Reforming Watershed Restoration: Science in Need of Application and Applications in Need of Science
Coastal and inland waters are continuing to decline in many parts of the world despite major efforts made to restore them. This is due in part to the inadequate role that ecological science has played in shaping restoration efforts. A significant amount of fundamental ecological knowledge dealing with issues such as system dynamics, state changes, context-dependency of ecological response, and diversity is both under-used by managers and practitioners and under-developed by ecologists for use in real-world applications. Some of the science that is being ‘used’ has not been adequately tested. Thus, restoration ecology as a science and ecological restoration as a practice are in need of reform. I identify five ways in which our ecological knowledge should be influencing restoration to a far greater extent than at present including a need to: shift the focus to restoration of process and identification of the limiting factors instead of structures and single species, add ecological insurance to all projects, identify a probabilistic range of possible outcomes instead of a reference condition, expand the spatial scale of efforts, and apply hierarchical approaches to prioritization. Prominent examples of restoration methods or approaches that are commonly used despite little evidence to support their efficacy are highlighted such as the use of only structural enhancements to restore biodiversity. There are also major gaps in scientific knowledge that are of immediate need to policy makers, managers, and restoration practitioners including: predictive frameworks to guide the restoration of ecological processes, identification of social-ecological feedbacks that constrain ecosystem recovery and data to support decisions of where and how to implement restoration projects to achieve the largest gains. I encourage ecologists to respond to the demand for their scientific input so that restoration can shift from an engineering-driven process to a more sustainable enterprise that fully integrates ecological processes and social science methods.
KeywordsWatershed Restoration Ecosystem Ecological theory Species diversity Heterogeneity Scale
Throughout much of the world, people have relied on the network of freshwater tributaries and tidal waters for recreation, clean drinking water, agricultural production, hydropower, and fisheries. Unfortunately, human activities have impacted watersheds worldwide such that countless rivers are polluted and many coastal oceans have degraded habitat or reduced fisheries. This is despite the tremendous efforts that have been devoted to managing and restoring watersheds and estuaries. Significant investments have been made in improving environmental conditions in waterways as diverse as those in the California Delta region of the United States, the Baltic region in Europe, the subcatchments of Moreton Bay in Australia, and the Mekong River Basin in Asia (Dudgeon 2005; Lotze et al. 2006; Bunn et al. 2007; CalFed 2008).
Why have rivers and coastal waters such as these continued to degrade despite large investments? A variety of responses to that question are commonly heard. Continued development and land use change are certainly contributing factors. Population growth and the movements of people typically lead to elevated pollutant loads and an increase in water extractions or diversions (Postel and Richter 2003; Walsh et al. 2005). Over-harvesting of living resources, particularly of keystone species such as bivalves, has also been blamed for the decline of a number of coastal watersheds (Grabowski and Peterson 2007). More recently, this decline has been attributed to narrow management approaches that do not rely on learning and adapting policies in an iterative fashion or that do not apply the precautionary principle (Boesch 2006). As Duarte and Conley (2008) point out, some of the blame should be placed on the scientific community, who in their desire to present a clear enough message to prompt management actions have probably over-simplified ecosystem dynamics. In addition to these explanations, I assert that current approaches to restoration are simply not sufficient and the need for input from scientists has never been greater.
I argue for a major reform in watershed restoration because much of the ecological science that has been produced is not being applied to restore aquatic ecosystems, some of the science that is being ‘used’ has not been adequately tested and scientists have been slow to produce use-inspired knowledge. By the latter, I mean ecological knowledge generated in response to the immediate needs of restoration practitioners—i.e. what Sarewitz and Pielke (2007) would call demand-driven science.
I outline elements of a more rigorous approach to restoration that requires scientists to identify when and how fundamental ecological knowledge is under-utilized, to scope out and test common restoration practices that have been inadequately evaluated or are based on faulty scientific logic, and to recognize the needs of practitioners then allow those needs to stimulate new science. I begin by focusing on three broad areas that encompass bodies of ecological knowledge pivotal to successful restoration: system dynamics, scale and context-dependency, and diversity. For each, I provide a brief overview of the relevant ecological theory, concepts, and results from research then describe the relevance to restoration in practice including the extent to which theories are being applied, ignored, or misused. My examples are admittedly biased toward my expertise, namely running waters that make up the many tributaries feeding coastal zones. Treatment of fisheries “restoration” is better left to those more knowledgeable who can meld principles of harvest management and stock assessment with ecosystem restoration.
Ecological Theory as a Foundation for Restoration Science
Strictly speaking, ecological restoration is an attempt to return a system to some historical state, although it is widely recognized that this is often difficult or even impossible to achieve. A more realistic goal may be to move a damaged system to an ecological state that is within acceptable limits relative to a less-disturbed system. In this sense, ecological restoration can be viewed as an attempt to shift ecosystem composition, structure, and function to within a range that is more desirable than current conditions.
The scientific underpinnings of ecological restoration can be found in ecological theory writ large. By the latter, I mean to include the interactions of physical processes (hydrological and geomorphic) with ecological processes and players. Ecological science has a long history of using theory to guide and advance knowledge and so today there is a strong and diverse body of theory addressing many aspects of ecological interactions (Weiner 1995; McGlade 1999; May and McLean 2007). These theories typically fall into two groups: mathematical descriptions of phenomena or conceptual descriptions of nature. Both are used to build and test predictions. I will focus on the broader category of ecological concepts and the extent to which they inform restoration in practice, but I acknowledge that many restoration efforts focused on single species have been informed by population dynamics modeling that was a direct outgrowth of ecological theory (Falk et al. 2006). For example, population viability analysis can be used to evaluate different restoration strategies (Maschinski and Wright 2006).
Ecological theory as it applies to restoration includes principles of population, community, and ecosystems ecology with the former two contributing insights mainly on the roles of recruitment, dispersal, and species interactions (Falk et al. 2006; Lake et al. 2007). At the ecosystems-level, restoration ecology has chiefly been influenced by research on the importance of natural disturbance, energy flow, and nutrient cycles (Naeem 2006; Kaushal et al. 2008). Together these bodies of knowledge have shaped the emergence of ecosystem restoration as a science with a particular influence played by at least three broad categories of theory, concepts, and data: (1) equilibrium dynamics and state changes; (2) scale and context dependency; and (3) diversity and heterogeneity. However, what is being done on the ground to restore ecosystems—ecosystem restoration in practice—does not necessarily reflect an acceptance of these concepts. Furthermore, some concepts have been adopted in a simplified form despite little evidence to support the way in which they are applied. Finally, even though restoration is fundamentally a human-directed enterprise, theoretical, and empirical treatment of these concepts with humans as integral components of ecosystems is seriously under-developed (Palmer et al. 2004).
Equilibrium Dynamics, Thresholds, and State Changes
The existence of definable states or conditions that systems exhibit has formed the basis of a great deal of theoretical work in ecology. The idea is that populations, communities, and entire systems might move away from a recognizable ‘normal’ state following a disturbance but would ‘tend’ back toward the state after some period of recovery time (Pimm 1991). The classic descriptions of population growth given limiting resources (e.g. logistic growth) assumed that there was some carrying capacity (k) or “equilibrium” population level for a given habitat at a given point in time; deviations from k suggest the resource base may have changed and/or the population is not under density-dependent control.
The equilibrium concept has also been applied extensively by community ecologists. For example, island biogeographic theory predicts an equilibrium number of species in a habitat as a function of the immigration and extinction rate (MacArthur and Wilson 1967). Succession theory, which has played a major role in plant community ecology (and in terrestrial restoration), also assumes directional change toward a single equilibrium point (Suding and Gross 2006). A great deal of work on ecosystem stability and persistence assumes equilibrium dynamics (Loreau and Behera 1999), as do many basic ecosystem models for nutrient cycling (Scoones and Toulmin 1998). In sum, a large amount of ecological theory has been dominated by deterministic perspectives in which internal feedbacks were assumed to keep systems in balance.
The possibility that multiple, stable states may exist is a concept that has been around for decades (e.g. Lewontin 1969); however, it has re-emerged as a topic of great interest because of concerns over ecosystems crossing “ecological thresholds” that result in abrupt state shifts that are difficult to reverse (Carpenter et al. 1999). Some scientists argue that most ecological systems have multiple stable states which they tend toward under different conditions (‘stable point attractors’) and these are separated by unstable transient thresholds (sensu Walker and Meyers 2004). Under this assumption, the key to restoration and management is not about keeping the system in the same state—this is not possible. Instead, the key is identifying what factors initiate changes that shift an ecosystem toward a state that is unacceptable (a societal decision), and then controlling those factors or enhancing ecosystem resilience to changes in those factors (Walker et al. 2004). Not all ecosystems exhibit threshold changes, but for those that do, the question becomes: can we identify a threshold before it is crossed? Answering this question is a central goal in the study of coupled human-ecological systems. While we currently have no theory or tools for predicting if and when a system nears a threshold, there is a great deal of work underway, including a joint effort by the Santa Fe Institute and the Resilience Alliance to build an empirical database on regime shifts in ecosystems and coupled socio-ecological systems (Walker and Meyers 2004).
Relevance to Restoration
The preoccupation with fixed reference sites ignores a great deal of ecological theory that has come along over the last 30+ years as ecologists began expressing great discomfort with the equilibrium concept (Parker and Wiens 2005). Problems include: application of the concept requires that there is some form of self-regulation that governs system dynamics; the concept typically does not apply at the small spatial and temporal scales in which most ecologists work; and the concept minimizes the importance of history, disturbance, and stochastic factors (DeAngelis and Waterhouse 1987; Wu and Loucks 1995). There is usually very little understanding of what, if any, internal regulating processes control the state of an ecosystem in need of restoration (Mayer and Rietkerk 2004). In fact, there is growing evidence that regulating processes may be ‘external’ to the nonhuman ecological system and reside at the level of the socio-ecological system—i.e. the ecosystem in its human context including its anthropogenic stressors and the feedbacks between ecosystem state and the sociopolitical management of its stressors (Liu et al. 2007).
With respect to spatial extent, restoration efforts typically involve taking only a small ‘piece’ of an ecosystem and assuming that given suitable treatments, it can recover to “equilibrium” (reference level) conditions despite its landscape/watershed context. Rarely are restoration actions taken over large temporal and spatial scales. Even if the desire is to restore an entire estuary, actions that are informed by ecological theory are typically implemented at the scales of hundreds of meters of shoreline, single oyster reefs, and sections of tributaries (Luckenbach et al. 1999; Zedler 2000; Bernhardt et al. 2005). Yet ecosystem processes like nutrient cycling, that involve the movement of many constituents, occur at much larger scales and in some cases over entire watersheds.
The important role that disturbances such as hurricanes and floods play in maintaining the ecological health of coastal wetlands, estuaries, streams, and river floodplains is extensively documented and yet countless restoration efforts attempt to minimize or eliminate geomorphic change in these ecosystems (Simenstad et al. 2006). While this is understandable if people and homes are at risk, efforts to suppress change in a system once it is restored are common even in areas where such risks are not an issue. For example, many stream restorations use a method called ‘natural channel design’ (Rosgen 1996) which is based on estimating channel dimensions that will prevent erosion or deposition of sediment even if the stream is on undeveloped land with ample room for channel migration (Kondolf 2006).
Scale and Context Dependency of Ecological Response
The dynamics of individual species, the composition of entire assemblages, and the rates and form of ecosystem processes vary spatially not only because of local conditions but because of the movement of individuals and the flux of materials. The arrangement of habitats across the landscape, spatial population structure, dispersal abilities, and the magnitude and direction of material flows all act to determine ecological patterns and processes, and so it is not surprising that ecologists have abundant evidence that species may perform differently in different environments—i.e. performance is context-dependent (Power et al. 1996). Environmental conditions determine if a species is physiologically capable of living in a particular region; subtle, within-region differences in these conditions along with biotic interactions may enhance the survival and reproduction of a species (i.e. its performance). These ‘optimal patches’ may serve as sources of colonists to other patches while suboptimal patches may act at sinks (high mortality); the degree of dispersal among these regions may determine population sizes and risk of extinction (Mouquet and Loreau 2003; Hanski and Gaggiotti 2004). For some species, there are critical thresholds of patch connectivity needed to ensure a supply of colonists (Turner and Gardner 1991).
Spatial and temporal environmental heterogeneity also influence the functional role species play in ecosystems (Wellnitz and Poff 2001). For example, both facultative feeding and changes in species interactions are known to be associated with changes in flow in both marine and freshwater systems (Cardinale et al. 2002; Cardinale and Palmer 2002; Hentschel and Larson 2005). This context-dependency of performance or interaction strength can influence ecosystem-level processes such as production and resource utilization (Cardinale et al. 2001; Stachowicz et al. 2007). Changes in such key processes can have cascading effects on nearby ecosystems, particularly in river networks where the movement of materials can effectively link very distant ecosystems (e.g. headwater streams to coastal marshes). For this reason, ecologists have increasingly focused research on broader spatial contexts and multi-scale processes. For example, some ecosystems receive substantial inputs of nutrients or organisms that act as ecological subsidies (Palumbi 2003). Nakano and Murakami (2001) showed that across-habitat prey flux accounted for 25.6% and 44% of the total annual energy budget of bird and fish assemblages in a Japanese stream. Subsidies such as these have been extensively studied both theoretically and empirically (e.g. Polis and Hurd 1996), but there are also unwanted forms of ‘ecological commerce’, particularly in the case of pollutants or other inputs that arise from human activities (Palumbi 2003; Palmer et al. 2004). Subsidies whether positive or negative, can exert tremendous control over ecological systems and for watersheds. Because the delivery of subsidies may be controlled by the dendritic network configuration of river channels the study of material fluxes and the persistence of species can be complex (Grant et al. 2007).
Relevance to Restoration
Given the potentially large spatial variability in a species performance in heterogeneous environments, the selection of restoration sites is critical. The position on the landscape or seascape may mean the difference between success and failure of newly established populations. Just as theory predicts there may be an optimal number and size of marine reserves that must be spatially arranged to maximize the persistence of a species (McCarthy et al. 2005), optimization of the size and position of a restoration project is important (Palmer et al. 1997; Craig et al. 2008). Smith and Lamp (2008) found that as urbanization proceeds in a watershed, insect taxa unique to headwater streams were at the greatest risk of extirpation and the authors suggest that once taxa are lost, the supply of dispersing adults from other headwaters may limit the success of headwater restoration projects. Unfortunately, we know very little about dispersal dynamics of freshwater and coastal species such as their maximum dispersal distance or preferred dispersal routes because faunal movements are so difficult to measure. It is therefore not surprising that decisions on the spacing and size of restoration efforts have only generally considered the potential for dispersal limitation (Lake et al. 2007).
Site selection for restoration projects is mostly opportunity based and not informed by factors such as metapopulation structure, ecological subsidies, or health of adjacent waterways. In fact, despite common acknowledgement among managers that watershed and landscape context are critical to restoration, only a small fraction of stream and river restoration projects throughout the USA have been initiated because of a broader river or watershed management plan (Bernhardt et al. 2007). For most projects, sites are selected based on land availability even though this does little to ensure project success, and in some cases results in the selection of sites that are clearly suboptimal. For example, restoration efforts to remove flow barriers and enhance the migration of anadromous fish in watersheds that are impaired, have contaminated sediments behind the barrier, or have non-native species that will extend their range must be viewed as foolhardy or, at best, as a trade-off (Stanley and Doyle 2003). While dispersal (migration) is important, the most basic elements of survival given the local environmental context can not be ignored.
Problems stemming from opportunity-based site selection are often compounded by a stove-piping approach to environmental management when it comes to agencies and funders that focus programs on specific habitat types not broad regions. Site selection and design of most restoration projects are done in a compartmentalized fashion such that tidal marshes, freshwater tributaries, and parcels of land are restored or conserved independently of their link to one another. Rarely are efforts coordinated at whole watershed scales to maximize environmental benefits even though the value of targeted watershed approaches is being increasingly recommended. Political pressure and jurisdictional issues remain huge obstacles to implementing the scale of efforts restoration scientists’ advise. Coordinated large-scale restoration is also impeded because of a mismatch in how freshwater vs. marine ecosystems are assessed and by lack of data on factors that link these ecosystems. For example, in the Chesapeake Bay region, measures of stream health for prioritizing restoration efforts are most often based on invertebrate diversity or loss of sensitive insect species (EPA 2006), yet assessments in large coastal rivers and tidal creeks largely focus on dissolved nitrogen and suspended sediment (e.g. CBP 2007). While diversity is undoubtedly important for stream ecosystems, it tells one little about the ability of a stream to reduce the downstream flux of nitrogen or sediment. There is no statistically significant relationship between insect diversity and stream nitrogen levels unless nitrogen concentrations reach levels rarely seen (EPA 2006), and the relationship between suspended sediment and insect diversity are only generally known (Berry et al. 2003; Matthaei et al. 2006).
This myopia or ‘my habitat’ focus flies in the face of widespread acknowledgement of the importance of watershed perspectives in management plans and agency documents. In fact, the extent to which the spatial scale (area) of a restoration effort influences restoration success has not been extensively studied even though the failure of some projects has been attributed, at least in part, to the size of the project (Ehrenfeld and Toth 1997; Hughes et al. 2005). In some cases, the area of restored habitat is simply too small compared to the size of nearby degraded habitat or the flux of water from upstream degraded habitat overwhelms efforts. There are numerous examples of stream and river restoration projects designed to reduce erosion but I know of no examples in which the net flux of sediment reaching tidal waters has been shown to be reduced through a stream restoration project. Identifying hotspots within watersheds that act as net sources or sinks of sediments would allow one to target small restoration sites, but such strategic, science-based selection of sites for restoration is simply not occurring today. Furthermore, it is worth questioning the assumption that all sediment movements are detrimental. Actually, sediment flux is a natural geomorphic process that is ecologically very important and further, reducing sediment flux below historic levels would be atypical with respect to how coastal receiving ecosystems evolved (Carter and Woodroffe 1994; Florsheim et al. 2008).
In sum, ecological understanding of the importance of material exchanges, spatial variability, and multi-scale interactions is significant but has had little influence on restoration in practice; it is not that the science is buried but that it is simply not being used. I briefly discuss socio-political issues that may be driving this later.
Diversity, Heterogeneity, and the Maintenance of Species Assemblages
The issue of what promotes diversity and why it matters ecologically has probably received more attention theoretically and empirically than any other issue in ecological science. Genetic diversity, species diversity, and habitat diversity have all been extensively studied. Genetic diversity allows organisms to adapt to environmental variation and reduces inbreeding depression thus contributing to the maintenance of populations (Lowe et al. 2004). Entire books have been written on this important topic so the reader can look elsewhere for an overview of theory and empirical work on the importance of genetic diversity.
Ecologists have argued that a primary reason that species diversity is important is because it may contribute to community stability and ecosystem function (Naeem et al. 1994; Tilman et al. 1996). The link between diversity and ecosystem function has been hotly debated, particularly over whether it is driven by one or a few species or if it is even a statistical artifact (Hooper et al. 2005). The current view on the link is that function probably does decline as species are lost but the reasons for this vary dramatically among systems and the magnitude of the effect is determined by the identity of the species lost (Cardinale et al. 2006). Thus, increasing attention has focused on the contribution of individual species to ecosystem functions, with the most focus on the role of species in rates of primary productivity. A meta-analysis of work in eight different European grasslands suggests that different species have a disproportionate impact on different functions so that maintenance of multi-functional ecosystems may require maintenance of high species diversity (Hector and Bagchi 2007). If both functional diversity and response diversity within functional groups are high, an ecosystem may exhibit a great deal of resilience in the face of environmental changes (Elmqvist et al. 2003).
Stability of ecosystem function does not mean that the species composition never changes. Whether or not more diverse ecosystems are also more stable in terms of exhibiting little change in species composition over time is debatable (McCann 2000). Current thinking is that community stability may be strongly associated with diversity because as long as species are only weakly interacting, diversity buffers communities from the impacts of a disturbance as species responses to the disturbance will vary. But the nature of the relationship between diversity and stability is complex and varies with system and environmental conditions (Shurin et al. 2007). In some settings or under some conditions, low diversity is associated with high levels of stability, not vice versa. The links between diversity and stability are difficult to study because temporal species turnover may be linked to environmental stability and seasonality, not diversity, and these can be hard to separate (Keitt 2008).
Working on the assumption that diversity is important to functionality and potentially to stability, ecologists have sought to understand what promotes diversity. Theoretical and empirical work on species diversity dates to the early days of ecological icons such as Hutchinson (1959) and MacArthur and MacArthur (1961). Of particular interest is the large body of ecological research on the link between species diversity and habitat heterogeneity (Tews et al. 2004) because this research has played an important role in the theory and practice of restoration (Palmer et al. 1997). In fact, central to many restoration efforts is the assumption that rehabilitation of physical habitat diversity will lead to the restoration of biological communities (Brown 2003; Spanhoff and Arle 2007).
The term heterogeneity has been used very loosely by ecologists—it may refer to habitat complexity (technically the spatial arrangement of patches), habitat diversity (the number of types of habitats in an area), and sometimes even environmental variability within a habitat over time (Li and Reynolds 1995). Species diversity has been shown to increase with habitat heterogeneity for a variety of species ranging from birds and mammals to insects and demersal fish (MacArthur and MacArthur 1961; Murdoch et al. 1972; Kaiser et al. 1999). The mechanisms are numerous and not necessarily mutually exclusive: habitat heterogeneity may provide more physical space, refuge, resource availability, and open niches for members of a community, thus promoting diversity. A great deal of work has focused on the role of niche fractionation in promoting diversity and the role of refuges in time and space minimizing competitive exclusion and mediating predator–prey interactions (Grabowski 2004; Willis et al. 2005).
Relevance to Restoration
The goals and approaches of many restoration efforts reflect the deeply held view that diversity is important—whether this is genetic diversity, habitat (patch) diversity, or species diversity. Restoration efforts are also largely based on the view that it is native species and habitats as well as wild genotypes that are desirable. Of course, this begs the questions, “desirable to whom?” and “what is the definition of native?” If ecosystems continue to perform functions that support valued ecosystem services such as the provision of clean water, perhaps the presence of non-natives is irrelevant. Dramatic events in which native species are lost due to the invasion of a non-native tend to be well remembered even though they are not necessarily very common relative to the number of non-natives that reach an area unnoticed. The use of non-native species in restoration efforts is not new and some have even argued that now is the time to consider moving species beyond their native range as a pre-emptive measure for species unable to disperse or adapt fast enough to keep up with climate change (Hoegh-Guldberg et al. 2008)
Restoration efforts for fisheries and wildlife populations, particularly endangered or threatened species, routinely consider a myriad of genetic issues and, in fact, the goal is often to enhance genetic diversity with the hope of reducing the risk of extinction (Machado-Schiaffino et al. 2007). But aside from valued fisheries or threatened species, there has been only scant research on the role of genetics in restoration success and this is mostly on plants. Travis et al. (2002) found that Spartina alterniflora that naturally colonized restored marshes had levels of genetic diversity as high or higher than reference sites. However, when plants are collected elsewhere and brought to a site, genetic diversity is not always as desired as was the case for eelgrass (Williams 2001). In the last decade, increasing attention has been given to the genetic composition of individual plants or animals that are brought to a restoration site, but the application of genetics theory to restoration is still in its infancy.
Interest in establishing diverse assemblages of species by restoration ecologists stems from the tenet that an ecosystem’s ability to withstand disturbances (i.e. be more stable) may be central to its long term survival following restoration, and that this ability is enhanced when species diversity, and thus functional redundancy, is high (Lake et al. 2007). This suggests that a restoration practitioner needs to carefully evaluate community level attributes and those factors that promote species diversity. Yet, the historical fishery-science approach of managing single stocks as well as societal preferences for certain iconic species have resulted in a huge number of freshwater and marine restoration efforts focused on only one or a few species (e.g. species of migratory salmon in the Pacific Northwest and oysters in the mid-Atlantic US). Despite their costs, these projects are common even where there is little evidence that efforts have resulted in measurable improvements (e.g. for salmon, Katz et al. 2007; for Eastern oysters, Mann and Powell 2007). As I elaborate on later, such projects are far more successful when they move away from structural fixes and focus on process-based restoration and identification of those factors limiting recovery of a fishery (e.g. as in Bottom et al. 2005).
When we turn our attention away from restoration of commercially important fisheries, species diversity as a stated restoration goal is actually quite common and is often assumed to be associated with increasing habitat heterogeneity (Palmer et al. 1997; Larkin et al. 2006). For example, an increase in the size and abundance of transient fish over oyster reefs in the Chesapeake Bay has been assumed to be due to an increase in the structural complexity of the habitat (Harding and Mann 2001). Topographic heterogeneity of freshwater wetlands has been implicated as critical to restoration of a variety of wetland species because small variations in elevation can shift hydrologic conditions and biotic responses (Vivian-Smith 1997). In tidal marshes, adding topographic heterogeneity in the form of multiple vegetative zones and mixtures of pools and creeks is believed to enhance restoration of invertebrates and fish (Larkin et al. 2008). Finally, a number of studies on sea grass ecosystems have suggested that the diversity of epifaunal communities increases as structural complexity of sea grass increases so long as the total habitat area is held constant (Sirota and Hovel 2006). While direct tests in an actual sea grass restoration project have not yet been published, there is a growing consensus that the arrangement of sea grass in space is critical (Bostrom et al. 2006).
These studies and many others suggest that the focus on habitat heterogeneity that has occupied ecologists for so long has largely been adopted in the field of restoration ecology. In some aquatic habitats, efforts to enhance physical habitat heterogeneity are so pervasive that hundreds of millions of dollars are spent annually on just the habitat aspect of the restoration process despite lack of evidence it works. For example, restoration of many streams has largely consisted of placing and sculpting in-stream structures in the name of ‘habitat improvements’, yet studies to date have not routinely confirmed that increasing habitat heterogeneity results in increasing species diversity or even restoration of individual taxa (Roni et al. 2002; Lepori et al. 2005). For example, stream fish are widely assumed to respond to an increase of in-stream habitat structures such a log jams, boulders, rock-filled gabions, and gravel, without definitive evidence this is true (Pretty et al. 2003; although see Schneider and Winemiller 2008). If increases in fish abundance are actually observed, they may be due as much to aggregation around the structures as to recovery of populations (i.e. successful reproduction and recruitment).
Despite the widespread focus on maximizing habitat heterogeneity and the suggestion that for wetlands and tidal marshes this may actually contribute to more diverse assemblages, it is not clear whether heterogeneity per se (i.e. habitat diversity and spatial arrangement of patch types) is what promoted restoration or it is just the presence of critical habitats or an increase in area (MacNally and Watson 2006). Just as there are keystone species in some ecosystems, there may be keystone habitats (Davidar et al. 2001) that are critical to the success (long-term persistence) of a restoration effort; such habitats need much more study by ecologists. Some of these habitats are the direct result of keystone species that act as ecosystem engineers. This is particularly obvious in aquatic systems in which oysters, coral, or beaver create complex habitat over extensive areas. In fact, these species are being increasingly used as ‘tools’ to enhance or stimulate restoration of bays, estuaries and wetlands (Bunt et al. 1993; Gorshkov et al. 1999; Nelson et al. 2004). However, it is not clear if desirable ecosystem-level effects that such keystone species produce are the result of increased habitat heterogeneity or other factors. Furthermore, we do not yet know if there are thresholds of habitat area and connectivity that must be restored to sustain populations and communities. Certainly, for stream restoration projects, the lack of any evidence to justify manipulations of habitat heterogeneity could be a threshold issue since most stream restoration projects are small in scope (Bernhardt et al. 2005).
Many scientifically well-accepted and fundamental ecological principles are under-utilized in restoration efforts while simultaneously some restoration practices in common use have not been adequately tested by the scientific community. The list is meant to be illustrative, not exhaustive. Examples of some of the most obvious consequences are also provided
Ecosystems often exhibit nonlinear, nonequilibrium dynamics
Can not assume an ecosystem will return to a former state or reference condition or that restoration trajectory is similar to degradation path
Processes critical to restoration typically reside at the level of the socio-ecological system
Manipulating just the physical environment will not usually lead to sustainable environmental change
Regional scale processes often swamp local processes
Unlikely that small areas embedded in a larger, degraded system can be restored in isolation
Ecological systems are spatially structured
Spatial context of a project is fundamental to restoration success; critical thresholds of connectivity among restored sites may exist
The dynamic character of natural systems is important to their health
Fixed structures typically fail or lead to a degradation of ecological processes and biota
Multiple, interactive factors control ecological systems
Single-fix approaches (e.g., restore habitat, re-stock species) rarely work
The exchange of material across systems (ecological subsidies) is critical to the health of many ecosystems
Fundamental changes in the natural movement of material (propagules, energy, organic matter) will interfere with restoration
Inadequately tested practices
Physical habitat heterogeneity promotes restoration of biological diversity
Diverts attention away from other factors that may be more important to restoring diversity; costly
The beneficial effects of many small restoration projects are additive; recovery will occur steadily as stressors are removed and sites restored
Many small projects may be easier to fund and implement but may not be as effective as larger projects; public expectations may go unmet
Created and restored ecosystems support rates of ecological processes similar to natural systems
If processes are not supported, environmental benefits of project will be limited
Streambank stabilization projects result in a net decrease in sediment flux to coastal waters
Projects are very costly, may have little impact on sediment flux but negative impact on stream biota
In some cases, managers simply need to be made aware of the latest science while in other cases scientists need to test methods that have not been scientifically evaluated but are widely used in restoration projects. Unless we can move ecological theory into more predictive modes, the most practitioners can do is apply basic ecological concepts in the general ways I outline below in “science in need of application”. Until these concepts are developed into relationships with predictive power, scientists must stress the importance of using the least interventionist and least expensive restoration approaches and then, only after weighing the potential risks vs potential benefits. There is also a great deal of new research that is critical if science is to play a useful role in restoring coastal watersheds and I outline some of these below in “applications in need of science”. For scientists to accomplish these tasks, routine tracking of methods and concomitant outcomes in restoration projects across diverse environmental and social contexts is required. Documenting failures is as important as successes and will accelerate the science and practice of ecological restoration.
Science in Need of Application: Well-accepted Concepts not being used
Focus on processes and limiting factors not structures or single species.
Add ecological insurance to all projects
Identify a probabilistic range of possible outcomes instead of a reference condition
Incorporating nonequilibrium dynamics into restoration planning suggests that many pathways are likely possible and may depend on restoration actions (Duarte and Conley 2008) and thus, it is important for restoration “targets” to be based on an array of possible outcomes or states. As Hughes et al. (2005) have argued, using reference systems can give a false sense of predictability of ecological outcomes. Rather than selecting reference ‘endpoints’, a desired trajectory should be defined that takes into account a range of values for key system attributes that are inherently variable; e.g. ranges of flow and sediment inputs, variability in the location and number of habitat types, and changes in the species composition of assemblages through time and space (Hughes et al. 2005).
Expand the spatial scale of restoration implementation
Apply a hierarchical approach to prioritize sites and choose restoration method
Science-based prioritization schemes that are consistent with specific goals should be developed to identify sites most in need of restoration and then, as necessary, take social and political issues into consideration in selecting from among the top priority sites. Currently in most settings the social and political issues take precedence over science. While we are far from identifying the relative importance each of multiple factors play in limiting the health of ecosystems, we can apply a logical, data-driven approach to choose among various restoration methods and to prioritize sites for restoration. As we build up a database on process-based responses to restoration treatments and combine this with data on a range of project characteristics, we can develop useful relationships between observable characteristics and probabilities of outcomes.
Prior work has shown that analysis of location characteristics conducted with GIS tools and available data can provide strong indicators of site-specific conditions (Brooks et al. 2004; Weller et al. 2007) that may determine probability of success given a particular method. For example, the level of watershed modification, adjacent land use, susceptibility to sea level rise, or hydrographic setting may be critical to project outcome. Restoration of wetlands in regions highly likely to be underwater in 25 years due to climate change is obviously not a wise investment. Similarly, habitat restoration in a tributary or coastal zone that is undergoing rapid urbanization makes little sense unless stormwater management is first implemented.
Applications in Need of Science: The Practice of Restoration Demands More Science
Test theories most relevant to the practice of restoration
Determine if physical manipulations can be used to manage ecological processes
Identify feedbacks between social and ecological systems that act to constrain science-based restoration
The most commonly cited reasons that science is not being used are socio-political and legal. An argument often heard is that the interests of stakeholders and the bureaucratic structures of regulatory and management agencies inhibit the implementation of truly science-informed restoration. Even if this is true, it does not mean that scientists have no role to play in overcoming the obstacles. The rapidly expanding literature on ecosystem based management (Leslie and McLeod 2007) suggests that new scholarship is needed to understand watersheds as socio-ecosystems—linked social and ecological systems in which there are complex, and often nonlinear, interactions between humans and the natural world (Walker et al. 2002) that determine management and restoration outcomes as well as social willingness to change (Berkes et al. 2003).
Determine how and when we can replace ecosystem functions using created ecosystems
Over the last several decades, the notion that ecosystems can be created has emerged and it has now gone wild—thousands of projects labeled as wetland creation occur every year, tidal marsh creation is now appearing, and even the possibility of stream creation has been raised in the most recent revision of the compensatory mitigation regulations in the USA (Federal Register 2008). The motive for creation projects is to replace or mitigate for aquatic resources that are lost and the social pressure for creation projects is clearly quite high. Creation projects are often referred to as a form of restoration; however, the science of ecosystem creation is in its infancy. There is no evidence that streams can be created de novo and results from wetland creation efforts suggest that while mitigation projects may meet compliance requirements, full ecological or functional success is low or unknown for most projects (Ambrose et al. 2007; Euliss et al. 2008; Mathews and Endress 2008). A 2008 review prepared for congress stated: “Both scientists and policymakers debate whether it is possible to restore or create wetlands with ecological and other functions equivalent to or better than those of natural wetlands that have been lost over time. [Yet] congress has repeatedly endorsed mitigation in recent years” (Copeland and Zinn 2008). Clearly, the social pressure is strong for ecosystem creation and scientists have an important role to play in determining the environmental implications. Ecologists, working in concert with hydrologists and engineers, need to determine the spatial contexts in which restoration projects are most likely to succeed and how local conditions influence functioning, especially biogeochemical.
Politics and social agendas will always influence the desired endpoints of a restoration effort, but the process by which restoration is done should be science driven. I have argued that science is not currently playing the role it should and that this is not just due to a lack of political or social will by the public. Instead, much of the science we have produced has not been effectively communicated to the potential users, or, more commonly, the science is not yet at the translation stage—i.e. implementing a general scientific finding in a specific restoration context is difficult. Scientists typically approach applied research by asking: how can my research be useful to managers and restoration practitioners? I suggest turning that question around by asking: what restoration science is being demanded? What are the applications in need of science? Is there quantitative evidence that methods currently being used are in fact ecologically effective? When science is viewed in this supply vs. demand framework (Sarewitz and Pielke 2007), restoration will become scientifically informed more quickly. I am also convinced that the science that is done in the process will contribute substantially to our fundamental understanding of watersheds. Many of the issues I have raised as central to reforming restoration are also central to ecological science—e.g. what is the relationship between ecosystem structure and function?
We are at an unusual time in the history of ecology in which our knowledge and theories are in great demand. Concerns over environmental sustainability are at an all time high, and there is a growing recognition that technological fixes won’t solve all the problems. Ecological science can provide unique insights into how nature solves complex problems and how we can capitalize on those methods. So, rather than allowing other scientists from other fields to co-opt the role of ecologists in advancing restoration science (e.g. eco-bio-engineers, and I don’t mean the beaver kind), let’s step up to the plate and take on the challenges.
I thank the organizers of the 2007 Estuarine Research Federation international meetings for inviting me to present a talk that stimulated this paper. I also thank those in my research lab who continually stimulate me with new ideas. Jim Cloern, Carlos Duarte, and Si Simenstad provided very helpful comments. This paper is Contribution number 4233 from the University of Maryland Center for Environmental Science.
- Ambrose, R.F., J. C. Callaway, and S.F. Lee. 2007. An Evaluation of Compensatory Mitigation Projects Permitted Under Clean Water Act Section 401 by the California State Water Resources Control Board, 1991–2002. Final report to the Water Resources Control Board of California.Google Scholar
- Berkes, R., J. Colding, and C. Folke (Editors). 2003. Navigating socio–ecological systems: building resilience for complexity and change. Cambridge University Press, Cambridge.Google Scholar
- Bernhardt, E.S., M.A. Palmer, J.D. Allan, G. Alexander, S. Brooks, J. Carr, C. Dahm, J. Follstad-Shah, D.L. Galat, S. Gloss, P. Goodwin, D. Hart, B. Hassett, R. Jenkinson, G.M. Kondolf, S. Lake, R. Lave, J.L. Meyer, T.K. O’Donnell, L. Pagano, P. Srivastava, and E. Sudduth. 2005. Restoration of U.S. Rivers: a national synthesis. Science 308: 636–637. doi: 10.1126/science.1109769.Google Scholar
- Berry, W., N. Rubinstein, B. Melzian, and B. Hill. 2003. The Biological Effects of Suspended and Bedded Sediment (SABS) in Aquatic Systems: A Review. An internal EPA report. http://www.epa.gov/waterscience/criteria/sediment/pdf/appendix1.pdf. Accessed 27 July 2008.
- Brush, G.S. 2008. Historical Land Use, Nitrogen, and Coastal Eutrophication: A Paleoecological Perspective. Estuaries and Coasts. doi: 10.1007/s12237-008-9106-z
- CalFed. 2008. The California Bay-Delta Ecosystem Restoration Program. http://www.delta.dfg.ca.gov/erp/. Last accessed 12-02-08.
- Cardinale, B.J., and M.A. Palmer. 2002. Disturbance moderates biodiversity—ecosystem function relationships: experimental evidence from caddisflies in stream mesocosms. Ecology 83: 1915–1927.Google Scholar
- Cardinale, B.J., M.A. Palmer, C.M. Swan, S. Brooks, and N.L. Poff. 2002. The influence of physical habitat on the rates of biofilm metabolism in a stream ecosystem. Ecology 83: 412–422.Google Scholar
- Carter, R.W.G., and C.D. Woodroffe (Editors). 1994. Coastal evolution. Cambridge University Press. Cambridge.Google Scholar
- Chesapeake Bay Program (CBP). 2007. What are tributary strategies? http://archive.chesapeakebay.net/pubs/tribstrats_backgrounder_final.pdf. Accessed 27 July 2008.
- Copeland, J.A., and C. Zinn. 2008. Wetlands: an overview of issues. Congressional Review Service publication RL33483, www.nationalaglawcenter.org/assets/crs/RL33483.pdf. Accessed online November 5, 2008.
- Craig, L.S., M.A. Palmer, D.C. Richardson1, S. Filoso, E.S. Bernhardt, B.P. Bledsoe4, M.W. Doyle, P.M. Groffman, B. Hassett, S.S. Kaushal, P.M. Mayer, S.M. Smith, and P.R. Wilcock. 2008. Stream restoration strategies for reducing nitrogen loads. Frontiers in Ecology and the Environment 6. doi: 10.1890/070080
- Duarte, C., and D.J. Conley. 2008. Return to Neverland: Shifting Baselines Affect Eutrophication Restoration Targets. Estuaries and Coasts. doi: 10.1007/s12237-008-9111-2
- Dudgeon, D. 2005. River management for conservation of freshwater biodiversity in monsoonal Asia. Ecology and Society 10: 15. [online] http://www.ecologyandsociety.org/vol10/iss2/art15/.
- Elmqvist, T., C. Folke, M. Nyström, G. Peterson, J. Bengtsson, B. Walker, and J. Norberg. 2003. Response diversity, ecosystem change, and resilience. Frontiers in Ecology and Environmental Science 1: 488–494.Google Scholar
- Environmental Protection Agency (EPA). 2006. Wadeable stream assessment: A Collaborative Survey of the Nation’s Streams. EPA 841-B-06-002. http://www.epa.gov/owow/streamsurvey/pdf/WSA_Assessment_May2007.pdf. Accessed 27 July 2008.
- Falk, D.A., C.M. Richards, A.M. Montalvo, and E.E. Knapp. 2006. Population and ecological genetics in restoration ecology. In Foundations of restoration ecology. 14-41, Washington, D.C.: Island Press.Google Scholar
- Federal Register 2008. U.S. EPA and U.S. ACOE Compensatory Mitigation for Losses of Federal Register. 2008. Aquatic Resources. Vol. 73 (70) April 10, 2008/Rules and Regulations. http://www.epa.gov/owow/wetlands/pdf/wetlands_mitigation_final_rule_4_10_08.pdf. Last accessed online November 5, 2008.
- Gorshkov, Y.A., A.L. Easter-Pilcher, B.K. Pilcher, and D. Gorshkov. 1999. Ecological restoration by harnessing the work of beaver. In Beaver protection, management, and utilization in Europe and North America, eds. Busher, P.E. and R.M. Dzieciolowski, 67–85. NY: Plenum Publishers.Google Scholar
- Grabowski, J.H., and C.H. Peterson. 2007. Restoring oyster reefs to recover ecosystem services. In Ecosystem engineers: concepts, theory and applications, eds. Cuddington, K., J.E. Byers, W.G. Wilson, and A. Hastings, 281–298. Amsterdam: Elsevier-Academic Press.Google Scholar
- Gutshall, M. 2004. Back to the Future: Stream Corridor Restoration and Some New Uses for Old Floodplains. A Landstudies, Inc. Report. http://www.landstudies.com/BacktotheFuture.pdf. Accessed July 16 2008.
- Hanski, I., and O. Gaggiotti. 2004. The ecology, genetics and evolution of metapopulations. San Diego: Academic Press.Google Scholar
- Harding, J.M., and R. Mann. 2001. Oyster reefs as fish habitat: Opportunistic use of restored reefs by transient fishes. Journal of Shellfish Research 20: 951–959.Google Scholar
- Hassett, B., M.A. Palmer, E.S. Bernhardt, S. Smith, J. Carr, and D. Hart. 2005. Status and trends of river and stream restoration in the Chesapeake Bay watershed. Frontiers in Ecology and the Environment 3: 259–267.Google Scholar
- Hooper, D.U., F.S. Chapin III, J.J. Ewel, A. Hector, P. Inchausti, S. Lavorel, J.H. Lawton, D. Lodge, M. Loreau, S. Naeem, B. Schmid, H. Setälä, A.J. Symstad, J. Vandermeer, and D.A. Wardle. 2005. Effects of biodiversity on ecosystem functioning: A consensus of current knowledge. Ecological Monographs 75: 3–35. doi: 10.1890/04-0922.Google Scholar
- Hughes, F.M.R., A. Colston, and J.O. Mountford. 2005. Restoring Riparian Ecosystems: The Challenge of Accommodating Variability and Designing Restoration Trajectories. Ecology and Society 10: 12. [online] URL: http://www.ecologyandsociety.org/vol10/iss1/art12/.
- Kaiser, M.J., S.I. Rogers, and J.R. Ellis. 1999. Importance of benthic habitat complexity for demersal fish assemblages in Fish habitat: essential fish habitat and rehabilitation. In American fisheries society, symposium 22, eds. Benaka, L.R., 212–223. Maryland: Bethesda.Google Scholar
- Kondolf, M. 2006. River restoration and meanders. Ecology and Society 11(2): 42. [online] URL: http://www.ecologyandsociety.org/vol11/iss2/art42/.Google Scholar
- Larkin, D.J., G. Vivian-Smith, and J.B. Zedler. 2006. Topographic heterogeneity theory and ecological restoration. In Foundations of restoration ecology, eds. Falk, D.A., M.A. Palmer, and J.B. Zedler, 142–164. Washington, DC: Island Press.Google Scholar
- Lewontin, R.C. 1969. The meaning of stability. Brookhaven Symposium in Biology 22: 13–23.Google Scholar
- Lowe, A., S. Harris, and P. Ashton. 2004. Ecological genetics: design, analysis, and application. Oxford: Blackwell Publishing.Google Scholar
- Luckenbach, M.W., M.R. Mann, and J.A. Wesson. 1999. Oyster reef habitat restoration: a synopsis and synthesis of approaches. Proceedings from the symposium, Williamsburg, VA. Virginia Institute of Marine Science Press, Gloucester Point, VA.Google Scholar
- MacArthur, R.H., and J.W. MacArthur. 1961. On bird species diversity. Ecology 42: 595–599.Google Scholar
- MacArthur, R.H., and E.O. Wilson. 1967. The theory of island biogeography. Princeton, N.J.: Princeton University Press.Google Scholar
- MacNally, R., and D.M. Watson. 2006. Distinguishing area and habitat heterogeneity effects on species richness: birds in Victorian buloke remnants. Australia Ecology 22: 227–232.Google Scholar
- Maryland Department of the Environment (MDE). 2004. List of Impaired Surface Waters [303(d) List] and Integrated Assessment of Water Quality in Maryland. http://www.mde.state.md.us/assets/document/Cover_TableofContents_2004-303d_Final.pdf. Accessed November 4, 2008.
- May, R., and A. McLean (editors). 2007. Theoretical ecology: principles and applications. Oxford University Press, 251 pp.Google Scholar
- McGlade, J.M. 1999. Advanced ecological theory: principles and applications. Oxford: Blackwell Publishing. 353 pp.Google Scholar
- Naeem, S. 2006. Biodiversity and ecosystem functioning in restored ecosystems: extracting principles for a synthetic perspective. In Foundations of restoration ecology, eds. Falk, D., M. Palmer, and J. Zedler, 219-237. Washington, DC: Island Press.Google Scholar
- Nelson, K.A., L.A. Leonard, M.H. Posey, T.D. Alphin, and M.A. Mallin. 2004. Using transplanted oyster (Crassostrea virginica) beds to improve water quality in small tidal creeks: a pilot study. Journal of Experimental Marine Biology and Ecology 298: 347–368. doi: 10.1016/S0022-0981(03)00367-8.Google Scholar
- Palmer, M.A., E. Bernhardt, E. Chornesky, S. Collins, A. Dobson, C. Duke, B. Gold, R. Jacobson, S. Kingsland, R. Kranz, M. Mappin, M.L. Martinez, F. Micheli, J. Morse, M. Pace, M. Pascual, S. Palumbi, O.J. Reichman, A. Simons, A. Townsend, and M. Turner. 2004. Ecology for a crowded planet. Science 304: 1251–1252. doi: 10.1126/science.1095780.Google Scholar
- Pimm, S.L. 1991. The balance of nature? Ecological issues in the conservation of species and communities. Chicago, Illinois: University of Chicago Press.Google Scholar
- Poole, G.C., J.B. Dunham, U.M. Keenan, S.T. Sauter, D.A. McCullough, C. Mebane, J.C. Lockwood, D.A. Essig, M.P. Hicks, D.J. Sturdevant, E.J. Materna, S.A. Spalding, J. Risley, and M. Deppman. 2004. The case for regime-based water quality standards. Bioscience 54: 155–161. doi: 10.1641/0006-3568(2004)054[0155:TCFRWQ]2.0.CO;2.Google Scholar
- Postel, S., and B. Richter. 2003. Rivers for life: managing water for people and nature. Washington, D.C.: Island Press.Google Scholar
- Pretty, J.L., S.S.C. Harrison, D.J. Shepherd, C. Smith, A.G. Hildrew, and R.D. Hey. 2003. River rehabilitation and fish populations: assessing the benefit of in-stream structures. Journal of Applied Ecology 40: 251–265.Google Scholar
- Rheinhardt, R.D., M.M. Brinson, and P.M. Farley. 1997. Applying wetland reference data to functional assessment, mitigation, and restoration. Wetlands 17: 195–215.Google Scholar
- Roni, P.T., J. Beechie, R.E. Bilby, F.E. Leonetti, M.M. Pollock, and G.R. Pess. 2002. A review of stream restoration techniques and a hierarchical strategy for prioritizing restoration. Pacific Northwest Watersheds North American Journal of Fisheries Management 22: 1–20.Google Scholar
- Rosgen, D. 1996. Applied river morphology. Wildland Hydrology. Lakewood: Colorado.Google Scholar
- Shurin, J.B. 2000. Dispersal limitation, invasion resistance, and the structure of pond zooplankton communities. Ecology 81: 3074–3086.Google Scholar
- Simenstad, C., D. Reed, and M. Ford. 2006. When is restoration not? Incorporating landscape-scale processes to restore self-sustaining ecosystems in coastal wetland restoration. Ecological Modelling 26: 27–39.Google Scholar
- Suding, K.N., and K.L. Gross. 2006. The dynamic nature of ecological systems: multiple states and restoration trajectories. In Foundation of restoration ecology, eds. Falk, D., M.A. Palmer, and J.B. Zedler, 190–209. Washington, DC: Island Press.Google Scholar
- Tews, J., U. Brose, V. Grimm, K. Tielbörger, M.C. Wichmann, M. Schwager, and F. Jeltsch. 2004. Animal species diversity driven by habitat heterogeneity/diversity: the importance of keystone structures. Journal of Biogeography 31: 79–92.Google Scholar
- Turner, M.G., and R.H. Gardner. 1991. Quantitative methods in landscape ecology. New York: Springer Verlag.Google Scholar
- Walker, B., S. Carpenter, J. Anderies, N. Abel, G.S. Cumming, M. Janssen, L. Lebel, J. Norberg, G.D. Peterson, and R. Pritchard. 2002. Resilience management in social-ecological systems: a working hypothesis for a participatory approach. Conservation Ecology 6(1): 14. [online] URL: http://www.consecol.org/vol6/iss1/art14/.Google Scholar
- Walsh, C.J., A.H. Roy, J.W. Feminella, P.D. Cottingham, P.M. Groffman, and R.P. Morgan. 2005. The urban stream syndrome: current knowledge and the search for a cure. Journal of The North American Benthological Society 24: 706–723.Google Scholar
- Wu, J., and O. Loucks. 1995. From balance of nature to hierarchical patch dynamics: a paradigm shift in ecology. Annual Review of Biology 70: 439–466.Google Scholar