ecology and Society 11(1): http://www.ecologyandsociety.
Kearney, F., R. R. J. McAllister, and N. D. MacLeod. 2012. Conservation
and grazing in Australia’s north-east: the bridled nailtail
wallaby. Pastoralism: Research, Policy and Practice 2:20.
http://www.pastoralismjournal.com (accessed September 2012).
Knight, A. T., R. M. Cowling, and B.M. Campbell. 2006a. An operational
model for implementing conservation action. Conservation Biology
Knight, A. T., R. M. Cowling, M. Difford, and B. M. Campbell. 2010. Mapping
human and social dimensions of conservation opportunity for
the scheduling of conservation action on private land. Conservation
Knight, A. T., R. M. Cowling, M. Rouget, A. Balmford, A. T. Lombard,
and B.M. Campbell. 2008. Knowing but not doing: selecting priority
conservation areas and the research–implementation gap. Conservation
Knight, A. T., et al. 2006b. Designing systematic conservation assessments
that promote effective implementation: best practice from
south africa. Conservation Biology 20:739–750.
Lambeck, R. J. 1997. Focal Species: a multi-species umbrella for nature
conservation. Conservation Biology 11:849–856.
Lee, K. N. 1993. Greed, scale mismatch, and learning. Ecological Applications
Leroux, S. J., F. K. A. Schmiegelow, S. G. Cumming, R. B. Lessard, and
J. Nagy. 2007. Accounting for system dynamics in reserve design.
Ecological Applications 17:1954–1966.
Lindenmayer, D. B., and G. E. Likens. 2010. The science and application
of ecological monitoring. Biological Conservation 143:1317–1328.
Lindenmayer, D. B., A. D. Manning, P. L. Smith, H. P. Possingham, J.
Fischer, I. Oliver, and M. A. McCarthy. 2002. The focal-species approach
and landscape restoration: a critique. Conservation Biology
Margules, C. R., and R. L. Pressey. 2000. Systematic conservation planning.
Margules, C. R., R. L. Pressey, and P. H. Williams. 2002. Representing
biodiversity: data and procedures for identifying priority areas for
conservation. Journal of Biosciences 27:309–326.
McAllister, R. R. J., B. Cheers, T. Darbas, J. Davies, C. Richards, C. J.
Robinson, M. Ashley, D. Fernando, and Y. T. Maru. 2008. Social
networks in arid Australia: a review of concepts and evidence.
Rangeland Journal 30:167–176.
Meir, E., S. Andelman, and H. P. Possingham. 2004. Does conservation
planning matter in a dynamic and uncertain world? Ecology Letters
Mills, M., R. L. Pressey, R. Weeks, S. Foale, and N. C. Ban. 2010. A
mismatch of scales: challenges in planning for implementation of
marine protected areas in the Coral Triangle. Conservation Letters
Moilanen, A., and A. Arponen. 2011. Administrative regions in conservation:
balancing local priorities with regional to global preferences
in spatial planning. Biological Conservation 144:1719–1725.
Murray—Darling Basin Authority. 2012. Proposed basin plan. Commonwealth
of Australia, Canberra.
Newman, L., and A. Dale. 2007. Homophily and agency: creating effective
sustainable development networks. Environment, Development
and Sustainability 9:79–90.
Nicholson, E., M. I. Westphal, K. Frank, W. A. Rochester, R. L. Pressey,
D. B. Lindenmayer, and H. P. Possingham. 2006. A new method
for conservation planning for the persistence of multiple species.
Ecology Letters 9:1049–1060.
O’Connell, M. J., R. M. Ward, C. Onoufriou, I. J. Winfield, G. Harris, R.
Jones, M. L. Yallop, and A. F. Brown. 2007. Integrating multi-scale
data to model the relationship between food resources, waterbird
distribution and human activities in freshwater systems: preliminary
findings and potential uses. Ibis 149:65–72.
Olsson, P., C. Folke, V. Galaz, T. Hahn, and L. Schultz. 2007. Enhancing
the fit through adaptive co-management: creating and
maintaining bridging functions for matching scales in the Kristianstads
Vattenrike Biosphere Reserve, Sweden. Ecology and Society
(accessed October 2012).
Pierce, S. M.,R.M.Cowling,A.T.Knight, A. T. Lombard, M. Rouget, and
T. Wolf. 2005. Systematic conservation planning products for landuse
planning: interpretation for implementation. Biological Conservation
Polasky, S. 2008. Why conservation planning needs socioeconomic
data. Proceedings of the National Academy of Science 105:6505–
Prell, C., K. Hubacek, and M. Reed. 2009. Stakeholder analysis and social
network analysis in natural resource management. Society & Natural
Pressey, R. L., and M. C. Bottrill. 2008. Opportunism, threats, and the
evolution of systematic conservation planning. Conservation Biology
Pressey, R. L., and M. C. Bottrill. 2009. Approaches to landscape- and
seascape-scale conservation planning: convergence, contrasts and
challenges. Oryx 43:464–475.
Pressey, R. L., M. Cabeza, M. E. Watts, R. M. Cowling, and K. A. Wilson.
2007. Conservation planning in a changingworld. Trends in Ecology
& Evolution 22:583–592.
Pressey, R. L., and V. S. Logan. 1995. Reserve coverage and requirements
in relation to partitioning and generalization of land classes: analyses
for western New South Wales. Conservation Biology 9:1506–1517.
Reyers, B., D. J. Roux, R. M. Cowling, A. E. Ginsburg, J. L. Nel, and P. O.
Farrell. 2010. Conservation planning as a transdisciplinary process.
Conservation Biology 24:957–965.
Rouget, M. 2003. Measuring conservation value at fine and broad scales:
implications for a diverse and fragmented region, the Agulhas Plain.
Biological Conservation 112:217–232.
Salafsky, N., R. Margoluis, K. H. Redford, and J. G. Robinson. 2002.
Improving the practice of conservation: a conceptual framework
and research agenda for conservation science. Conservation Biology
Sarkar, S., et al. 2006. Biodiversity conservation planning tools: present
status and challenges for the future. Pages 123–159 in Annual Review
of Environment and Resources. Annual Reviews, Palo Alto,
Simberloff, D. 1998. Flagships, umbrellas, and keystones: Is singlespecies
management passe in the landscape era? Biological Conservation
Spellerberg, I. F. 1994. Monitoring ecological change. Cambridge University
Press, Cambridge, United Kingdom.
Stem, C., R. Margoluis, N. Salafsky, and M. Brown. 2005. Monitoring
and evaluation in conservation: a review of trends and approaches.
Conservation Biology 19:295–309.
Tulloch, A., H. P. Possingham, and K. Wilson. 2011. Wise selection
of an indicator for monitoring the success of management actions.
Biological Conservation 144:141–154.
Tyler, T. R. 2006. Psychological perspectives on legitimacy and legitimation.
Pages 375–400 in Annual review of psychology. Annual
Reviews, Palo Alto, California.
Van Houtan, K. S. 2006. Conservation as virtue: a scientific and social
process for conservation ethics. Conservation Biology 20:1367–
Vance-Borland, K., and J. Holley. 2011. Conservation stakeholder network
mapping, analysis, and weaving. Conservation Letters 4 DOI:
Walters, C. J. 2007. Is adaptive management helping to solve fisheries
problems? Ambio 36:304–307.
Waudby, H., T. How, D. Frazer, and C. Obst. 2007. South Australian
recovery plan review 2007: findings, patterns and recommendations.
Report. Federal Department for Environment and Heritage,
Wilson, K. A., J. Carwardine, and H. P. Possingham. 2009. Setting conservation
priorities. Year in Ecology and Conservation Biology 2009
Wilson, K. A., M. F. McBride, M. Bode, and H. P. Possingham. 2006.
Prioritizing global conservation efforts. Nature 440:337–240.
Wondolleck, J. M., and S.L. Yaffee. 2000. Making collaboration work:
Lessons from innovation in natural resource management. Island
Press, Washington, D.C.
Wyborn, C. 2011. Landscape scale ecological connectivity: Australian
survey and rehearsals. Pacific Conservation Biology 17:121–131.
Young, O. R. 2002. The institutional dimensions of environmental
change: fit, interplay, and scale. MIT Press, Cambridge, Massachusetts.