September 2018
Advances in Conservation Ecology: Paradigm Shifts of Consequence for USACE Environmental Planning, Management and Conservation Cooperation
2018-R-05 The Institute for Water Resources (IWR) is a U.S. Army Corps of Engineers (USACE) Field Operating Activity located within the Washington DC National Capital Region (NCR), in Alexandria, Virginia and with satellite centers in New Orleans, LA; Davis, CA; Denver, CO; and Pittsburg, PA. IWR was created in 1969 to analyze and anticipate changing water resources management conditions, and to develop planning methods and analytical tools to address economic, social, institutional, and environmental needs in water resources planning and policy. Since its inception, IWR has been a leader in the development of strategies and tools for planning and executing the USACE water resources planning and water management programs.
IWR strives to improve the performance of the USACE water resources program by examining water resources problems and offering practical solutions through a wide variety of technology transfer mechanisms. In addition to hosting and leading USACE participation in national forums, these include the production of white papers, reports, workshops, training courses, guidance and manuals of practice; the development of new planning, socio-economic, and risk-based decision-support methodologies, improved hydrologic engineering methods and software tools; and the management of national waterborne commerce statistics and other Civil Works information systems. IWR serves as the USACE expertise center for integrated water resources planning and management; hydrologic engineering; collaborative planning and environmental conflict resolution; and waterborne commerce data and marine transportation systems.
The Institute’s Hydrologic Engineering Center (HEC), located in Davis, CA specializes in the development, documentation, training, and application of hydrologic engineering and hydrologic models. IWR’s Navigation and Civil Works Decision Support Center (NDC) and its Waterborne Commerce Statistical Center (WCSC) in New Orleans, LA, is the Corps data collection organization for waterborne commerce, vessel characteristics, port facilities, dredging information, and information on navigation locks. IWR’s Risk Management Center is a center of expertise whose mission is to manage and assess risks for dams and levee systems across USACE, to support dam and levee safety activities throughout USACE, and to develop policies, methods, tools, and systems to enhance those activities.
Other enterprise centers at the Institute’s NCR office include the International Center for Integrated Water Resources Management (ICIWaRM), under the auspices of UNESCO, which is a distributed, intergovernmental center established in partnership with various Universities and non-Government organizations; and the Conflict Resolution and Public Participation Center of Expertise, which includes a focus on both the processes associated with conflict resolution and the integration of public participation techniques with decision support and technical modeling. The Institute plays a prominent role within a number of the USACE technical Communities of Practice (CoP), including the Economics CoP. The Corps Chief Economist is resident at the Institute, along with a critical mass of economists, sociologists, and geographers specializing in water and natural resources investment decision support analysis and multi-criteria tradeoff techniques.
The Director of IWR is Dr. Joe D. Manous, Jr., who can be contacted at: [email protected]. Additional information on IWR can be found at: http://www.iwr.usace.army.mil. IWR’s NCR mailing address is:
U.S. Army Engineer Institute for Water Resources 7701 Telegraph Road, 2nd Floor Casey Building Alexandria, VA 22315-3868 Advances in Conservation Ecology: Paradigm Shifts of Consequence for USACE Environmental Planning, Management and Conservation Cooperation
2018-R-05
September 2018
Richard A. Cole, USACE Institute for Water Resources
USACE Institute for Water Resources
U.S. Army Corps of Engineers
Recent advances in conservation ecology are changing the way federal agencies in the United States plan, manage and cooperate to achieve their missions. These advances have contributed to major changes in the widely accepted working concepts—or paradigms—that underlie widely held assumptions of ecological management. Many of these paradigm shifts occurred since Corps environmental policy and technical guidance was written and are unevenly understood among ecological managers (eco-managers). The historic highlights of these changes and some of their implications are presented here for the use of the U. S. Army Corps of Engineers Civil Works Program (Corps) as it adapts to climate and other environmental change.
One of the most challenging aspects of Corps and other agency adaptation to new environmental legislation was protecting species and their support systems from increasing rates of extinction and restoring them to sustainable states as the environment was rapidly changing. The management demands and conflicts often seemed overwhelming, even before global climate change was widely accepted as real. The concept of ecosystem management emerged in the 1990s as an attractive alternative to species-based management, but, in its earlier incarnations in the 1990s, ecosystem management tended to rely on assumptions of long-term ecological stability and integrity that are now largely dismissed as unrealistic by ecologists and leading eco-managers in large part because of increasing awareness of past and potential climate-change effects.
Other ecological paradigms were already shifting, but have shifted more quickly since global climate change has been widely accepted as real. While many eco-managers in the 1990s continued to accept a tightly organized, deterministically resilient concept of community integrity, most now believe that many species redistribute individualistically and often uncertainly in continuously changing community assemblages. Yet leading eco-managers continue to believe that many of the threatened species elements of ecosystems can be protected at and restored to more naturally sustainable abundances somewhere on the continent, if not in previous, “more natural” ecosystem assemblages and locations.
Once thought to be separate from nature, ecologists and many eco-managers now generally accept human effects as pervasive and often inseparable attributes of nature (although some environmental advocates find the separation politically useful). Leading eco-managers now argue that most, if not all, ecosystems are humanly-altered and quite resistant to holistic protection and restoration of past ecosystem conditions. The old paradigms allowed a locally independent and deterministic certainty in management that has been replaced with wide acceptance of the need for forward looking, objective-guided and cooperative adaptive management. Most tenets of contemporary ecosystem management concept are generally accepted; particularly, the importance of sustaining biodiversity, defining management objectives in terms of ecosystem services, and considering species population needs and management at a wide range of ecological scales. But ecosystem management is accepted mainly in support of population-based management. Corps practitioners can do much to accommodate these changes, including possible revisions of policy and new technical guidance.
Disclaimer: The contents of this report have been developed and reviewed for factual accuracy, logic and clarity, but remain the author’s interpretations and views and do not necessarily reflect the views of the U. S. Army Corps of Engineers, the Institute for Water Resources, or any other agency or organization.
Advances in Conservation Ecology: Paradigm Shifts of Consequence
Preface and Acknowledgements During the past couple of decades, conservation ecology has rapidly advanced with new understanding of ecological processes and management needs among professional conservationists, other professional ecological managers, and ecological scientists. Since the 1990s, the profoundly disturbing reality and implications of widespread human-caused environmental change, including global climate change, has penetrated federal management of ecological resources and contributed greatly to major changes in management paradigms. Federal concern over the sustainability of the nation’s diverse plant and animal heritage became all the more worrisome during this time. Preventing its further erosion and the unsustainable use of ecological resources is a huge challenge that no single organization can cope with alone.
To effectively address this problem, agencies and other organizations must communicate, cooperate and collaborate more effectively within their organizations and with other organizations through concerted efforts modeled after groups like the Landscape Conservation Cooperatives (LCCs) set up by the Department of the Interior from 2009 to 2017. One challenge for future eco-management cooperatives, in whatever form they may take, is simply getting all participants communicating in universally understood ecological concepts. During a previous study of Corps participation in the LCCs, it became clear that knowledge of advances in ecological science and management was uneven, both within and among organizations, and there was no digest of information available for Corps ecologists and eco-managers. The digest provided here was directed toward reducing that deficiency.
For their helpful reviews and comments, my thanks go to Drs. Jae Chung, Forrest Vanderbilt and Paul Wagner of IWR, Dr. Tony Lyn Morelli of the U. S. Geological Survey, and Ms. Jeanette Gallihugh of IWR. Financial support was provided through the IWR Landscape Conservation Cooperative project under the guidance of Director Robert Pietrowski. Especially appreciated is the strong administrative support of Dr. Robert Brumbaugh, Deputy Director of IWR, and Dr. Paul Wagner, IWR Group I Manager.
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Table of Contents
PREFACE AND ACKNOWLEDGEMENTS ...... I
1. EXECUTIVE SUMMARY ...... 1
2. INTRODUCTION ...... 9
2.1 Purpose ...... 9
2.2 A Brief Environmental Policy History of the Corps ...... 10
2.3 A Few Words about Paradigms ...... 13
2.4 Method and Organization ...... 14
3. CLIMATE CHANGE PARADIGMS ...... 15
3.1 Science Paradigms ...... 15
3.1.1 Historical Climate Was Changing ...... 15
3.1.2 Global Climate Is Rapidly Changing Largely From Anthropogenic Effects ...... 16
3.1.3 Local Climate Change Is Happening At Uncertain Speeds and Directions ...... 18
3.1.4 Sea Levels Are Rising ...... 19
3.1.5 Life Processes are changing as Global Climate Changes ...... 19
3.2 Management Paradigms ...... 21
3.2.1 Species and Support-System Protections from Climate Change Are Uncertain ...22
3.2.2 Species-Specific Objectives Require Species-Specific Climate Adaptation ...... 23
3.2.3 Climate Change Adaptation Requires a Regional Planning Perspective ...... 25
3.2.4 Climate Change Adaptation Requires Integrated and Adaptive Management ...... 27
4. POPULATION PARADIGMS ...... 28
4.1 Science Paradigms ...... 28
4.1.1 Populations Are Identifiable Evolutionary Units of Species Membership ...... 29
4.1.2 Populations Occupy Unique Ecological Niches and Habitats ...... 31
4.1.3 Population Members Compete for Ecological Resources ...... 34
4.1.4 Population Size is regulated by Density-Dependent and Density-Independent Factors 35
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4.1.5 Increasing Environmental Instability Favors r-Selection over K-Selection Strategies 38
4.1.6 Ecological Instability Increases Survival Dependency on Dispersal and Filtration 40
4.1.7 Populations below a Minimum Size are Vulnerable to Instability and Random Events 42
4.1.8 Population Distributions, Size and Number Indicate Species Vulnerability ...... 45
4.2 Management Paradigms ...... 46
4.2.1 Populations Are Basic Eco-management Units That Require Specific Assessments ...... 46
4.2.2 Sustaining Populations Requires Sufficient Habitat Quality, Size and Connectivity 49
4.2.3 Sustaining Population Integrity Requires Genetic Viability ...... 51
5. COMMUNITY PARADIGMS ...... 52
5.1 Science Paradigms ...... 52
5.1.1 Community organization is maintained by forces that reduce competitive exclusion 53
5.1.2 Community Compositions are Always Changing Locally ...... 54
5.1.3 Community Functional Stability Depends On Disturbance Level ...... 57
5.1.4 Community Stability Increases as Species Numbers Increase ...... 59
5.1.5 Anthropogenic changes are causing an exceptionally rapid decline in biodiversity 61
5.2 Management Paradigms ...... 63
5.2.1 Uncertain Community Composition Requires an Adaptive Population Approach .63
5.2.2 Multi-species Surrogates Are More Reliable Indicators than Single Species ...... 65
5.2.3 Depending on Composition, High Species Richness Sustains Desired Species ..68
5.2.4 Species Sustainability Needs Management from Local to Regional Scale ...... 70
6. ECOSYSTEM AND LANDSCAPE PARADIGMS ...... 72
6.1 Science Paradigms ...... 72
6.1.1 Ecosystems Form Indistinct and Changing “Functional Units” ...... 72
6.1.2 Ecosystems Recycle Essential Materials Imperfectly at Local Scales ...... 75
6.1.3 Ecosystems Are Sustained By Flows of Energy ...... 77
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6.1.4 Ecosystems Are Regulated by External Forces and Internal Feedbacks ...... 79
6.1.5 Certain Patterns of Ecosystem Development are Generally Predictable ...... 81
6.1.6 Ecosystem Interactions Depend on Landscape Patterns, Populations, and Materials ...... 83
6.1.7 Ecosystems Are Hierarchically Organized and Regulated...... 86
6.2 Management Paradigms ...... 88
6.2.1 Ecosystems Include Humans and cannot be Protected in or Restored to a “Natural” State ...... 89
6.2.2 Ecosystem Management Emphasizes Sustainability, Broad Scale and Adaptive Flexibility ...... 94
6.2.3 Ecosystem-based Management Units Have Value When Carefully Defined and Updated 97
6.2.4 Ecosystem Management Should Complement Species-Based Management .... 101
6.2.5 Ecosystem Goods and Services Inform Ecosystem Management ...... 102
6.2.6 External and Internal Forcing Functions Often Need Joint Management ...... 104
6.2.7 Restoration of Ecosystem Elements is a Legitimate Complement to Protection . 106
6.2.8 A Landscape Approach Is an Essential Aspect of Ecosystem Management ...... 109
7. SPECIES RICHNESS REGULATION PARADIGMS ...... 111
7.1 Science Paradigms ...... 111
7.1.1 Regional Species Richness Decreases as Latitude Increases ...... 113
7.1.2 Total Energy Availability Limits the Maximum Species Richness of Regions ..... 114
7.1.3 Species Richness Increases as Competitive Exclusion Is Reduced ...... 116
7.1.4 Evolutionary Time May Contribute to Limiting Regional Species Richness ...... 118
7.1.5 Area Size and Isolation Contribute Largely to Species Richness Regulation ..... 120
7.1.6 Habitat Heterogeneity and Moderate Disturbance Increase Species Richness .. 123
7.2 Management Paradigms ...... 125
7.2.1 Species Richness Management Requires an Integrated Approach ...... 125
7.2.2 Managing for Species Composition Requires a Species Approach ...... 127
7.2.3 Managing for High Biodiversity Requires a Regional Approach ...... 128
7.2.4 Manage For Historically Moderate Disturbances and Against Extremes ...... 130
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7.2.5 Manage Local Areas for Appropriate Habitat heterogeneity ...... 131
8. IMPLICATIONS OF PARADIGM SHIFTS FOR THE CORPS ...... 134
8.1 The Relevance of the Paradigm Shifts to the Corps ...... 134
8.2 Strategic Adaptation of the Corps to Paradigm Shifts ...... 135
8.2.1 Focus on Sustaining Local and National Biodiversity ...... 135
8.2.2 Emphasize Species-Based Planning at All Ecological Scales ...... 137
8.2.3 Accept New Restoration and Protection Paradigms ...... 139
8.2.4 Adapt Management to Constant Environmental Change ...... 141
8.2.5 Cooperate More Effectively With Others ...... 143
8.2.6 Make Use of Policy Guidance Flexibility and Consider Future Revisions ...... 145
9. SUMMARY AND CONCLUSIONS ...... 148
10. REFERENCES ...... 150
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List of Figures
Figure 1. Global temperature records from three sources...... 16 Figure 2. Global mean sea level change...... 19 Figure 3. Changes in marine species distribution associated with global climate change...... 20 Figure 4. General framework for species-based risk assessment suggested by Glick et al. (2011)...... 25 Figure 5. Basic elements of adaptive management for climate change effects...... 27 Figure 6. Metapopulations are sustained by habitat connectivity between subpopulations...... 30 Figure 7. Species niche defined simply by the geographical presence and absence of a species with respect to two variables...... 32 Figure 8. The logistic and exponential models for population growth...... 36 Figure 9. Attributes of r and K selected species (from Pianka 1966)...... 38 Figure 10. Unconstrained radial and constrained linear dispersal, and generalized dispersal distance model of Wolfenbarger and Kettle...... 40 Figure 11. Conceptual model for processes of ecological filtering of species as they assemble in the local community from the regional species pool (after Belyea 2004)...... 42 Figure 12. Proposed relationship between number of individuals in a species population and the likelihood of extinction...... 44 Figure 13. Effect of small population size on genetic variation after population recovery...... 51 Figure 14. Predation and disturbance reduce competitive exclusion by separating niches...... 54 Figure 15. Under stress, ecosystems can shift to a different stable state quite resistant to restoration...... 58 Figure 16. Relationship between functional stability and species richness (from Tilman et al. 2006)...... 60 Figure 17. The carbon cycle, including human effects through fossil fuel emission...... 76 Figure 18. Conceptual model of a complex system, such as an ecosystem at various scales from populations to ecoregions (from Parrott 2002)...... 87 Figure 19. Island biogeography theory relates the size of islands and distances of islands from mainland to the balance between immigration and extinction that determines the species richness on the islands...... 121 Figure 20. Integrated theory of species richness regulation in continental and oceanic ecosystems (modified from Ricklefs 1993)...... 126 Figure 21. A focus on sustaining biodiversity starts with an understanding of species richness regulation...... 136
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U.S. Army Corps of Engineers vii Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence
1. Executive Summary Introduction
Recent advances in conservation ecology are changing the way the federal agencies of the United States plan, manage and cooperate to achieve their missions. These advances have contributed to major changes in the widely accepted ecological concepts—or paradigms—that underlie widely held assumptions of ecological management (eco-management), which more specifically applies to the management of and for living resources than the broader concepts of natural resources management and water resources management. The history of these paradigm changes and some of their management implications are presented here primarily for the use of the U. S. Army Corps of Engineers (Corps) as it adapts to climate and other environmental change in pursuit of its environmental and ecosystem restoration missions.
The eco-management activities of the Corps Civil Works Program—including much of its environmental impact mitigation, ecosystem restoration, stewardship of Corps lands and waters, and regulatory activities—are, in effect, conservation activities aimed at ecological-resource sustainability. As the term is used here, eco-managers manage ecological structure, function, and other processes and have titles like wildlife biologist, fisheries biologist, conservation biologist, forester, range manager, ecologist, environmental planner and many other variations. They include professionals who plan, implement, and maintain projects and programs directed at the sustainability of ecological resources and biodiversity. Conservation biologists are most focused on maintenance of biodiversity while resource managers are most focused on sustaining resource use while protecting the most vulnerable biodiversity.
Corps eco-management is institutionally guided by the understanding of ecological science and management concepts reflected in agency policy guidance written largely during the 1990s. Important advances in the science and management of conservation ecology have occurred since then, with numerous implications for Corps environmental protection and restoration programs. The advances contributed to major shifts in professional paradigms—the widely accepted working concepts of ecological scientists and managers—but had yet to be summarized for the Corps before this work began. The purpose of this study was to review advances in ecological understanding and create a digest of the prominent paradigms and paradigm shifts in conservation ecology. The targeted audience is professional ecologists and eco-managers in and outside the Corps who participate in Corps management of ecological problems, but with particular attention paid to the maintenance of sustained use and biodiversity in lake, river, wetland, riparian, and coastal settings.
The immediate impetus for this review was the results of a study of Corps participation in the Landscape Conservation Cooperatives (LCCs) set up by the U. S. Department of the Interior from 2009 to 2017. The LCCs are a prime example of inter-organizational cooperatives that apply ecological principles for sustainability outcomes. The LCCs were self-directed organizational partnerships focused on developing and sharing knowledge for landscape-scale management of climate-change effects and other environmental effects on fish, wildlife, and cultural resources. This large-scale, regional approach to conservation planning applies many of the new paradigms to management of ecological resources and species vulnerable to extinction. The study revealed that the new paradigms of ecological science and management are not uniformly understood or accepted by all participants in ecological cooperatives despite their general commitment to conservation ideals. The reasons are understandable: the literature is huge and diverse, people are busy with immediate concerns, few have the time to keep up with advances, some have not had a thorough educational grounding, organizational leaders
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may not be ecologically well-informed or concerned, and there are few digests designed specifically to serve organizational needs as this study is intended for the Corps.
Corps policy guidance and emphases in technical guidance have changed little since the 1990s. Between 1970 and 1996, the Corps had to respond to major changes in environmental law and federal water resources policy guidance related to the protection and restoration of environmental quality (EQ). It faced a steep corporate learning curve complicated by increasing uncertainty and paradigm shifting in the ecological sciences. Corps eco-managers largely accepted the old and managerially convenient stability, stationarity, tight community-unit, and natural ecosystem paradigms as a basis for a new federal emphasis on ecosystem management with a sustainability goal. This emphasis greatly influenced development of Corps environmental policy guidance during the 1990s, which has changed little since then. While that guidance is general enough to implicitly allow different interpretations, it fails to explicitly reflect important shifts in ecological paradigms. Since the guidance is frequently open to interpretation, Corps practitioners can view it with the paradigm shifts in mind and adjust their approach accordingly. They may also have opportunities to influence future revisions of policy guidance and the development of new technical guidance. Many Corps practitioners will have already recognized at least some paradigm shifts and have adjusted their approaches to eco- management consistent with those changes. The intent here is to comprehensively review all of the contemporary paradigms relevant to Corps planning and active management and present it as an integrated digest for additional practitioner considerations.
Paradigms of ecological science and management do not necessarily shift in unison or quickly follow key advances in ecological understanding, nor do they necessarily influence different branches of management within the same organization at the same pace. Leading ecologists and eco-managers generally accept new paradigms first, and then influence broader acceptance largely through the professional literature. Paradigms often shift incrementally and gradually as new information accumulates, but sometimes momentous events accelerate shifts. That happened for many ecologists and eco-managers as the high probability of human-caused global climate change gained paradigm status and was recognized for the broad influences it most likely would have on ecological structure, function, and management. This has happened more rapidly than some past paradigm shifts and some dependent paradigms are continuing to shift toward broader acceptance as ecologists and eco-managers assimilate the change.
Because of climate influences, this review starts with the climate paradigms of ecologists (see the table of contents for short titles of each paradigm). The remaining paradigms touch upon all aspects of the ecological applications of importance to the Corps. The history of paradigm change is divided into periods before and since the mid-1990s, when much of the contemporary environmental policy guidance of the Corps was developed. This organizational split highlights the implications of the paradigm shifts for Corps policy and technical guidance. Paradigms are examined at the population, community, ecosystem, landscape and ecoregion levels, which first appeared about in that order in the history of ecological thought and determine the organization of this report. Some of the paradigms have changed little in recent decades, but, with climate change in mind, nearly all are now considered in a different light and many with greater perceived relevancy for the Corps.
The review of paradigms in ecological science included readings of paradigm-shifting journal articles, ground-breaking papers in scientific anthologies, and major texts on basic ecology and conservation biology. Textbooks in particular are among the most complete records of disciplinary paradigms and how they shift through time. Eco-management paradigms are identified largely in fish and wildlife management, conservation biology, and ecosystem management texts and papers. No attempt was made to cover all ecological and eco-
U.S. Army Corps of Engineers 2 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence management paradigms. The ones covered here were thought to be among the most relevant for the Civil Works environmental program. Other paradigms may also be relevant and revealed with further assessment.
Climate Change Paradigms
Ecologists once assumed climate to be stable and generally invulnerable to human influence, but, by the 1990s, most had become uncertain about climate stability and were growing very concerned about the ecological implications. They had not seen much clear evidence of anthropogenic climate change or significant ecological effects from climate change. But much enlightening research has taken place since then and ecologists in general now believe past climate has changed significantly and recent global climate change is exceptionally rapid, largely anthropogenic, and largely responsible for rising sea levels. They believe climate is changing as a consequence of increased greenhouse gas emissions and reduced storage of organic carbon in the ecosphere. After much research, they also believe recent climate change has already altered numerous species distributions and life processes, and at least some future change is inevitable.
Many eco-managers found past assumptions that climate is stable and stationary convenient for generally independent management of properties within fixed management boundaries. Most well-informed eco-managers now recognize that management requires cross-boundary cooperation and coordination to adapt effectively to climate change. They generally recognize that species often respond to habitat changes uniquely, requiring specific attention as much as feasible (practicality demands some careful use of “indicator” species), and that management requires considerations of ecological influences on objectives at larger regional scales. They are coming to realize that projections of climate change and ecological responses are uncertain and require adaptive management as well as a scale of planning effort that requires cooperative integration across a wide range of public and private management organizations.
Population Paradigms
Ecologists long accepted the concept of species populations as demographic units; often defined by arbitrary boundaries; with characteristic birth, death, and dispersal rates; and with other attributes that determine population numbers. They believed populations are largely regulated by density dependent factors, such as competition and predation, in generally stable environments, but can also be regulated by density independent factors, such as severe weather, in less stable environments. They believed, and continue to believe, populations occupy unique ecological niches and habitats where they compete with other species populations for ecological resources. Once thought to exist in a “natural balance” in generally stable settings, ecologists now believe populations and species are exposed to more destabilizing environmental change that favors opportunistic generalists over specialists adapted to stable settings. Extinction was once believed to be rare and balanced with the rate of species origin, but has been, for decades, widely believed to be accelerating as a consequence of increasingly widespread human effects. Because of extinction concerns, many population ecologists now focus more on populations as evolutionary units with unique genetic characteristics and boundaries that depend on the degree of separation from other populations. A new concept in the decades before the 1990s, ecologists now generally recognize that some species persist in metapopulations composed of semi-isolated subpopulations, which die out over time and reestablish depending on the suitability of habitat connectivity for population maintenance.
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Eco-managers have long accepted the species population as a fundamental management unit. Population ecology has played a minor role in past Corps eco-management, but widespread acceptance of probable global climate-change effects has greatly elevated its relevancy. Resource managers continue to accept the concept of populations as demographic units with boundaries determined by either political or biological means, but all eco-managers now place much more emphasis on populations as evolutionary units with minimum viable sizes and habitat connectivity among subpopulations in metapopulations. They rely more on population viability analysis to assess and prevent populations decline to a genetically less viable and demographically unstable condition vulnerable to extinction when they decline below a minimum viable size. Eco-managers now generally accept the need for population-specific information about the niche requirements of greatly imperiled to vulnerable species, including community interactions, minimum habitat size requirements, and habitat connectivity for meta-population maintenance and dispersal adaptation to changing environments. Well-informed eco-managers accept some of the tenets of “ecosystem management” produced in the 1990s (detailed later) as complementary to population-based management and accept the need to assess and manage populations for sustainably at many geographical and ecological scales.
Community Paradigms
Ecologists practicing early in the 20th century accepted a dynamic but generally stable, geographically stationary, and compositionally rigid concept of a coherent community unit based largely on the behaviors of common and often dominant species. Following temporary disturbances, community units were believed to recover from temporary disruption through a deterministic process of community succession. That paradigm acceptance gradually began to change through the remainder of the 20th century and accelerated with increasing knowledge of climate-change and other environmental effects. Most basic ecologists now believe that except for obligatory species relationships (e.g., parasite-host), species populations assemble, interact, and disassemble individualistically in continuously changing community associations. Acceptance of a deterministic process of community succession and community unity has been replaced with an understanding of community resilience that is much less compositionally deterministic and more likely to result in compositional change. Ecologists now generally accept the premise that, except under extreme stress, most changes in species composition contribute to the resilience and stabilization of gross community structure and function (e.g., primary production, total biomass).
Ecologists also are much more inclined to believe that sustaining species richness depends on predation and moderate environmental disturbance to prevent competitive exclusion of species. They now recognize that rapid climate change, greatly increased nutrient enrichment, loss of top carnivores and other keystone species, invasion by exceptionally aggressive species, and other exceptional change can cause the rates and qualities of community structure and function to shift into new, often less diverse states that may become more resistant to change. Many ecologists now believe that increasing species richness, with the exception of particularly aggressive invasive species, contributes significantly to maintenance of gross functional and structural stability, such as the stability of community production, biomass, and dominant life forms (e.g., trees, grasses, oysters).
Eco-management before the 1990s was largely focused on the needs of particular species and paid little attention to other community composition. Many of those eco-managers in the minority who were concerned about community condition largely assumed communities respond as coherent units to “ecosystem restoration” and other holistic management. That paradigm shifted rapidly after the 1990s, when this belief became known as the “if you build it, they will come” myth. Most eco-managers now recognize that the individualistic behavior of many species
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targeted for management typically requires more specific population assessment and tailored management actions than assumed by a holistic approach alone, while also recognizing the need to sustain important community functions. They also recognize the practical need for indicator species, but increasingly accept the greater reliability of multi-species surrogates for other species’ needs over single-species surrogates, as well as careful monitoring for indicator reliability.
Well-informed eco-managers have generally believed that, within limits, gross community functions and structure, such as community productivity rates and broadly adapted dominant species, can be managed collectively for desired game and other resource outputs. Conservation biologists have come to generally accept management for high species richness, including apex predators, as a means for bolstering community support for imperiled species while resource managers are more slowly warming to that idea. But as recognition of the reality of global climate change and its ecological effects has grown, many eco-managers, including resource managers, have come to accept the need for taking a much broader regional view of factors influencing local community and population dynamics.
Ecosystem and Landscape Paradigms
Early ecologists conceptualized “natural” ecosystems as generally closed systems with definable boundaries until the 1950s to 1970s when they were progressively revealed to be open systems with imprecisely definable and porous boundaries through which energy continuously flows and species populations and materials continuously move. Ecosystems were generally considered dynamic but generally stable in space and time until the late 20th century when ecologists began to accept that their form and function could shift quickly, and often dramatically, under extreme stress. Research into the ecological effects of global climate change also reinforced the growing belief that ecosystems are continuously changing in space and time as abiotic properties, individual species elements, and associated functions respond differentially to environmental change and random events. However, when carefully considered, ecologists continue to accept the theoretical utility of an ecosystem-unit concept, defined mostly by gross structural and functional attributes observed in landscape patterns and the general predictability of ecosystem development following fires, storms or other temporarily destructive events.
Ecologists generally assumed until the late 20th century, that the structural and functional characteristics of ecosystems were determined largely by “bottom-up” forcing functions, such as energy inputs and flows that drive functions, sustain structure, and recycle scarce materials. Ecologists now generally accept a much bigger role for “top-down” consumer effects in joint regulation of community structure, function, and abiotic conditions with bottom-up forcing functions. In about the same time frame they also came to accept the hierarchical organization of and interactions among landscapes, ecosystems, communities, and populations at different ecological scales. None of these characteristics depends on stable species compositions, which ecologists now believe change continuously and often significantly over relatively short time frames at the local ecosystem level, contributing to gross functional and structural resilience, while retaining greater stability in the larger landscape and ecoregion, as long as habitat suitability and connectivity are sustained.
Eco-managers were once focused largely on the local habitat needs of species, but now are much more likely to emphasize assessing species management needs comprehensively from local population habitat to landscape and ecoregional scales to help contend with climate and other environmental change. In addition to a past focus on shaping hydrology, geomorphology, vegetation structure and other “bottom up” forcing functions, informed ecosystem managers
U.S. Army Corps of Engineers 5 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence also accept the need to manage for provision of “top-down” community feedback effects, including top carnivores and other keystone and dominant species. Eco-managers are rapidly replacing the old concept of restoring and protecting “natural” ecosystems free of human effect with the concept of managing “novel” ecosystems, infused with human influence and highly resistant to sustainable ecosystem protection and restoration, for improvement of desired ecosystem services and achievement of clearly defined objectives. The old paradigm that was fixed on restoration of more natural conditions at historical locations is quickly reorienting toward increasing the viability of unsustainable species somewhere in an increasingly less natural larger region through a broader landscape consideration of future habitat and connectivity needs. As many species shift into previously uninhabited areas, the focus on restoring “native” species in particular areas is rapidly losing meaning. Once quite deterministic in their thinking and independently minded, modern eco-managers now generally accept the need for adaptive management of uncertain outcomes and collaboration with others.
Species-Richness Regulation Paradigms
Ecologists proposed many different hypotheses for species richness regulation before the 1990s when more coherent theory began to take form. Based largely on observations of decreased species richness with increased latitude, ecologists now generally believe that maximum species richness of a local ecosystem is set largely by total solar and chemical energy influx, sufficient time for diversification through evolution, and the balance that results from immigration into the ecosystem from the regional species pool and emigration from and extinction within the local ecosystem (consistent with island biogeography theory). In this context, regions are broadly defined by latitude (e.g., tropical, subtropical, temperate, subpolar, polar) and large-scale geographic variation (e.g., continents and oceans). The species richness of local ecosystems is never more than the regional richness and usually much less because a variable fraction of the species is ecologically filtered from the immigrant stream by blockages and mortality before a subset reaches the local ecosystem. Species are further filtered from the local ecosystem pool (driven to local extinction) by both biotic and abiotic factors related to ecosystem size, habitat suitability and heterogeneity, and major disturbances (e.g., floods, fires, droughts). The details of the filtration process are site- and species-specific.
Eco-managers concerned about biodiversity conservation now generally accept this broad theory of species richness regulation. Once focused tightly on native species richness, some leading eco-managers now believe that, with some exceptions, the ecosystem stability needed to support threatened species is a function of total species richness, regardless of species origin. The exceptions are aggressively invasive species that can reduce local species richness, especially in isolated island and island-like environments. Eco-managers generally believe that many variables affecting ecosystem immigration and extinction rates can be managed through careful targeting of the needs of desired species at local to landscape scales. They generally believe that habitat heterogeneity appropriate for the targeted species can be managed through proper selection of management areas and through various interventions, such as those that simulate different stages of succession and mixes of water depths and upland elevations. They also accept the need for restoring and protecting moderate disturbance more or less like the conditions under which species evolved, as well as a frequent need to mitigate effects of extreme disturbances through design and selection of larger management areas, inclusion of refuge habitats, and restoration and protection of corridors suitable for species dispersal.
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Implications of Paradigm Shifts
The ecological paradigm shifts described here have profound consequences for federal resource management for environmental sustainability and biodiversity conservation. A major goal of national and international environmental policy is developing and maintaining the benefits from environmental use while also maintaining biodiversity across human generations. This goal contributes largely to the ideals of all federal resource management and protection agencies, including the Corps, where it is often expressed broadly through declared dedication to environmental and ecological sustainability. The paradigm shifts greatly complicate eco- management for sustainability, starting with a need for rethinking objectives, geographical perspectives, management of risk and uncertainty, appropriate approaches to objective achievement, and adaptation of policy and technical guidance.
In light of these paradigm shifts, Corps practitioners may want to reexamine strategies for achieving Corps sustainability goals and consider the need for changes in policy and technical guidance. While many Corps practitioners already recognize the need for updating policy and technical guidance, revision takes time and resources, both of which are scarce. While a clarification of Corps guidance that reflects the paradigm shifts is ultimately desirable, Corps policy guidance often allows broad enough interpretation for Corps practitioners to interpret it in ways largely consistent with the paradigm shifts described here and elsewhere.
A major apparent inconsistency of policy guidance with contemporary paradigms is the implication that past ecosystems can be partially or fully sustainably restored and/or protected in place given the effects of locally unmanageable environmental change. But guidance can be interpreted differently. The old assumptions of ecological stability, stationarity, community-unit integrity are no longer accepted, making sustainable ecosystem restoration in historic locations impractical in the long run. To sustain national biodiversity, eco-managers need to pay more attention to the individual needs of the most desired ecosystem elements and to manage for their long-term sustainability somewhere in the nation, if not necessarily in historical locations.
Restoring and maintaining species sustainability requires deep consideration of species needs in ever-changing ecosystem and landscape contexts. Eco-managers need to consider community, ecosystem and landscape dynamics with individual species populations in mind. Since many species have already redistributed independently of one another, and are likely to continue to redistribute into new ranges, consistency with the paradigm shifts also requires Corps practitioners to rethink what “natural” and “native” means in policy and technical guidance. While Corps practitioners, in cooperation with others, can plan and manage for more natural abundances of species elements at regional and national scales based on a different interpretation of policy guidance, they are likely to find it much more challenging, if possible at all, to holistically restore ecosystems to long-term sustainable states in previously occupied spaces, or any other spaces. The Corps has institutionally accepted adaptive management in principle, but, in practice, Corps practitioners and their nonfederal project sponsors need to adopt a much broader eco-regional approach to project planning, more regular monitoring of indicators of success, and appropriate modification of projects as needed to sustain achievement of clearly defined objectives. Leading eco-managers believe that doing this effectively requires a greater focus on identifying, improving, and protecting the human services appropriate to the Corps restoration and/or protection missions.
The paradigm shifts away from unrealistic assumptions about environmental and ecological stability, stationarity, and holistic integrity to an increased focus on the recovery and protection of species elements as they independently redistribute within the broader regional and national context previously considered necessary. That will require more information, intricate planning,
U.S. Army Corps of Engineers 7 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence and adaptability than the old paradigms indicated. The new paradigms will also require more cooperation among organizations to bring greater coordinated and collaborative effectiveness to managing conservation challenges that are chronically thwarted by less than necessary funding. The LCCs have been prominent examples of this new approach, but the technical challenges are daunting and many social-political barriers hobble a more comprehensive and integrated approach to management.
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“Our job as faceless bureaucrats…was not to do science, but to summarize the frontier of science.” Michael Greenstone
2. Introduction
2.1 Purpose The frontiers of ecological science and eco-management have been changing rapidly. Recent advances in conservation ecology are changing the way the federal agencies of the United States plan, manage and cooperate to achieve their missions. A prime example is the network of Landscape Conservation Cooperatives set up by the U. S. Department of the Interior in 2009. These advances have contributed to major changes in the widely accepted ecological concepts—or paradigms—that underlie widely held assumptions of ecological management (eco-management), which more specifically applies to the management of and for living resources than natural resources or water resources management. The history of these paradigm changes and some of their policy and management implications are presented here for the use of the U. S. Army Corps of Engineers (Corps) as it adapts to climate and other environmental change.
The eco-management activities of the Corps Civil Works Program—including its project impact mitigation, ecosystem restoration, stewardship of Corps lands and waters, and regulatory activities—are, in effect, resource conservation activities aimed at the sustainability of ecological resources and biodiversity. As the term is used here, eco-managers manage ecological structure, function, and other processes and have titles like wildlife biologist, fisheries biologist, conservation biologist, forester, range manager, ecologist, environmental planners and many other variations. They include professionals who advise the Corps or actually plan, implement, and maintain Corps projects and programs directed at the sustainability of ecological resources and biodiversity.
Corps eco-management is institutionally guided by the understanding of ecological science and management concepts reflected in agency policy guidance written largely during the 1990s. Important advances in the science and management of conservation ecology have occurred since the 1990s, with numerous implications for Corps practitioners and programs. No digest of these changes existed when this effort began. This report responds to that need by summarizing some of the most relevant changes in the widespread conceptual beliefs, or paradigms, of ecological science and eco-management, which underlie accepted management assumptions and practices. These “paradigm shifts” could greatly influence execution of Corps environmental programs, adaptation to environmental change, and cooperation with other environmental and resource management organizations. As discussion of the results reveals, the paradigm shifts indicate a probable need for adaptation, including interpretation of existing Corps policy and technical guidance in ways consistent with the changes and possible clarifications in future revisions of Corps guidance. These points appear to be consistent with the Corps climate change adaptation plan (USACE 2014).
The audience targeted for this report is professional ecologists, eco-managers, and other practitioners involved in Corps environmental and ecosystem restoration programs, in and outside the Corps, who are asked to guide the Corps in its application of ecological concepts to the management of ecological problems typically associated with rapid loss of desired ecosystem services. These professionals provide a human resource foundation for linking
U.S. Army Corps of Engineers 9 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence contemporary advances in ecology to state-of-the art application of policy and technical guidance.
The immediate stimulus for this review was a recent study of Corps participation in the Landscape Conservation Cooperatives (LCCs) set up by the U. S. Department of Interior in 2009 (Cole et al. 2018). It revealed wide variation in appreciation for contemporary ecological paradigms and the implications of recent paradigm shifts. The disciplinary paradigms generally accepted among well-informed ecologists and eco-managers provide a common foundation for environmental planning and management necessary for essential coordination and collaboration among diverse organizations. Effective cooperation requires generally consistent understanding of those paradigms of ecological science and eco-management.
Despite future uncertainties in their makeup and fiscal support, the LCCs may be the best recent example of the type of cooperative planning needed to conserve ecological resources in rapidly changing environments. They were created to facilitate cooperative development of new information and a nationally coordinated vision for large-scale, landscape-level adaptation to climate and other environmental change (Cole et al. 2018). This new approach to conservation planning applies state of the art assessments and models to the management of species populations, ecosystems and landscapes for sustainable resource use and maintenance of ecological heritage for future generations. The LCCs focused particularly on landscape-scale research products and extension to potential product users, but the dynamics of local-scale species populations, communities, and ecosystems are integral. The Corps study revealed that contemporary paradigms of conservation ecology are not uniformly understood by all cooperators. The reasons are understandable: The literature is huge and diverse, people are busy with immediate concerns, few have the time to keep up with advances, some do not have a thorough ecological grounding, organizational leaders may not be ecologically well informed or concerned, and few contemporary digests have been assembled to serve agency needs.
Significant changes in organizational practices reflect recognition of paradigm shifts; most notably those shifts related to climate change. But the recognition has not been uniform across organizations, members of organizations, or ecological paradigms. For the Corps, the absence of a digest on paradigm status may contribute to this variation and to the paucity of ecological technical guidance for Corps activities. Even before much of the Corps environmental policy and technical guidance was written, mostly in the 1990s, some paradigm shifts were too late to be incorporated and some well-established paradigms were not emphasized for their importance.
For the Corps environmental program, the practitioners most affected by paradigm shifts are ecological scientists and eco-managers (including planners) responsible for land and water management, regulation of the discharge of fill materials and other physical modification of the Nation’s waterways, mitigation of Corps-project environmental impacts, and ecosystem restoration projects. All of these program areas are concerned with the condition, protection, and restoration of ecological structure and function consistent with law.
2.2 A Brief Environmental Policy History of the Corps The Corps of Engineers probably reached the peak of its storied reputation in mid-20th century based largely on its success “taming nature” for the good of the Nation. Fairly or not, its reputation began to suffer during the environmental movement of the 1950s and 1960s when the Corps and other water resources development agencies were often placed by critics among the worst environmental offenders in federal government. Congress responded dramatically to the environmental movement. The environmental laws passed and amended starting in the
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1970s had unprecedented influence on the planning and other environmental management activities of the Corps and other federal agencies.
Few laws before or since insinuated themselves as deeply into the activities of federal water resource agencies as the National Environmental Policy Act (NEPA), Clean Water Act (CWA), and Endangered Species Act (ESA). The ESA is particularly demanding because it requires all federal agencies to use their authorities to promote the goal of the law, which is to sustain the Nation’s fish and wildlife heritage. The Corps was transformed by these laws and a sequence of other laws defining its authorities. They broadened the Corps relationship to the environment from one narrowly focused by its water resources development and emergency management missions to one including environmental protection and ecosystem restoration missions. The rapidity and complexity of these revolutionary changes imposed major challenges for developers of Corps environmental policy and technical guidance.
Corps environmental policy guidance had to rapidly adapt to the new laws and federal policy guidance to assure careful stewardship of Corps-owned lands, mitigation of environmental impacts by Corps project implementation and operations, administration of regulatory responsibilities for permitting and mitigating material-discharge impacts on aquatic ecosystems, and implementation of new environmental quality (EQ) improvement authorities, including one restricted to the use of ecosystem restoration and protection measures. The Corps responded with major changes in its policy guidance between 1973 and 2000.
Corps project planning was particularly challenged by changes in law and federal policy guidance. NEPA was signed into law in 1970, establishing the concept of EQ and environmental protection and improvement policies. Later in 1970, the Water Resources Planning Act (WRPA) of 1965 was amended to include congressional objectives that accounted for EQ protection and improvement effects. These two laws formed the legal foundation for Civil Works project planning, with a much clearer emphasis on environmental protection than in previous federal law. The first federal guidance for WRPA implementation was published three years later (WRC 1973) when national economic development (NED) and EQ improvement and protection were declared to be the two objectives of federal water resources project planning. NED included any resource use improvements measurable in monetary terms and EQ included all environmental attributes valued in ways that were not acceptably monetized and expressed in long-term environmental protection, such as the ability of ecosystems to sustain species threatened with extinction. Then a more detailed section on EQ evaluation was added in two revisions before the entire document was rewritten in 1983 (WRC 1983). At that time, the federal objective of water resources planning was changed to a single NED objective constrained by EQ protection. This made sense, since the water resources agencies had never been authorized by Congress to improve EQ and it simplified the planning process. That new sensibility evaporated when, in a rapid turnaround in 1986, the Corps was authorized by Congress to improve EQ at Corps projects. Then, in 1996, Congress authorized the Corps to restore and protect EQ anywhere nationally, but limited to an ecosystem and restoration approach and contingent on demonstrated national benefit and partnership with a nonfederal sponsor.
During these changes, the federal project planning guidance remained unchanged for reasons beyond Corps control. Despite the Corps’ new EQ improvement authority, NED continued to be the sole federal objective through 1999 (USACE 1999). In attempting to make the new EQ improvement authorities sensible in light of the NED objective, EQ improvement was treated more or less like post-project impact mitigation. This was a particularly awkward arrangement for its new aquatic ecosystem restoration authority. A year later, the Corps revised its project planning guidance in large part to better accommodate the new authority (USACE 2000). The new guidance for Corps project planning included two sub-objectives in its federal objective
U.S. Army Corps of Engineers 11 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence statement: one for NED and another for national ecosystem restoration (NER). The Corps did not return to the old EQ objective of the earliest federal guidance. It viewed EQ improvement as a broader authority than restoring and protecting the EQ held in ecosystems. Thus NER is in effect a sub-objective of the EQ protection and improvement objective of WRPA (which continues to apply). The NER objective is to restore unsustainable aspects of ecosystems to a sustainable status indicated by species desired in greater abundance for unique biological attributes that contribute to sustaining high plant and animal diversity (Cole 2014a, b). Except for some minor details, the project planning guidance for ecosystem restoration has not changed since 2000.
Starting in the 1970s, the Corps hired numerous environmental professionals to oversee policy implementation. Their professional preparation was highly variable. Many had no previous exposure to environmental law and policy guidance. Most had only modest exposure to the now outdated ecological paradigms of that time. Relatively few had advanced ecological training and more nuanced understanding of paradigm dynamics and uncertainties. Despite these limitations, the demands for EQ improvement were substantial and the Corps expanded rapidly into environmental improvement activities during the 1990s. Its aquatic ecosystem restoration program grew rapidly. Growth leveled off only recently as the Corps redirected more of its attention to other issues, including repair and rehabilitation of aging projects.
This period of Corps policy, employee, and technical change was also a time of increasing uncertainty and paradigm shifting in the ecological sciences, which is the main topic of this report. By the 1990s, well established ecological paradigms were beginning to disassemble in light of new evidence and understanding. Informed by early IPCC studies, state of the art analysts in and outside the Corps were quite aware of the need to prepare for and adapt to climate change possibilities (Stakhiv and Major 1997). Much Corps environmental guidance was written during the mid to late 1990s either before or about the time the old paradigms began to shift rapidly, including climate-related paradigms. While open to interpretation, the ecosystem restoration guidance implied that the old paradigms, based largely on the assumption of climatic and ecological stability, still held. The assumptions of climatic and ecological stability also influenced the Corps approach to environmental protection on lands managed by the Corps and through compensatory mitigation, both in its civil works projects and its regulatory program established by the Clean Water Act.
Corps environmental policy guidance has changed little since the turn of the millennium despite major paradigm shifts. Technical reports and manuals have not fully embraced the needs of ecological restoration and protection either, focusing largely on hydrology and associated geomorphology. Independent of Corps guidance, the awareness of paradigm shifts varies widely among personnel within the Corps. The stability of policy guidance was recently disturbed by a major revision of the federal planning guidance, but, as a consequence of Congressional actions, it is not yet clear how the revision will apply to the Corps and affect its project planning guidance. Many competing demands continue to delay revisions of ecosystem restoration and environmental policy guidance, which take significant time. However, any eventual revision of Corps environmental policy guidance can provide an opportunity to respond to the new paradigms in guidance. Meanwhile, Corps practitioners have some policy interpretation latitude that allows incorporation of most new paradigms into field applications.
This brief history of the Corps environmental program and policy guidance indicates that the 1990s was a watershed decade for Corps adoption of its contemporary environmental policy guidance. This is reflected in the organization of this report, which recounts paradigm histories in sections before and since 1996.
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2.3 A Few Words about Paradigms The modern concepts of scientific paradigm and paradigm shift are closely linked to Kuhn (1962). The paradigm concept has been widely accepted and applied in and out of science with various degrees of specificity. A broad interpretation of paradigm is used in this review. For its purposes a paradigm is a widely accepted belief in a philosophy, theory, hypothesis or other conceptual model in knowledge-based disciplines, sub-disciplines and associated occupations. Some might call them principles, but, regardless of terminology, these widely accepted concepts often shift with advances in understanding. The changes may be evolutionary or revolutionary.
Obvious examples of paradigms include the general theory of relativity in physics and the theory of organic evolution by natural selection in biology. Some reserve the concept of paradigm for major theories like these, but for the purposes served here, paradigms exist at many levels of specialization in ecological science and management. That does not mean that paradigms are equal in effect. Sub-disciplinary paradigms are generally consistent with disciplinary paradigms and respond to major shifts in disciplinary paradigms, but shifts at a sub-disciplinary level influence the details of a disciplinary paradigm. A revolutionary shift in the evolution paradigm of biology, for example, would initiate myriad shifts in sub-disciplinary paradigms, but as genetics, population biology, and other disciplinary paradigms evolved, the details of the evolution paradigm have also changed. Many paradigms shift incrementally and gradually, while fewer shift rapidly, typically in response to some momentous event. Response to the widespread scientific acceptance of global climate change as real is one such keystone sequence of events.
Paradigms provide a framework for following the course of ecological science and its influence on eco-management paradigms. The formation of paradigms allows practitioners to interact from a common understanding of the discipline, communicate more efficiently and to manage resources more cost-effectively while also acknowledging some uncertainty and the possibility of future paradigm shifts. Cooperation, communication, and collaboration may suffer when paradigm understanding varies significantly among practitioners. Ecological paradigms form a foundation for commonly understood disciplinary concepts. In the Corps, ecological paradigms guide the application of ecological science in Civil Works project planning, implementation, and management, as well as the Corps regulatory program. The most common summations of the state of paradigms are found in textbooks. Neither they nor paradigm shifts are typically identified as such, however. They are much more often presented as principles.
Ecological variability and uncertainty are never entirely eliminated by paradigm acceptance and some scientists and eco-managers continue to test their validity. Early ecologists were more likely than contemporary ecologists to accept paradigms as inviolable. Simberloff (1980), for example, was highly critical of certain 20th century paradigms because they were so often based on deterministic models that allowed no room for uncertainty and random events, and were most likely influenced by non-scientific ideology.
The paradigms of ecological science typically shift as new understanding of ecological process overwhelms past thinking. A shift to a new paradigm occurs when an alternative hypothesis or theory stands up to repeated empirical testing and refinement better than the established paradigm, despite the residual uncertainty. It then serves as the new model for scientific beliefs and assumptions until further testing leads to greater acceptance of an alternative. Time lags between the origin of new science and applications are commonly acknowledged, but rarely measured. Morris et al. (2011) found a mean estimate of about 17 years for medicine, but concluded that the reasons for the lags are not fully understood. A lag in new science application by government eco-managers may result in part from slower acceptance of scientific understanding by governing bodies. Knowledge advances with acceptance of new hypotheses
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and theories that better reflect real-world behavior. The theory of organic evolution, for example, has become a master paradigm of biological science, including ecology. But regardless of how well it has stood up to testing and how widely it is accepted by biologists, it continues to motivate more testing and continuous refinement.
Paradigms often persist despite much convincing evidence that a shift is appropriate. Both ecological scientists and eco-managers can become invested in an old paradigm because it is convenient. Paradigm shifts often cause disruptive reorientation of research activities and management applications. Eco-managers are often exposed to the full breadth of scientific paradigms for the first and last time during academic preparation. Management paradigms may lag behind shifts in scientific paradigms because in part many practitioners do not closely follow the “state of the art”.
Because ecology remains a relatively new and complex science, its contemporary paradigms are more likely to shift dramatically than paradigms of some other sciences. Ideally, ecological paradigms should always be applied with caution and with continuous monitoring of applications and changes in contemporary scientific thinking. In reality, this adaptive approach to eco- management has been difficult to adopt in public-service agencies, perhaps, most fundamentally, because the public and their governing bodies tolerate little uncertainty in management results.
2.4 Method and Organization The review of paradigm development and shifts for this digest included readings of ground- breaking papers in journal articles and scientific anthologies as well as major texts on basic ecology and conservation biology. They are extensively cited in the report and identified in the “References” section. The history of paradigm development was largely traced back from references highlighted in more recent literature. Management paradigms were identified largely in fish and wildlife management, conservation biology, and ecosystem management texts and papers. Corps policy and technical guidance was also consulted to assess its status with respect to paradigms. Particular attention was paid to influential literature relevant to the Corps eco-management activities; especially its ecosystem restoration program. No attempt was made to cover all ecological and management paradigms. The ones covered here were considered among the most relevant for the Corps environmental program and its interactions with other organizations. Other paradigms may also be relevant and revealed with further assessments.
The choice of report organization was influenced by the level of complexity in ecological structure and process, the history of paradigm evolution, and paradigm shifting events. Sometimes something momentous happens in a discipline that stimulates a broad review of existing paradigms. That happened as the high probability of rapid global climate change was recognized for the profound ecological influences it could have. Over the past decade, environmental agencies and nongovernment organizations have been striving to adapt and with it has come considerable reexamination of old science and management paradigms. In large part because of that influence, the review starts with the major shifts in climate paradigms. Being abiotic, climate is also lowest in ecological complexity. Climate paradigms are followed by paradigms examined at species population1, community, ecosystem, landscape and larger ecoregion levels. The order of paradigm description is based largely on concepts of ecological
1 Throughout this report the biological definition of species is used instead of the legal definition in the U. S. Endangered Species Act.
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organization, but is also the approximate sequence of their appearance in the history of ecological paradigm development. The report also is divided into periods before and since the mid-1990s (1996 specifically), when much of the Corps environmental policy was developed.
3. Climate Change Paradigms
3.1 Science Paradigms Overarching Ecological Paradigm: A rapid rate of global climate change has altered and is likely to continue altering ecological structure and function in uncertain and potentially profound ways.
Starting this review with climate change paradigms is a reflection of the great ecological importance of climate and widespread recognition by ecologists that climate is changing rapidly with potentially profound ecological effects. The origins of climate science can be traced back centuries along with observations of ecological responses. Ecologists have continuously accepted the great influence of climate on geophysical and ecological process from the earliest years of the ecological science to modern times (Merriam 1894, Cowles 1899, Clements 1905, Colinvaux 1993, Brewer 1994, Jackson 2010, Cain et al. 2011, Ricklefs and Relyea 2014). 3.1.1 Historical Climate Was Changing Paradigm: In the past, climate was continuously changing and sometimes changed quickly and dramatically. 3.1.1.1 Before 1996 Ecological study of past climate change is most detailed for the late Pleistocene and Holocene, a geological period that spans roughly 15,000 years since the greatest advance of the last continental glaciations. The seminal paleoecological studies of Sears (1935) and Deevey (1944, 1949) began to reveal the instability of climate during the Holocene and its significant ecological effects. Allee et al. (1949) reported those observations in a major textbook of the period, but also believed contemporary times were climatically stable. There was much discussion in the ecological literature about whether short term variation in weather was random or cyclic, but no indication that anyone believed global climate was warming (Allee et al.1949).
Many meteorologists continued to believe that the global climate was variable but stable over many hundreds to thousands of years, and that humans had little meaningful effect. This paradigm was summarized in one meteorology textbook published in 1965:
“In this sense, climate is about as stable as anything we know on earth, about as permanent as the hills” and “the climates of the world appear not to have changed progressively in one direction within the period of history…We have no reason to suppose that any of them ever has or ever will change suddenly or appreciably within a few hundred or even a few thousand years…Further, it is evident that the activities of man cannot influence these major controls of climate.” (Blair and Fite 1965)
Meanwhile, a paradigm shift was in the making. Additional paleoecological research revealed the extent to which climate and ecosystems had changed during the Holocene (e.g., Davis 1969). Researchers studied preserved pollen and remains of marine organisms in sedimentary deposits, geologically recorded water level changes, and oxygen isotope ratios in ice cores (Emiliani 1966, Broecker and van Donk 1970, Dansgaard et al. 1971). Some of the estimated
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changes in temperature were quite rapid (Heusser 1966). By the 1970s, many atmospheric scientists accepted as fact that climate had changed significantly over past millennia (Anthers et al. 1978, Lutgens and Tarbuck 1979). Within two decades, many climatologists and basic ecologists generally accepted the historical instability of regional and global climate (IPCC 1990a and b, Colinvaux 1993, Ricklefs 1993, Brewer 1994, IPCC 1996). 3.1.1.2 After 1995 The climatic instability paradigm has been reinforced by subsequent evidence. Basic ecologists accept the historical instability of climate and recent changes in climate (Odum and Barnett 2005, Cain et al. 2011, Ricklefs and Relyea 2014). Eco-managers are most concerned about recent and future changes, and have devoted significant sections of their texts on evidence of such changes (Van Dyke 2008, Krausman and Cain 2013, Primack (2014) 3.1.2 Global Climate Is Rapidly Changing Largely From Anthropogenic Effects Paradigm: Anthropogenic effects on greenhouse gasses and biosphere carbon storage are causing rapid global warming, greater global precipitation, and acidification of aquatic ecosystems. 3.1.2.1 Before 1996 Over a century ago, Arrhenius (1896) proposed that increased atmospheric carbon dioxide from fossil fuel emissions could increase global warming of the atmosphere based on the physics of gasses and rapidly increasing industrialization. Callendar (1940) concluded that both carbon dioxide and temperature may be increasing based on his analysis of the limited climate data available. Plass (1956) summarized the theory of how carbon dioxide causes climate change and its likely connection to human-caused emissions from fossil fuel use. The New York Times published a summary of his work that same year, reaching a broad audience (Corcoran 2015). But others believed that atmospheric carbon dioxide concentrations were stable, or at least had not changed significantly (Hutchinson 1948, Odum 1959). Much of the uncertainty originated in inconsistent measurement of carbon dioxide before measurement began in 1957 at the Mona Loa Observatory on the island of Hawaii.
The belief of climatologists that contemporary Figure 1. Global temperature records from three climate is stable began to shift with growing sources. confidence in data indicating that human fossil fuel emissions, atmospheric carbon dioxide concentrations, and global temperatures were increasing simultaneously (Figure 1). Matthews et al. (1971) edited a major report by climate scientists that concluded “it is well known that the CO2 content of the global atmosphere has been rising due to the burning of fossil fuels and it is expected to go up another 20 percent by 2000 AD”. Machta (1973) developed a model based on fossil fuel use that projected a rise in concentration to about 385 parts per million by volume (ppMv) in 2000 (the projection turned out to be close to the actual recorded average Source: CSIRO, Australia website of about 370 ppMv). But atmospheric scientists
U.S. Army Corps of Engineers 16 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence generally remained cautious during the 1970s. While Stringer (1972), for example, believed that atmospheric carbon dioxide and temperature had probably increased as a consequence of increased fossil fuel use and reduction of plant cover, he avoided discussion of potentially significant climate change. Other atmospheric scientists of the time were aware of many theories about climate regulation and change, including possible human effects, but remained reluctant to speculate about the degree of human influence (Anthers et al 1978, Lutgens and Tarbuck 1979).
Government-sponsored climate-change research increased rapidly as the concerns of scientists grew. The National Academy of Sciences completed a review of research (NRC 1983) and noted that climate models projected (correctly) a significant decrease in the areal extent of polar ice caps with a relatively small increase in mean global temperature. The models also signaled a need for research along numerous avenues. They concluded that global climate change was probably increasing rapidly but required more research to reduce the uncertainties.
During the 1980s and 1990s, the global climate stability paradigm was quickly shifting to a climate-change paradigm. A major review by an international body of scientists quite confidently concluded that documented changes in the atmospheric climate revealed a progressive and significant global warming over the past century and that the rate of change was higher than changes that had occurred during the Pleistocene and Holocene (IPCC 1990a and b, 1996). The increasing rate of change was linked to increasing rates of fossil fuel use, but the contribution of human effect remained less certain than the change itself. The IPCC also concluded the global hydrologic cycle was changing as a consequence of global warming based on model projections and documented increases in the global average precipitation, stream flow, and intensity of precipitation events (Lins and Michaels 1994, Karl et al. 1995). The rates and amounts of glacial, polar ice cap, and permafrost melting were not well quantified, however (Torn and Chapin 1993). Because methane is a much more powerful greenhouse gas then carbon dioxide, its release from frozen soils was of substantial concern, but also had not been quantified.
By the 1990s, the hypothesis that atmospheric carbon dioxide and global warming are increasing rapidly was gaining wide acceptance in ecological circles, but the ecological effects and degree of human influence on the change remained uncertain (Colinvaux 1993, Ricklefs 1993, Brewer 1994.). 3.1.2.2 After 1995 Based on rapid scientific advances, the IPCC 2013) became much more certain of the high rate of global climate change and the clear human contribution to climate change. They stated that recent global warming is “unequivocal” and “unprecedented over decades to millennia”. Consistent with that warming, the average precipitation over land at middle latitudes has increased since the 1950s, when good records began. Just as important, the extreme temperatures and precipitation are now known to be increasing, causing short-term instabilities that may exceed the tolerances of some species. The IPCC (2013) added that the anthropogenic increase of greenhouse gasses is “the highest in history” and fossil fuel emissions are “extremely likely” to be the main cause of warming since the 1950s. Evidence indicating that increased carbon dioxide contributed to past mass extinctions of marine life is growing (Wingall 2004). Ocean acidity has increased by 25 percent since the beginning of the industrial age because of increased carbon dioxide and further increase will challenge the capacity of many calcifying species of marine organisms to sustain their essential structures (Kump et al. 2009, Albright et al. 2016).
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The certainty of effects on methane production from permafrost melting has grown, adding significantly to the global warming effect of greenhouse gasses (Hodgkins et al. 2013). Even if fossil fuel emissions were dramatically reduced, permafrost melting and greenhouse gas release may continue for decades (Diffenbaugh and Field 2014, Schurr et al. 2015, Davis 2016)
Atmospheric carbon dioxide concentrations and temperature have varied widely during geological history, but usually more gradually than recent history. Recent advances in tracking the geological history of volcanism have generally confirmed volcanic emissions and chemical erosion of volcanic rock as primary drivers of long-term climate change (Kump 2016). However, the generally accepted primary causes of recent climate change are fossil fuel combustion, deforestation, accelerated decay of organic matter, and other ecological changes, which are now believed to have caused a greater rate of climate change than at any time during the past 65 million years (Diffenbaugh and Field 2014, Zeebe et al. 2016). IPCC (2013) projected further warming, more extreme heat waves and precipitation events, and more acidic oceans regardless of management effort, but varying depending on the speed of corrective actions. Most ecologists and many eco-managers now generally accept the evidence of an increasing rate of global climate change caused largely by fossil fuel emissions and human modification of ecological carbon storage (Odum and Barrett 2005, Van Dyke 2008, Cain et al. 2011, Primack 2014, Krausman and Cain 2013, Ricklefs and Relyea 2014). They also believe that these changes will cause many ecological changes and threaten the survival of numerous species in ways that will become more apparent later in this review. 3.1.3 Local Climate Change Is Happening At Uncertain Speeds and Directions Paradigm: Regional climates are changing, but the speed and direction are uncertain and may be greater or less than global change. 3.1.3.1 Before 1996 The IPCC (1990a and b, 1996) projected local climate changes in response to global climate change, but the degree and direction of local change may vary significantly. General circulation models generated much less certain projections of regional climate change because of smaller data sets and larger effects of topographic variation. Empirical data also varied (Karl et al 1995). The IPCC was more confident of differences in the rates of warming along latitudinal gradients, with the least change in the tropics and greatest change in Polar Regions. But at smaller regional scales, model projections were much more data limited and sensitive to changes in assumptions. Uncertainty was even greater for precipitation projections. 3.1.3.2 After 1995 Research since the 1990s generally confirms and refines past conclusions about climate change. More research has been invested in improving global circulation models and linking them to regional climate models through dynamic modeling and statistical techniques (IPCC 2007, 2012). Improvements were facilitated by increased computing capacity and enhanced data streams (Bader et al, 2008). Regional climate models add a level of uncertainty to the significant uncertainty that exists in global circulation models, especially for precipitation, but provide estimates of a range of potential changes at a practical level of land and water resources management (Bader et al. 2008). The uncertainty associated with projected climate change and its effects, including regional variation in climate response to global change is acknowledged by applied ecologists (Van Dyke 2008, Krausman and Cain (2013) and implied by ranges of projected change (Primack 2014).
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3.1.4 Sea Levels Are Rising Paradigm: The ocean volume is expanding and sea levels are rising due to global warming.
3.1.4.1 Before 1996 Figure 2. Global mean sea level change. Sea levels were projected to increase from thermal expansion as the global climate warmed, but the mean global rate of sea level rise (Figure 2) was uncertain until Douglas (1991) accounted for differences in land subsidence and uplift at measured sites. Within a short time, there was little doubt that global sea levels were rising, consistent with projections. 3.1.4.2 After 1995 Over two decades after its first report on climate change, the IPCC (2013) was Source: NASA website much more certain that human activities indirectly contribute to increased mean sea levels as a consequence of global warming. Sea- level rise over the past century can be accounted for by thermal expansion, ice melt, and change in liquid water storage on land. It is “virtually certain” to continue to rise for centuries to come even if emissions ceased immediately (IPCC 2013). Sea-level extremes will also increase and be accompanied by water quality changes. The most saline areas of the oceans are likely to become saltier and the fresher areas less salty. Both applied and basic ecologists are now well aware of sea-level change associated with global climate warming and the effects those changes could have on coastal ecosystems and the species that inhabit them (Odum and Barrett 2005, Van Dyke 2008, Cain et al. 2011, and Primack 2014, Krausman and Cain 2013, Ricklefs and Relyea 2014) 3.1.5 Life Processes are changing as Global Climate Changes Paradigm: Species distributions and abundances, community composition, and ecosystem structure and function are changing in response to global climate change. 3.1.5.1 Before 1996 The importance of climate in determining species distributions has been extensively documented since the early days of ecology, when latitudinal and altitudinal changes in plant and animal occurrence were first linked to climate by Alexander von Humboldt in the early 1800s (Jackson 2010). Merriam (1894) and Grinnell (1917) later built on his ideas. Whitaker (1975) characterized distributions of terrestrial species using plots of mean annual precipitation against mean annual temperature (climographs). The effects of climate were recognized to be variable but assumed in general to be temporally stable and geographically stationary. By the early 1990s, numerous ecologists had warned about the possible destabilizing effects of climate change on species functions, distributions, and interactions (e.g., Curry 1977, Peters 1988, IPCC 1990b, Behrensmeyer et al. 1992, Firth and Fisher 1992, Primack 1993). An increasing number believed that climate change had already caused changes in species distributions and timing of life cycle events, both in the distant and recent past (Behrensmeyer et al. 1992, Wilson 1992).
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However, research reports in general remained more tentative about the degree of climate change, its causes, and its effects, emphasizing the need for more research (Firth and Fisher 1992). This was also evident in major ecological textbooks of the 1990s. Neither Colinvaux (1993), Ricklefs (1993) nor Brewer (1994) mentioned the groundbreaking IPCC (1990b) study of global climate change effects on life processes. They basically thought global climate was warming, and could have important ecological effects, but remained cautious about present rates of climate change, the relative amount of human contribution, and the ecological effects. 3.1.5.2 After 1995 Most ecologists are now much more certain Figure 3. Changes in marine species distribution that measurable increases in global associated with global climate change. From IPCC (2014) warming, global precipitation, sea levels, and and Schwartz (2014). ocean acidity have already occurred because of human activities, and that the effects are likely to continue for decades or more even if the causes of the problem are eliminated immediately. Studies of species distribution and seasonal timing of critical functions and other physiological changes accelerated rapidly after 1995, confirming a pattern of ecological change consistent with global warming effects. Most of the studies documented changes in population distributions and the timing of critical functions (Figure 3).
After reviewing the literature (e.g. IPCC 1996, Parmesan and Yohe 2003, Rosenzweig et al. 2008, Chen et al. 2011, Burrows et al. 2011). Odum and Barrett (2005), Cain et al. (2011), IPCC (2013), Ricklefs and Relyea (2014), and Primack (2014) all emphasized the rapidly growing evidence of worrisome climate effects on species distributions and timing of events affecting species life cycles. Primack (2014) believed that many species may not adapt to the present rate of climate change because it is much faster than changes they were exposed to during their evolution. Mean rates of climate change are often misleading indicators of threat to species viability because changes in the extremes are more likely to determine when species tolerances are exceeded. He also argued that many range adjustments are likely to be impeded by natural barriers to movement (e.g., isolated lakes and mountain tops) as well as by habitat fragmentation caused by human land and water use.
Based on present rates, the distributions of numerous species are likely to shift at least 100 km during the next century and 500 to 1000 km is possible (Diffenbaugh and Field 2014). Past episodes of climate change indicate that species are likely to assemble in new community combinations as climate changes (Cain et al 2011, IPCC 2013, Zarnetske et al. 2014). Alley (2016) reviewed the effects of exceptionally rapid climate change caused by elevated atmospheric carbon dioxide during the Paleocene-Eocene thermal maximum, 55.9 million years ago. While the rate of change was less than recent rates, it was associated with major changes in terrestrial and marine communities, including exceptional extinction rates. He concluded that continued anthropogenic trends could have worse consequences.
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Much recent research on climate-change threats to species has turned to habitat suitability models, which have been criticized for relying too much on projected temperatures and discounting the effects of precipitation, competition, predation, symbiosis, and other interactive factors (Zarnetske et al. 2014). Because of altered precipitation patterns, for example, Chen et al. (2011) and Harsch and Lambers (2014) found that significant fractions of plants and animal species moved down slope and to lower latitudes contrary to projections based solely on warming effects. In a review of the literature, Moritz and Agudo (2013) emphasized the uncertainty associated with forecasts and extinction vulnerability, and the need for more careful species-specific analyses, such as historical analysis of climate and species relationship (Moritz et al. 2008). The collective evidence points to the destabilizing effects of climate change. Contemporary basic and applied ecology texts extensively document changes in life processes that have already occurred and are expected to continue in the future (Odum and Barrett 2005, Van Dyke 2008, Cain et al. 2011, and Primack 2014, Krausman and Cain 2013, Ricklefs and Relyea 2014). Estimates of possible future extinctions caused by climate change alone range between 0 and 54 percent of all species, with a most recent meta-analysis estimate of about one in six species worldwide (Urban 2015).
Many other anthropogenic changes have occurred, which may interact cumulatively with climate change to impact ecological structure and function. Humans have profoundly altered their environment through massive changes in the distributions of nutrients; intentional and accidental introduction of thousands of foreign species; acid deposition, intense harvest, pest control, and other killing, particularly of large species; contamination of land and waters with thousands of toxic and possibly toxic chemicals, and widespread conversion of habitats for agricultural, industrial, commercial, domestic, transportation, energy production and transmission, recreational, and other uses of lands and waters (Miller and Spoolman 2013). But the acceptance of anthropogenic climate change as paradigm has influenced other ecological paradigm shifts more than the recognition of other anthropogenic changes in the environment. A comprehensive report on climate change impacts in the United States is provided by the National Climate Assessment (http://nca2014.globalchange.gov/report).
3.2 Management Paradigms Overarching Paradigm: Achieving ecological management goals and objectives in a rapidly but uncertainly changing climate requires climate-change adaptation.
Most leading eco-managers have accepted the evidence that management goals must somehow accommodate a rapidly changing and uncertain global climate, which is a major paradigm shift since the 1990s. At that time, many eco-managers assumed their local climate and ecological conditions were variable, but generally stable and stationary, despite growing awareness of possible global climate change. The acceptance of new management paradigms has had many implications for eco-management, which built a strong foundation on the assumption that climate and ecological process were at stable equilibrium. Those implications imposed a huge inconvenience, which continues to delay full acceptance.
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3.2.1 Species and Support-System Protections from Climate Change Are Uncertain Paradigm: Exposed to climate change, the sustainability of many species populations and support systems is uncertain everywhere and improbable in many locations that rely on traditional approaches to species and systems protection. 3.2.1.1 Before 1996 Resource management is typically authorized for fixed locations and property ownership. Past management guidance and long-term planning assumed ecological stationarity and stability within property boundaries. This approach applies generally to the management of fish and wildlife resources for their use value as well as to species that require habitat and other protection from consumptive use because of their tenuous sustainability. Any redistribution of environmental variables and species that occurs in response to climate change will affect long- term management planning and objective achievement. Significant changes would have profound effects.
The potential far-reaching effects of global climate change were unevenly recognized in the fish and wildlife resource management texts of the mid-1990s and climate change adaptation in general was yet to be called for. Bolen and Robinson (1995) made no mention of climate change. Scalet et al. (1996), on the other hand, were well aware of the “potential” effect of increasing concentrations of greenhouse gasses on global temperatures. Furthermore, they were aware of the limitations imposed by habitat fragmentation on species adaptation to climate change. They also knew that community interactions may be disrupted by differential species responses to climate change. Their adaptation recommendation, however, was tepid and limited largely to continued monitoring. Primack (1993) also believed that the potential effects of climate change on biological communities would be profound, but provided little guidance for adaptation.
A highly influential National Research Council (NRC) report advising federal agencies on aquatic ecosystem restoration policies and practices (NRC 1992) advised agencies to consider the potentially serious threat of climate change during ecosystem restoration planning. The NRC was particularly aware of the proposed association of sea-level rise with global warming and the foreboding implications for species trapped in coastal wetlands between human development and the sea. Park et al. (1989) had estimated a 65-percent loss of U. S. coastal wetland with a 1-meter rise in sea level even before sea-level measurement uncertainty was generally resolved (Douglas 1991). 3.2.1.2 After 1995 Applied ecologists and managers of ecological variables gradually absorbed the reality of climate change and began to incorporate it into planning considerations for natural resource management (e.g., Michaels et al. 1995, Stakhiv and Major 1997). Groves et al. (2002) accepted the likelihood of rapid climate change, but relied largely on the traditional nature- reserve approach to extinction management and biodiversity preservation. They believed that eliminating local human impacts on reserves would reduce the cumulative stress on individual species imposed by climate change. Their proposed framework also emphasized the need to establish a network of interlinked and representative conservation areas to sustain biodiversity. They believed that approach would help species adapt to climate change, but provided no other insights into a more adaptive planning process.
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Groves et al. (2003) built on the 2002 framework for a guide to drafting conservation plans. They greatly expanded their discussion of climate change and suggested characterizing the future locations of conservation targets using climate criteria and global circulation models. They suggested that managers consider high-priority protection investment for locations projected to change the least wherever they served as refuges from climate change for targeted species, the concepts of which have since been further developed and demonstrated (e.g., Ashcroft 2010, Morelli et al. 2012, Morelli et al. 2016). The uncertainty associated with projections of global circulation models may have been understated, but the main management message was clear; conservationists could no longer assume that climate is stationary during the design of biodiversity reserves. Harris et al. (2005) accepted the high probability of significant global climate change and assessed the serious implications it had for ecosystem restoration. More recently, the IPCC (2013) believed that adaptation to climate change can reduce risks, but there are limits to its effectiveness. They advocated a long-term planning perspective with an eye to sustainable development while emphasizing that adaptation delays may reduce future options.
Contemporary conservation biologists now largely accept the management concepts espoused by Groves et al. (2003). Van Dyke (2008) and Primack (2014) proposed protection of areas projected to be suitable for many species in the future, especially along elevation gradients and north-south corridors in both freshwater and terrestrial environments. Both Van Dyke and Primack knew that much uncertainty still existed about how species will respond to changes. Other than reducing the effects of local environmental impacts, they offered no management strategy for the growing ocean acidity that threatens many species. They advised monitoring management areas and intervening adaptively when needed to meet objectives. Krausman and Cain (2013) dedicated an entire chapter to the “overwhelming” effects of climate change on wildlife. They emphasized the need for bioclimatic modeling, which take a species-specific regional perspective to climate adaptation, and assessments of wildlife vulnerability to climate change. They also recognized that local rates of climate change vary, are much less predictable than global estimates, and should be closely monitored. Speaking for inland fisheries management, Kwak and Freeman (2010) unequivocally believed that “some degree of climate change is certain” but the specifics are uncertain. 3.2.2 Species-Specific Objectives Require Species-Specific Climate Adaptation Paradigm. When management objectives are species-specific, the approach to climate adaptation must be species-specific but considered in an ecosystem and landscape context. 3.2.2.1 Before 1996 This paradigm began to shift toward acceptance of a holistic ecosystem approach to management in place of a species-based approach before the mid-1990s, but has since retreated to a hybrid paradigm that emphasizes species-specific considerations in an ecosystem and landscape context. Early ecological management focused largely on the habitat and other needs of selected species, which relied on species-specific strategies and tactics, whether the objectives were for improved species use or for restoring and protecting species vulnerable to extinction. Leopold’s (1933) seminal book on game management established a species-oriented perspective. But that management paradigm began to shift as some leading ecologists argued for a more holistic and integrative approach to ecology with indications for a more holistic approach to management (Levins 1968, Odum 1977). But the potential danger as seen by some was that the approach was not holistic, but rather a less effective, much coarser scale of management than population-based management. Simberloff (1980), for example, realized that
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the management convenience implied by much greater reliance on an ecosystem approach was very attractive, but he warned that its acceptance in lieu of finer-scaled species-based analyses was inconsistent with the uncertainty and variability associated with many internal dynamics at the population level. A truly holistic approach included simultaneous management of numerous populations while maintaining a broader systems perspective.
Despite Simberloff’s warnings and the predominance of population focus in ecological management, the NRC (1992) report on aquatic ecosystem restoration promoted what amounted to a coarse-scaled approach to management in lieu of fine-scaled population considerations. They deemphasized population-based management and the different vulnerabilities and needs of species while relying on now outdated community and ecosystem paradigms (addressed in later sections). While the NRC (1992) report emphasized the importance of restoring ecosystem structure as well as function, it hardly considered restoration of species composition (structure) and unsustainable species consistent with the concerns of conservation biologists. The approach was much more generally oriented toward restoration of ecosystem-level functions (e.g., primary production, nutrient cycling) and services (e.g. anadromous fisheries and water treatment). The report by the NRC appeared at the leading edge of a new movement in support of ecosystem management and had significant influence on certain aspects of federal resource management. However, the emerging emphasis on ecosystem management had relatively little effect on the predominant population-oriented approach to management, which continued into the 1990s (Bolen and Robinson 1995, Primack 1993, Scalet et al. 1996). 3.2.2.2 After 1995 The ecosystem approach to management rapidly gained advocates among academics and numerous management agencies in the mid to late 1990s (Boyce and Hayney 1996, Vogt et al. 1997, Meffe et al. 2002). The Corps was particularly attracted to the convenience of relying on this approach in its new ecosystem restoration program. Many conservation biologists resisted a total embrace of ecosystem management to the exclusion of species-based management. Groves et al. (2003) promoted the use of ecosystem typologies and surrogate species to represent the collective needs of all species while recognizing the individualistic response of many species to climate and other environmental change. Possible shifts in indicator-species distributions were to be characterized in terms of projected climate shifts using general circulation models (IPCC 2007, 2013). Relying on indicator species was thought to be a more practical and less risky way to preserve the greatest biodiversity than attempting to forecast the range shifts of every species. The use of indicator species was still oriented toward sustaining all species to the extent possible, but selecting the best indicators for preventing the most extinctions was challenging.
Groves and his coauthors recognized the need to manage the risks imposed by climate change as it interacted with other variables. They recommended conserving more locations where conservation targets occur and more widely spaced occurrences than would be needed if climate were stationary; i.e., setting aside more total land and water area for preservation in all parts of all ecoregions. How much was needed to assure sustainability was unclear. They thought that the number and spacing of conservation targets should be considered across their entire historical range of the targeted species as well as the range shifts that may occur as climate changes. They also believed that risks would be managed somewhat by minimizing threats from sources other than climate change. This conclusion was recently reiterated with more evidence of effectiveness by Scheffer et al. (2015)
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In a major report with interagency support, Glick et al. (2011) addressed adaptation to climate change primarily through assessment of species, habitat, and ecosystem vulnerability associated with exposure, sensitivity, and capacity to adapt to climate-related risks. In practice, the report emphasizes the need to evaluate risk largely at the level of species characterizing ecosystems and habitats (Figure 4). Krausman and Cain (2013) emphasized the species- specific effects of increasing rates of climate change and the need to adapt by using species- specific methods such as bioclimatic models. Figure 4. General framework for species-based risk assessment suggested by Glick They also believed that et al. (2011). many wildlife managers mistakenly put off Species Vulnerability Risk Source Attributes Risk climate change adaptation thinking its Narrowly Specialized Unsuitable Habitat In Project Area Project Area Habitat threats were far off in Degradation Disturbance Intolerant Unsuitable Habitat in Project Area the future. Kwak and Matrix Intolerant Unsuitable Habitat In Project Area Freeman (2010) Low Mobility Failure to Establish in Project Area emphasized the need Ecosystem Division for adaptive and Isolation Unable to Move Through Matrix Failure to Establish in Project Area management to Low dispersal ability Failure to Establish in Project Area contend with the “high Narrowly Specialized Failure of Support Species to Establish uncertainty” of future Species interaction disruption climate change. Low competition tolerance Failure to Adapt to Project-Area Community Species biological Primack et al. (2014) disruption High Biological Complexity Failure to Adapt to Ecosystem Conditions also recognized that the Low Population Density Failure to Recover from Event vast amount of data Destructive Stochastic Events demonstrating the Limited Population Distribution Failure to Recover from Event ecological effects of global climate change are based on individual species range and physiological changes, and that species varied in their vulnerability. He accepted rapidly growing evidence that species vulnerability depended uniquely on their diverse population attributes. 3.2.3 Climate Change Adaptation Requires a Regional Planning Perspective Paradigm: Climate adaptation requires larger regional planning perspectives consistent with rates of climate change and the redistribution requirements of species and support systems. 3.2.3.1 Before 1996 Because of lingering uncertainties, adaptation to climate change was only tentatively considered by eco-managers during the 1990s. The most common advice was to think more expansively, both spatially and temporally. The need for a larger regional perspective showed up in recommendations for conservation planning (Peters 1988) and for federal ecosystem restoration activities (NRC 1992). Yet, the need for a wider regional planning perspective was not widely accepted in the 1990s. Primack 1993, Bolen and Robinson (1995), and Scalet et al. (1996) did not address planning adaptation needs for conservation biology or sustained use of fish and wildlife resources.
The main management emphasis of conservation biologists in the 1990s was the design, selection and management of nature reserves, and that was largely focused on the needs of
U.S. Army Corps of Engineers 25 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence species on an unsustainable track toward extinction (Primack 1993). The design and selection of nature reserves based on the scarcity of entire ecosystems had yet to gain much conservation interest. Other than to indicate that bigger preserves were in general better than smaller ones, Primack (1993) had little to say about adapting to climate change by taking a regional approach to species conservation and climate change adaptation played no explicit role in the prioritization and selection of areas for preservation.
The potential for global change in climate was just beginning to be incorporated into the planning of federal agencies in the early 1990s (e.g., Naiman 1992). The authors of the NRC (1992) report on ecosystem restoration recognized the need to factor climate change into ecosystem restoration plans. They reiterated Peters’ (1988) three strategies for avoiding future species extinctions as climate changes: increased reserve size to incorporate a wider range of habitat types, moving reserves as climate changes, and establishing new reserves in other climate zones in anticipation of shifting habitat suitability. Yet the report provided little detail about how to adapt to climate change in the context of ecosystem restoration planning, particularly at the level of individual species. The agencies were left largely to their own devices. Subsequent policy guidance for ecosystem restoration in the Corps skipped the issue of climate change.
The NRC report on ecosystem restoration was also inconsistent in its approach to the climate- change issue. While some authors apparently believed that climate profoundly influenced ecosystems, and those conditions would likely shift, the report was based on the concept of sustainable ecosystem restoration. Nothing was written about the increasing improbability of holistically restoring previous conditions or considering climate change adaptation during project planning. The main recommendation for sustaining coastal wetlands also was inconsistent with the restoration objective of the NRC report. It advocated preservation of a buffer strip and creation of conditions that would allow wetland migration inland. While logical, this recommendation created a new ecosystem condition further inland. It seemed apparent that the challenges posed by global climate change had not been fully assimilated into the ecosystem restoration thinking of the time. Whether resource development, ecosystem restoration or environmental protection was their primary concern, most eco-managers continued to assume climate is stable and stationary. 3.2.3.2 After 1995 Contemporary eco-managers generally accept a need to account for and adapt to climate change in their management plans. Managers of resource wildlife have in general been slower to adopt a regional approach to climate adaptation than conservation biologists, but Krausman and Cain (2013) described the use of bioclimatic models as one of the main avenues toward climate adaptation. Reflecting contemporary conservation biology, Primack (2014) proposed that global climate change needed to be incorporated into nature reserve design planning and management, mainly by creating new reserves and corridors for species range adjustments as change occurred. The details have yet to be worked out, however. For example, Groves et al. (2012) only recently proposed that new reserve designs be based on the diversity of fixed geophysical features that are more likely to resist climate change than existing biological indicators of reserve suitability. Others have reviewed the paleoecological evidence of climatic refugia and proposed a way to forecast the locations of future climate microrefugia based on interactions between climate and local terrain (Dobrowski 2010). The issues involved with biological and geophysical approaches have yet to be thoroughly evaluated.
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3.2.4 Climate Change Adaptation Requires Integrated and Adaptive Management Paradigm: Climate change adaptation requires cooperative organizational integration and adaptive management planning and execution. 3.2.4.1 Before 1996 As more national emphasis was placed on environmental quality protection and improvements, uncertainty in eco-management outcomes became an increasingly important issue. Adaptive management was proposed as one means for dealing with it (e.g., Holling 1978) and later for facilitating climate change adaptation. A formal approach to adaptive management (Walters 1986) was new to natural resources management in the 1990s and was inconsistently recognized for its importance in the management community. The informal approach was little more than an ad hoc trial and error process that failed to improve institutional knowledge very effectively. The formal approach built later management flexibility into planning, incorporated pilot experimentation when needed, required rigorous research-grade monitoring, used planning models that could be improved with learning from the monitoring results, and responsively adjusted management actions to achieve objectives (Walters 1986). Adaptive management was intended to be a continuous process of monitoring and adjustment as needed throughout the project life-cycle (Figure 5). Figure 5. Basic elements of adaptive management for climate change effects. Gunderson et al. (1995) identified numerous barriers Assess to more effective integration of diverse management Objectives authorities into a more focused approach to Problem achieving environmental sustainability. “Piecemeal” policies and policy differences were frequent barriers, Adjust Design which in government can be traced back to legislative process. They emphasized that there was too much Adaptive Management “command and control in natural resource management, which often excluded “citizen” Cycle involvement. Other barriers relate to unclear Evaluate Implement statement and understanding of objectives and what each partner in the process is to bring to their achievement. Monitor Adaptive management was just beginning to be recognized for its potential in the federal eco- management community, particularly in the U. S. Forest Service (e.g., Stankey and Schindler 1997). The NRC (1992) report on federal aquatic ecosystem restoration mentioned the suitability of adaptive management for ecosystem restoration because of the uncertainties involved. It also recognized the need for inter-organization cooperation for large and complex projects. Some federal agencies were beginning to integrate adaptive management into their planning process, but not specifically to address climate change effects. However, neither Primack (1993) nor Bolen and Robinson (1995) and Scalet et al. (1996) mentioned the need for adaptive management or more inter-organization cooperation, coordination, and collaboration. 3.2.4.2 After 1995 Adaptive management, in various forms, became widely accepted as a fundamental need in ecological management, including climate change adaptation, over the last two decades, but the importance of a more fully integrated organizational approach has not been so consistently
U.S. Army Corps of Engineers 27 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence recognized. Primack (2002) first mentioned the utility of adaptive management in the third edition of his textbook on conservation biology. The NRC (2004) recommended that federal water resources management agencies incorporate adaptive management into project planning, including ecosystem restoration and protection planning. Adaptive management, as it applies to climate-change adaptation is defined by IPCC (2012 page 3) “as the process of adjustment to actual or expected climate and its effects, in order to moderate harm or exploit beneficial opportunities.” For wild, self-regulating ecosystems, adaptation is “the process of adjustment to actual climate and its effects; human intervention may facilitate adjustment to expected climate” (IPCC 2012 page 3).
Cortner and Moote (1998) identified the same barriers to collaborative integration mentioned by Gunderson et al. (1995). Many of the problems remain fundamental in the legislative process, which often encourages cooperation while establishing conflicting authorities and responsibilities within and among organizations. Some organizations let competition for scarce funding interfere with cooperation. While the problems are generally clear, the solutions are not, given the many differences in public attitudes towards the environment and the role of government.
Because of the uncertainty of climate change projections and biological responses, Krausman and Cain (2013) stressed the importance of monitoring, the utility of adaptive management for managing uncertainty, and the need for regional collaboration. They referred to and described the conservation partnership formed at landscape and regional scale by the LCCs as a national example (the LCCs are described in the Introduction and Cole et al. [2018]). Several chapters in Hubert and Quist (2010) emphasize the importance of managing the fisheries of inland waters in the context of their watersheds and including watershed management among measures used when practical. Both Cain et al. (2011) and Primack (2014) incorporated a detailed but generally applicable model of adaptive management process into their texts. They thought the potential effects of climate change should be carefully monitored along with other environmental changes that need to be adaptively managed. In general, the paradigm shift to adaptive management as a means for dealing with the uncertainties of climate change and other environmental change has advanced rapidly and is quickly becoming a standard approach to uncertainty management. The need for organizational collaboration is often taken for granted in eco-management texts, but its importance is frequently revealed in case studies and other examples of real-world eco- management.
4. Population Paradigms
4.1 Science Paradigms Overarching Paradigm: Species are composed of one or more discernible populations made up of species members with characteristic demographic, genetic, and ecological attributes that fit them uniquely into community and ecosystem structure and functions.
Population ecology has played a minor role in past Corps eco-management, but recognition of global climate change has elevated its relevancy, which should become clearer in later paradigm discussions. The roots of contemporary population ecology are found in early actuarial analyses of human populations; the theory of organic evolution through natural selection; exploration of species interactions in biotic communities; and early fish, wildlife, and pest population management. Later in the 20th century, growing concern over accelerated rates of human-caused extinction joined the major interests in population ecology.
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A coherent population concept did not fully form until after ecological concepts of species communities were first proposed in the late 1800s. Population ecology grew rapidly in the early 20th century out of quests to explore the theories of community ecology and genetic variation using more quantitative and experimental approaches. The idea that populations of each species sustained stable equilibriums with the conditions in their biotic community became an ecological paradigm that fit with and was influenced by a similar paradigm that developed in community ecology (described in the next section).
The paradigms of community and ecosystem structure and function cannot be fully appreciated without some understanding of population paradigms, and many of those paradigms have shifted in response to shifts at the ecosystem level. Some of the important contemporary paradigms of population ecology were generally accepted by ecologists by the 1990s, but are included in this review because of their relevancy to other eco-management paradigm shifts, which in general reflects the importance of population-based management in fish and wildlife resource management, pest management, and biodiversity conservation. 4.1.1 Populations Are Identifiable Evolutionary Units of Species Membership Paradigm: Populations are spatially discrete evolutionary units of species membership maintained by the balance of birth, death, emigration and immigration rates; they often differ genetically, demographically, and ecologically. 4.1.1.1 Before 1996 The population concept has long been a demographic concept, having to do with the statistics of population structure and dynamics. It is also an evolutionary concept determined largely by genetic attributes having evolutionary significance as well. The boundaries of demographic units may be defined by any means, including evolutionary-unit and political boundaries. Demographic units derive much of their ecological meaning from population birth, death, immigration and emigration rates; age and sex structure; and numerous other variables. The boundaries of evolutionary units are defined by degrees of past and present spatial separation, which influence genetic mixing. The effect of spatial isolation on genetic variation from other populations became the generally accepted explanation for how species originate after prolonged isolation (Mayr 1970). Populations may or may not remain genetically isolated once spatial isolation ceases depending on the degree and types of genetic and phenotypic changes that take place while spatially isolated.
The demographic basis of the population concept originated centuries ago in analysis of human population dynamics (Halley 1693, Malthus 1826). Pearl and Reed (1920) were among the first to clearly articulate a generally applicable demographic-unit concept of populations and a theory of population change and regulation. The demographic unit soon became the foundation for fish and wildlife resource management. The concept of a population as an evolutionary unit was first implied by Darwin in his theory of species origin through natural selection of heritable traits (Darwin 1859), but was not fully formed until the fields of population genetics and evolutionary ecology were established decades later. The evolutionary unit is the basic unit of interest for conservation biologists and of growing interest for resource managers since compliance with the Endangered Species Act (ESA) became an important aspect of their activities.
Until the 1920s, species and communities were the fundamental units of ecological study and the population link between them was conceptually vague. The rise of population ecology was associated with an increasing interest among ecologists in quantification, experimental analysis,
U.S. Army Corps of Engineers 29 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence evolutionary process, and development of a more coherent theory on community interactions. Population boundaries were often determined by the demographics within a defined space and were often defined by a laboratory container for small animals. Most of the basics of demographic population ecology were generally accepted by ecological scientists and eco- managers by mid-20th century (Allee et al. 1949, Odum 1953). Population life tables were widely used to track and forecast changes in population numbers (Ricklefs 1993, Brewer 1994).
The scientific notion of an evolutionary unit, defined as a genetically discrete group of organisms that successfully reproduce with each other much more than with individuals outside the group was greatly advanced by the works of early population geneticists (e.g., Haldane 1927, Fisher 1930, Wright 1942) and was increasingly accepted in principle by ecologists of the 1940s and 1950s (Allee et al. 1949, Odum 1959, McCollough 1996). The delineation of evolutionary units based on genetic differences became more explicit as techniques improved and conservation biologists began to focus more on preventing genetic loss (Primack 1993).
Andrewartha and Birch (1954) were perhaps the first to recognize that many populations Figure 6. Metapopulations are sustained by habitat delineated by reproductive interactions were connectivity between subpopulations. composed of incompletely isolated subpopulations that interacted reproductively at sub- travel corridor sub- rates depending on degree of isolation. Levins population population (1969) called those reproductively interactive 1 2 subpopulations “metapopulations” and developed a model that could be used to examine the balance between subpopulation extinctions in suitable habitats (which happens frequently) and habitat patch colonization by repopulation members dispersing from other subpopulations local (Figure 6). Primack 1993) accepted the extinction metapopulation concept and its importance for conservation biology. It and the concept of METAPOPULATION population evolutionary units were recognized, but were less widely considered important by fish and wildlife resource ecologists (Bolen and Robinson (1995) and Scalet et al. 1996). 4.1.1.2 After 1995 The broad concept of demographic units remained important in fish and wildlife resource management while the relative emphasis on the evolutionary concept of population units continued to grow, partly in response to the cumulative effects of climate and other environmental change. Odum and Barrett (2005) described populations as energy-flow subsystems in ecosystems, providing a critical conceptual link between models of population demographics and ecosystem process models. Pope et al. (2010) included genetic separation as well as spatial and demographic separation as important means for delineating population units in fishery management. Perhaps more than earlier ecologists, Cain et al. (2011) recognized that it is sometimes hard to ecologically determine the spatial extent of a population unit or even to determine individuals in some cases. This is particularly true for many plant and fungal species. Ricklefs and Relyea (2014) recognized that basic ecologists were more likely to consider populations units of community organization and evolutionary understanding while political boundaries were often used by managers to define populations demographically. In their wildlife management book, for example, Krause and Cain (2014) defined wildlife populations as “individuals of species occupying a defined area for which it is meaningful to
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refer to birth and death rates, sex ratios, abundances, and age structures”. They also recognized the importance of species conservation in increasingly instable ecological settings as well as its evolutionary basis for population management.
Conservation biologists Van Dyke (2008) and Primack (2014) were quite explicit about the importance of limited gene flow and the genetic distinctiveness of evolutionary units. Van Dyke (2008) identified spatial, genetic, and demographic disjunction as the main ways that the biological boundaries of population units are determined. Spatial separation is the obvious indicator of reproductively separate populations and the potential for the other two disjunctions to exist. Differences in genetics and in sex ratios, age structure, and other demographic attributes become increasingly important as population numbers decline (Primack 2014). Since spatial separation is not a sure indictor of genetic or demographic differences, their disjunctions need to be separately assessed when they are significant. 4.1.2 Populations Occupy Unique Ecological Niches and Habitats Paradigm: Distinct populations of each species uniquely occupy ecological niches and habitats. 4.1.2.1 Before 1996 Natural historians have been fascinated by the diversity of species since Aristotle, who developed the first classification scheme for animals. Taxonomists have used a standard approach to classifying species based on trait differences since Carl Linnaeus developed the universally accepted binomial classification system in 1735. Darwin’s (1859) treatise on the origin of species through natural selection was his explanation for how those differences arose through organic evolution and contributed to the development of many ecological concepts, including the population concept. Numerous natural historians began to observe and record differences in species life histories.
Grinnell (1904) is often credited with the origin of niche theory when he proposed one species would inevitably crowd out another species when their food habits were similar. Elton (1927) believed each species occupied a unique “functional niche” within community food webs. The concept of niche that is now most favored was also influenced by the earlier concept of environmental tolerance (Shelford 1913). Shelford proposed that species were not only limited by too little of some environmental factor, as Sprengel and Leibig had proposed much earlier (Van der Ploeg et al. 1999), but were also limited by too much. The range tolerated by a species brackets the conditions where a species can survive. Shelford recognized that environmental tolerance varied among species and uniquely determined the limits of species ranges. His proposal soon became a paradigm as empirical data on species range and tolerance differences accumulated. Odum (1959) identified some corollary principles which also became widely accepted. Based on empirical evidence, he believed species may broadly tolerate some factors while narrowly tolerating others, that species with consistently wide tolerance ranges are likely to be most widely distributed, that less than optimal conditions for one or more factors may reduce the tolerance range for other factors, and that tolerances may vary seasonally and geographically among different populations.
MacFadyen (1957) and, particularly, Hutchinson (1957) proposed what has become the contemporary concept of species niche, which is based on resource and environmental requirements, including the forms of suitable nutritional energy and their vulnerability to factors other than food limitation. Hutchinson (1957) built on the tolerance concept of Shelford to develop a multi-dimensional concept of niche “hyper-volume” defined by the range of diverse conditions tolerated by a species. In this concept, environmental extremes are more important
U.S. Army Corps of Engineers 31 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence determinants of tolerance limits than averages. Hutchinson recognized a difference in the fundamental niche, defined largely by the abiotic extremes tolerated in the physical environment, and the realized niche, which included the effects of competition, predation, obligatory symbionts and parasites, and other biotic variables that derive from community interactions with the abiotic variables. The Figure 7. Species niche defined simply by the geographical abiotic variables of the fundamental niche presence and absence of a species with respect to two are relatively easy to measure (Figure 7) variables. compared to the biotic variables of the realized nice, which are often indicated by the consequences of biotic interactions, such Niche as the survival of competitors and prey, although various morphological and behavioral measures have been used. Shelford’s tolerance concept was universally accepted by the 1990s (Ricklefs 1993, Species presence Colinvaux 1993, Brewer 1994, Bolen and Temperature Species absence Robinson 1995, Scalet et al. 1996, Primack 1993). The niche concept was generally accepted, but less explicitly so by some applied ecologists (Kohler and Hubert 1993, Primack 1993). Precipitation The concepts of habitat and niche were closely linked but different. Elton (1927) defined habitats as the places where individuals of a species are located—the place that provides all of their food, cover, reproduction, locomotion, and other functional needs as well as stressors. He implied that habitat included other species as well as the nonliving environment. Odum (1959) echoed Elton’s concepts of habitat and niche. The interactions between niche and habitat varied within species and probably among different populations of the same species. Wildlife resource managers of the 1990s continued to see habitat primarily through the lens of a single species “address” in the environment (Bolen and Robinson 1995, Scalet et al. 1996).
Plant ecologist Tansley (1935) defined habitat a bit differently—as the physical setting of entire communities. His concept of habitat differed from Elton’s by being limited to the nonliving environment. The interactions between community and its physical environment, or habitat, became the basis of Tinsley’s (1935) concept of ecosystem. While a community is made up of species, the concepts of community identity at the time were based largely on vegetation structure and dominant plant species, which were also believed to indicate a coherent community unit. The sense of habitat as strictly physical and supportive of communities as well as species was also common in fisheries management (Orth and White 1993). However, plants, which form plant communities in the minds of plant ecologists and managers, become a physical setting or cover for animal ecologists and managers (Bolen and Robinson 1995, Scalet et al. 1996). The differences are relevant to niche-concept applications, which are species- specific and probably impossible to apply at the collective community level.
When applied to species, the physical concept of habitat was easily linked to Hutchinson’s concept of the fundamental niche while the broader, biophysical concept of habitat accepted by many zoologists linked with Hutchinson’s concept of the realized niche. Both concepts of habitat and niche persisted into the 1990s and ecologists often did not clearly distinguish between them. Ricklefs (1993) for example, mentioned that the definition of habitat depends on the species the habitat serves. But he also mentioned that ecologists spent much time classifying
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“habitat types”. As pictured in Ricklef’s text, habitat types include red cedar forest, coral reef, and acacia woodland, all of which are considered community or ecosystem types by other ecologists. In sum, the habitat concept was contextual and varied depending on perspective. If the concept was linked to the species niche, a species perspective was most appropriate. 4.1.2.2 After 1995 Basic ecologists and eco-managers now widely accept the utility of niche theory and differentiate it from the species concept of habitat that is now most widely accepted. Van Dyke (2008) presented niche assembly theory as the accepted paradigm for conceptual understanding of species assembly into communities. Cain et al. (2011) described how niches often change with life stage, demonstrating the complexity of complete niche definition and modeling. Niche models have become a staple in evaluating the destabilizing effects of climate and other environmental change on species distributions and viability (Cain et al. 2011, Primack 2014, Ricklefs and Relyea 2014). Eco-managers also widely accept the niche approach to management as a management fundamental (Morrison 2009, Krause and Cain 2014, Primack (2014). Primack (2014) believed that narrow niche width is a reliable indicator of population vulnerability to extinction. The vulnerability of many species is likely to increase as global climate change accentuates local weather extremes, such as heat waves and droughts (IPCC 2013).
The habitat concept remained somewhat elusive and dependent on perspective, but the concept of community habitat has faded in importance. Morrison et al. (1998) recognized this and, unlike many who used the term without clearly defining it, defined habitat as “an area with a combination of resources (like food, cover, and water) and environmental conditions (temperature, precipitation, presence or absence of predators and competitors) that promotes occupancy by individuals of a given species (or population) and allows those individuals to survive and reproduce”. In this now widely accepted view, habitat is that part of an ecosystem that interactions with a species of interest. While Morrison et al. (1998) knew that the notion of habitat as the physical environment of a community was still accepted among some plant ecologists (as a capacity to produce vegetation of the same general type), they emphasized a species-specific concept that included all variables in the species’ environment for practical management purposes.
Pegg and Chick (2010) tried to accommodate all views by defining habitat broadly “as the area where an organism, population, or community occurs in the environment”. The definition appears to include both abiotic and biotic attributes of the area’s environment. They found it difficult to separate concepts of habitat and ecosystem, and they considered ecosystem restoration a form of habitat management. The prospect that habitat suitability for each species is unstable and nonstationary in changing climates is now widely accepted by contemporary ecologists and leading eco-managers (Cain et al. 2011, Ricklefs 2014, Primack 2014, Krausman and Cain 2013). There is no similar consensus regarding shifts in habitats for community units (this is addressed further in the Section on community paradigms).
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4.1.3 Population Members Compete for Ecological Resources Paradigm: When niches overlap, members within populations and populations of different species compete for and partition use of limited energy, nutrient, space, and other resources as competitive advantages vary with environmental change. 4.1.3.1 Before 1996 The idea that individuals compete within their own population and with members of different species populations for resources in limited supply was a widely accepted paradigm by the (Colinvaux 1993, Ricklefs 1993, Brewer 1994), but the degree to which competitive advantages vary with environmental change was not so clearly understood. Darwin (1859) believed competition is a major force in natural selection. Ecologists widely accepted the idea that resources are partitioned among species and competition occurs whenever the individuals of the same or different species seek the same limiting resources, or when species harm one another in pursuit of the same resources regardless of resource supply (Andrewartha and Birch 1954, Colinvaux 1993).
The concept gained theoretical rigor after Gause (1934) developed and refined simple two- species models for testing interspecific competition hypotheses. He, as well as others, observed one species die out as the other species thrived on a food consumed by both species (e.g., Park 1948, Allee et al 1949). Interspecific competition was demonstrated in the wild less frequently, but the ecological importance of interspecific competition became a widely accepted paradigm because it explained many observed community changes (Colinvaux 1993, Brewer 1994). The evidence for competition was gathered largely under controlled laboratory or generally stable field conditions consistent with prevalent assumptions about ecosystem stability. The competitive advantages of one species over another were generally assumed to be stable while the intensity of interspecific competition was believed to vary depending on niche similarity. Species within “guilds” of species with similar niches (Root 1967) were believed to compete more intensively with each other than with species in other guilds. By the 1990s, interspecific competition was generally accepted as an important key to explaining the failure of population members to fully occupy their fundamental niche in a suitable physical environment (Colinvaux 1993, Brewer 1994).
Odum (1953, 1959) used population energetics to explain how species partitioned resources to determine their population numbers. He characterized population respiration, biomass production, and population functions and structure in caloric units to enable integration of population dynamics with ecosystem energy-flow models. Detailed analysis of individual population energetics began with the works of Richman (1958) and Slobodkin (1960) on laboratory populations of small crustaceans. Partitioning energy and other resources varied among species and required coefficients to simulate competition in models (Schoener 1974). These coefficients were usually assumed to be constant, in keeping with prevailing ecological stability assumptions, but would later be recognized as variables. 4.1.3.2 After 1995 Basic ecologists now generally accept the premise that interspecific competition alters fundamental niches, but the competitive dominance of a species can change in response to environmental change (Cain et al 2011, Ricklefs and Relyea 2014). Conservation biologists Van Dyke (2008) and Primack (2014) accepted the importance of competition primarily with respect to the potential effects of foreign invasive species on the distribution, abundances, and extinction probabilities of native species. However, Taylor (2004) found little evidence in the fossil record that competition with invasive species was severe enough to cause much
U.S. Army Corps of Engineers 34 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence extinction of resident species. Except for human effects on habitat use, the extent of competition’s role in population and species extinction remains unclear.
In an important shift influenced by widespread recognition of rapid climate change, contemporary ecologists recognize the importance of environmental stability in determining the consistency of competitive advantages. In general, competition between species is now viewed as more dynamic in unstable environmental settings than once perceived and less likely to be a consistent driving force in population extinction (Cain et al 2011, Ricklefs and Relyea 2014). Ecologists generally believe that the consistency of competitive advantage can be altered by abiotic changes associated with climate and other environmental change, changes in biotic interactions such as predation and other sources of competition, and other disturbances of ecological equilibria. Alexander et al. (2015), for example, concluded that the potential effects of competition in alpine plant communities that reassemble as climate changes may require closer attention to changing interactions to more accurately predict species responses.
The generally accepted influence of environmental stability on competition has important implications for community-level modeling. Odum and Bartlett (2005), for example, built on the energy flow concept of Odum (1959) to further integrate individuals, populations and communities into a universal systems model for energy flow. They conceived of populations as open systems within community systems and ecosystems in keeping with population energetics models (e.g., Kitchell et al. 1974). Models of this type often used resource partitioning coefficients (Schoener 1974) that relied on estimates of competition effects under stable conditions, but the assumption of stability has become less tenable with increased doubts about ecological stability. 4.1.4 Population Size is regulated by Density-Dependent and Density- Independent Factors Paradigm. Maximum population numbers are regulated by density-dependent factors in species-unique ways that establish a dynamic equilibrium at the carrying capacity of a relatively stable environment while density independent factors contribute more to population fluctuations in less stable environments. 4.1.4.1 Before 1996 Early natural historians were fascinated by the apparently consistent abundances of various species in undisturbed environments. Influenced greatly by Darwin (1859), species extinction was believed to be a very slow process balanced by the gradual origin of new species. This stability of populations and species was taken as evidence of natural harmony or the “balance of nature” that results from the stability of species interactions in community associations (Botkin 1990). If and why this “balance” actually occurred was not closely examined before the 1920s when more coherent theories of population regulation began to emerge.
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First articulated by Pearl and Reed (1920), Lotka (1925) and Gause (1934) further developed and refined the theoretical concepts of biotic potential and environmental carrying capacity. They presented them in the general form of a Figure 8. The logistic and exponential models for logistic model of population growth in a stable population growth. environment (Figure 8). From a few initial members with unlimited resources, the modeled density-dependent population number increases slowly at first then more Carrying capacity (K) rapidly, reaching a maximum rate of increase (the intrinsic rate of natural increase or biotic K-selected potential) just before environmental species resistance begins. Then, with increasing r-selected species LOGISTIC GROWTH CURVE resistance to further growth, the rate slows EXPONENTIAL and gradually comes into equilibrium with the GROWTH environmental carrying capacity. Smith (1935) CURVE was the first to use the term “density- INTERMEDIATE GROWTH
dependent” regulation. The intrinsic rate of Population Size increase is defined as r and the carrying capacity is K. In the experimental environments used to test the proposed Time model, a single-species population was typically initiated in a food- and space-limited environment. Under those conditions, intraspecific completion was the obvious density- dependent factor, but interspecific competition, predation (including cannibalism), disease, and intrinsic intolerance of crowding were also hypothesized to play roles in wild environments. The environmental carrying capacity was assumed to be stable enough to allow population equilibration. At the time, both climate and mature biotic communities were believed to be stable over quite long periods of time (e.g., Clements 1916, Weaver and Clements 1928).
Leopold (1933) recognized the value of the logistic model and the concept of density-dependent regulation for game population management. But his field experience also revealed “the balance of nature” at carrying capacity was not static; population numbers typically fluctuated, sometimes dramatically. The growth parameters of the observed species often differed from the standard logistic model. Leopold also observed that reduced predation was associated with a less tightly regulated rate of deer population growth, which often overshoots and temporarily depresses carrying capacity by reducing food abundance. This produced a J-shaped exponential growth curve that peaked, precipitously collapsed, and then continued to fluctuate through time around the level supported by the depressed carrying capacity. He accepted a concept of dynamic equilibrium over a static equilibrium, with some species populations fluctuating more than others.
Others independently observed that insect pest populations also exhibited J-shaped growth curves, Uvarov (1931) and Andrewartha and Birch (1954) proposed that the dramatically fluctuating population dynamics occurred with changes in weather. Optimal weather allowed rapid growth of populations, which crashed precipitously when droughts occurred. Andrewartha and Birch (1954) found consistent density-independent effects of dramatic weather events among several insect species and claimed it was the predominant form of population regulation among all species. On the other hand, Nicholson (1933) and Lack (1954) argued that density- independent factors could temporarily contribute to sharp population changes, but density dependent mechanisms ultimately set the limits for the highest sustainable population numbers of all species. Nicholson (1933) believed the logistic model was too simplistic for many
U.S. Army Corps of Engineers 36 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence populations and he proposed some alternative models to include a wider variety of variables. But Lack (1954) and others persisted in the belief that the logistic model exhibiting nearly static equilibrium was the usual condition exhibited by populations.
The net result of a large number of studies since then indicated that some groups of species (such as insects) are less likely to sustain stable populations than others (such as vertebrates), even when the environment appears stable. Park (1948) paid more attention to the different outcomes of experiments with meal worms, revealing the variable nature of population growth curves. But the exponential and logistic models continued to be the widely accepted parameters for characterizing the range of population growth conditions. Weigert (1974) argued that the population growth of many species populations was actually intermediate between exponential (density-independent) and logistic (density-dependent) models (Figure 8). Because many ecosystems are characterized by environmental changes and all environments are at least somewhat variable, ecologists in the 1990s generally believed that few populations are so tightly regulated by density dependent mechanisms that they remain static and at least some density- independent influence is probable in all but the most stable environments (Colinvaux 1993, Brewer 1994).
All of the accepted models were deterministic. Simberloff (1980) believed deterministic models were unrealistic based on evidence showing the effects of random events on growth curves within the same populations, exposed to the same conditions. Simberloff (1988) especially emphasized the importance of random events in his description of demographic, genetic, and environmental stochasticity. Yet, the logistic and exponential models of population growth continued to be the accepted population growth paradigms used to explain much about the regulation of population numbers by density dependent and density independent factors (Colinvaux 1993, Ricklefs 1993, Brewer 1994). However, they also recognized the variable behavior of populations within a range of conditions generally defined by the logistic and exponential models. 4.1.4.2 After 1995 As past assumptions about climate and other environmental stability proved unreliable, assumptions about the relative importance of density dependent and independent regulation has shifted in favor of logistic models with dynamic equilibria and density independence models. The emphasis of random events on population regulation paradigms also continued to increase after the 1990s. Random events may originate outside or within populations (Rockwood 2006). Rockwood (2006), Cain et al. (2011), Ricklefs and Relyea (2014), Krausman and Cain (2013), and Primack (2014) generally agreed that density-dependent and density-independent factors interact to play important roles in regulating population numbers, but the specific responses of populations to environmental and population conditions is unique. They agreed that the older deterministic models were less realistic than contemporary stochastic models that incorporated random events, which are particularly important in the dynamics of small populations. Few ecologists believe that the numbers of any populations are absolutely stable for any appreciable amount of time.
Cain et al. (2011) placed greater explicit importance on the density-dependent regulation of population numbers than others, but also recognized that environmental conditions and community interactions change, often leading to continuous re-equilibration with changing carrying capacities. Ricklefs and Relyea (2014) believed carrying capacity often fluctuates and some exponentially growing populations overshoot environmental carrying capacity and may reduce it. Krausman and Cain (2013) accepted generic models as useful as long as researchers and managers realize that the specific population growth rates and limits for each species vary
U.S. Army Corps of Engineers 37 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence widely and are influenced by demographic, genetic, and environmental stochasticity. These observations imply that populations of density-dependent species adapted to dynamic equilibriums are more vulnerable than density-independent species to climate and other environmental change. The differences among population growth behaviors have major implications for species assembly into communities as ecosystems change. 4.1.5 Increasing Environmental Instability Favors r-Selection over K- Selection Strategies Paradigm: Increasing environmental instability favors species with a large number of broadly adaptive traits associated with a high intrinsic rate of increase, r, over species with a large number of narrowly adaptive traits favored by relatively stable environmental conditions at carrying capacity K. 4.1.5.1 Before 1996 Figure 9. Attributes of r and K selected species (from Pianka 1966). Cole (1954) integrated early studies of population survival strategies into a more coherent theory of Characteristics of r- and K-selected population growth. Then Pianka (1966, 1970) Species established the concepts of survival strategy most accepted by the ecologists of the 1990s. Pianka’s Characteristic r-selected K-selected concept reflected acceptance of the logistic equation Maturation time short long of population growth, being based on intrinsic rate of Lifespan short long population increase, r, and adaptation to equilibrium Death rate usually high usually low conditions, K, in stable environments (Colinvaux Offspring many few 1993, Primack 1993, Brewer 1994). Many ecologists Reproductions/life once several had come to believe that populations with high First reproduction early in life late in life intrinsic rate of increase and relatively poor Size of offspring small large adaptation to generally stable environmental Parental care none extensive conditions exhibit characteristics that enable rapid population growth following disturbance (Colinvaux 1993, Brewer 1994) (Figure 9). Pianka (1970) referred to them as r-selected or opportunistic species. Species with low biotic potential and adaptation to conditions at the environmental equilibrium were believed to adapt to more-or-less stable ecosystem conditions through specialization. Pianka (1966, 1970) referred to them as K-selected or equilibrial species.
Pianka (1970) believed the data he compiled indicated that species were bimodally distributed between r and K-strategies, but others came to believe that many species exhibited mixed characteristics and could not be simply categorized by any single criterion (e.g.,Weigert 1974). Colinvaux (1993) believed that the competitive strength of species exhibiting different “strategies” varied along an opportunist-equilibrial continuum. Extreme opportunists rely largely on high biotic potential (r) while extreme equilibrial species rely on especially good adaption to the equilibrium condition (K). Equilibrial species compete more effectively for resources under stable conditions while opportunistic species quickly take advantage of less stable states with newly available resources before equilibrial species can become established.
Opportunists typically dominate in more unstable, often dramatically disturbed habitats with changing attributes much less favorable for equilibrial species. Opportunists are the weed species. Their populations are more likely to grow exponentially in ecologically immature ecosystems while the populations of equilibrial species grow more slowly toward equilibrium and dominate in relatively stable, ecologically mature ecosystems. Opportunists often have wide fundamental niche dimensions, are dietary generalists, and have high dispersal capacities and a
U.S. Army Corps of Engineers 38 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence short life span. Equilibrial species often have narrow fundamental niche dimensions, are dietary specialists, and have low dispersal capacity and a long life span. Because opportunists do not compete well with equilibrium species in relatively stable situations, their realized niches become much narrower as ecosystems mature and become more stable. The theory of optimal feeding strategies lent testable structure to these concepts (Schoener 1971).
By the 1990s, the prevailing paradigm continued to emphasize a bimodal model of the distribution of r- and K-selected species with predominantly opportunistic and equilibrial traits, respectively. The biomodality, however, was not believed to be numerically balanced, favoring, instead, a substantially larger number of equilibrial species over opportunists. These observations had implications for species viability in increasingly unstable environments. Equilibrial species were believed in general to be more vulnerable to extinction than broadly adapted opportunists, and, because they were believed to be numerous, implied disproportionately large species vulnerability under rapidly changing environmental conditions (Primack 1993, Brewer 1994). The result of accelerated extinction of specialists was believed to be a regionally more homogeneous composition dominated more by opportunistic species. 4.1.5.2 After 1995 With some modifications, this paradigm has generally withstood the tests of time that have important implications for management decisions including climate change adaptation. Working with the fossil record, Taylor (2004) confirmed that generalists were less vulnerable to extinction than specialists in past marine communities. Most ecologists believe that the traits of predominantly r-selected species make them better adapted to rapid global climate change than predominantly K-selected species. Cain et al. (2011) and Ricklefs and Relyea (2014) generally accepted the diversity of life strategies, but placed more emphasis on the continuum of life- history characteristic combinations between r- and K-selection extremes. A bias toward a larger number of species with primarily K-selected traits was still generally accepted, however, in more stable environments. Cain et al. (2011) and Ricklefs and Relyea (2014) emphasized that life strategies often involve survival tradeoffs between r- and K-selection characteristics.
Ecologists continue to accept the characteristics as described by Pianka (1966, 1970) and have added others. In addition to high intrinsic rate of increase, Cain and his coauthors identified small body size, rapid development, early maturation, high reproduction rates, and low parental care as r-selection strategies. K-selected species were believed in general to be larger, live longer, develop and mature more slowly, reproduce more slowly, and invest more in the survival of fewer and larger offspring.
Primack (2014) emphasized the importance of life strategies in assessing species vulnerability to extinction, which included species with specialized niches and species adapted to stable ecosystem conditions. Vulnerability is expected to vary with the balance of equilibrial and opportunism traits. The predominance of species with K-strategy traits in the tropics and other relatively stable environments may help to explain the higher numbers of extinctions observed in those environments. Cain et al. (2011), Ricklefs and Relyea (2014) and Primack (2014) all believe that the destabilizing effects of environmental change favor r-selected species. As the increasingly destabilizing effects of global climate change interact with numerous other anthropogenic environmental changes, the bimodality of r- and K-selection traits is likely to become more balanced as an increasing extinction rate takes a bigger bite of K-selected species. From a management perspective, however, the greater emphasis on mixed traits and tradeoffs also implies that simplistic categorization of species into one or the other selection strategy is less appropriate than careful examination of each species adaptive attributes.
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4.1.6 Ecological Instability Increases Survival Dependency on Dispersal and Filtration Paradigm: Ecological instability increases the dependency of species survival on population dispersal capacity and tolerable intensities of ecological filtration. 4.1.6.1 Before 1996 Early animal ecologists recognized that species often “spread” after environmental changes following human settlement or the introduction of foreign nonnative species (Leopold 1933). Generally though, ecologists believed native populations free of human impact were quite stationary even though individuals moved significant distances and some, such as lemmings, dispersed dramatically when exceptionally abundant (Elton 1927). Early ecologists generally recognized that dispersal of population members contributed to the recovery of decimated populations after fire, flood and other disturbances through a general process called succession (e.g., Clements 1916, Clements and Shelford 1939). But the study of population dynamics continued to emphasize relative birth and death rates over population immigration and emigration effects (Allee et al. 1949) until Odum (1959) pointed out their importance in some circumstances and Levins (1969) emphasized their importance in metapopulation maintenance.
By the 1990s, the importance of population was consistently recognized by Ricklefs (1993), Primack (1993) and Brewer (1994) as the means by which populations expand geographically, share genes with other populations, and sustain subpopulations in metapopulations (more will be said about this concept later). The r-selected species in a region were believed to have greater dispersal capacities (greater numbers, mobility, and survival) than K-selected species (Brewer 1994). Based on fossil records of marine snails, species with high dispersal capacity tended to have larger geographical ranges and lower extinction vulnerability (Hansen 1978). Dispersal appears to be triggered by crowding and competition within and among species populations. Dispersing members are often younger and smaller than the population average. Depending on their adaptations, they may be directionally unconstrained, radiating in all directions, or highly Figure 10. Unconstrained radial and constrained linear dispersal, and generalized constrained within dispersal distance model of Wolfenbarger and Kettle. suitable habitats (Figure 10). Dispersal distance depends on rates of movement and mortality, which depend on species traits and environmental conditions. By the 1990s, the concepts of fine- and coarse- filter approaches to conservation were emerging (Hunter et al. 1990, Scott et al. 1990). The fine-filter approach had been
U.S. Army Corps of Engineers 40 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence the more traditional species-specific approach to conservation of rare species and the coarse filter approach was a broader ecosystem approach used primarily for more common species (Bourgeron and Jenson 1994). Keddy (1992) built on the older concepts of environmental resistance and dispersal barriers to develop the concept of ecological filtration of species from community assembly in a landscape context (a fine-filter approach). Because the effectiveness of dispersal barriers varies among species, barriers are said to be “porous filters”, which hold back dispersing members of different species at different rates.
At the fine-filter level, ecological filters may be biotic or abiotic and lethal (predation) or non- lethal (physical impediments to movement) and they may be encountered during dispersal or during establishment in a biotic community. Population establishment in an area depends on dispersal rates from other populations and ecological filtration rates along the way. Filtration rates depend on specific adaptations and the conditions encountered along the dispersal paths. Filters include biotic and abiotic sources of mortality, such as predators and temperature extremes, but are influenced by conditions that alter the exposure of dispersing members to sources of mortality, such as an environment with fewer refuges from predation and temperature extremes.
Wolfenbarger (1946) and Kettle (1951) proposed that the percent of surviving dispersants to various distances from point of origin decreased at a uniform rate where filtration rates are constant along the dispersal path (Figure 10). The slopes of the curves differ among species and environments depending on species traits and filtration intensity. By the 1990s, the Wofenbarger-Kettle model was widely accepted as the general pattern for dispersal (Brewer 1994). But the smooth curve of the model is often altered by the effects of ecological filters that are not uniformly distributed along the dispersal path and more sophisticated models take those effects into account.
The entire regional species “pool” is the source of species dispersal and immigration into local ecosystem areas. It sets the limit on the maximum number of species in any specific ecosystem area within its region (Keddy 1992). The distribution of predation, for example, may be uniform, changing the Kettle model little, or clumped with periodically sharp drops in the rate of decline along the dispersal path. Within a newly colonized area, abiotic filters eliminate immigrant species with inadequate fundamental-niche requirements. Additional realized niche requirements impose a final filter on those species that cannot coexist with the mix of competitors, predators and parasites present in the ecosystem.
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4.1.6.2 After 1995 Ecologists increasingly accepted the importance of dispersal and ecological filters for species survival as the reality of climate change grew more obvious (Van Dyne 2008, Cain et al 2011, Primack 2014). The important role of species in the dispersal of other organisms—seed dispersal by birds, for example—is increasingly appreciated and considered (Soons et al. 2016). Dispersal and ecological filtration are now basic considerations in contemporary conservation biology for metapopulation analysis and projections of species redistribution in changing environments (Van Dyne 2008, Primack 2014). The concepts of dispersal and filtration are the basis of filter-based community assembly Figure 11. Conceptual model for processes of ecological models useful in basic as well as applied filtering of species as they assemble in the local community from the regional species pool (after Belyea 2004). ecology. Belyea (2004), Menninger and Palmer (2006), and Morrison (2009) explored the degree that porous barriers filter populations of different species at points of origin, along dispersal pathways and in the colonized location (Figure 11). The conceptual models they and Keddy (1991) developed for species dispersal and filtration generally guide the development of specific quantitative models.
Cain et al. (2011) and Ricklefs and Relyea (2014) continued to accept dispersal as the means by which species expand range and sustain populations in suitable habitats. They also accepted the general effects of ecological filters operating on dispersal from the regional species pool to reduce immigration into local community assemblages and on species survival within the community. Brashares (2010) described the complexities of filter interactions and the need for species-specific assessments to identify probable causes of future extinctions. Primack (2014) identified species with poor dispersal as among those most vulnerable to extinction and linked poor dispersal with the filtration effects of habitat fragmentation. Species with high dispersal capacity (capable of successfully passing through numerous types of ecological filters) often dominate in fragmented ecosystems. 4.1.7 Populations below a Minimum Size are Vulnerable to Instability and Random Events Paradigm: Below a minimum population size, populations lose genetic diversity, adaptability, and viability, and become extremely vulnerable to extinction caused by environmental instability, restricted habitat, and random genetic and demographic events. 4.1.7.1 Before 1996 The processes of extinction grew to prominence in ecological research and management after the ESA was passed in 1973. The new field of conservation biology responded to the nationally expressed need to preserve species diversity (Soulé 1986, Simberloff 1988). Conservation biologists developed and applied methods for population viability analysis (PVA) designed to assess species threats, population viability probabilities, minimum viable population sizes (MVPS), and required habitat sizes and qualities (Boyce 1992, Primack 1993, Beissinger and McCullough 2002).
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When long-term viability of a population is an objective, any protective or restorative management of species in an ecosystem context must consider the MVPS of the species desired in the area. Shaffer (1981) first proposed the concept of MVPS and defined it as the “smallest isolated population having a 99% chance of remaining extant for 1000 years despite the foreseeable effects of demographic, environmental, and genetic stochasticity, and natural catastrophes." For greater practicality, that criterion was soon relaxed to 95 % chance of remaining extant for 100 years (Primack 1993). Restoration and protection objectives will fail to be achieved when the restored and protected area is too small to sustain the desired set of species. The first estimate of an MVPS large enough to fully sustain genetic integrity was a population large enough to maintain about 500 reproductively interactive adults (Franklin 1980). When the number fell below about 50 adults, normal variations in births and deaths started to cause erratic fluctuations in population numbers that markedly increase the likelihood of extinction on a downswing (Menges 1992). To remain viable, the area of suitable habitat has to be at least large enough to support the MVPS.
The decline of species toward global extinction progresses with the extinction of each population. The vulnerability of a population to local extinction increases rapidly once populations reach a MVPS. Small populations are especially vulnerable to erosion of genetic integrity, which reduces the genetic heterogeneity that supports the widest range of adaptive traits (Primack 1993). As populations shrink, environmental instability and random genetic, environmental and demographic events play increasingly important roles in determining their fate.
PVA generally relies on stochastic models of population behavior. Various weather and biological disturbances are stochastic, including exposure to a major flood or fire, or the chance invasion by a new predator, competitor, or disease organism. At small population sizes, demographic variation in birth and death rates causes random fluctuations that may result in population extinction. Unique genes randomly drop out with the deaths of unique individuals. Random changes begin at the MVPS and accelerate very quickly as numbers drop below the effective population sizes of 50 breeding adults (Franklin 1980, Lande and Barrowclough 1986)
Since most populations have much smaller adult population sizes than full population sizes (juveniles greatly outnumber adults), the MVPS needs to be considerably larger than the breeding population (Bartley et al. 1992). Because of age, health and other reasons, many members do not participate in reproduction and the censused population size was estimated to be several times an effective population of 500 individuals (Kinnaird and O’brien 1991). By the 1990s, population ecologists generally agreed on the dramatic effects of small population size on extinction risk. But opinions about the practicality of estimating MVPS differed and it was not used much in conservation planning. Brewer (1994 was concerned about MVPS estimation uncertainty while Primack (1993) and Colinvaux (1993) accepted its potential utility while acknowledging the estimation uncertainties. 4.1.7.2 After 1995 During the past two decades the concepts of small-population dynamics, genetics, and extinction caused by random events, environmental instability and restricted habitat were further researched and generally accepted by ecologists (Van Dyke 2008, Cain et al. 2011, Primack 2014). Increasing awareness of the rapidity of climate change helped crystalize the threats imposed by increasing environmental instability. PVA is widely practiced and the MVPS concept has become more widely accepted and applied (Traill et al. (2007, 2010). Van Dyke (2008) noted that a probabilistic approach to population size was needed to account for the variability in estimated sizes.
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Analysis of studies on over 200 species (mainly vertebrates) determined that the estimated median MVPS is about 4,000 individuals, most falling within the range of 3,000 to 5,000 individuals (Traill et Figure 12. Proposed relationship between number of individuals in a species population al. 2007, Flather and the likelihood of extinction. 2011, Clements et al. 2011, Figure 12). A review of several studies by Traill et al. (2010) revealed a median estimate closer to 5,000 individuals. Primack (2014) believed that population variability contributed to the variation and thought that 10,000 individuals may be a better estimate for extremely variable population sizes. More recent From UC Santa Barbara Department of Geography http://www.geog.ucsb.edu/events/department-news/1189/extinction-is-the-rule- research indicates survival-is-the-exception/ that an effective population size of several thousand reproductive adults is needed to prevent genetic loss over long time spans (Frankham et al. 2014), supporting Primack’s contention. Theoretically, the MVPS needed to sustain a population for a predetermined period should be a function of generation time, which is highly correlated with species size (e.g. Lande and Barrowclough 1986), but Traill et al. (2007) found no correlation between MVPS and species size in their meta-analysis.
An important issue facing conservation biologists is the size of an area needed to preserve species and supporting biodiversity. Primack (2014) accepted the utility of the MVPS for estimating the minimum area of suitable habitat and referred to it as the minimum dynamic area (MDA). The MDA can be estimated in a number of ways, including density estimates and home ranges (Pé er et al. 2014). For terrestrial mammals, estimated sizes of these areas vary between 100 and 1,000,000 hectares depending on average densities and home ranges. Krausman and Cain (2013) also emphasized the importance of metapopulation dynamics in resource species as well as imperiled species of conservation concern.
The concepts of metapopulations, MVPS and MDA are clearly critical and presently well accepted for planning reserve restoration and protection sizes. It is also clear that the estimation of the population size needed to sustain genetic integrity varies among populations and the estimated median size has increased significantly over the last 30 years of research. The uncertainty of the estimated MVPS requires an adaptive approach to population management.
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4.1.8 Population Distributions, Size and Number Indicate Species Vulnerability Paradigm: Species vulnerability is indicated by distributions, numbers and sizes of populations within ranges. 4.1.8.1 Before 1996 Variation in species range sizes and member densities were among the earliest differences in species attributes noted by early ecologists (Merriam 1884, Elton 1927), but they showed little theoretical understanding of population concepts or species vulnerability to extinction. Once population theory was better developed (Allee et al 1949, Odum 1953) and the theory of fundamental and realized niches was proposed by Hutchinson (1957), ecologists had a construct for identifying the theoretical limits of species ranges and evaluating why populations and members of populations were often limited to parts of apparently suitable habitat.
Field studies of populations in wild settings revealed random, uniform, and clumped distributions of individuals within populations and identified some probable causes, including territoriality for regular dispersion patterns, solitary and wide-ranging members of species for random dispersion patterns, and herding, schooling or habitat patchiness for clumped dispersion (Allee et al. 1949, Odum 1953, Andrewartha and Birch (1954).
As concerns about species extinction grew, more attention was paid to range size, the number of populations in a species range, how populations were distributed within the range, how members were distributed within populations, and the effects of density on reproductive success. The individuals of many species are clumped within numerous discrete populations and metapopulations throughout the species range. Primack (1993) noted that the vulnerability of those species generally decreases as range size, population size, and the number of populations increases. Species with small range sizes or population concentrations within a small fraction of the range are more vulnerable to storms, fires, floods and other large-scale events that extend throughout much of the inhabited area. The members of some species are distributed widely throughout large ranges, forming few discrete populations within the species range. Their large range sizes and widespread distributions compensate for the low number of populations.
Primack (1993) also was aware of the potential conservation importance of member distribution within populations. Populations with uniform or random distribution of members may be less vulnerable to widely distributed lethal events (e.g., fire, disease) than populations with clumped distributions. However, uniformly distributed animal species frequently have large home ranges, making them more vulnerable to habitat fragmentation than more sedentary species. Clumped dispersion may provide reproductive, predator-avoidance, or other advantages that compensate for vulnerability to other ecological filters.
For the many species that exhibit some form of clumped distributions of subpopulations in metapopulations, subpopulation size and isolation largely determine subpopulation viability and metapopulation status. Subpopulation clumping is typically determined by the distributions of suitable habitat. Gilpin (1987) proposed that three rules from the theory of island biogeography (MacArthur and Wilson 1963, 1967) applied to metapopulations (the theory of island biogeography is described later in the review). He argued that large patches of suitable habitat produce more emigrants than smaller ones, increasing the chances for colonizing suitable habitat elsewhere. Consistent with the Keddy (1992) model, the probability of patch colonization decreases exponentially with distance from the source population. Also, extinction rate in a
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patch is inversely related to patch size. These rules had not been widely accepted by the 1990s, but as understanding of metapopulation processes improved, more ecologists accepted their potential importance in viability determinations (Primack 1993, Ricklefs 1993, Brewer 1994). 4.1.8.2 After 1995 Contemporary ecologists continue to accept the long held belief that species and population distributions are determined largely by adaptations, habitat suitability, time required to disperse into new suitable habitat and dispersal barriers (Cain et al. 2011. Ricklefs and Relyea 2014). Conservation biologists are fully aware of the importance of population number and distribution attributes in determining extinction vulnerability, and are particularly attuned to climate change effects (Van Dyke 2008, Primack 2014).
Most ecologists now widely accept Levins’ (1969) concept of a metapopulation and the importance of metapopulation distributions in assessing the effects of habitat fragmentation on species viability (Cain et al, 2011, Krausman and Cain 2013, Ricklefs and Relyea 2014, Primack 2014). Cain et al. (2011), Van Dyke (2008), and Primack (2014) included PVA and metapopulation modeling in the basics of population ecology application. Metapopulation analysis has grown in importance as landscapes and their habitats have become more fragmented and the environment less stable. The importance of habitat fragmentation and metapopulation dynamics grew rapidly when ecologists recognized how the effects of increasingly rapid climate change could interact with habitat fragmentation to drive populations toward extinction.
4.2 Management Paradigms Overarching Paradigm: The common focus of ecological management is on maintaining the desired abundance and viability of species populations regardless of the scale of the approach.
Virtually all ecological management by government agencies is currently concerned about population sustainability for both the sustained use or resources and biodiversity maintenance. Their approaches may be broad, such as ecosystem management, but the focus remains on desired levels of population and species viability for sustained use of resources and for heritage preservation. Eco-managers are challenged with determining the levels of desired output for a wide variety of management circumstances. The following management paradigms are particularly applicable to the Corps environmental program. 4.2.1 Populations Are Basic Eco-management Units That Require Specific Assessments Paradigm. Populations are basic units of eco-management that require specific assessments of size, structure, distributions and dynamics; niche characteristics; adaptive strategies; regulatory factors; and other population-specific parameters. 4.2.1.1 Before 1996 Fish and wildlife management started out early in the 20th century focused on species-specific resource management and continued to be so focused in the 1990s despite growing emphasis on ecosystem approaches to management. In the earliest years, game species were treated like crops to be cultivated and protected from pests and over-exploitation, planted (stocked) when necessary, and ultimately harvested, (Leopold 1933, Nielsen 1993). This approach relied
U.S. Army Corps of Engineers 46 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence heavily on stocking to augment or replace natural reproduction, harvest rate regulation, some specific habitat improvements, and control of predators and competitors.
Recognition of the applicability of emerging population paradigms to wild resource populations took fish and wildlife management to a new level of sophistication (Leopold 1933, Lagler 1952, Nielsen 1993). In his seminal book on game management, Leopold (1933) accepted a “theory of population” and identified factors that contribute to population regulation. His approach to populations was strictly demographic, including roles of intrinsic rate of increase and habitat limiting factors. He was not concerned with evolutionary concepts of populations, but realized that each species is unique and that generic management prescriptions had to be tailored to fit species-specific harvest-regulation and habitat needs.
The early approach to fish population management was also strictly demographic. In one of the first comprehensive books on freshwater fishery management, Lagler (1952) accepted the principle of maximum sustained yield developed for commercial fisheries based on a simple but unrealistic interpretation of fish population dynamics (Graham 1935, Russel 1942, Van Den Avyle 1993). The concept was based on the intrinsic rate of population growth and the logistic growth model, fit to the growth characteristics of each species population. The approach largely ignored the situational uniqueness of species population dynamics, environmental instability and natural population fluctuations, economic variables, community-level effects, and other species- specific considerations. In later years, population analyses relied more on species-specific information and less on old population stability assumptions (Beverton and Holt 1957, Nielsen 1993, Van Den Avyle 1993). The maximum sustained yield management model gave some conceptual ground to an optimum sustained yield model (Roedel 1975), which considered more species-specific information about population dynamics, social and economic factors, the sustainability of the production system and multi-species consideration of fisheries interactions. But the concept proved difficult to apply.
In contrast, conservation biologists were concerned with evolutionary units, population-based species vulnerability assessments, viability analyses, and protective measures (Soulé 1986). While traditional resource management was focused on managing a sustainable resource- consumption “surplus” substantially above the minimum required for species sustainability, conservation biologists were concerned about identifying and eliminating threats, often through protection of existing suitable habitat. Resource managers were also less concerned about the reproductive interactions and genetics of population units. For example, stocking of different genetic strains on top of native populations continued apace (Kohler and Hubert 1993). Conservation biologists were more concerned with the genetic uniqueness and integrity of populations. The sizes of nature reserves set up to protect the genetic integrity of imperiled species were determined largely by these population-based criteria.
Species vulnerability assessment and PVA became the analytical core of conservation biology. The information useful for determining species and genetic vulnerability included the number of populations, population sizes, distributions and densities of members within populations, distributions of populations within species ranges, range sizes, niche widths, combinations of r- and K-selection attributes, dispersal capacities, MVPSs, and MDAs. PVA also used information on population genetics to assess long-term genetic integrity. Conservation biologists of the 1990s generally agreed that this information was important, but they also agreed that collecting it for all unsustainable species would be challenging, if not impossible. The use of surrogate species to represent the conservation needs of other species was proposed in the 1980s to address this problem (Noss 1990), but Primack (1993) did not discuss it as an option.
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The authors of the NRC (1992) report on aquatic ecosystem restoration must have believed the restoration of individual populations was an important aspect of ecosystem restoration since they emphasized complete restoration of all structure and functions in their definition of ecosystem restoration. Yet the discussion and case studies of the report focused instead on restoration of broadly defined community types (such as the wetland types identified by Cowardin et al.1979) with an emphasis on eliminated sources of water pollution and restoration of past hydrology and geomorphology. This emphasis on abiotic restoration of ecosystems was consistent with the restoration goal of the Clean Water Act (CWA) and other environmental legislation that focused largely on gross function and structure and not so much on the species- specific needs of biodiversity maintenance. Underrepresentation of restoration needs for the species recovery goal of the ESA may have reflected an emphasis on habitat protection by federal agencies and conservation biologists. The report’s approach could easily leave the impression that thorough restoration of hydrology and geomorphology would almost automatically result in reestablishment of the original species composition.
The focus of the NRC (1992) and other early ecosystem management concepts on gross function and structure (see later sections) diverted from the population focus of eco- management of the 1990s. The paradigm had continued to shift away from a simplistic agricultural model for management toward management with a broader community and ecosystem perspective, but always with a focus on specifically targeted populations and their needs (Bolen and Robinson (1993, Primack 1993, Scalet et al. 1996). Despite the growing interest in ecosystem management alternatives, the population paradigms of basic ecology were quite thoroughly integrated into ecological management texts of the 1990s while emphasizing the need to manage target species as dictated by their unique attributes (Bolen and Robinson 1995, Scalet et al. 1996, Primack 1993, Kohler and Hubert 1993). 4.2.1.2 After 1995 Despite the movement toward acceptance of a coarse-scaled, ecosystem approach to management in some federal agency settings, including the Corps of Engineers, acceptance of existing population-centered paradigms for management did not change much among leading eco-managers after the 1990s (Primack 2002 and 2014, Van Dyke 2008, Hubert and Quist 2010, Krausman and Cain 2013). Hilderbrand (2005) identified five myths prevalent in ecosystem restoration activities. The one most often remembered is the “if you build it, they will come” myth, which pointedly referred to the common practice in aquatic ecosystem restoration of restoring hydrology and geomorphology to some previous condition without serious consideration of species-level population ecology.
As the reality of global climate change became accepted, more attention had to be paid to species vulnerability and adaptive capacities (Glick et al. 2011). Brashares (2010) reviewed studies that together emphasized the site- and species-specific drivers of wildlife declines and a need to depend less on broad coarse-scale approaches to assessment and management and more on species- and community-specific approaches. The concepts of populations as genetic units were more widely accepted in fish and wildlife resource management and texts began to integrate the conservation of imperiled species with wildlife resource management (e.g., Morrison 2009, Krausman and Cain 2013).
Fisheries managers broadened their concentrated attention on population dynamics and regulation to include more emphasis on habitat limitations at all life stages, including ecologically disruptive invasive species. The concept of optimum sustained yield is now captured more specifically in clear declaration of management goals and objectives as well as a
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more holistic consideration of population interactions in marine ecosystems (Hubert and Quist 2010), but the maximum sustained yield concept remains attractive because of its simplicity.
In summary, with some exceptions in the ecosystem restoration community, the primary concerns of eco-management over the last two decades has remained centered on populations, even as some aspects of ecosystem management were more widely accepted. Largely as a consequence of climate change acceptance, much emphasis has been placed on developing improved species niche and population-dispersal models, including the effects of ecological filters (Morrison 2009, Krausman and Cain 2013). Primack (2014) accepted consideration of ecosystem approaches while continuing to emphasize that “conservation biologists should try to answer as many questions as possible” about population environments, distributions, biotic interactions, morphology, physiology, demography, behavior, and genetics. The need for population-specific information continued to grow as the importance of genetic diversity and integrity was recognized and the threat of climate and other environmental change grew. Regardless of the amount of information gained, however, appreciation of management uncertainties continued to grow with increasing emphasis on adaptive management (Van Dyke 2008, Primack 2014). 4.2.2 Sustaining Populations Requires Sufficient Habitat Quality, Size and Connectivity Paradigm. Sustaining populations and metapopulations for present and future possible use requires knowledge of niches, minimum viable population size, density, and dispersion to determine required habitat quality, size, and connectivity. 4.2.2.1 Before 1996 Before the modern concept of species niche emerged, early wildlife resource managers recognized that the individuals of targeted species needed enough space and habitat heterogeneity of appropriate quality to complete their life cycles (Leopold 1933). Most managers considered habitat qualities required of individual organisms and the total number of organisms that might be supported by habitat of proposed size, but not the minimum number required for population viability. For example, the Habitat Evaluation Procedure developed for fish and wildlife mitigation purposes (FWS 1980) used a habitat suitability index for the qualities required by an indicator species and multiplied it times acres to indicate the relative abundance of the indicator species (FWS 1980). It did not, however, include a minimum habitat size requirement. Most of the selected indicator species were common, widely distributed species with recreational or commercial value. The MVPS of most recreational and commercial species is not a general concern because they are usually secure from extinction (Bolen and Robinson 1995, Nielsen 1993).
On the other hand, conservation biologists worried about the fate of populations too small and fragmented to remain viable because of habitat fragmentation and poor habitat quality (Primack 1993). Nature reserve selection, design, and maintenance were the main strategies used in the 1990s to protect biodiversity (Primack 1993). The developing theory of metapopulations played a prominent role. In theory, they needed to know the habitat quality, MDA and corridor connectivity required to support the MVPS of the species populations that were to be sustained in the reserves. Early models of fundamental niches were developed to assess habitat quality. Once the targeted species populations were identified, one of the first questions asked was how big do protected and restored areas need to be to sustain population numbers and preserve population genetic integrity. They also needed to know what ecosystem qualities were required
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to sustain the targeted populations, how many discrete population occurrences should be sustained, and how far apart and connected the occurrences should be for long-term viability.
Generalized prescriptions were unacceptable because the needs of each species differ. But this required a huge and growing amount of information as the list of vulnerable species grew longer. Some conservation biologists began to advocate for the use of surrogate species to represent the needs of other species (Noss 1990), but Primack (1993) failed to mention it as an important concern (the use of surrogate species is addressed in the next report section).
Primack (1993) was not satisfied with simply protecting vulnerable species in conservation refuges through prevention of illegal take, wildfire prevention, air and water pollution control, and other threats to the viability of desired populations. He also advocated more active management intervention, including the transfer of population members from poor habitat to better habitat and managed enhancement of habitat using controlled burns, grazing, and other measures that mimic disturbances. A major emphasis was placed on slowing and reversing habitat fragmentation, which was believed to be the primary threat faced by many species.
Primack (1993) spent few words on the use of ecosystem restoration as a means of reversing habitat fragmentation and expanding connectivity between suitable habitats. Conservation biologists remained skeptical about the effectiveness of ecosystem restoration for conservation at the individual species level. Population-targeted management of habitat was preferred. In some cases, restoring connections among fragments was thought to be more of a liability than the existing fragmentation. In fact, further fragmentation was sometimes used as a conservation measure. Carlson and Muth (1993), for example, described use of small dams to block predator and competitor invasion of streams with vulnerable fish species.
In general, while the theories of metapopulations, MVPS, MDA, species niche, indicator species, and connectivity and other habitat restoration were growing more acceptable among eco-managers of the 1990s, there remained many population-specific questions about their management utility. 4.2.2.2 After 1995 Empirical evaluation lagged behind development of metapopulation theory and it was not applied much in the early 1990s, but application has advanced rapidly since then (Van Dyke 2009, Primack 2014). Responding to species conservation needs, McCollough (1996) tried to improve upon past descriptions of metapopulations and how they differed from other populations. The theory gradually became more accepted in management as the importance of gene-flow maintenance became more widely recognized. On top of many previously recognized threats, the potential threats of global climate change increased the interest in applications of metapopulation theory. Van Dyke (2008), Cain et al. (2011), Krausman and Cain (2013) and Primack (2014) accepted metapopulation theory and models as an important means for reducing species vulnerability to climate and other change where populations had been exposed to habitat fragmentation. They also accepted PVA as the main analytical approach to conservation biology, including assessments of MVPS, MDA, and niche models. Species- targeted habitat rehabilitation (not necessarily restoration to a previous condition) is widely acknowledged as essential wherever fragmentation is a major threat to species population viability (more is said about some of these concepts in the next report section).
The emphasis on PVA and metapopulation analysis has not carried over as consistently to management of resource species (Krausman and Cain 2013), perhaps because many of these species occupy niches broad enough to be significantly less affected by climate change and
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habitat fragmentation. PVA and metapopulation analysis may be applicable to resource species distributions limited to specific habitat conditions that have become quite fragmented, such as forested habitats of black bears and rough grouse and the river habitats of anadromous fish species. 4.2.3 Sustaining Population Integrity Requires Genetic Viability Paradigm. The genetic viability of populations must be protected and restored to sustain species for present use and future possible use. 4.2.3.1 Before 1996 Genetics played a minor role in wildlife resource management in the early years. The management importance of population genetics grew rapidly after the ESA was passed because it defined species to include “distinct” populations with significant genetic variation from other subspecies and populations within a species (Waples 1995). Conservation biology refined its species focus on extinction to the prevention of genetic loss at below the MVPS (Figure 13). The decline of specific west coast salmon populations drove this trend among fish ecologists (Nielsen and Powers 1995). Genetics became more relevant in fishery resource management as well as endangered fish management (Carlson and Muth 1993).
Figure 13. Effect of small population size on genetic By the 1990s, the importance of genetics in all variation after population recovery. aspects of renewable resource management was MVPS generally recognized. Scalet et al. (1996), for example, believed that genetic and phenotypic variation within resource species often indicated distinct populations and that the ability of any population to adapt depended on the genetic variation present in the population gene pool. Gene extinction Both Bolen and Robinson (1995) and Scalet et al. (1996) recognized the management significance of losing genetic variation as populations become small.
The concept of MVPS was based largely on the viability of genetic integrity within the population. However, estimating the MVPS for each one of a rapidly growing number of unsustainable species was judged impractical. Based on previous research on a number of species, Primack (1993) believed that at least 500 breeding individuals were needed to assure genetic variability. He also believed that enough estimates had been made to assume a population of 1000 should assure the MVPS for many vertebrates, but that for highly variable species, 10,000 would be a better estimate.
Simberloff et al. (1992) recognized the potential for protecting and establishing corridors between nature reserves to sustain gene flow among subpopulations of metapopulations, but he also knew that the concept remained largely untested. While Primack (1993) also recognized the potential importance of corridors, he did not link them directly to concepts of metapopulation management and did not emphasize them as much as other aspects of landscape ecology.
Species introductions into previously occupied habitats were an accepted part of conservation biology in the 1990s (Primack 1993) and continued to be practiced in fish and wildlife resource management. But genetic issues became more evident. Scalet et al. (1996) acknowledged a number of genetic issues pertaining to the success of species introductions from a remote source. The small size of many population introductions was a major concern. The members of
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a small population are likely to incompletely represent the heterozygosity of the population from which they came and undergo genetic drift after introduction, further decreasing heterozygosity. Disproportionate selection of closely related family members may result in inbreeding, which also reduces genetic variation. Introductions drawn from distant populations may not be well adapted to the specific conditions in the new location. Because of these risks, introductions were reserved for those situations when the alternatives were even more risky.
By the 1990s, extinction managers were beginning to use metapopulation concepts and models to assess genetically intact population viability based on the effects of events contributing to local extinction and factors determining colonization of habitat patches by members from surviving subpopulations (Primack 1993). However, acceptance of the metapopulation concept and models for management purposes remained limited (McCullough 1996), as was the use of niche models and other such tools (Primack 1993). But the concerns of conservation biologists were beginning to shift more toward a focus on maintaining genetic viability in populations. 4.2.3.2 After 1995 Contemporary conservation biologists widely accept maintenance of genetic diversity as a foundation for biodiversity sustainability (Van Dyke 2007, Primack 2014). Realization of the potential threat of climate change markedly elevated management concerns over genetic diversity down to the population level. Conservation biologists are well aware of recent advances in population genetics and incorporate them into estimates of MVPS and reserve design. The metapopulation concept is now the primary approach to management of population and genetic viability for numerous species. The importance of considering species-specific habitat suitable as dispersal corridors also became more widely accepted after two decades of research, as long as their potential for allowing access to threats is carefully considered (Hilty et al. 2006, Ricklefs and Relyea 2014, Cain et al 2011, Primack 2014, Krausman and Cain 2013). The restoration and protection of corridor connectivity is now a fundamental strategy for climate change adaptation.
5. Community Paradigms
5.1 Science Paradigms Overarching Paradigm: Species populations assemble and interact in temporally and spatially changing community associations that tend to maintain stable community-level functions in equilibrium with the abiotic and biotic environment, but can change dramatically with enough environmental change.
Community ecology originated in the late 19th century and had a significant influence on population paradigm development. Abiotic elements were progressively incorporated into the community concept (Elton 1927, Clements and Weaver 1928) until Tansley (1935) defined interactive complexes of communities and abiotic environments as ecosystems. From then on, community studies concentrated on species interactions and ecosystem studies concentrated on more holistic interactions between communities and their abiotic environments. Community characterization was based largely on associations of species populations indicated by the structurally dominant species (e.g., oyster, cattail, cypress-tupelo, bass-bluegill, beech-maple- hemlock, salt grass).
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5.1.1 Community organization is maintained by forces that reduce competitive exclusion Paradigm: Communities are organized and maintained by species assembly, interspecific competition, predation, environmental instability, and other forces that reduce niche overlap and competitive exclusion. 5.1.1.1 Before 1996 Theories about species interactions in communities grew out of interest in how diverse populations maintain community organization. Interspecific competition was prominently included in most theories. By the 1990s, ecologists generally accepted the proposition that interspecific competition for resources was a primary means of community interaction and that competitive exclusion, if unchecked, could limit species numbers and disrupt community organization (Colinvaux 1993, Ricklefs 1993, Brewer 1994). They also generally accepted the theory that more populations could continue to coexist in communities where species niches are sufficiently separated by predation and other forces that reduce competitive exclusion. This acceptance did not mean that species composition could not change, since different species in the same functional guild could more or less replace each other in the niche.
The reductionist approach to understanding community organization through niche assembly theory differed dramatically from the more holistic approach of Clements (1905, 1916), Clements and Weaver (1927), Clements and Shelford (1938), which assumed communities behaved largely as coherent units without much species change. Niche assembly theory proposed that communities maintained functional coherency as a consequence of niche interactions, even when the species occupying the niches changed.
Early niche theory was advanced by Gause’s (1934) investigations of what would later be called “competitive exclusion” by Hardin (1960). Gause’s controlled experiments consistently revealed that only one of two competitors survives when they are both limited by the availability of a resource. Others verified those results in laboratory settings and, to lesser extent, in the field (Lack 1947). Hardin (1960) minced no words: “Complete competitors cannot coexist.” This succinct statement of a principle was widely accepted by the 1990s (Colinvaux 1993, Ricklefs 1993, Brewer 1994). It left room for the occurrence of partial competition, which was also widely accepted. The competitive exclusion principle is an important corollary to the resource competition and partitioning paradigm described in the section on population paradigms.
Since Hardin (1960) also believed that no two species are ever completely identical, the more important issue was the extent niches have to differ to allow coexistence. He argued that small differences in niches in very stable environments would result in one species ultimately excluding the other. In contrast, MacArthur (1957) and Hutchinson (1959) hypothesized that species niches did not overlap at all—that species neatly partition resources and niches abutted each other in trophic levels like a “broken stick”. MacArthur (1958) provided evidence of this separation by careful measurement of how warbler species partitioned resources spatially in separate parts of the same tree. If that hypothesis were generally true, interspecific competition should be rare. Later studies showed, however, that, as a consequence of some niche overlap, interspecific competition is important in determining the relative abundances of species in communities (e.g., Connell 1980).
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Hutchinson (1959, 1961) offered two important mechanisms for reducing competitive exclusion enough for new niches to evolve (Figure Figure 14. Predation and disturbance reduce competitive 14). He first hypothesized in 1959 that exclusion by separating niches. predation reduced competitive exclusion by reducing competitor abundances and Competitive exclusion selecting for new adaptations that reduced by predation and separated the niches and opened new environmental disturbance opportunities for additional niches. He then added environmental disturbances to the hypothesized forces that could reduce competitive exclusion. In the 1990s, both predation and environmental disturbances were accepted means by which Niche 1 Niche 2 competitive exclusion is reduced (Colinvaux 1993, Ricklefs 1993, Brewer Resource Use Resource 1994). Interspecific competition was Competition generally accepted as an important force in Niche shaping communities, but ecologists overlap disagreed about the comparative importance of environmental factors. Niche Dimension 5.1.1.2 After 1995 In his “unified neutral theory of biodiversity and biogeography” Hubbell (2001) proposed that interspecific competition is weak to non-existent, much as MacArthur had done. Odum and Barrett (2005) also accepted the concept of competitive exclusion and the role of predation in reducing it. They believed that species often compete for and share resources without population extinction and that the energy assimilated by the trophic levels in a community is partitioned among species niches and into energy for individual population respiration and biomass production. That energy is then partitioned among population growth, reproduction, storage, and excretion. The resource partitioning principle (Schoener 1974) is generally accepted as the basis of detailed community energetics simulation models. Despite Hubbell’s proposition, acceptance of the significant role of interspecific competition as well as the competitive exclusion principle have held up well with supporting evidence that predation and disturbances interfere with competitive superiority (Cain et al. 2011, Ricklefs and Relyea 2014). Cain et al. (2011) mentioned speciation as the ultimate mechanism for reducing competitive exclusion. 5.1.2 Community Compositions are Always Changing Locally Paradigm: Small groups of species associations individualistically come and go in continuously changing communities. 5.1.2.1 Before 1996 Early ecologists believed that “natural” communities formed tightly organized, self-regulating and closed units, or metaphorical “superorganisms”, separated from other such units (Clements 1905 and 1916, Pickett and Ostfeld 1995). They believed community composition was stable for long periods of time. Some early ecologists believed communities were the perfect creations of divine origin (Pickett and Ostfield 1995). Early ecologists often believed humans corrupted “natural” communities and focused their studies on uncorrupted communities.
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Karl Mobius (1877) first articulated a concept of a biological community in his description of an oyster reef “society”. He proposed that oyster communities were self-regulating units with spatial boundaries delineated by the distributions of the dominant species, a practice that persisted into contemporary times. Merriam (1894) presented the different plant and animal associations in mountain “life zones” as discrete units fixed through their common responses to temperature variation. Shelford (1913) defined an animal community as the animal portion of a “biota” that included all plants and animals adapted to a particular habitat. The geographical dimensions of communities varied depending on the rates at which habitat qualities vary and the specific interests of the ecologists. The organisms in a small wetland swale might be defined as a community separate from the upland forest around it or incorporated with the forest community, but the boundaries of community units at any scale were generally clear.
Plant ecologists added the process of community succession to the community-unit concept. Frederick Clements in particular had a prolific and influential career, affected significantly by his early years on the American tall-grass prairie (Worster 1977). Clements (1905, 1916) and Weaver and Clements (1928) believed community units contained all of the species necessary to recover the same climax-condition functions and species abundances following disruptive disturbances. This recovery through a succession of species composition changes behaving as a community unit was likened to an organism healing after a wound. However, their paradigm was based largely on the observed behaviors of a common and often dominant species. Clements and Sheldon (1939) assumed that species assembled as a community-unit because all of the species were adapted to the same geophysical conditions as the common ones and species interactions evolved to become cohesively tight.
The community-unit paradigm, and the idea of community integrity, was challenged by Gleason (1917, 1926), who believed plant species respond “individualistically” to environmental changes. His evidence was significant, but limited and largely ignored. Nevertheless, the community-unit paradigm began to erode. Tansley (1935) thought community organization was only analogous to organism organization. Tansley also did not separate humans from nature, believing they were just as much members of communities as any other species. Yet, years later, Leopold (1949) considered wilderness the exemplar of ecological “health” and “most perfect norm” able to maintain itself as an “organism”. As expressed in their encyclopedic textbook, Allee et al. (1949) believed that “a combination of factors in the ecosystem has exerted selection pressures guiding the evolution of organisms toward adaptation to the system as a unit” and accepted the feasibility of the community as “supraorganism”.
About the same time, empirical evidence was breaking down the organismic concept of community units with clearly defined boundaries. Cain (1947), Egler (1947), and Mason (1947) provided empirical evidence in three different situations calling into question the community-unit concept in favor of Gleason’s concept of species individualism. This was followed by detailed studies of Whittaker (1953, 1956, 1957) and Curtis (1955), who deduced that the tight cohesion of species in a community unit would not permit gradual integration of one community into another as they had observed. Their careful mapping of plant locations along gradual elevation and soil gradients produced no evidence of discrete and discernible terrestrial community units. A more or less clear boundary seemed to exist only when a sharp change occurred along an abiotic gradient, as often seen at the edges of some wet depressions. Later studies indicated that aquatic species also distribute individualistically along a depth gradient (e.g., Sheldon and Boylen 1977).
Despite growing evidence that holistic beliefs in community integrity and discrete community units were rare, if they existed at all, the paradigm faded slowly. Numerous ecologists continued to believe that “natural” communities maintained a dynamic equilibrium and a “balance of
U.S. Army Corps of Engineers 55 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence nature” (e.g., Milne and Milne 1960) as a coherent unit. Odum (1959) accepted the proposition that ecosystems sustained a kind of “balance of nature” and “homeostasis”, if those concepts are defined by mechanisms “that resist change in the system as a whole”. But he and most other ecologists stopped short of accepting a holistic “supraorganism” concept. The paradigm was beginning to shift more rapidly as new studies questioned the old paradigm. A number of textbooks of that era generally rejected the community-unit concept (e.g., Krebs 1972, Colinvaux 1973). Further acceptance of species individualism then got a big boost from studies of pollen preserved in lake and wetland sediments, which generally revealed the individualistic response of species range changes to glacial retreat since the last ice age (e.g., Davis 1983, Delcourt et al. 1983, Graham 1986, Overpeck et al. 1992).
By the 1990s, the community-unit paradigm was shifting rapidly in favor of the individualistic assembly paradigm, but with some caveats. Colinvaux (1993) thought the community-unit concept was thoroughly debunked and accepted the individualistic assembly concept first proposed by Gleason (1926). Ricklefs (1993) echoed that belief, emphasizing that “community composition varies continuously over environmental gradients.” But Colinvaux and Ricklefs also recognized that species often exhibit obligatory associations with other organisms, such as parasites and Brewer (1994) was impressed by the apparent high fidelity of some plant species associations. Thus the paradigm evolving in the 1990s emphasized the importance of species individualism while allowing for coevolution of close and often obligatory associations of two or more species, most clearly in parasitism, specialized trophic interactions, and various symbiotic relationships. In these close relationships the largest species is often, but not always, the individualistic species, which determines the distributional behaviors of the dependent species. 5.1.2.2 After 1995 The vestiges of belief in the community-unit paradigm continued to fade as new data confirmed how much communities change (Delcourt 2002, Jackson 2013). Odum and Barrett (2005) firmly rejected the idea that ecosystems, including communities, are “supraorganisms”. While they emphasized that systems organization and behavior are “one thing held in common” by organisms and ecosystems, the systems-unit concept they accepted is much more functional than structural. Cain et al. (2011) believed “communities are always changing” as physical factors interact in new ways and species locally come and go. They concluded that the mechanisms of community succession “are diverse and context-dependent, resulting in different successional pathways and alternative states”. Ricklefs (2014) believed that many species assemble individualistically and different species often replace others in the same guild. He noted, however, that some species may play an irreplaceable role in simple communities of extreme environments where functional redundancy is limited. A species loss in those communities would more dramatically affect community functions.
From the perspective of conservation biology, Primack (2014) did not discuss how species assemble, but he understood that species respond to environmental changes differently and some go locally extinct while others invade and cause composition changes. In the well-studied and iconic ecosystems of Yellowstone National Park, Marris (2014) and Morell (2015) described the complexity of interactions and compositional changes among wild species in response to human and other effects.
In sum, Gleason’s (1917, 1926) extreme view of each species’ individualism is not accepted by contemporary ecologists. While communities are always changing, not all species behave individualistically. Many species are involved in obligatory associations with other species (Odum and Barrett 2005, Primack 2014, Ricklefs and Relyea 2014, and Cain et al. 2011). The contemporary view is that many species redistribute continuously and individualistically, but
U.S. Army Corps of Engineers 56 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence often in close associations within small groups. In many situations, communities appear to continuously change composition without losing species diversity (Pandolfi and Lovelock 2014). 5.1.3 Community Functional Stability Depends On Disturbance Level Paradigm: Gross community function and structure generally come into a dynamic equilibrium with the environment following local and temporary disturbances while extreme and prolonged disturbances cause instability and difficult-to-predict shifts to different functional and structural states. 5.1.3.1 Before 1996 The stability of communities and ecosystems (and what was later called “resilience”) has been of great interest to ecologists since before the field of ecology was named. In his seminal book on the effects of watershed condition on water quality and quantity, George Perkins Marsh (1864) claimed that nature, left undisturbed, establishes an “almost unchanging permanence of form, outline, and proportion” and restores itself to the former condition following rare disturbances. He was referring to the gross structure of watershed vegetation —forest verses grassland, for example. Mobius (1877) noted the persistence of oyster reef communities over long periods of time and Forbes (1887) marveled over the “steady balance of organic nature, which holds each species within the limits of a uniform average number, year after year, although every one (sic) is always doing its best to break across its boundaries, on every side.” Clements (1905, 1916) formally established a community stability paradigm that persisted for many decades. He more clearly articulated Marsh’s views in his concept of a functionally and structurally stable climax state that is at equilibrium with the regional climate.
Through the process they called succession, Clements (1905, 1916) and Clements and Weaver (1928) defined community stability by the resistance of the climax community form and function to disturbance and the predictable recovery of the same climax community composition and functional condition following a disruptive disturbance. They believed climate and the communities defined by climate were stationary and stable for thousands of years. Population ecologists inferred from the community stability paradigm that populations within communities were also “balanced” at equilibrium with other populations in the community and stabilized as the community recovered through a deterministic succession of changes following a destablizing event. The concept is now known not to apply to species composition, but for gross community function and structure, such as primary production and biomass, the stability paradigm is more complex. The concept of community equilibration following destructive disturbance would later be called community or ecosystem resilience (Holling 1973).
Most ecologists believed there was only one climax condition for a climatic region—the Clementian monoclimax at equilibrium with climate (Clements and Shelford 1939). Tansley (1935), however, believed that most communities exhibited a polyclimax condition revealed in a mosaic of different community compositions uniquely adapted to various soils, microclimates, moisture levels, and other attributes of the physical environment. He continued to accept Clements’ thesis that climax communities were functionally and compositionally stable. The stable polyclimax hypothesis explained the mixed patterns of plants and animals within a region exposed to the same climate well enough that Allee et al. (1949) accepted it. Whittaker (1953) proposed a variation called pattern climax in which the hard edges of polyclimax communities became less distinctive and more gradually transitional from one climax condition to another. Odum (1959) accepted the polyclimax concept, which then became widely accepted by other ecologists.
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Based on the accumulating evidence, Odum (1959) rejected the idea of a consistent species composition, but accepted the long-term stability of gross functions in the climax community. A decade later, Odum (1969) continued to believe that gross functional and structural recovery following a disturbance was directional, reasonably predictable, and culminated in a stable state. Based on numerous studies, he also proposed that, as succession progressed, the biomass, production, species richness and related community attributes all increased toward a condition more like the state that existed before disturbance, even as species compositions changed. Whittaker (1975), however, continued to believe polyclimax communities had relatively stable species compositions as well as steady-state functions that are “essentially permanent if undisturbed”.
General acceptance of the polyclimax stability paradigm set up the possibility that communities could shift to new states of gross function and structure as local conditions changed. These shifts typically involved sometimes dramatic changes in species composition and growth form. Lewontin (1969) and Sutherland (1974) proposed that multiple stable states may occur in communities as they adjust to local environmental changes. Holling (1973) accepted Lewontin’s premise and linked it to differences in system resilience when systems are disturbed. Holling suggested that ecological resilience be defined by the capacity of communities to resist change to another system state. The resilience of a community often returned it to something like the same gross functional and structural state following moderate disturbance and the time required to do that was a quantitative measure. But extreme stress could shift an ecosystem to a persistently Figure 15. Under stress, ecosystems can shift to altered state of gross function and structure even a different stable state quite resistant to restoration. after the environment returned to its previous state (Figure 15). In that event, the resilience is the Threshold amount of disturbance the ecosystem can withstand and recover from before it shifts. An example is the Stressors shift in vegetation type and productivity level that follows a fire severe enough to destroy the organic matter and reduce the water retention capacity of the soil. In addition, the capacity of life to conserve at least some of the gross function and structure as ecosystems shift to new forms and functions is also a more resilient response to extreme environmental change than total collapse of life processes. Ecological State 1 Ecological State 2 The multiple steady-state concept was well defined by the 1990s (Holling 1996), but remained an increasingly popular hypothesis open to further investigation. Yet the older, relatively rigid concept of climax stabilization and stability was rapidly losing support. Botkin (1990) concluded that “Scientists know now that this view is wrong at local and regional levels….Change now appears to be intrinsic and natural at many scales of time and space in the biosphere.” This was particularly true at the species level. However, the gross vegetation and trophic-level structure of many terrestrial and aquatic ecosystems was often observed to be resilient, once destabilizing stresses ceased.
Colinvaux (1993), Ricklefs (1993) and Brewer (1994) presented Clementian succession theory with a mixed appraisal of its continued validity. They appeared to accept the resilience and general stability of gross community function and structure, such as productivity and biomass, at dynamic equilibrium with environmental conditions, but rejected the notion that species composition was stable. Thus the acceptance of a community stability concept depended on the
U.S. Army Corps of Engineers 58 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence community measure used. It continued to be accepted for gross function and structure, but not for fine-scale structural and functional changes. 5.1.3.2 After 1995 The gross functional stability and resilience of biotic communities remains widely accepted although weakened by the realization of how rapidly biotic communities are changing as climate changes, nonnative invasive species and synthetic chemicals proliferate, and human exploitation of living resources continues. Odum and Barrett (2005) believed that communities and ecosystems maintained functional stability through many complex feedback, redundancy, and rate-regulating interactions (homeorhesus). When moderately disturbed, they believed communities demonstrate some degree of resilience and generally accepted, as Odum (1969) had proposed, that species richness, biomass and productivity of communities were commonly sustained even as species composition changed and contributed to maintenance of gross functional resilience.
Odum and Barrett (2005) discovered no evidence to support the theory of community stability at equilibrium with climate and believed the issue was academic anyway because some “climatic, geologic or anthropogenic force would likely intervene.” They accepted the evidence that past climates had changed significantly, that rapid global climate change is a contemporary reality, and that it is causing changes in species assemblages. Based on observed rates of climate and species range change (Diffenbaugh and Field 2014), the species composition of many communities now viewed as climax could be very different within 50 to 100 years. Mellilo et al. (2014) recently reported that global warming is already severe and causing a large variety of physical and community changes. They and many others now believe that climate will continue to warm and cause other changes. Ecological resilience and the related concept of ecological resistance to change (Nimmo et al. 2015) have become important conservation strategies for climate change adaptation.
Ecologists now largely accept the reality of increasing community exposures to rapidly accumulating environmental stressors and uncertainty in future community responses, especially at the species level (Odum and Barrett 2005, Cain et al. 2011, Primack 2014, and Ricklefs and Relyea 2014). The classic concept of a deterministic succession pathway to community maturity has been modified to accommodate the present belief that communities can follow different succession paths to various alternative stable states (Cain et al. 2011). Ecologists also generally accept the difficulty of predicting thresholds for shifts in functional stability in communities and ecosystems (Odum and Barrett 2005, Cain et al. 2011, Primack 2014, and Ricklefs and Relyea 2014). Over the long run, continued global climate change is believed to be a profoundly destabilizing condition that is likely to result in both gradual and dramatic shifts in gross function and structure, as well as contributing to species composition change in many terrestrial and aquatic environments. 5.1.4 Community Stability Increases as Species Numbers Increase Paradigm: As the number of species increase, community stability and functional redundancy increase. 5.1.4.1 Before 1996 The relationship between the number of species in a community and the functional stability of the community was long debated into the 1990s. Early ecologists claimed that communities rich in species were resistant to invasion by new species and other disturbances, implying greater stability (Weaver and Flory 1934, Elton 1958). Weaver and Flory (1934) defined stability based
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on the persistence of dominant species in the community; effects on less abundant species were often overlooked. Allee et al. (1949) noted that biotic barriers to invasion were stronger on continents (rich in species) than on islands (with fewer species).
MacArthur (1955) hypothesized that community stability was related to the complexity of community food webs, which is associated with high species richness. He argued that the effects of any abnormal event could be absorbed without as much stress on any one species in the web in species rich communities. Species richness often was associated with spatial complexity (e.g., coral reefs and forests), which was believed to increase the stability of predator-prey interactions (Huffaker 1958). Odum (1979) noted close correlations between community stability and species richness during succession following a destructive disturbance, but did not presume cause and effect. Then May (1973 and1974) raised doubts about diversity driving stability by showing mathematically that complexity can generate instability in general systems models. But May did not rule out the possibility that ecological systems were an exception to the general rule because of the natural selection that occurs in communities. Huston (1979) observed that periodically destabilized ecosystems tended to support higher biodiversity than very stable ecosystems, implying that some instability caused diversity. Carson and Barrett (1988) observed that fertilizing old field ecosystems reduced diversity and stability, but did not draw conclusions about cause and effect. Vitousek (1990) believed certain invasive nonnative species destabilized community function while increasing species richness.
Until the 1990s, the empirical evidence for believing diversity caused stability was correlative, it did not clearly establish whether stability enabled greater diversity or vice versa. The first Figure 16. Relationship between functional stability and strong experimental evidence for a positive species richness (from Tilman et al. 2006). cause and effect relationship between species numbers and functional stability was produced during the 1990s. Tilman and Downing (1994) added species incrementally to experimental grasslands and observed increased stability of community biomass exposed to various drought intensities (Figure 16). They also noted that the stability increased most rapidly with the addition of the first few species to a monoculture and later species additions add progressively less to the stability in the short term. By the 1990s, the belief that species diversity increases community functional stability was becoming more prevalent, but remained equivocal (Colinvaux 1993, Brewer 1994). 5.1.4.2 After 1995 McCann (2000) reviewed the diversity-stability debate in depth and concluded that the bulk of the evidence supported the positive effect of species number on functional stability, with the possible exception of certain aggressively invasive species. Much experimental evidence since then confirmed his conclusion (Naeem 2002, Loreau et al. 2002, Hooper et al. 2005, Tilman et al. 2006, Thompson and Starzomski 2007, Ives and Carpenter 2007, Naeem et al. 2009, Zavaleta et al. 2010, Isbell et al. 2011, Hautier et al. 2015, Isbell et al. 2015, Figure 16). Adding to the findings of others, Hector and Wilby (2009) found that, in the short run, the most productive species were nearly as effective stabilizing community productivity as a more diverse
U.S. Army Corps of Engineers 60 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence species group and a more diverse community increased stability in the long run as functionally redundant species replaced other species in changing community compositions.
Cain et al. (2011) and Ricklefs and Relyea (2014) generally accepted the stabilizing effect of diversity while also recognizing that climate and other abiotic instability more fundamentally limits the effect of species richness on stability. Ricklefs and Relyea (2014) particularly pointed out the increased stability of herbivores and predator abundances when plant species numbers were experimentally increased. But the evidence for the most part did not reveal the complexity of the interactions. A multivariate statistical model used by Grace et al. (2016) indicates that productivity is driven largely by variation in climate and fertility while species richness is driven largely by variation in climate variables and plant biomass. They concluded from their analysis that species richness increases and stabilizes productivity and biomass as biomass increases in response to increased productivity. 5.1.5 Anthropogenic changes are causing an exceptionally rapid decline in biodiversity Paradigm: As human effects on their environment have accumulated, species extinction rates have accelerated and global biodiversity is declining exceptionally rapidly compared to most other periods in the geological record. 5.1.5.1 Before 1996 The knowledge of naturalists and ecologists has come a long way since 1805 when, based on fossil bones found in Virginia, Thomas Jefferson directed Lewis and Clark to look for living mammoths. Jefferson could hardly think otherwise, since the concept of species extinction was not yet recognized by science (Kolbert 2014). The possibility that species could become extinct was first proposed in the published scientific literature during the early 1800s (Cuvier and Rudwick 1997). When Darwin (1859) published his theory of on the origin of species through natural selection, extinction was generally accepted in the scientific community. While many early paleontologists believed in catastrophic episodes of extinction based on the spotty record of fossil finds at the time, the belief in a uniform rate of geological change (the principle of geological uniformity) prevailed for decades. The rates of species extinction and origin were generally believed to be uniform and balanced.
Even though several contemporary extinctions of species at the hands of human predators had been well documented in North America by the early 1900s (Primack 1993), contemporary ecologists considered them human corruptions of the more general rule that “natural” communities are stable, stationary, and coherent. After decades of additional research, that proved not to be true for the communities indicated by the geological record. Newell (1967) identified what he believed were five mass extinction events, which stood out against much more stable (but variable) background rates. By the mid-1990s, the high likelihood of past mass extinction events was generally accepted while their causes remained debatable (Glen 1994).
In the early 1900s, the trends toward further extinctions from overhunting in the United States were stemmed by influential sportsmen, who lobbied successfully for more effective state and federal regulatory laws (e.g., Reiger 2000). Habitat contributions to species decline and extinction were recognized for at least some species at that time, but it was not until the second half of the 20th century that the exceptional importance of habitat-caused extinction and threat was widely accepted. The data on species extinctions were still sparse, however, and generally indicated a low percentage of the total number of species. Forecasts of future extinction rates
U.S. Army Corps of Engineers 61 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence were typically based on the relationship of species richness to geographical area and the projected rates of ecosystem conversion to human use, largely in the terrestrial tropics where most species occur. Concerns about human-caused mass extinction grew after the rates of tropical ecosystem conversion increased rapidly following World War II.
Based on forecasts of ecosystem deforestation in the tropics, Myers (1979) was among the first to call for public concern over what might be the beginnings of human-caused mass extinction. Estimates of potential future extinctions were variable, but in general quite high, based on projected conversion of tropical rainforests to agricultural and other use (Wilson 1989, Reid and Miller 1989). By the 1990s, conservation ecologists had become quite certain that human- caused extinction was increasing rapidly (Wilson 1988, Wilson 1992, Primack 1993, Brewer 1994, Pimm et al. 1995) and required much greater research attention (Lubchenco et al 1991). Leaky and Lewin (1995) referred to this human-caused acceleration as the beginning of the “sixth extinction”.
The foci of concerns were terrestrial rainforests and oceanic islands, where most documented extinctions had occurred. But, in the United States, it was becoming increasingly clear that the greatest continental extinction concern was in freshwater ecosystems where fish, mussel, and crayfish extinction and endangerment rates were especially high (Miller and Williams 1987, Williams et al. 1989, Williams et al. 1993, Taylor et al. 1996, Ricciardi and Rasmussen 1999). Potential species losses in the freshwaters of the United States were far from trivial. The fish species richness of the Mississippi River basin is particularly high (about 360 fish species based on range maps in Page and Burr 1991) for a temperate zone river system, more than half of that of the tropical Mekong River basin (599 fish species, Winemiller et al. 2016). The Mississippi River basin also supports exceptionally large diversities of freshwater mussels, gilled snails, and crayfish, many of which are endangered or extinct. 5.1.5.2 After 1995 As the reality of the rates and complexity of anthropogenic environmental change were better documented, most applied ecologists began to accept as fact that global extinctions rates were rapidly accelerating. However the rates of estimate overall loss varied from as low as 0.1% per decade (Lomborg 2001) to 8% per decade (Reid 1997) implying significant uncertainty. Yet even the most conservative estimators accepted a much higher rate of extinction for certain taxonomic groups (Lomborg 2001), more like the 1.2 to 2.5% per decade estimated by May (1995) for birds and mammals, or roughly 1000 times background rates. Ricciardi and Rasmussen (1999), Stein et al. (2000) and Cole (2009) confirmed the already exceptionally high extinction rate of freshwater species in the continental United States, which is about five times the continental terrestrial rate and about equal to the rate in tropical rain forests. Cole (2009) found declining rates of identified terrestrial extinction and increasing rates of freshwater extinction in records of last observations of extant freshwater species in the United States.
The high estimates forecast for future extinctions encourage comparisons to the five mass extinctions, particularly the documentation of a rapidly increasing amphibian extinction rate, raised the possibility of a sixth mass extinction in progress (Wake and Vredenberg 2008). Moritz and Agudo (2013) reported that numerous models based on projected climate shifts alone predicted catastrophic endangerment and extinction during the 21st century. Kolbert (2014) vividly described the trends toward a possible sixth mass extinction for the lay public.
Cain et al. (2011), Ricklefs and Relyea (2014), and Primack (2014) all accept as facts the exceptionally rapid rate of recent species extinctions and the overwhelming predominance of human causes. They generally accept the work of paleontologists indicating there have been
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five mass extinction events occurring over relatively short periods in the past and requiring millions of years for biodiversity to recover to the previous level after each event. Primack (2014) in particular believes that the sixth mass extinction event is underway and is caused by human impacts starting about 30,000 years ago. Quantification of the sixth extinction at the scale of past mass extinction events requires about a 75 percent loss of species, which would occur hundreds of years from now at contemporary rates of extinction (Barnosky et al. 2011). Yet the belief that a sixth mass extinction is in progress has grown with increasing documentation of the rapidity of global climate change and its potential for interacting with natural features and anthropogenic changes to trap species in intolerable environments (Urban 2015).
Mass extinction appears to be approaching more quickly in freshwater ecosystems where extinction rates are disproportionately large. If the combined effects of climate and other environmental change are not addressed fairly quickly, the high rate of past extinctions and present imperilment of species in freshwater aquatic ecosystems of the United States (Ricciardi and Rasmussen 1999, Stein et al. 2000, Cole 2009), as well as rapid world-wide development of river basins elsewhere (Winemiller et al. 2016), signals the approach of a freshwater mass extinction somewhat sooner..
5.2 Management Paradigms Overarching Paradigm: Gross community functions, such as primary productivity, often can be managed for desired rates and qualities, within limits, but species often behave individualistically and must be individually assessed and adaptively managed when targeted for management.
Past assumptions about management of communities as units were based on old paradigms that have since shifted significantly. The paradigm shifts have large implications for management of Corps properties, environmental impact mitigation actions, and ecosystem restoration and protection practices. 5.2.1 Uncertain Community Composition Requires an Adaptive Population Approach Paradigm. Managing uncertainty in community composition requires an adaptive approach to population management that is focused on the needs of desired species. 5.2.1.1 Before 1996 The community-unit and community stability paradigms allowed eco-managers the convenience of assuming ecological consistency and deterministic planning over long periods of time. Acceptance of the paradigms greatly reduced the amount of information needed for management. Community maps could be assembled and counted on to indicate the needs of all community members for years to come without frequent updating. While the community unit and stability paradigms of ecological scientists shifted in the 1990s, eco-managers lagged behind. Community classification schemes and maps persisted (e.g., Bailey 1996, Mitch and Gosselink 2007) and continued to be used in forestry and other natural resources management (Holechek 2002, Comer et al. 2003).
Structurally dominant and typically broadly adapted and long-lived species have often been relied on to indicate community types in various classification schemes. These community types
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have been used to indicate gross community structural and functional differences, associations with other species, and general abiotic attributes (Leopold 1933, Hubbs and Eschmeyer 1938, Lagler 1952, Bolen and Robinson 1995, and Scalet et al 1996). These classification schemes are likely to be more reliable for indicating the continued occurrence of broadly adapted species that live in many different community types than the continued occurrence of narrowly adapted and often rare species. However, the individualist responses of species to environmental change, suggest that major changes could occur among many rarer species without much apparent effect on the species used to indicate community type. Narrowly adapted species are often more vulnerable to ecological changes over short time intervals.
Before the rapidity and pervasiveness of environmental change was widely recognized, the use of more refined terrestrial community (and ecosystem) typologies became quite attractive to some conservation biologists concerned about the viability of narrowly adapted and rare species (e.g., Noss 1983, Noss et al. 1995). Noss (1983) proposed that biodiversity reserves could be selected to preserve all species based on the scarcity of community-ecosystem types typically indicated by species dominants. The idea rested basically in the acceptance of the community- unit and community stability paradigms. That approach allowed the needs of a few dominant and well known species to serve as surrogates for the needs of all species in the community. Scalet et al. (1996) assumed that habitat typology was prerequisite to identifying the needs of unsustainable species. A major effort to identify land and water protection needs by exposing the gaps between protected habitats and the locations of imperiled species (Scott et al. 1993) implicitly assumed community stability and stationarity.
Stability and stationarity assumptions were also implicit in reserve design and selection for species conservation purposes. Because property boundaries are stationary in law, sustainable preservation in a nature reserve requires community stationarity and resilience following destabilization to sustain the full range of suitable conditions for all species within the boundaries of the preserve. However, Primack (1993) recognized that reliance on wild ecosystem processes may have to be supplemented with active management and careful monitoring to assure that the needs of targeted species are met despite environmental changes and potentially threatening self-regulating processes.
The classical approach to ecosystem restoration also relied on community stationarity and resilience assumptions. In this regard, the report of the NRC on federal aquatic ecosystem restoration (NRC 1992) was replete with inconsistencies. The idea of ecosystem restoration rests squarely on the concept of predictable community succession and reestablishment of a stable and stationary community following elimination of human effects. The authors of the report, however, suggested that community succession should not be assumed to be “unidirectional”. Yet they also resisted complete endorsement of Niering’s (1989) conclusion that ecologists should drop the unidirectional concept of succession entirely. They accepted the use of community typologies for identifying historical community conditions and deemphasized the more specific targeting of desired populations needed to contend with unpredictable succession pathways. Adaptive management of uncertain results was recommended in a section on wetland restoration, but was not universally embraced for all types of ecosystem restoration projects. The confusing inconsistencies of the NRC report appeared to reflect incomplete digestion of paradigm shifts in progress. That confusion may have affected later planning policy guidance for the Corps ecosystem restoration program. 5.2.1.2 After 1995 During the past two decades, less emphasis was placed on the utility of community typologies even as an ecosystem approach was more widely accepted as a complement to species-based
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approach to conservation biology. This has accompanied growing recognition of the uncertainty in a management process that no longer could assume climatic and ecological stability and that populations would behave as predictable community units. The wildlife resource managers contributing chapters to Krausman and Cain (2013) accepted adaptive management as a basic tenet of natural resource management that is necessary for managing outcome uncertainty to the extent it is possible. They emphasized a population approach to wildlife resource management, but also accepted a need for a broader ecosystem view of influential processes to help manage the uncertainty in population outcomes. They also recognized that species composition has and will continue to change in most community settings and that habitat needs to be specifically assessed and managed to reduce the uncertainty of responses by targeted species.
Among conservation biologists, Van Dyke (2008) embraced an ecosystem management approach complementary to an even greater emphasis on a species-based approach. His emphasis was on population genetics, demographics, habitats, conservation methods, and other species-specific data required to conserve them. He accepted monitoring and adaptive management as essential aspects of population and ecosystem management. Primack (2014) also concentrated on species and population-based approaches to conservation planning and management, but gave more ground to complementary integration with community and ecosystem approaches than he did two decades earlier (Primack 1993). He also accepted a need for adaptive management of populations at risk of extinction as well as supporting aspects of ecosystems. The contemporary management paradigm points to ecosystem management (both protection and restoration) that is guided by the needs of the most threatened ecosystem elements: populations of vulnerable to imperiled species 5.2.2 Multi-species Surrogates Are More Reliable Indicators than Single Species Paradigm: Multi-species surrogates are more reliable indicators of the needs of other species in communities than single-species surrogates, but no use of surrogates is fully reliable. 5.2.2.1 Before 1996 Eco-managers have long known that they cannot identify all species in community assemblages let alone learn enough about each of them to manage in a fully informed way. Eco-management has always relied on knowledge of a subset of species ranging from one to dozens, depending on the objectives. The subsets were typically resource species valued for commodities and recreation.
The use of surrogate species to assess the condition of aquatic ecosystems for all species became popular following passage of stronger environmental laws in the 1970s (Washington 1984). A suite of “indicator species” were typically chosen to indicate habitat suitability for them and a wider array of other species. In response to fish and wildlife mitigation guidance developed for the Fish and Wildlife Coordination Act, the Fish and Wildlife Service developed the Habitat Evaluation Procedure based on a single- or multi-species indicator of habitat attributes to guide replacement of habitat damaged by water resources projects (USFWS 1980, 1981). Conservation biologists also began to consider use of surrogate species as a practical alternative to targeting each imperiled species needs individually. Noss (1990) developed a taxonomy for “surrogate” species of conservationist utility. Others thought the indicator species approach is unreliable (Landres et al.1988, Landres (1992).
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The use of surrogate species was still provisional and controversial in the 1990s. For the most part, Primack (1993) wrote little about the use of species surrogates, focusing directly on the needs of each unsustainable species. He mentioned that some species may serve as “umbrella species” with needs so comprehensive that providing for them would also provide for the needs of others. Primack, however, believed that the assumption should not be accepted without evaluation and, if necessary, subsequent management. 5.2.2.2 After 1995 After 1995, the surrogate species concept received a lot of attention, almost entirely for species conservation planning purposes, but remained controversial. Because of the convenience where information was often lacking, the use of surrogate species in management planning came into widespread use despite little empirical evidence to support it. Caro (2010) reviewed much of the literature generated for application of surrogate species up to 2009 and came to several important conclusions. The common classification terms that are ecologically most relevant include indicator, umbrella, focal, keystone, and dominant species. Another group, flagship species, is largely used for public relations purposes. Flagship species usually have high public recognition and value.
Indicator species are single, non-target species selected to represent the needs of conservation targets (Caro 2010). Indicator species are often used because more information is available about their needs than the needs of any of the targeted species. The assumption—that a single species can be a reliable surrogate for targeted species—continues to be accepted by various practitioners even though the evidence indicates they are frequently unreliable (Caro 2010). This lack of reliability reflects the unique niches and individualistic responses of species to environmental change, which was accentuated by the growing acceptance of rapid global climate change as reality. Water resources agencies often rely on indicator species in endeavors to restore and protect ecosystems (FWS 1980).
Umbrella species are variously defined, but Caro (2010) determined them to be the species targeted for protection that require the largest amount of habitat area needed to support viable populations. As a conservation target, they differ from indicator species, but serve a similar purpose and have similar potential limitations. The umbrella-species concept assumes that the variety of possible conditions that the umbrella species requires provides all the needs of other species, which is questionable. The concept is attractive, but has yet to be generally accepted because of insufficient evidence of reliability (Caro 2010).
Caro (2010) accepted Lambeck’s (1997) definition of focal species as a “suite of species, each of which is used to define different spatial and compositional attributes that must be present in a landscape and their appropriate management regimes” (Lambeck 1997, page 849). Focal species may be targeted species, but also may be non-target indicators of important needs of the targeted species. Caro (2010) found the use of focal species to be the most reliable. Cain et al. (2011) believed that the use of a suite of indicator species was often preferred because they indicate a more complete array of management needs than any single species. Krausman and Cain (2013) echoed the general sentiment that single species surrogates are unreliable and accepted a combination of multispecies and coarse-filter ecosystem approaches to guiding ecological restoration. The focal species concept has not worked in some contexts because the focal species relationships to the targeted species were not always reliable (Caro 2010). They must be chosen with great care to assure the most complete representation of targeted-species needs. The multi-species focal group is becoming more widely accepted as a management paradigm for indicating species needs, but its application is hampered by greater cost than a single-species approach.
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Keystone species may be targeted for management because of their scarcity, but are often important support species included with other focal species as indicators of habitat needs. When used alone, they have the same limitations as indicator species, although they may have advantages other indicator species lack. Keystone species have disproportionate community effects for their population abundance (Caro 2010). They differ from dominant species, whose influences are more proportional to their abundance (Primack 2014). Top carnivores are frequently keystone species (and often umbrella species too) with major influence on community structure and functions at lower trophic levels (Terbough and Estes 2010). Many of them are also vulnerable to extinction. Ecological “engineers” are often keystone species (Jones et al. 1994). Gopher tortoises (Gopherus polyphemus), beavers (Castor canadensis), and alligators (Alligator mississippiensis), for example, structure habitats essential for the survival of other species (Brewer 1994).
Dominant species (or foundation species) are rarely targeted directly for conservation because of their great abundance, but are often species that essentially support targeted species and should be considered for inclusion among focal species groups. Dominant species are often ecosystem engineers having significant structural influence on the community (Jones et al. 1994). Dominant species were defined early in ecological history as species with the greatest mass and influence on the community environment (e.g., Clements and Shelford (1939). They are frequently plants. Animals are less commonly dominants, but, before it was harvested so intensively, the American oyster (Crassostrea virginica) often dominated estuarine community structure and clarified the water by filtering out large quantities of suspended algal and detrital particles, a role frequently played in fresh water by unionid-mussel species.
The extent to which some invasive species become troublesome depends largely on how much they demonstrate dominant or keystone traits. Relatively few nonnative species become troublesome invasive species, but the ones that do can become costly problems (Williamson 1996). One of the better examples is the zebra mussel (Dreissena polymorpha), which occurs in freshwaters, where it can profoundly alter bottom structure, water clarity, nutrient cycles, trophic dynamics, and the relative abundances of species in the community (Karatayev et al. 2002). Cheat grass (Bromus tectorum) is a nonnative invasive plant species that dominates the biomass of many grassland communities. It often increases fire frequency and extent in the western United States and alters nutrient cycles (D’Antonio and Vitousek 1992, Billings 1994, Evans et al. 2001). The Burmese python (Python bivittatus) shows traits of an invasive nonnative keystone species in the Everglades of south Florida (Dorcas and Willson 2011). It could have great effect on food web interactions if not controlled. Some of the most damaging invasive species are disease organisms. An invasive fungus species, Batrachochytrium dendrobatidis, is the predominant cause of accelerated decline and extinction of many frog species (Fisher et al. 2009). A similar fungus species threatens salamanders (Martel et al. 2014).
The outstanding example of a non-native dominant and keystone species, as well as the “ultimate invasive species” (Levy 2011), is Homo sapiens. Humans are the major vector for contemporary nonnative species invasion. The rapid decline and extinction of North American megafauna and “herbivory release” about 13,000 years, and the ecosystem changes that followed, is believed to have been caused by human hunting and use of fire interacting with climate change soon after humans invaded the continent (Gill et al. 2009, 2012, Levy 2011). The effects of human predation on extinction of the largest species continues to grow (Payne et al. 2016).
Krausman and Cain (2013) did not address the use of surrogate species in management planning. Their focus is directly on the species of concern, whether for resource management or
U.S. Army Corps of Engineers 67 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence species conservation purposes. Van Dyke (2008) recognized the practical use of surrogate species, but questioned their reliability for representing the needs of species targeted for conservation. He pointed out that the use of rare species as surrogates for indicating the needs of other species is very likely to fail because rare species often inhabit entirely different habitat conditions within ecosystems and landscapes. On the other hand, common species are often so broadly adapted they too are poor indicators of other species needs. With reservations, Primack (2014) accepted the use of indicator species for identifying areas to set aside in reserves. He emphasized the need for research to assure that indicator species actually work, and that it is likely to be most effective to use a group of focal species to indicate the needs of targeted species. In sum, the use of surrogate species remains controversial, but appears in many instances to be the only practical approach to conservation planning for which multi-species focal groups are most reliable (Caro 2010). Ordinarily, focal groups should include dominants and keystone species as well as species actually targeted for conservation. 5.2.3 Depending on Composition, High Species Richness Sustains Desired Species Paradigm: Depending on the composition, management for high species richness, possibly including many nonnative species, generally supports target species sustainability by increasing the output and stability of productivity and other community functions. 5.2.3.1 Before 1996 Fish and wildlife resource managers did not manage for high species numbers until after the ESA was passed. Neither Leopold (1933), nor Hubbs and Eschmeyer (1938), or Lagler (1952) mentioned it. They managed habitat and human predation for a small set of resource species and essential support species consistent with their perceived objectives, and often attempted to control non-human predators and competitors. In fisheries management, the target was explicitly for maximum sustained yield of selected resource species (Van Den Avyle 1993). Public resource managers began to shift to a broader ecosystem approach after the ESA required broader consideration of species needs (Bourgeron and Jensen 1994). A broader ecological perspective, including community productivity and food-web interactions, contributed to the acceptance of optimum sustained yield in fisheries management (Nielsen 1993).
By the 1990s, resource managers had begun to accept the idea that high native species richness, including predators and competitors, contributed to a community stability that provided long-term support for resource species production and yield (Bolen and Robinson 1995, Scalet et al. 1996). Nonnative species were uniformly excluded from the emerging paradigm, however. For example, terrestrial wildlife managers paid much more attention to leaving most of the remaining old-growth forest intact and more carefully plan forest and range use patterns to reserve habitat for increasingly imperiled native species. In aquatic ecosystems, fishery managers began to restrict harvests as an alternative to stocking and discontinued stocking with nonnative species where adverse effects on native species were apparent. The rate of new species introductions decreased rapidly, and more careful introduction was emphasized. Predation and competition were more often viewed as essential aspects of community self- regulation and stability that sustained resource species carrying capacities, and moderated outbreaks of pests and disease. The shift came slowly in part because of the costs incurred by some resource users.
The authors of the NRC (1992) report on ecosystem restoration for federal agencies only weakly considered population and community-level dynamics that might require specific restoration attention to restore community integrity and functional stability. They defined
U.S. Army Corps of Engineers 68 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence ecosystem restoration as a structurally and functionally complete reestablishment of a past ecosystem integrity while inconsistently recognizing that restoration is usually an incomplete and approximately “naturalistic” rendering of past conditions. They emphasized the need for clear objectives, but failed to recommend connecting the objectives to the essential elements of communities or the need to target those elements more carefully. The authors briefly mentioned that community reestablishment should consider the size and longevity of the species involved, the use of local genotypes when introductions are necessary, and habitat heterogeneity and connectedness. However, the case studies emphasized manipulation of abiotic variables, such as hydrology and vegetation structure, but generally neglected management of the uncertainty inherent in restoring specific populations to a desired condition. Little was said about niche analysis, population size regulation, minimum viable population size, habitat size requirements, population viability and vulnerability to extinction, metapopulation management, or the importance of predation and competition in sustaining community structure and functions. This simplistic “if you build it, they will come” approach was later dismissed as management myth (Hilderbrand et al. 2005).
By the 1990s, preserving high native species richness with an eye on long-term sustainability was a basic management strategy of conservation biologists (Primack 1993). Managing for high native species richness typically includes many rare species, which are most often among the species vulnerable to extinction. By the 1990s, conservation biologists also emphasized support species, especially in keystone and dominant roles, which they increasingly believed stabilized species interactions and sustainability (Noss 1990). Nonnative species were largely regarded as threats that should be controlled when feasible (Primack 1993). Resource managers were beginning to accept the idea of a more holistic biodiversity management as well (Scalet et al. 1996), but conflicts over the costs to specific resource yield continued to encourage resistance. 5.2.3.2 After 1995 Contemporary managers of ecological resources generally respect the idea that relatively high species richness in a management area is desirable for the long term stability of resource populations (Krausman and Cain 2013). They generally accept the importance of maintaining global biodiversity for future generations at perhaps some cost in resource species abundance. Many eco-managers are now as likely to promote predator control to protect globally imperiled species as they are to promote greater abundance of resource species, but also recognize the importance of sustaining predators (Krausman and Cain 2013). Zedler et al. (2012) believed that managing for increased biodiversity in general is likely to increase ecosystem resilience.
Ricklefs and Relyea (2014) demonstrated the importance of species richness by describing an experiment in which zooplankton were removed by a pesticide applied at a dose too low to affect other species, including the leopard frogs of management concern. The frogs do not depend directly on zooplankton, but the tadpoles failed to metamorphose because the phytoplankton normally consumed by the zooplankton increased to such a high density it shaded out the bottom algae the tadpoles depended on for nutrition. Primack (2014) accepted preservation of biodiversity at all levels of ecological organization to maintain diversity of species and their genes. He argued that by preserving representative ecosystems within landscapes, the species composition, including imperiled species, could be stabilized at the landscape level while it fluctuated at the individual ecosystem level. Removing local anthropogenic stresses to encourage species richness is a commonly voiced strategy for mitigating the effects of climate change on vulnerable species.
Reflecting the sentiments of many others, Kennedy et al. (2013) believed that highly invasive nonnative species can endanger native species through predation, competition and other
U.S. Army Corps of Engineers 69 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence destabilizing interactions. They recognized, however, that only a small fraction of nonnative species become invasive threats and argued that many ecosystems now include nonnative species that are or could be critical support for threatened native species. Elimination of most nonnative species may not only become an unnecessarily costly endeavor, but one that actually thwarts conservation efforts (Schlaepfer et al. 2011) as species redistribute into new areas in response to environmental change.
Regardless of species origin, species richness is a key to food-web and other community-level stabilization, but also depends on the species composition. In light of climate and other environmental change, many contemporary eco-managers realize that species-based understanding of key community interactions and functions is typically essential to conserve desired species (e.g., Dunwiddie et al. 2009). Leading eco-managers also believe that reliance on restoring and protecting only native species is increasingly unrealistic (Kennedy et al. 2013) and could actually become a conservation risk, since climate change is likely to result in many species becoming nonnative in many locations (e.g., Chen et al. 2011, Burrows et al. 2011). This shift in paradigm may not be thoroughly accepted by all practicing eco-managers as yet, but the new management paradigm is consistent with projected future changes in species distribution changes and management that requires species redistribution into other locations simply to assure their survival. 5.2.4 Species Sustainability Needs Management from Local to Regional Scale Paradigm: The number of species that can be sustained generally increases with the size of the area restored and protected, but requires attention to species needs at local to regional scales. 5.2.4.1 Before 1996 Informed sportsmen of the late 19th century knew species could be driven to extinction by hunting but were not nearly as aware of extinction threats caused by habitat loss. Most early wildlife refuges set up by government agencies were designed as refuges from over hunting and disturbances during critical periods, such as nesting. Refuge sizes were typically based on the needs of resource species with little regard for the total number of species protected or community interactions that support desired species. Recognizing these shortcomings, Shelford (1933) proposed a national plan for conserving representative biotic communities, but without serious consideration until the 1950s (Primack 1993, Groves 2003).
After the ESA was passed, conservation biologists generally accepted some foundational concepts centered on designing nature reserves of appropriate size and quality. These included the species-area relationship (Arrhenius 1921, Rosenswieg 1995), equilibrium theory of island biogeography (see last section on regulation of species richness), and, a bit later, the MVPS. Many conservation biologists concluded that they should design for the largest reserve size possible to protect the most species (Primack 1993). The size limits of reserves were often determined by the land acquisition cost for each per species protected. Spatially heterogeneous areas with a greater variety of fundamental niche possibilities were frequently preferred. Another consideration was the role of the umbrella species requiring the largest habitat areas to sustain viable populations. The needs of keystone species were also considered. But Primack (1993) emphasized that the needs of each species vulnerable to extinction had to be considered to ensure they were actually served by the conditions indicated by reserve size and heterogeneity as well as the needs of umbrella and keystone species.
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By the 1990s, leading ecologists selected global change, biological diversity, and sustainable ecological systems as the three main research priorities of ecological science for informing “the wise management of Earth's resources and the maintenance of Earth's life support systems” (Lubchenco et al. 1991). The selection was driven largely by the rapidly accumulating human impact on essential resources, the potential impacts of global climate change, and accelerating rates of species extinction. Rates of species decline and extinction were widely recognized indicators of unsustainable practices. 5.2.4.2 After 1995 Scott et al. (1999) emphasized the importance of considering a broad range of local habitat to ecoregional scales during the design of nature reserves. They reflected the prevalent belief that “reserves must be of a size and design appropriate to meet the needs of the target life forms or systems and their functionally associated species, at all life stages, at all seasons, and in the face of all reasonably likely environmental perturbations.” In other words, planning should consider the needs of each targeted species at appropriate scales. While populations of small species may appear to require small habitats, their essential community support may require populations that need much larger habitats to remain viable. Scott et al. (1999) also emphasized the importance of considering all influential events over a sufficiently large time scale. Fires, floods, droughts and other moderate disturbances of historic frequencies often maintain necessary habitat qualities, but, less frequently, can have disastrous effects on subpopulations of metapopulations. Designing reserves of the greater size and heterogeneity typically needed to accommodate the effects of extreme as well as moderate events increases the likelihood that other subpopulations will survive and eventually repopulate devastated areas. Hansen et al. (2003) emphasized improvement of ecosystem resilience to better weather climate change.
Building on the work of Noss (1983) and others, Groves et al. (2002) presented a 7-step framework for conservation planning. The general strategy ultimately was to preserve representative landscapes and ecosystems everywhere. It incorporated consideration of biodiversity needs at species-population, ecosystem, and landscape scales. Since this could not be achieved quickly, determining the priority of investments was based on some measure of existing scarcity and vulnerability to future threats. The extent to which each level was relied on to guide reserve size and selection priority depended largely on the availability of information at the species level. Information on species scarcity was preferred, but when it was not available, the scarcity of different ecosystem and landscape structures identifiable from remote sensing information was used to approximate species scarcity as well.
Groves et al (2002) recognized that scarcity and vulnerability of species are not always indicated well by coarser-scale measures of scarcity. They recommended the incorporation of all available species information to increase the likelihood of identifying exceptional species scarcity. Biodiversity at the species level continued to be the driver of nature reserve priority settings in areas where species scarcity information was available, such as in many developed nations. When both the species richness and conservation status of many species are generally well known, as they are in the United States, the number of species of vulnerable conservation status play a very important role in determining where to invest in nature reserves. The availability of species-based information determines the extent to which the species-based fine- scale and ecosystem-based coarse-scale approaches are used to identify priority conservation areas.
Recognition that larger landscape areas may need to be set aside from destructive use to assure the viability of all species has been reinforced by evidence that reveals how rapidly global climate is changing. The emphasis on planning management from local to regional scale
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and designing for the largest reserves feasible has grown as a consequence (Primack 2014). Krausman and Cain (2013) dedicated an entire chapter to consideration of appropriate scale in wildlife resource management and conservation. The implications for project-level planning for species conservation purposes are numerous. Projects that for various reasons are limited to small geographical size should be viewed with skepticism. A small project area may be justified if the scale of unique biodiversity needs is also small. Investing in conservation of a stream community that supports a number of rare species may seem justified, but is likely to be riskier than a larger project area that includes a number of streams that could support the rare species in the same river basin (a metapopulation approach). A project that stands alone as an island in an otherwise unsuitable area is likely to fail sooner or later.
6. Ecosystem and Landscape Paradigms
6.1 Science Paradigms Overarching Paradigm: Ecosystems and landscapes are open, hierarchically organized and continuously changing systems maintained by inputs of energy and essential materials, and by complex feedbacks within and between biotic communities and abiotic environments.
Tansley (1935) coined the term ecosystem to bring more clarity to the community concept, which often included interactions with the physical environment. Since communities are major parts of ecosystems, the paradigms of community ecology also pertain to ecosystem ecology. Odum (1953) was the first author of a major ecological textbook to incorporate the ecosystem concept presented as the fundamental “functional unit” of ecology. Ecosystems occupy areas where biotic communities and nonliving environments interact and exchange materials in structurally and functionally consistent ways (Tansley 1935, Odum 1953, Brewer 1994). Individual lake environments and their inhabitants are examples. Landscapes typically occupy larger areas characterized by generally consistent patterns of ecosystem structure and function embedded in a larger “matrix” ecosystem (Forman and Godron 1986). The rocky boreal forest matrix ecosystem embedding the many lakes on the Precambrian Shield of Canada is a good example.
With the ecosystem concept came other paradigms. Later ecosystem ecology focused largely on energy flow, material cycling, and community effects on physical process while community ecology dealt with food-web and other species interactions. The ecosystem concept evolved with general systems theory (Bertalanffy 1950) and computer-facilitated development of systems analysis (Patten 1971). Recognition of ecosystem patterns in ecological landscapes grew rapidly into a new field of geographically based landscape ecology (Forman and Godron 1986, Forman 1995), which is now an integral part of conservation planning. It evolved rapidly with remote sensing technology and geographical information systems. The following scientific paradigms provide a foundation for eco-management paradigms described later. 6.1.1 Ecosystems Form Indistinct and Changing “Functional Units” Paradigm: Ecosystems spatially and temporally intergrade, forming indistinct, changing, open and highly interactive functional units, often with boundaries determined for reasons other than ecological function. 6.1.1.1 Before 1996
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The ecosystem concept was implied in the works of Mobius (1877), Forbes (1887), and Sheldon (1913), but was not defined by them as a “unit” of ecological study different from the biotic community. Tansley (1935) defined the term, in brief, as a complex organization of biotic and abiotic elements that results from the interactions of a community with its abiotic environment. Tansley conceived of ecosystems as discrete units of land or water that supported discrete community units. Their boundaries were usually determined by the dominant members of the community. Ecosystems were believed to be stable and their stability was maintained through the process of succession following a destabilizing disturbance. Odum (1959) emphasized his view of the ecosystem as the basic “functional unit” of ecology. This functional concept of an ecosystem unit was generally accepted and persisted despite some issues. The boundaries defined for these functional units could range from microbial dimensions to the entire ecosphere of the earth, depending on research or management perspective.
Since ecosystem units are often defined by their community attributes, they shared the unit- concept and stability problems of communities. The unit boundary had been a sticking point for many decades because ecologists were in general well aware that ecosystem boundaries are rarely precise or static. This awareness was growing among applied ecologists relatively early in the evolution of systems thinking (Leopold 1933). It was greatly reinforced by the works of Curtis (1955), Whittaker (1957, 1975), Vannote et al. (1980) and others who showed that many ecosystems changed gradually along spatial continuums and what we call a community or ecosystem “is really just an arbitrary subdivision of a continuous gradation of local species assemblages”. Terrestrial ecologists called these areas of gradation “ecotones” and where ecotones were relatively narrow, they were often called “edges” and species were frequently categorized as edge and interior species (Primack 1993).
Even when defined by abiotic boundaries determined, for example, by the distributions of surface water, spatial boundaries of ecosystems were easily shown to be functionally blurred. Many aquatic and terrestrial species, for example, depend on food sources outside their resident ecosystems. Riparian and stream ecosystems are among the best examples of open systems and the ecosystem continuum concept (Naiman and Decamps 1997, Vannote et al. 1980). Stream organisms frequently depend largely on terrestrial sources of nutrition. As streams flow to rivers and rivers to estuaries most ecologists believed they form a continuum of ecosystem changes without distinct boundaries (Vannote et al. 1980). Similar gradations occur in many terrestrial and marine ecosystems. Functional units are often subjectively determined by the relative influence of cross-boundary fluxes compared to internal energy processing, nutrient recycling, and other ecosystem-level functions.
Ecosystem-level interactions became the domain of landscape ecology, which emphasizes the effects of different patterns of ecosystem distributions on energy, material, and population movements among and within ecosystems. Landscape ecology was defined as the study of open-system ecological dynamics at large geographical scale (Forman and Godron 1986). It grew rapidly with the technological development of remote sensing and geographical information systems. Landscape ecology reinforced the concept of discrete ecosystem units, identifiable in aerial images, while at the same time emphasizing trans-boundary fluxes of living organisms and other materials at ground level (Forman and Godron 1986).
Landscape ecologists examine the patterns defined by ecosystem patches and corridor connections within a larger “matrix” ecosystem for their effects on material and population flux and other interactions. The matrix, being the dominant and most internally connected ecosystem, largely defines the landscape boundary. A predominant matrix ecosystem in the Everglades of southern Florida is sawgrass wetland, which surrounds islands of semi-tropical forest. Landscapes were first described as kilometers-wide mosaics “over which ecosystems
U.S. Army Corps of Engineers 73 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence recur” (Forman 1995, Turner et al. 2001), but more recently, landscape ecologists have defined landscapes from quite small to grand scales (Turner et al. 2001). Since both landscapes and ecosystems can be defined at a very wide range of geographical scales, the two concepts overlap geographically. An important difference between the two is the focus of the researcher or manager on the location of functional performance, whether within the ecosystem or across ecosystem boundaries.
The central paradigms of ecosystem science described here were largely developed for general application. Some of the early works that defined these themes incorporated ecosystem boundary effects (e.g., Odum and Odum 1955, Teal 1957). But the more usual focus was on functions within one ecosystem type or another (e.g., coral reef, cold-water spring). Then research shifted to more attention on cross-boundary import and export processes, such as work on terrestrial and aquatic “coupling” (Likens et al. 1970, Hasler 1975). The temporal consistency of ecosystems began to be questioned about the same time when several ecologists became more fully aware of changes in ecosystems beyond the old concept of succession (Lewontin 1960, Holling 1973 and 1978, Sutherland 1974). Despite strong arguments from leading ecologists (DeAngelis and White 1994), the idea that ecosystems are discrete and generally stable, functional units prevailed into the 1990s. Yet, ecologists were well aware that ecosystems intergrade and form indistinct unit boundaries, especially in river and coastal systems. And they knew that ecosystem boundaries were often defined based on research or management objectives that might selectively exclude consideration of some attributes (Colinvaux 1993, Ricklefs (1993), Brewer (1994). The ecosystem-unit paradigm was shifting. 6.1.1.2 After 1995 The idea that ecosystems form distinctive units has continued to wane since the 1990s as the community-unit paradigm faded and the implications of anthropogenic climate change were integrated into conceptual thinking. Golly (2000) made this point quite clearly:
“The boundaries of the ecosystem are not fixed or rigid. As a consequence, we say that ecosystem boundaries are fuzzy, that is, they are imprecise, changing and dynamic.
Odum and Barrett (2005) broadly dismissed the need for any predetermined boundary protocol. They stipulated that ecosystem boundaries can be “whatever is convenient or of interest”. Thus the ecosystem-unit described by them is largely a study or management convenience that may be based on property boundaries, somewhat arbitrarily defined “natural” boundaries, or other artifice. The functions within the unit are the defining attributes of the unit, but full definition of the ecosystem must include all of the transboundary functions as well—the flux of materials and populations between systems. Populations change within the unit as the functions are generally preserved or, under extreme conditions, shift to another functional state. Depending on how arbitrarily ecosystem boundaries are defined, transboundary dynamics may completely overwhelm any sense of internal functional coherency. This is common in the river and coastal ecosystems where the Corps pursues its civil works missions.
Ricklefs and Relyea (2014) emphasized the functional attributes of ecosystems. They implicitly recognized the difficulty in defining ecosystem units by acknowledging the difficulty in defining community boundaries. Yet, they accepted the utility of the unit concept by accepting the ecosystem approach as a complement to other approaches to selecting reserves for conservation. Because communities were believed to be constantly changing, most ecologists also believed that ecosystems were always changing as well (Odum and Barrett 2005, Primack 2014, Ricklefs and Relyea 2014, and Cain et al. 2011) even as ecosystems tended to conserve
U.S. Army Corps of Engineers 74 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence gross functions and structure. They also recognized as a consequence of accepting the reality of climate change that the physical and chemical attributes and their interactions were also changing. Thus, the concept of ecosystems as gross functional units is generally accepted by ecologists with full recognition that temporal and spatial boundaries are arbitrary and that composition and functions at the species level are always changing. These changes may be gradual responses to climate and other change or they may be dramatic, reacting in “highly nonlinear ways” once some threshold is reached (Scheffer et al. 2015, Figure 14). 6.1.2 Ecosystems Recycle Essential Materials Imperfectly at Local Scales Paradigm: At scales below the entire ecosphere, ecosystems incompletely retain and recycle materials essential for sustaining community structure and function. 6.1.2.1 Before 1996 The practical rudiments of biogeochemical cycling were generally understood in 18th century agricultural practice (soil to plant, to cow, to manure, to soil). But the first articulation of scientific principles is often credited to Vernadsky (1924). Elton (1927) conceptualized nutrient cycling in his notion of food cycles. While Tanzley (1935) coined and defined the term ecosystem, he did not explicitly describe nutrient cycles. In the United States, Hutchinson (1948) assumed a leading role in developing the concept of nutrient cycles. Odum (1953, 1959) established biogeochemical cycling as one of the defining characteristics of ecosystem function. The description of material cycles as if they were complete implied that ecosystems were functionally closed systems. That concept works best at a global scale and misrepresented processes at smaller ecosystem scales.
Odum (1959) split biogeochemical cycles into gaseous and sedimentary cycles. Those who studied the carbon, oxygen, nitrogen and other gaseous-cycle nutrients typically approached them globally, treating the atmosphere as a global source and sink. Because phosphorus and numerous sedimentary-cycle nutrients at a global scale required geological uplift that took millions of years to complete, their cycles were largely studied within water bodies (Hutchinson 1948, Wetzel 2001). The early emphasis was on the completeness of nutrient retention and recycling at both local- and global-ecosystem levels (Hutchinson 1948, Odum 1953). Much less emphasis was placed on the importance of inter-ecosystem nutrient flow, which began to emerge more clearly in the 1950s and 1960s. Teal (1957), for example, showed that a small amount of nutrient leakage from terrestrial ecosystems was enough to develop moderately productive and diverse stream communities, largely through leaf and other plant litter deposition. Then the export rates from undisturbed and experimentally disturbed terrestrial ecosystems were carefully documented and described by whole system experimentation (Bormann and Likens 1967 and Likens et al. 1977). In a textbook of the time on ecosystems, Clapham (1973) recognized that “inputs from and outputs to other ecosystems” typical of ecosystems and that boundaries were difficult to define, but described types of ecosystems mostly as if they were discrete and closed systems.
As more research accumulated, ecologists came to accept the importance of ecosystem nutrient import and export in sustaining diverse but closely connected ecosystems as a consequence of imperfect recycling and nutrient loss. Even the most nutrient retentive climax stage terrestrial ecosystems leaked nutrients enough to sustain communities elsewhere. Aquatic ecologists in particular were very aware of aquatic-ecosystem dependency on terrestrial ecosystem exports and the role of humans in accelerating nutrient loss from ecosystems through land disturbance and fertilization (Wetzel 2001). Applied ecologists also recognized how
U.S. Army Corps of Engineers 75 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence changes in nutrient cycle rates and efficiencies could dramatically alter the functions and species compositions within aquatic ecosystems (Rohlich (1980) and terrestrial ecosystems (Galloway et al. 1995) adapted to inflows of certain levels of nutrients.
In aquatic ecosystems, the efficiency of cycling and retention is greatly influenced by physical variables (Wetzel 2001). Seasonal stratification and oxygen depletion have major effects on nutrient cycling in lakes and downstream export to other aquatic ecosystems. Flow rate and turbulence are major influences in rivers. Nutrients repeatedly “spiral” between bottom and water as streams and rivers flow through their channels (Webster and Patten 1979). The influence of human effects on these and other aspects of materials cycling was Figure 17. The carbon cycle, including human effects through becoming better documented and fossil fuel emission. recognized for its importance, including human effects on structural and functional disturbance of ecosystems, including the major global effects of fossil fuel emissions on the carbon cycle (Figure 17).
By the 1990s, ecologists had accepted the concept of biogeochemical cycling as paradigm and widely recognized how thoroughly aquatic ecosystems in particular are connected to terrestrial ecosystems and other aquatic ecosystems by nutrient and other material flows (Colinvaux 1993, Ricklefs 1993, Brewer 1994). There would be no aquatic ecosystems without material leakage from incomplete terrestrial cycles. The sustainability of aquatic ecosystem functions depends on the quality and quantity of that leakage. 6.1.2.2 After 1995 In general, the paradigms for biogeochemical cycling have changed little since the 1990s, but understanding of rate regulations and human effects has improved substantially. Cain et al. (2011) noted that the rates of nutrient cycling are largely regulated by decomposition rate, which is determined by climate and the chemical composition of the organic matter. Ecosystems in colder climates generally cycle nutrients slower than warmer systems and more organic matter accumulates in them. Ricklefs and Relyea (2014) believed that recycling rates are controlled by moisture as well as temperature and accepted the importance of the hydrologic cycle as a means of nutrient and other material transport from one ecosystem to another.
Contemporary ecologists are more aware of the highly influential role of humans in changing nutrient cycling rates and distributions in many ways that contribute to new ecosystem conditions over most of the Earth (Cain et al. 2011, Ricklefs and Relyea 2014). Just a few examples are mentioned here. Recognition of the causes of global climate change has amplified this awareness. Anthropogenic sources of carbon dioxide and methane add to the atmospheric reservoir at a faster rate than they can be recycled and contribute greatly to anthropogenic climate change through the greenhouse effect (IPCC 2007). The effect on global temperature is causing changes in hydrological cycles and rates of other nutrient cycling (e.g., decomposition and nutrient-uptake rates are temperature dependent). The nitrogen deposited in wild ecosystems has nearly doubled as a consequence of human fertilization, with widespread
U.S. Army Corps of Engineers 76 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence biological consequence. The augmented soil nitrogen is recycled by temperature-influenced microbial process to the atmosphere where it is widely dispersed and precipitated on remote landscapes. Doughty et al. (2015) estimate that the recycling of phosphorus to land from the deep ocean via the combined effects of large marine animals, anadromous fish and sea birds was significant, but that humans have reduced its effectiveness by more than 96 percent by overharvest and habitat impacts. Clark et al. (2007) have documented some of the nutrient- cycle effects of human activities on the composition and number of species in ecosystems, but ecological ramifications remain far from completely understood. 6.1.3 Ecosystems Are Sustained By Flows of Energy Paradigm: Ecosystem structure and function are sustained by flows of energy from solar and chemical origins. 6.1.3.1 Before 1996 The nutrient cycling paradigm was incomplete without a complementary concept of ecosystem energy flow to explain what drove the cycles. Lotka (1922) proposed that energy flowed through “the system of organic nature” and maximized energy embodied in the total mass of the system through natural selection. Elton (1927) recognized that plants and animals derived energy from food sources and the number of individuals decreased at each feeding level along a “food chain”, forming a “pyramid of numbers” as well as a “food cycle”. But the flow of energy through ecosystems, and its progressive loss from the system through community respiration were not explicitly described and quantified until Lindeman (1942) published his seminal paper on “trophic-level dynamics”.
Lindeman assumed that ecological energetics were consistent with the laws of thermodynamics. His energy flow hypothesis explained why food chains were short and ecological pyramids of numbers and biomass often decreased at each step. He proposed that many species could be categorized by the number of energy transformations through trophic levels that occurred before the species transformed their food into biomass production. First-level producers were plants and other “primary producers”. Secondary producers (primary consumers) consumed primary producers and tertiary producers consumed secondary producers, and so on until energy was totally dissipated. The total number of observed levels in later studies was rarely more than four. At each level, species populations partitioned the energy resources available at their level into their own production and biomass depending on specific adaptations and complex competitive, predatory and other interactions within the community (later clarified by Odum, 1959).
These concepts provided the theoretical foundation for later ecosystem simulation models using energy flow as a driver of ecosystem process. Estimates of the percentage of energy loss at each trophic level transformation was a critical requirement for realistic models. Under equilibrium conditions, Lindeman estimated that the production of each trophic level formed an energy pyramid created by a 90-percent respiration loss at each level. That left 10 percent of the previous level’s production available for consumption at each subsequent level. For example, first-level consumers feeding on a net primary production of 1000 Cal/m2/yr produced about 100 Cal/m2/yr while second-level consumers produced about 10 Cal/m2/yr and the third level produced about 1 Cal/m2/yr. This trophic-level efficiency has also been called food-chain efficiency or ecological efficiency (which is favored here). Since each population in a trophic level required some minimum amount of energy to survive, the production of the community theoretically determined the maximum number of species populations that could survive in the ecosystem. The concept was well received. Odum (1953) introduced his text with the energy
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Lindeman’s study stimulated much subsequent evaluation and testing of the 10-percent hypothesis because of its importance in ecosystem modeling, resource yield limitations, and understanding of how species richness is regulated. Connell and Orias (1964) argued that ecological efficiency should decrease with greater ecosystem instability because more energy would be diverted from production to respiration. Slobodkin (1960) estimated the ecological efficiency of his laboratory population to be close to 10 percent and hypothesized it would turn out to be a constant for consumers. Later studies at first supported this proposal (Kozlovsky 1969) even though the theory was weak and only aquatic ecosystems had been studied. But Barnes (1980), using data from Whittaker (1975), showed that the ecological efficiency of herbivores generally decreased as primary production increased. Using Cushing’s (1971) data, Parsons (1980) concluded that the ecological efficiency of herbivorous marine zooplanktonic varied from 15 percent at low primary productivity to 5 percent at high primary productivity. Then Colinvaux and Barnett (1979) found a much lower efficiency of 1 percent in a terrestrial ecosystem where the consumers were ectotherms unlike the endotherms in the aquatic ecosystems. In 48 studies of aquatic ecosystems, Pauly and Christensen (1995) found the ecological efficiency averaged about 10 %, but varied between 2 and 24 percent. Despite the few studies that indicated environmental causes for different ecological efficiencies, there was no consensus about what caused efficiencies to vary—whether strictly measurement error or also some response to environmental rigor or instability.
The variation in the trophic-level efficiency is significant. The third consumer level of an ectothermic community with 24-percent efficiency would be about 1,700 times as productive as a community with 2-percent efficiency. Assuming endotherms are 1 percent efficient, an aquatic community entirely dominated by fish at 10-percent efficiency would be 10 times as productive at the third consumer level as one made up entirely of birds. These results have very practical ramifications for estimates of food production. While the general concept of ecosystem energy flow was well accepted by the mid-1990s (Colinvaux 1993, Brewer 1994), the 10-percent rule of ecological efficiency remained questionable without a theory to explain why it should be a constant except for the difference between endotherms and ectotherms. 6.1.3.2 After 1995 In general, research interest in holistic ecosystem process waned as attention shifted back to more specific food web interactions and population dynamics. Except for more explanation of the differences between terrestrial and aquatic ecosystems, the status of the energy flow paradigm has changed little since the 1990s (Odum and Bartlett 2005, Cain et al. 2011, Ricklefs and Relyea 2014). Terrestrial herbivore species may be less than half as efficient as aquatic herbivores because the terrestrial ingestion efficiency is depressed by lower food quality than occurs in aquatic ecosystems (Cebrian and Lartique 2004, Cain et al 2011). Terrestrial primary producers require more structural support made up of materials resistant to digestion than fully aquatic primary producers. Cain et al. (2011) and Ricklefs and Relyea (2014) believed higher trophic levels should be more efficient because of better food quality, but also reiterated a hypothesis of Pimm and Lawton (1977) that higher trophic levels in frequently disturbed ecosystems may be less efficient because of lags in disturbance response. This is a generic explanation for the previous observations of Barnes (1980) and Parsons (1980) on the apparent effects of primary production on secondary trophic-level efficiency. Given the influence of temperature on rates of nutrient uptake, food ingestion, decomposition, and metabolism (Cain et al. 2011, Ricklefs and Relyea 2014), global warming associated with climate change is expected to influence changes in rates of energy flow in ecosystems.
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Ricklefs and Relyea (2014) emphasized how important shifts in community composition can be in redirecting energy flows to different populations and contributing to further change. It also seems likely that organisms under stress divert resources from net production into respiration (Laureno et al. 2008, Agnieiszka and Stachowicz 2013). An increase in trophic-level respiration and a decrease in trophic-level efficiency would be expected as the number of species responding to environmental stress increases. However, the effects of community-wide stress on trophic efficiencies have not been well examined. Outside of differences between terrestrial and aquatic ecosystems and endotherms and ectotherms, there are few data or agreement about the sources of variation in ecological efficiency. 6.1.4 Ecosystems Are Regulated by External Forces and Internal Feedbacks Paradigm: The “bottom-up” effects of external environmental forcing functions on biotic communities interact with the “top-down” effects of internal feedbacks to regulate ecosystem functional and structural stability and diversity. 6.1.4.1 Before 1996 Integrating decades of previous research, Ricklefs (1993) characterized the regulation of ecosystem functions as the product of external “forcing functions” and internal “control feedbacks”. In his concept, forcing functions operate from outside the ecosystem to force (or drive) functional and structural dynamics and diversification. Major forcing functions include solar radiation, chemical sources of energy and the availability of essential inorganic materials, including nutrients and water. Climate, geophysical heterogeneity, and other abiotic variables greatly affect the ecological influence of the major forcing functions. Variations in forcing functions are major contributors to species niche variation and potential for diversification. Internal-control feedbacks originate in species responses to their environment and to other species in the community, which are influenced by the diversity of species functions and structures. The levels, diversity, and stability of community production, biomass formation at different trophic levels, and biogeochemical cycling are regulated by the complementary and reciprocal effects of forcing functions and internal feedbacks.
The reciprocal interactions of organisms and abiotic environment were generally recognized early in the progress of ecological science through emerging concepts of climate zones, biogeochemical cycling, community succession, and species bioengineers. But acceptance of Leibig’s concept of abiotic limiting factors and the widely recognized influence of atmospheric temperature and precipitation on life processes sustained a generally held belief that the abiotic environment affected communities far more than communities affected the environment (Elton 1927, Clements and Sheldon (1939). Lindeman’s (1942) presentation of the trophic dynamic concept, for example, mentioned nothing about regulation of trophic-level production by anything other than abiotic factors. Odum’s (1959) treatment of ecosystems is largely based in assumptions of bottom-up control.
Acceptance of nearly complete bottom-up control of life process rates, amounts, stability and diversification began to weaken significantly during the late 1950s. Redfield (1958) illustrated how important biological process was in controlling chemical forcing functions in the environment, such as nitrogen levels in the ocean and oxygen levels in the atmosphere. Hutchinson (1957) proposed that the diversification of niches in each trophic level is regulated not only from below by the amount of energy entering the community, but also from the top through predation. Hairston et al. (1960) hypothesized that the earth remains green (functionally stable) in large part because herbivores are kept in check by predators, implying that plant
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abundance, species composition, and functional stability was regulated by both bottom-up and top-down processes. Brooks and Dodson (1965), Paine (1966), Zaret and Paine (1973), and Shapiro et al. (1975) demonstrated the effects of predators on species composition and biomass. “Megaherbivores” were recognized for the outsized influence they can have on plant community structure and function (Owen-Smith 1988). Food web models began to be developed indicating that ecosystem stability depended on food chain length and predator-prey interactions (Terbough and Estes 2010).
Carpenter et al. (1985) and Kitchell and Carpenter (1993) proposed cascade effects on trophic level biomass and production when top trophic levels (apex predators) are added or eliminated, including physical and chemical effects. They hypothesized that these top-down effects complement bottom-up effects, but do not replace them. When apex predators are added to a lake community without them, they predicted that the production of the trophic level immediately below it would decrease, allowing the biomass and production of the underlying herbivore level to increase, and the primary production to decrease. The effect on primary production was predicted to increase the depth of light transmission as a consequence of reduced phytoplankton biomass and potentially cause a food-chain shift to species more dependent on rooted plants and benthic algae. Data supporting the theory demonstrate that the attributes of an ecosystem shift markedly when the top trophic level occupied by apex predators is eliminated or restored.
By the 1990s, the theoretical sharing of bottom-up and top-down effects on community structure and productivity crystalized by Ricklefs (1993) was becoming more acceptable for certain ecosystems, but not universally so (Colinvaux 1993, Brewer 1994). Many ecologists still believed that consumers had little persistent effect on communities and their physical habitats. If that belief held true, the elimination of apex predators and megaherbivores from biotic communities would have little influence on abiotic attributes of ecosystems as well as community structure and function. The possible destabilizing effects of global climate change, a profoundly important variable among influences on external forcing functions, were just beginning to be considered. 6.1.4.2 After 1995 One of the main points that has emerged from studies of ecosystem regulation since the mid- 1990s is that understanding of ecosystem structural and functional diversity and stability is incomplete without knowledge of the stabilizing effects of internal feedbacks from the biotic community; in particular, the roles of apex predators and climate change. Odum and Barrett (2005) accepted the premise that top-down biotic agents of change influence the primary production, biomass, and composition of communities, and Cain et al. (2011) accepted the trophic-cascade hypothesis as paradigm. Based on empirical evidence, they believed that both bottom-up and top-down controls operate simultaneously and that both may operate in species- rich ecosystems as well as species-poor ecosystems.
The universality of top-down effects in numerous ecosystems has been documented in many terrestrial and aquatic ecosystems, as well as in small predators, such as terrestrial salamanders, as well as large organisms, such as killer whales (Orcinus orca) (Ritchie and Johnson 2009, Terbough and Estes 2010, Chipps and Graeb 2010, Estes et al. 2011, Best and Welsh 2014, Payne et al. 2016). Yet, the conditions governing the relative contribution of top- down and bottom-up processes to ecosystem stabilization and diversity maintenance remain only partially explained; especially the role of global climate change.
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Estes et al. (2011) concluded that a rich literature revealed greater than expected effects of trophic cascades on many biotic and abiotic processes, including wildfire, carbon sequestration, biogeochemical cycles, invasive species, and disease. Estes and his coworkers revealed that the severe reduction of apex predation through human agencies interacts both additively and synergistically with climate and land use change, pollution, habitat loss, and other human impacts on the environment. These changes commonly result in dramatic changes in patterns of energy and material flux.
Most studies of apex predators have involved large predators (Terbough and Estes 2010, Estes et al. 2011) and large herbivores (Hempson et al. 2015), but small species and species in intermediate trophic levels can have profound effects as well. In a study of invertebrate-eating woodland salamanders, Best and Welsh (2014) demonstrated experimentally that the exclusion of woodland salamanders from forest floor environments was followed by increased abundances of litter-eating invertebrates, which substantially reduced the sequestration of carbon in soil humus. Because of the high biomass of woodland salamanders, they concluded that the carbon sequestration resulting from salamander predation could be a significant management consideration. Among apex predators, humans exhibit unsustainable “superpredator” behaviors that are profoundly disruptive in most ecosystems (Darimont et al. 2015) and have major impacts on ecosystem functions (e.g. Doughty et al. (2015).
The addition or deletion of a top carnivore or other keystone species to the food chain is likely to significantly alter community composition and critically influence the establishment and viability of numerous species. Cain (2011), Ricklefs and Relyea (2014) and Primack (2014) all accept the importance of top-down biotic regulation and cascade theory in ecosystems as well as bottom-up regulation, but the extent and complexity of these effects is still undergoing active investigation (e.g., Marris 2014). Conservation biologists in particular have often focused on conservation of large animals because they are often scare and keystone species that often serve as umbrella-species indicators of other species needs species (Van Dyke 2008, Primack 2014). These species in particular could play important roles in adaption to climate change. 6.1.5 Certain Patterns of Ecosystem Development are Generally Predictable Paradigm: Certain patterns of structural and functional change are generally predictable as ecosystems develop following a destructive but temporary disturbance. 6.1.5.1 Before 1996 Clements (1916) and Clements and Weaver (1928) described community development during succession largely in generic terms that prevailed for decades despite numerous advances in the scientific literature. Odum (1969) pulled these advances together in a highly influential synthesis paper entitled The Strategy of Ecosystem Development, which was intended to have important implications for the way people managed ecosystems. He summarized 24 hypotheses for how the gross functional and structural attributes of ecosystems change predictably during community succession. Many of the hypotheses relate to ecosystem stabilization and diversification. Most were well supported by empirical data and the general predictability of ecosystem development patterns, as described by Odum, was accepted by most ecologists.
At the population level, Odum noted that average niche specialization and the mean size of population members increase as ecosystems develop and population life cycles become more complex. Conditions in immature ecosystems favor r-selection and population production of high
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quantity while K-selection and more moderate production increasingly predominate in more mature ecosystems.
At the community level, the ratio of gross production to community respiration approaches 1 as ecosystems mature while the ratio of gross production to biomass decreases and the biomass per unit of energy flow increases. Food chains become more complex, longer, and more web- like as detritus feeding becomes more important. Species richness, the evenness of species abundances, biochemical biodiversity, and symbiotic interactions of species increase as ecosystems become more mature. Stratification and spatial heterogeneity generally become more recognizable.
At the ecosystem level, total organic matter in the system increases while inorganic nutrients redistribute through uptake from the physical environment to organic matter. Mineral cycles become tighter (less loss from the system) and conserve nutrients more effectively while nutrient exchange rates between organisms and the environment slow down. Detritus becomes increasingly important as a source of nutrient regeneration compared to geological sources. Ecosystems become more resistant to external perturbation and more stable as they mature.
Odum does not describe anything like a community-unit concept in the paper. The trends he described are largely gross functional and structural properties that emerge with ecosystem development under historical levels of stress. While Odum recognized that species richness and biochemical diversity increase and stability increase, the diversity may be composed of different species than were present in the ecosystem before it was disturbed. Different species from the same guilds may occupy the niches. Odum recognized that these patterns were highly interrelated and believed that further research would reveal much more about cause and effect relationships underlying the patterns.
Ecological concepts of resilience are implied in these hypotheses, but they do not take into account the effects of major shifts in ecosystems from one stable state to a quite different other stable state (Lewontin 1969, Holling 1973, Sutherland 1974). Yet, by the 1990s, Odum’s general proposal was mostly accepted, although some of the predictability of some of the specific patterns he proposed was not mentioned in Colinvaux (1993), Ricklefs 1993, and Brewer (1994). 6.1.5.2 After 1995 With some minor modifications, contemporary ecologists continue to generally accept Odum’s observations on predictable patterns of change in ecosystem condition as succession proceeds, but with possible destabilizing shifts in ecosystem states to different states of relative stability (Cain et al. 2011, Ricklefs and Relyea 2014). Climate change is now widely recognized as a potential source of such destabilizing effects, but the effects of rapid climate change on Odum’s more specific observations have not in general been considered in much depth. For reasons pertinent to management of global climate change and its potential effects on global biodiversity, more attention has been paid on the capacity of ecosystem development to sequester carbon in biomass and to resist the destabilizing effects of climate change on species viability than on the long-term effects of climate change on successional pattern predictability.
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6.1.6 Ecosystem Interactions Depend on Landscape Patterns, Populations, and Materials Paradigm: As open and dynamic systems, ecosystems interact with and influence each other at rates determined by landscape patterns and persistence, as well as population and material attributes. 6.1.6.1 Before 1996 Forbes (1887), Clements (1916), Clements and Weaver (1928), and Tansley (1935) focused intently on the internal processes of biotic communities as if they were closed and stationary systems. This may have been a matter of study convenience (they recognized that the source of energy for photosynthesis was outside community boundaries) , but the dependency of ecosystems on the cross-boundary movements of abiotic materials and populations was not clearly articulated or widely studied until after Odum (1953) highlighted the ecosystem concept. Odum (1959) further emphasized importance of open-system dynamics in sustaining the gross structural and functional attributes of ecosystems following a number of pioneering investigations of aquatic ecosystems (Odum and Odum 1955, Odum 1957, Teal 1957). The long-assumed stationarity of ecosystems gradually became more questionable with increasing recognition of past and present climate change (e.g., IPCC 1990a and b, 1996).
The importance of landscape patterns on the movements of abiotic materials and populations among ecosystems became a foundational principle of landscape ecology (Forman and Godron 1986). The ecosystems within rivers were often considered among the most open systems (Vannote et al. 1980), but by the 1990s, ecosystems in general were assumed to be open to and greatly dependent on cross-boundary “flows” of kinetic and potential energy, nonliving inorganic and organic materials, and species populations (Colinvaux 1993, Ricklefs 1993, Brewer 1994).
An early marker for the analytical origin of what was to become landscape ecology was perhaps the first mathematical analysis of the species-area relationship by Arrhenius (1921). It served as a foundation for the equilibrium theory of island biogeography proposed by MacArthur and Wilson (1963, 1967). Other outstanding theoretical and empirical contributions to the field included Watt (1947), Evans (1956), Odum (1957), Teal (1957), Likens et al. (1970), and Delcourt et al. (1983). Landscape ecology was first recognized as a separate discipline in Europe during the 1960s (e.g., Troll 1968) and became firmly established in the United States during the 1980s (Forman and Godron 1986). Studies in landscape ecology grew rapidly in number as the technology underlying remote sensing and Geographical Information Systems (GIS) advanced.
Forman and Godron (1986) defined an ecological landscape as a “heterogeneous land area composed of a cluster of interacting ecosystems that is repeated in similar form throughout”. The scale of major interest was large, ranging upward to a “kilometers-wide mosaic over which ecosystems recur” (Forman 1995). Its principles are based on the patterns of different ecosystem elements, or patches and corridors, in the larger ecosystem context, or matrix condition (Forman and Godron 1986, Forman 1995). Landscape boundaries were and continue to be determined by a combination of geophysical and biological attributes with the latter typically most important. Landscape ecologists faced the realization that all ecosystems intergrade to some extent, but where changes are sharp enough to recognize a boundary, the resulting patterns could reveal generalities about material and population flux rates when combined with ground knowledge. Landscape ecologists demonstrated that ecosystem interactions were influenced by the size, separation distances and shapes of ecosystem
U.S. Army Corps of Engineers 83 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence patches as well as their location on slopes and orientation to wind and other abiotic conditions (Forman and Godron 1986, Forman 1995).
Forman and Godron (1986) and Forman (1995) were well aware of landscape-level instability associated largely with human effects, which was a major research thrust in landscape ecology (Forman and Godron 1986, Forman 1995) and in conservation biology (Primack 1993). Landscape fragmentation was determined to be a major generic cause of species declines in many settings (Primack 1993). Over longer time frames, landscape ecologists accepted the growing paleoecological evidence showing quite dramatic changes in landscapes, particularly during the early Holocene, following retreat of continental glaciers (Forman 1995). The existing and potential effects of global climate change were not yet well assimilated, however (Primack 1993, Forman 1995), and many of the early principles of landscape ecology generally assumed landscape stability and stationarity in areas generally free of local human effects on the gross structure of ecosystems.
While sharing some common characteristics, watersheds were not considered ecological landscapes (Forman and Godron 1986) because their boundaries are entirely geophysical. However, watershed slope and drainage features contribute profoundly to understanding of on- the-ground flows of water, other materials, and organisms across ecosystem boundaries. The high degree of connection between watershed condition, stream flow, and sediment load carried in streams was well documented by hydrologists by the 1940s (Hursh 1948) largely before material and energy flows between ecosystems had been scientifically investigated. But before Odum (1959) was influenced by watershed studies that demonstrated the openness of ecosystems to material movements, ecologists in general operated in a different intellectual sphere. After disciplinary bridging, aquatic ecologists began to more carefully integrate watershed process into ecological assessments of materials import, retention, and export dynamics (Likens et al. 1970, 1977). By the 1990s, the importance of watershed process in aquatic ecology was widely accepted; much more than the principles of landscape ecology.
In terrestrial ecology, metapopulation concept development was closely linked to development of landscape concepts. The sizes of subpopulations are limited by the size of inhabitable patches, but because members of subpopulations can disperse with some success through the matrix ecosystem or through connecting corridors, the populations remain genetically connected and replenished by emigration and immigration. Yet the association of the metapopulation concept with landscape ecology was not widely recognized by ecologists in the 1990s. Colinvaux (1993), Ricklefs (1993) and Brewer (1994) accepted metapopulation concepts without explicitly connecting them with the emerging principles of landscape ecology.
The general relationship of species niches to landscape patterns and metapopulation dynamics was becoming more widely recognized among conservation biologists by the 1990s. Primack (1993) recognized that the population viability of species in narrow niches depends largely on size of suitable habitat fragments, or patches, and the quality of habitat corridors linking suitable patches within otherwise intolerable landscapes (Primack 1993). The distributions and abundances of species occupying broad niches are less likely to be constrained by fragmented landscapes. Intermediately adapted species might tolerate the landscape matrix condition for a limited time while moving to more optimum conditions within ecosystem patches. Primack (1993) generally believed that metapopulation models need to be specifically calibrated to the unique needs of each species.
The emerging principles of landscape ecology were not widely recognized by many ecologists outside the narrow field of landscape ecology in 1990s. There was, however, growing interest in metapopulation ecology among conservation biologists (Primack1993), and some leading
U.S. Army Corps of Engineers 84 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence academic ecologists. Colinvaux (1993), Ricklefs (1993) and Brewer (1994) generally accepted the importance of population and materials movement among ecosystems within landscapes. Yet the paradigms of ecosystem ecology remained conceptually dominated by the internal dynamics of trophic interactions, energy flow and nutrient recycling. 6.1.6.2 After 1995 Turner et al (2001) considered landscape ecology a work in progress with no well-developed theory of ecosystem function that is spatially explicit. They applied principles of landscape ecology to geographical scales ranging downward from many square kilometers to a few square meters. New ideas in landscape ecology have not advanced much more in textbooks since then, but practical applications have been honed (e.g., Lindenmayer and Fischer 2006, Farina 2007). Growing belief that global climate was changing rapidly amplified interest in application of landscape- scale concepts—particularly connectivity—to species adaptation assessments (Soulé et al. 2004). In the last edition of a textbook first published in 1953, Odum and Barrett (2005) introduced a new chapter on landscape ecology incorporating a summary of primary concepts, including the theory of island biogeography. General ecology texts enlarged their treatment and acceptance of landscape ecology principles. Cain et al. (2011), Ricklefs and Relyea (2014), and Primack (2014) generally accept the basic concepts of landscape ecology as well as the large body of evidence confirming the effects of landscape pattern on various ecological processes, species distribution, and species adaptation to climate and other environmental change.
More specifically, Cain et al. (2011) believed that ecological process at smaller scales influences landscape patterns at large scales; for example, the effects of animal movement and herbivory on seed dispersal and vegetation development. They also believed that disturbances both cause and respond to landscape heterogeneity. Fires and floods, for example, both affect and are influenced by landscape patterns. The variation in the responses of species to habitat fragmentation is now more widely appreciated (Lindenmayer and Fischer 2006, Primack 2014). Highly mobile species, for example, are less likely to be isolated by ecosystem fragmentation than less mobile species. Among species narrowly adapted to ecosystem conditions, isolation by fragmentation into smaller ecosystem patches is generally believed to be tolerated better by smaller species, which can generally sustain an MVPS in smaller ecosystem fragments than larger species. Larger or more abundant species are more likely to influence landscape patterns at larger scales than smaller or less abundant species, either directly through ecosystem engineering (Jones 1994) or indirectly through predation effects (Estes et al. 2014).
Relatively sharp ecosystem boundaries (edges), frequently caused by human land use, have numerous ecological effects (Van Dyke 2008, Cain et al. 2011, Primack 2014). Habitat fragmentation alters edge conditions in ways that favor some species over others. As fragment size decreases, the amount of edge increases and predator-prey and competitor interactions typically change. Invasive species tend to be well adapted to the interfaces between patch and matrix conditions providing them competitive or predatory advantages over native species less adapted to edge conditions. Fragmentation also alters evolutionary processes associated with small population sizes including genetic drift and the consequences of increased inbreeding. Microclimate often differs between patches and matrix, the extent of difference depending in part on patch size (Chen et al. 1999).
The influence of rapid climate change on landscape stability is a critical issue that has yet to be well researched. As temperature, precipitation, and gas concentrations change and interact with geology and physiography, abiotic aspects of landscape patterns are expected to change, causing changes in vegetation structure and other biotic aspects. Species adaptation to these changes is believed to depend in part on the rate of change. Some landscape changes could be
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dramatic, if the matrix ecosystem condition shifts rapidly into a new and quite different structural state. Those aspects of landscape science are underdeveloped leaving many questions about the future of species even in areas that are relatively free of local human impacts (Primack 2014). 6.1.7 Ecosystems Are Hierarchically Organized and Regulated Paradigm: Ecosystems are hierarchically organized and regulated, and events and interactions across hierarchical levels have population and other effects across many scales at the same time. 6.1.7.1 Before 1996 When Tansley (1935) coined the term ecosystem, he implicitly accepted the concept of a self- regulating ecosystem unit, largely isolated from other ecosystem units based on the community- unit concept championed by Clements and others (Clements 1916, Clements and Weaver 1928). The isolated ecosystem-unit concept began to break down as the openness of ecosystems became clearer and the influences of larger ecosystems on smaller ones were more clearly demonstrated. Teal (1957), for example, demonstrated the dominant influence of a deciduous forest ecosystem on a small cold-water spring ecosystem. The hierarchical theory of ecosystem organization emerged out of observations such as this to conceptually deal with the complexity confronted in attempts to understand and manage ecosystems at different scales from individual through populations to landscape and larger regional scales.
Evans (1956) had implied that ecosystems interacted hierarchically among the different levels, three decades before O’Neill et al. (1986) provided a theoretical foundation for the hierarchical organization of ecosystems, building upon the work of Allen and Starr (1982) and others. Hierarchy theory clearly dismisses the old concept of ecosystems as either closed or self- regulating. According to the theory, ecological processes are organized hierarchically from smaller systems to larger systems. An ecosystem at one level can be viewed and more effectively understood through the influences of the levels above and below the ecosystem level of interest (Urban et al. 1987). Functional units respond to lower-level units and are controlled by higher-level units. For example, the production of a number of terrestrial consumers responds to wetland food production as the wetland production is controlled by imported nutrients and the top-down effects of the terrestrial consumers. Levin (1992) believed that whatever population or other phenomenon was of interest it was simultaneously affected across a wide range of scales. Inherent in the paradigm is the belief that the stability of ecological structure and function generally increases with increased geographical area.
One of the more practical aspects of the hierarchical ecosystem paradigm from the standpoint of biodiversity management is the effect of spatial scale on system structure, functions, and stability. At the smallest spatial scale, say a subpopulation of snails in a small wetland area, the structure and functions of the subpopulation and habitat are vulnerable to local impacts and are often highly instable. A weather event, school of fish, invasion by a new species, visitation by snail-eating birds, or other destabilizing events may drive the snail subpopulation to local extinction and alter local structural patterns. At the larger spatial scale of the entire wetland complex, the numbers of existing subpopulations and habitats gain greater stability even though specific patterns are always changing within and among the ponds and the terrestrial ecosystem matrix. Through metapopulation dynamics, snail subpopulations may reestablish following local extinction in those areas where habitat suitability and connectivity exist or in new areas. Wetland areas that are permanently obliterated by a flood event are often replaced by newly formed wetland areas suitable for snail establishment. Based on these general relationships and
U.S. Army Corps of Engineers 86 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence local assessments, uncertainty about species viability can be managed by preserving at spatial scales most likely to provide the greatest stability in population numbers and long-term sustainability.
The hierarchical paradigm postulates linear relationships between the spatial and temporal scales of disturbance regimes, ecosystem processes, environmental constraints, and vegetation patterns. In those relationships, the scale of forcing factors and responses are related. In an application to ecosystem restoration, Holling et al. (1994) illustrated the relationship between the size of landscape features affected by disturbances—ranging from invertebrate habitats to land and ocean zones—and the dimensions and frequencies of atmospheric events—ranging from microbursts and lightning strikes to droughts and hurricanes. For example, a particular floodplain wetland area close to the river may be significantly altered by frequent small flood events, which incrementally fill it with sediment over time. Yet a much larger flood event reworks much of the floodplain area filling many old wetlands, restoring some filled ones, and creating entirely new ones, maintaining roughly the same balance between wetland ecosystem patches and the floodplain matrix.
The paradigm of hierarchical ecosystem organization had become a significant part of the theoretical core of landscape ecology by the mid-1990s. Landscape ecologists were beginning to accept the hierarchical theory of ecosystems (Forman 1995), but the concept had yet to be mentioned in contemporary general text books (Colinvaux 1993, Ricklefs 1993, Brewer 1994). 6.1.7.2 After 1995 Turner et al. (2001) emphasized the importance of scale and hierarchy theory in landscape ecology and defined landscape ecology to be “scale neutral” and just as applicable to areas a few meters across as to very large regions. They defined a landscape “as an area that is spatially heterogeneous in at least one factor of interest.” They, Parrott (2002) and others proposed broad theoretical frameworks for Figure 18. Conceptual model of a complex system, addressing how the attributes at one level in the such as an ecosystem at various scales from populations to ecoregions (from Parrott 2002). hierarchy emerge from the next lower level in the hierarchy (Figure 18). This definition focused more on heterogeneity than on repeatable patterns and indirectly on boundary interactions between ecologically different areas. It facilitated a more comprehensive integration of processes at the finest spatial scales of ecosystem dimensions and interaction (a small vernal pool in a forest, for example) to the regional interactions at the global scale of Earth’s ecosphere. It also allowed ecologists at all but the most extreme spatial dimensions to examine the interactions of their ecosystems of interest at whatever scale chosen from the perspective of ecosystem organization at lower and higher spatial levels. While Van Dyke (2008), Cain et al. (2011), Primack (2014), and Ricklefs and Relyea (2014) did not explicitly describe hierarchy theory, all of them recognized the importance of scale in ecological organization and scalar effect in ecological analysis. Yet many uncertainties
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6.2 Management Paradigms Overarching Paradigm: Ecosystem management is a comprehensive approach to protection, restoration and enhancement of ecosystem services that can and should complement a species-based approach to establishment and survival of specifically targeted populations.
In the past, eco-management was so focused on individual species that the viability of other species sometimes deteriorated. After numerous environmental laws were passed in the 1970s, federal management paradigms began to shift from maximizing output of specific goods and services to an approach intent on sustaining all existing and potential ecological resources found in biodiversity. Many of the paradigms in this section grew out of this transition. The ESA was a particularly influential law that contributed to the pressure placed on land and water management agencies to do what they could within their authorities to sustain endangered species and to prevent significant decline of other species. Federal policy emphasized that species targeted for enhanced use were to be managed for a level consistent with the long term sustainability of virtually all native species. Conservation biologists also began to shift their orientation from one solely focused on protecting imperiled species to one complemented by more holistic protection, restoration, and other management of ecosystems.
From the standpoint of objective setting in federal agencies, two of the paradigms standout, one based on the separation or integration of nature and humanity and the other on the concept of ecosystem services. As shown in this section, conservation biologists and other eco-managers at first emphasized the protection of ecosystems and associated biodiversity, but have since responded to a need in many situations to increase the likelihood of species viability through various management actions. In this process, the concepts of nature and natural have played a central role, as in protecting “natural ecosystems” from further degradation (implied unnatural conditions caused by human kind) and, when necessary, restoring “natural ecosystems”, including their associated biodiversity. The description of this paradigm shift shows how the separation of nature from human effects is more political than scientific, and follows out of a long history of philosophy and theology that separated man from nature (Uggla 2010).
The other paradigm pertains to the reorientation of management objectives toward provision of specific human services provided by ecosystems (ecosystem services). Two broad types of ecosystem services are central to government decisions about using or protecting ecosystem attributes from use. One type includes all service to present-day resource users, including commodity, recreational-aesthetic, floodplain, water-based commerce, water supply, hydropower and other services based on active resource use. These ecosystem services have use value (NRC 2005) demonstrated by the willingness of users to pay for them. Most use value is measurable in monetary units using widely accepted economic methods. The other type of service maintains use options associated with maintaining biodiversity for future generations of potential users in response to a national imperative to do so (such as the Endangered Species Act). It has “nonuse” value, which is only monetizable using questionable techniques that are not accepted by the Corps (USACE 2000). The third type of value—intrinsic value—is believed to be held in plants, animals, or inanimate things independent of any human service (NRC 2005). Those things are not valued as existing or potential resources, but may be protected from destructive use if enough people claim they believe in their “rights” to exist.
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6.2.1 Ecosystems Include Humans and cannot be Protected in or Restored to a “Natural” State Paradigm: Because of widespread ecological management and unintended human impacts, essentially all ecosystems include the effects of Homo sapiens and cannot be protected in or restored to a “natural” state. 6.2.1.1 Before 1996 This paradigm continues to evolve rapidly as recognition of the global extent and effects of global environmental change spreads among ecologists and eco-managers. In ecology’s early years, most Americans were theologically influenced to believe humans were special creations set apart from nature, which was created for human use. This separation implied that humans and their effects are not natural (Oelschlaeger 1991) or perhaps unnatural. Common language continues to reflect this sense of separation deeply, such as natural resources verses cultural resources. American ecologists also separated biotic communities into “natural”, meaning free of human influence, and communities altered by humanity (Pickett and Ostfeld 1995). They typically believed communities were in “balance” except when impacted by humans. Natural also meant normal, while human effects were abnormal. While linguistically convenient, the separation when taken as real had by the 1990s been recognized as an impediment to effective conservation and other eco-management (Curry 1977, Oelschaeger 1991). The paradigm shift from viewing nature as separable from human kind and its influences is particularly relevant for the Corps, which has embraced restoration of a more natural ecosystem condition as the operational objective for planning its ecosystem restoration projects (USACE 2000).
Whether any ecosystem studied by early ecologists was totally free of human impact is doubtful and grew more doubtful as human effects proliferated and scientific documentation of effects expanded. The details are typically unknown and may be unknowable, but humans probably have profoundly altered ecosystem structure and function in many places, including much of the United States, for thousands of years (e.g., Martin, 1967, Gill et al. 2009 and 2012, Levy 2011, Rule et al. 2012). The difficulty sorting natural process from anthropogenic effects escalated as agricultural, urban, transportation, water resources, and utility development spread; many more nonnative species were widely introduced, synthetic chemicals permeated the environment; intense harvest of keystone and dominant species greatly altered internal trophic dynamics; biogeochemical cycles were globally altered; and climate was changed as a consequence of human actions. By the 1990s, ecologists continued to speak of managing “natural” resources, protecting natural” areas, and restoring more “natural” ecosystems, but were increasingly aware of how difficult it was to sort out the underlying interactions between human effects and natural forms and processes.
The paradigm shift away from a belief in separate natural and human ecosystems happened slowly. Representing a professional minority, Tansley (1935) believed humans were just as much natural members of communities as any other species. Even several decades later, however, Clapham (1973) entitled his textbook “Natural Ecosystems”, splitting his concept of natural ecosystems from the “human ecosystems” addressed in another textbook. Then, in an early discussion about ecosystem restoration, Curry (1977) declared “All ecosystems are natural systems, with or without the presence of man.” In that scientific view of the material world, there is no such thing as more or less natural; the material world simply changes in ways that may require management to meet societal objectives. Curry tried to reorient restoration practitioners from what he regarded as a largely futile focus on restoring past “more natural” conditions toward achievement of specific objectives, regardless of the degree of human effect.
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Curry’s view was slow to be accepted, however. While the old paradigms were weakening, restoration professionals actually increased their focus on restoring past “natural” conditions, as exemplified by NRC (1992).
Certain federal agencies had practiced what was called ecosystem restoration without much scientific guidance until the National Academy of Sciences was commissioned to study the issue for aquatic ecosystems in the early 1990s (NRC 1992). The approach of the Academy’s study panel was broad and top-down. It held to the “objective” of restoring “natural” ecosystems”, but was also influenced and confused by growing skepticism about the feasibility of ecosystem restoration to a previous “natural” condition. The objective held by the NRC panel influenced the Corps policy guidance for ecosystem restoration project planning (USACE 1999, 2000), while the skepticism about the feasibility of ecosystem restoration was generally ignored.
The authors of the NRC report defined aquatic ecosystem restoration as the “reestablishment of predisturbance aquatic functions and related physical, chemical, and biological characteristics” with the aim to “return an ecosystem to a former natural condition” or a “close approximation of the condition prior to disturbance”. The report made it clear that achieving the ecosystem restoration objective meant restoring all original function and structure, including species composition. While the report emphasized the importance of setting clearly stated goals and objectives, such as improved water quality, it conflated the ecosystem restoration approach to objective achievement with actual objective achievement: the “objective” was “to emulate a natural, self-regulating system that is integrated ecologically with the landscape in which it occurs” (NRC 1992, page 18). A reader could conclude that naturalness itself is the valued objective and that relative naturalness of referenced ecosystems is the standard by which investment decisions should be made.
At least some of the authors of the NRC report suspected that restoring a fully natural condition by removing all human effect was impractical. They believed only a “naturalistic” approximation of previous conditions could be achieved (in the sense of Berger 1990). In addition to incompleteness, naturalistic implies artificiality—an artificial copy of nature as it appears in some “natural” reference. Removing a dam, for example, is often an impractical way to achieving the actual objective, such as successful fish passage to spawning areas or releasing flows from dams that are suitable for freshwater mussel survival. Building a fish ladder or releasing water from a dam to produce more naturalistic conditions are more practical alternatives, but are just as artificial as the dam and require continued human maintenance. In that context, the objective of ecosystem restoration stated in the report made little sense. What was actually required for practicality was the creation of new ecosystem conditions that were thought to emulate some previous ecosystem qualities. Corps policy guidance and practice embraced this “naturalistic” approach to ecosystem restoration (USACE 1999).
The NRC (1992) report signified the persistence of the man-verses-nature view of ecosystem management while that view was rapidly losing acceptance. The paradigm was shifting away from the notion that natural and unnatural ecosystems were scientifically separable. Pickett et al. (1992) revisited some earlier assessments and offered a “new paradigm”, which accepted the roles of humans in ecosystems as natural. In that new paradigm, ecosystem naturalness is irrelevant, ecosystems are not “in balance”, and local ecosystems are not fully self-regulating. They are instead ever changing and indistinct units bridged and in large part regulated by numerous cross-boundary processes originating from outside sources. This view could be defended with many empirical observations, but had little effect on the development of environmental policy guidance in the Corps or the ecological restoration community as a whole.
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6.2.1.2 After 1995 Many ecologists have since joined the lists of those who see futility in attempts to separate natural from unnatural ecosystems. That does not mean the word natural, as in natural resources or natural behavior, is becoming less acceptable in everyday language, but it does mean leading ecologists and eco-managers now doubt the use of a “natural” standard for restoring and protecting entire ecosystems. In an early textbook on ecosystem management, Vogt et al. (1997) accepted humans as integral to ecosystems. Chapin and Starfield (1997) introduced the idea of “novel ecosystems” to reflect how much humans had brought about ecosystem change and limited management options to approaches other than classical ecosystem restoration. Hobbs et al. (2006) later defined novel ecosystems as humanly altered systems that strongly resist restoration to their historical states. The novel-ecosystem concept owed much to the generally accepted concept of multiple stability states in ecosystems (Holling 1973, 1996). Once viewed as localized in effect and separable from nature, many ecologists now believe that at least some human effects are virtually everywhere within management reach (e.g., Jackson and Hobbs 2009, Hobbs et al. 2013). In no way has this been made clearer than by recognition of anthropogenic climate change of global scale and increasing acceptance of a new geological period called the Anthropocene, based on the extent of persistent human effect in the geological record (Kareiva et al. 2012).
Some leading eco-managers concerned about the restoration and protection of threatened biodiversity began to reject the idea of “natural” ecosystems entirely in favor of “wild” ecosystems and “rewilding” (Soulé and Noss 1998, Simberloff et al.1999, Foreman 2004, Monbiot 2014). Rewilding is often linked with reestablishment of the functions of large animals, but can apply anywhere ecosystems reestablish at least some degree of regulation free of human controls and disruptive impacts. Still, many believe there is a role for protection and restoration of gross ecosystem functions and associated services, if not for precisely the same species of earlier ecosystems. While the possibility of restoring the fine structure and function associated with native species composition is generally in doubt, a belief persists based on substantial evidence that some ecosystems can be restored to something close to what was believed to be “natural” rates of gross function (e.g., primary production) and forms of gross structure (forest biomass) (Hobbs et al. 2013).
Restoring gross ecosystem function and structure to a previous self-regulating condition independent of human control may be more tenable than restoring previous species compositions and associated fine structure and functions, but it remains a challenging effort for novel ecosystems that have shifted into another functional and structural state. It may require eliminating human regulation of landscapes and watershed processes including elimination of past agricultural, travel, urban, and water resources development designed basically to regulate nature. Even if technically feasible, such efforts are likely to be economically infeasible in many situations. Zedler et al. (2012) believed that anthropogenic changes are likely to be irreversible at any local project site situated in landscapes that remain significantly altered by humans.
Harris et al. (2013) distinguished novel ecosystems from humanly altered “hybrid” ecosystems that are to some extent restorable. They emphasized the importance of determining the restorability of desirable processes and structure consistent with clearly defined objectives before implementing expensive restoration measures. Since most ecologists now accept that biotic communities are always changing, and climate change has contributed even more to that acceptance, a holistic approach to ecosystem restoration (e.g., removal of dams or agriculture) when the objective is support for particular species is less likely to be successful than a species- based approach to determining ecosystem needs.
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Many ecologists have concluded that. In the sense of restoring a historical ecosystem condition, essentially all ecosystems are now irreversibly novel ecosystems (Vitousek et al. 1997, Ellis et al. 2010, Harris et al. 2013, Mascaro et al. 2013, Perring and Ellis 2013, Dornelas et al. 2014) or soon will be as climate changes in response to human activity. Mateos (2013) suggested that the changes are so universal; the term “novel ecosystem” applies essentially to all ecosystems and “we can dispense with its use.” These conclusions also imply that no ecosystem is “natural” and no reference condition can serve as a reliable indicator of a natural condition. The acceptance of human caused climate change among ecologists is quickly eroding any remaining acceptance of a management paradigm based on separating human effects from nature.
Appreciation for the depth of human impact on ecosystems is continuing to grow. Jackson (2013) reviewed numerous papers documenting the frequent changes in North American ecosystems before European settlement, especially during the period toward the end of the last ice age when humans first arrived. An increasing number of paleoecologists believe that humans had profound effects on ecosystems since first invading the Americas (Gill et al. 2009, 2012; Rule et al. 2012). Evidence is growing in support of the hypothesis that numerous large keystone species with great influence on ecosystems apparently were driven to extinction by human effects interacting with the effects of climate change (Levy 2011, Cooper et al. 2015). Losses of those species alone make complete restoration of what would have existed without human effect impossible.
There remain, however, numerous claims of more or less successful ecosystem restoration of gross function and structure to some recent “more natural” standard, such as less disturbed reference ecosystems (Hobbs et al. 2013). This coarse-filter approach to restoration is linked largely to improved use services such as those that result from soil erosion control or water treatment and supply. It is much less reliably linked to maintenance of species and genetic biodiversity, which requires more species-specific attention in both protection and restoration actions.
There appears to be no consistently reliable way to tell when an ecosystem is a somewhat reversible hybrid verses a novel ecosystem (Harris et al. 2013). Sustainable restoration of species composition has never been convincingly demonstrated. Since the nonuse values associated with maintaining biodiversity for present and future generations is linked closely to species composition, the likelihood that unsustainable species can be restored to a sustainable state by a coarse-filter approach alone is much less probable than a more targeted population approach to restoration.
The belief that many ecosystems are novel and changing in response to climate and other environmental change is causing conservation biologists and restoration scientists to reconsider the practicality of ecosystem protection and restoration on fixed properties and to “intervene” instead “with an eye to the future and toward managing for future change” (Hobbs et al. 2013, Heller and Hobbs 2014). Clearly identifying service-based objectives that indicate desired future conditions is important to this group. Managing for a “more natural” unit condition has lost practical meaning because ecosystems are constantly changing and human effects within them are so pervasive and difficult to discern, or impossible to eliminate when discernable. Starzomski (2013) noted that historically representative reference conditions used to guide restoration are scarce. He predicted there will be greater reliance on models that forecast species responses to climate and other changes in place of restoration models based on indicators of past conditions. These observations are particularly relevant to those concerned about the recovery of unsustainable species and long-term sustainability of biodiversity.
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Acceptance of humans as inseparable parts and processes of ecosystems has an important implication for management investment evaluations based on the scarcity of the desired outputs from management, whether it is economic scarcity or other measure of scarcity. Measuring the scarcity of “natural” and continuously changing ecosystems is much more problematic than measuring the scarcity of individual elements, such as species. While ecosystem elements may be sustainable in some combination somewhere, specific ecosystem configurations are unsustainable in the long run.
Many advocacy groups continue to resist the paradigm shift, believing that much of the scientific community has “given up on nature” (Goode 2015). Some contemporary eco-managers argue that we must “save” ecosystems to save individual species, but careful assessment of positions reveal their concept of ecosystem protection and restoration is based on maintenance of gross structure (e.g., prairie grasslands); not on specific community compositions (e.g., Ackerman 2016). Description of the “unnatural history” of human impacts is sometimes used as a device designed to appeal to the romantic notions of the lay public (e.g., Kolbert 2014). But, as concluded by Uggla (2010), there are managerially “dubious for at last four reasons: it may result in futile boundary-setting between humans and nature; it may be counterproductive to environmental protection; it assigns responsibility in a narrow way; and it proves an artificial dichotomy between humans and ‘pristine’ nature.”
Some of the more scientifically based and most influential advocacy groups have generally accepted these and other observations and the paradigm shift toward acceptance of humans as integral parts of nature with the power to alter nature, including the human parts, for sustainable human benefit. The positions taken are important because they determine how effectively governments and advocacy groups will spend scarce funds on protection and restoration actions, whether for “futile negotiations around how to define and distinguish natural and human impacts” (Uggla 2010), or to focus on achieving important sustainability objectives for this and future generations.
This paradigm shift has had two particularly relevant points for federal managers constrained by their authorities to specific categories of objectives. Whenever the objective of a restoration investment is the restoration and protection of species to a sustainable state, as it appears to be in Corps planning guidance (USACE 2000), attempting to restore a more natural or “naturalistic” ecosystem condition without considering the needs of desired species is less likely to succeed than appropriate long-term planning and management for the colonization, community, and habitat needs of the desired species. In contrast, for those managers concerned about sustainable use, restoring for typically common resource species—such as trees for lumber and recreationally valued wildlife—or for gross structure and function—such as floodplain ecosystems and flood impact moderation—is more likely to succeed without as much species- specific attention.
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6.2.2 Ecosystem Management Emphasizes Sustainability, Broad Scale and Adaptive Flexibility Paradigm: Ecosystem management for service-improvement objectives is constrained by a sustainability imperative, which requires consideration of management measures over a broad ecological scale, protection of rare and threatened ecosystem elements, continuous monitoring of service outputs, and adaptive flexibility to accommodate uncertainty in management outcomes. 6.2.2.1 Before 1996 The length of this management paradigm reflects the multi-dimensional complexity of the ecosystem management concept. Some elements of ecosystem management appeared in early ecological writings (Shelford 1933, Leopold 1949) long before the contemporary concept was developed in the 1990s. Whole-system management of reservoir levels and discharges was used to increase fishery yields (Nielsen 1993, Summerfelt 1993). Wildlife managers often applied forestry and agricultural techniques to enhance energy and material flow through a few game species. But few of these activities were self-regulating or self-sustaining. Odum (1969), among others, concluded that much agricultural, urban and other developmental management of ecosystems had reduced soil and nutrient retention, biological diversity, and ecological stability. The ecosystem concepts highlighted in Odum’s (1953, 1959) earlier texts set the stage for an ecosystem approach to management. The National Environmental Policy Act (NEPA) created a federal policy encouraging harmony between man and environment and increased understanding of ecosystems, leading Caldwell (1970) to call for an ecosystem approach to federal land use.
These early views laid a conceptual foundation for future federal concepts of ecosystem management, which included adaptive management (Holling 1978, Walters 1986) and management integrated over a broad ecological scale ranging from the habitats of individual populations to large ecoregions (Peterson and Parker 1998). The focus of ecosystem management shifted to ecosystem sustainability as the requirements of the ESA were assimilated by federal management agencies (Van Dyke 2008). Ecosystem sustainability included maintenance of all species while providing the social and economic outputs mandated by federal authorities. The evolution of the ecosystem management concept diverged from and merged with new concepts of sustainable development (WCED 1987), integrated resource management (Kidd and Pimentel 1992), conservation biology (Soulé 1986, Primack 1993), ecological economics (Costanza 1991, Costanza and Daly 1992), and adaptive management (Holling 1978, Walters 1986).
Adaptive management ultimately became a particular important and widely accepted aspect of ecosystem management before the concept was also accepted as essential to climate change adaptation (as described in more detail for an earlier paradigm in the climate section). As formally presented by Holling and Walters, it was a demanding approach to ecosystem management based in rigorous hypothesis testing, pilot studies, ecosystem models, scenario- based planning, and greater flexibility in the planning and management process than had been previously accepted in government agencies. It required an intimate integration of research science with management planning and implementation. The adaptive aspect of ecosystem management was generally conceptualized as a continuous, open-ended, and cyclic process that requires consistent monitoring and repeated adaptive management as needed for the expected outcomes far into the future (Bourgeron and Jensen 1994). See Figure 5 for a simplified diagram of the adaptive management cycle.
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Driven by political controversies involving resource management and the demands of the ESA, the U. S. Department of Agriculture fully embraced ecosystem management in the early 1990s as a strategy for resolving major management issues. The Forest Service announced its land management reorientation from a specific resource production focus to a more holistic consideration of and management for the long-term sustainability of all ecosystem outputs while compatibly providing goods and services for more immediate human wants and use (Overbay 1992, Jensen and Everett 1993). Other federal agencies soon followed suit, the Corps included, after the Clinton administration arranged a Memorandum of Understanding (MOU) among agencies about an ecosystem approach to management.
Politically, this “new” concept of management was basically a more comprehensive systematic socio-ecological framework for complying with NEPA, ESA, CWA, and other federal environmental laws as agencies pursued their congressionally authorized missions. But the ecosystem management concept quickly became many faceted and more confusing as it was diversely applied in different situations. Trying to make sense of it, Grumbine (1994) distilled a number of themes, including an ecosystem sustainability goal, a broad “systems perspective”, inclusion of humans in the ecosystem concept, defining ecosystem management boundaries, inter-organization collaboration, improved data use and management, adaptive management, and incorporating human values.
The meaning of ecosystem management diverged in the different interpretations of water and land management agencies. The Corps and other water resources agencies were influenced greatly by the emphasis of the CWA and NRC (1992) on the restoration and protection of the Nation’s waters, which the agencies pursued using hydrologic techniques to reestablish “more natural” attributes of gross process, structure and function (e.g., hydrology, dominant species, productivity). The naturalness of the result was the predominant indicator of social interest served. Other than using some dominant, common species as an indicator species, they generally ignored the fine-scaled needs of most other species. The land management agencies, led by Forest Service philosophy (Overbay 1992 and Jensen and Everett 1993) and influenced more by the ESA, emphasized a more comprehensive assessment of and consideration for the full diversity of socially significant ecosystem outputs. The naturalness of outcomes and the measures used were less important than sustaining the diversity of ecosystem outputs, often at a species-specific scale, consistent with societal goals. They accepted integration of fine-scaled, species-based management with coarse-scaled ecosystem management.
Ecosystem management philosophies were rapidly becoming established in the early 1990s, largely before the ecological community at large could fully assess the ecosystem management concept. Their rapid acceptance by federal government preceded development of any texts on the subject. In general, the concept was not well reflected in the major ecological and eco- management texts of the time (Ricklefs 1993, Colinvaux 1993, Primack 1993, Bolen and Robinson 1995, Brewer 1994, Scalet et al. 1996). Despite the federal acceptance, a commitment to ecosystem management, whatever it was, had yet to become a paradigm reflected widely in eco-management thinking outside federal government. 6.2.2.2 After 1995 Acceptance of certain aspects of ecosystem management progressed rapidly in the late 1990s. Boyce and Haney (1996) edited one of the first volumes dedicated to the idea that ecosystems can be managed to provide services to present users while sustaining biodiversity for future generations. The authors of this anthology emphasized how little was known about ecosystems, including the possible effects of climate change. They largely accepted the ecological concepts of resiliency, stability, constancy and persistence as “fundamental properties” of ecosystems,
U.S. Army Corps of Engineers 95 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence even as those paradigms were shifting significantly. They also believed local ecosystem integrity could be sustained with comprehensive planning: Temple (1996), for example, declared: “An ecosystem that fails to accommodate the needs of its rarest elements cannot be considered complete and reflects inadequate planning and management”. Some accepted the idea of ecosystem endangerment and extinction, much like the supraorganism concept of Clements and others.
In their book on ecosystem management, Vogt et al. (1997) elected to emphasize the management process independent of a sustainability goal or any other specific goal. They were exceptional, however. Of 13 definitions of ecosystem management described by Meffe et al. (2002), all but one clearly expressed an ecosystem sustainability goal or close approximations in ecosystem protection and maintenance statements. The apparent exception (Lackey 1998) implied a sustainability goal. Many of the definitions either implicitly or explicitly linked the sustainability goal to human service now and in the future. The roots of this thinking are diverse, but were probably significantly influenced in federal government by NEPA, the ESA, and the United Nation’s efforts in developing a conceptual basis for an international sustainable development ethic (WCED 1987).
Other themes, including adaptive management, were mentioned less consistently. Except for the consistent inclusion of sustainability and, to lesser extent, ecosystem services, the definitions of ecosystem management were so diverse that the concept verged on becoming meaningless (Bean 1997). Brussard et al. (1998) tried to make sense out of implementing ecosystem management by identifying seven critical steps: 1) delineate the ecosystem, 2) define strategic management goals, 3) develop comprehensive ecosystem understanding, 4) obtain socioeconomic data, 5) link socioeconomic and ecological data in a model, 6) implement experimental management actions, and 7) monitor management results for success or failure.
Dale et al. (2000) enlarged on these steps by distilling ecosystem- and landscape-level principles for sustainable resource management. Given the complexity and uncertainties of ecosystem dynamics, they recommended that managers plan for the unexpected, think at regional and landscape levels, preserve rare landscape and species elements, preserve connected areas for key species and processes, conserve resources, minimize species introductions, avoid and repair human impacts, and determine ecological compatibility with land use in advance of its use. Meffe et al. (2002) also believed ecosystem management was “an approach to maintaining or restoring the composition, structure, and function of natural and modified ecosystems for the goal of long-term sustainability”. They made points of genes being the “blueprints for life” and gene extinction leading to a lower level of ecosystem sustainability (see the paradigm on the genetic viability of populations in the population section). In that view, PVA and estimating MVPS are essential aspects of ecosystem management.
From the perspective of conservation biology, Van Dyke (2008) believed the concept of ecosystem management produced more confusion than effective application, but the emphasis on sustainability and human service, as well as the steps and principles offered by Bruzzard et al. (1998) and Dale et al. (2000), began to be more generally accepted and refined by conservation biologists. Consistent with the sustainability emphasis, the contemporary concept of ecosystem management provides a “safe operating space” for species vulnerable to climate and other global change by managing local threats as tightly as possible (Scheffer et al. 2015). Ecologists began to accept the role of ecosystem management complementary to species- based management bringing with it more emphasis on understanding the relationships of individual species to the whole. Although conceptually attractive, the idea that ecosystem
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management could replace species-based assessment and management has not been adopted by leading eco-managers (Hobbs et al 2013, Krausman and Cain 2013, Primack 2014).
The uncertainties in ecosystem management were highlighted by Vogt et al (1997) and Meffe et al. (2002) who advocated planning flexibility and adaptive management, but with less emphasis on the rigorously scientific approach advocated earlier by Holling (1978) and Walters (1986). Adaptive management became much more widely accepted as an essential part of sustainability-focused ecosystem management. The need for adaptive management is now accepted in principle in conservation biology and wildlife resource management (Primack 2014, Krausman and Cain 2013). Adaptive management generally includes a scenario approach to planning, monitoring management results, and responsive management flexibility (Peterson et al. 2003, Schindler and Hilborn 2015). Brashares (2010) emphasized the importance of continuously monitoring species-and community-specific conditions of protected areas to identify and manage undesirable changes in ecological filters. Hulvey et al. (2013) believed uncertainty can be managed by scenario planning and structured decision making using decision trees for which levels of uncertainty are assigned to each decision point (Polasky et al. 2011, Gregory et al. 2012).
McFadden et al. (2011) described two schools of thought about adaptive management. Both approaches address uncertainty and involve stakeholders, prediction models, monitoring, and a reiterative process. The resilience-experimentalist school emphasizes ecosystem resilience, stakeholder involvement in process and hypothesis development, experimentation, monitoring, and learning adaptively at several levels of complexity. Everglades management is used as an example (Gunderson and Light 2006). The decision-theoretic school emphasizes managing uncertain outcomes using an explicit decision-theoretic framework for the problem. Uncertainty is managed without experimentation by monitoring and learning from management outcomes; then adjusted as needed. Harvest management of waterfowl is an example (Nichols et al. 2007). Adaptive management can draw from both schools.
Most contemporary texts now accept adaptive management as a valid approach to management for sustainability at any ecological level, while wrestling with its complexity (Van Dyke 2008, Kwak and Freeman 2010, Cain et al. 2011, Krausman and Cain 2013, Primack 2014). However, the ecosystem management concept is not accepted exclusively or even predominantly. Eco-managers continue to focus on the needs of valued populations, whether for their use value or their biodiversity sustainability value. Primack (2014), for example, continues to be species centric while accepting the complementary principles of ecosystem management. 6.2.3 Ecosystem-based Management Units Have Value When Carefully Defined and Updated Paradigm: Ecosystem and landscape management units can be delineated usefully when they are carefully specified for management objectives and, in the case of ecologically determined units, boundaries are regularly updated to reflect environmental and demographic change. 6.2.3.1 Before 1996 Well into the 1990s, many eco-managers continued to manage areas as if the closed, self- regulating, and balanced community-unit and stability concepts of early ecologists still applied (Pickett and Ostfeld 1995, Meffe et al. 2002). Except for obvious influences, the old paradigms allowed areas to be managed quite independently of their surroundings. They could generally rely on the existing condition to indicate future conditions for decades to come. Community-unit
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maps provided a reliable foundation for management forecasting. Following the lead of Noss (1983), Jensen and Bourgeron (1994), Noss et al. (1995), and many others began to accept an ecosystem-based, coarse-filter approach to eco-management, which required delineation of management-unit boundaries. As first described, the ecosystem approach implied acceptance of old ecosystem unity, stability and stationarity paradigms.
Plant community-unit mapping began early in the history of professional forestry and continued in diverse forms for various purposes. A variety of typologies were developed at national scales over the years leading up to the mid-1990s (e.g., Kuchler 1964, Daubenmire 1966, Bailey 1996). These were primarily ecoregion classifications at relatively large scales less vulnerable to major boundary changes than smaller units, but also of less management use than classifications mapped at a more refined community scale. The 1992 NRC report on aquatic ecosystem restoration for federal agencies explicitly identified Omernik’s (1987) ecoregion classification and the more detailed wetland classification by Cowardin et al. (1979) as potentially useful for determining historical states of aquatic ecosystem restoration areas and suitable reference ecosystems for indicating “more natural” conditions. Kaufman et al. (1994) described conceptual criteria for boundary placement based largely on the strengths of ecological interactions.
But the concept of management units remained difficult to master because the open-ecosystem concept itself is difficult to translate into management-unit boundaries. Noss et al. (1995) found existing maps of “original” ecosystems (i.e., Kuchler 1964) to be too coarse and inaccurate, especially for wetland and riverine systems. They realized that species distributions were always changing and expressed a need for a classification scheme based on “enduring” physical features, such as landform and soil. They finally relied on professional judgment to delineate units while admitting the weaknesses of that approach. Scalet et al. (1996) warned of the difficulties inherent in boundary setting while implicitly accepting the concept of an ecosystem-based management unit defined largely by the ranges and needs of targeted fish and wildlife. 6.2.3.2 After 1995 Views on management boundaries diversified as ecosystem management evolved conceptually. Most accepted a need for management boundaries (Christiansen et al. 1996, Boyce and Haney 1996, Vogt et al. 1997, Meffe et al. 2002). The appeal of management units with well-defined boundaries was consistent with management that had long depended on maps of ecologically defined units. But defining units ecologically came up against the growing realization that ecosystems are continuously changing (Pickett and Ostfeld 1995, Jackson 2013). More eco- managers accepted boundaries based on property ownership and management jurisdictions in large part because they seemed to be not that much more arbitrary than ecological boundaries. As acceptance of rapid climate change grew, more ecologists realized that assumptions of ecosystem stationarity were unreliable.
The most consistent theme among assessments was selection of unit boundaries based on management goals. Christiansen et al. (1996) believed ecosystems are too dynamic for any single spatial or temporal scale. They thought management units ought to be defined based on goals and flexibility instead of ecological attributes. Vogt et al. (1997) advised delineating boundaries at various scales based on management purposes and selecting an appropriate scale for specific management goals and objectives. Meffe et al. (2002) believed that “strict boundaries” were irrelevant because ecosystem management is not a geographical place but an approach to problem solving. Kwak and Freeman (2010), however, reflected the widely accepted belief that the relatively fixed boundaries of watersheds define “natural units” for
U.S. Army Corps of Engineers 98 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence freshwater ecosystem management. However, given widespread anthropogenic changes in land use, climate, nutrient cycling, species distributions, and predation, aquatic ecosystem attributes often change dramatically within watershed boundaries, which also limit species adaptation to rapid climate change and other environmental change.
Advances in systems modeling (e.g., Grant et al. 1997) provided a possible means for circumventing the need for ecologically determined management boundaries. Regardless of where boundaries are placed on modeled ecosystems, the rates of energy and material inputs and outputs can be modeled in connection with dynamics within the unit boundaries. Yet at some point, input-output dynamics may overwhelm management of processes within units making it necessary to reexamine the choice of unit size or management objectives. Ecosystem restoration of segments in rivers and coastal zones is particularly influenced by the input-output dynamics of water and water-transported materials.
Despite the difficulties, conservation biologists in the United States now generally accept the ecosystem and landscape approaches to biodiversity protection promoted by Noss et al. (1995) and others as a complementary approach to species-based approaches. Groves et al. (2002) developed a 7-step framework for conservation planning that incorporated considerations of diversity at the landscape and ecosystem level as well as the species level. The extent to which each level is relied on to guide conservation planning depends largely on the availability of information at the species level. When relatively little species information is available, the assumption is that the scarcity of different ecosystem and landscape types is an imperfect indicator of scarcity at the species and population levels, but better than nothing. More useful, however, where species information exists, is an emphasis on identifying species-based management needs at a wide range of ecological and geographical scales.
Despite significant challenges identified by O’Neill (2001), Comer et al. (2003) developed a hierarchical classification scheme “that does not rely on a fixed landscape map unit and defines ecosystems using both biotic and abiotic criteria. They linked the “system units” to existing plant communities as defined in the IVC/USNVC (Grossman et al. 1998). While they recognized that the community-unit concept was no longer accepted, Comer et al. (2003) argued that “many combinations of species and habitats do indeed recur” and that “this viewpoint—one that is perhaps intermediate between the ‘community-unit concept’ and the ‘continuum concept’—has been widely used in guiding ecological classification.” This view was based largely on the distributions of relatively common species.
The approach of Comer et al. (2003) is less reliable for the many rare species, which can come and go individualistically in response to environmental and demographic changes, including random events. They admitted that “there are no unambiguous boundaries between plant communities or ecological systems in nature and species assemblages or ecosystem processes are not entirely predictable” and that any classification method “must be somewhat arbitrary with multiple acceptable solutions”. They also admitted to a need for periodic adjustment in response to change. However, their 50 to 100-year estimate for species composition stability at a “meso- scale” of tens to thousands of hectares are inconsistent with contemporary estimates of climate shifts (e.g., Diffenbaugh and Field 2014). Very frequent updating may be needed for the fine- scaled classification schemes of most value to many project-level management actions, particularly when the objective is to restore and protect the sustainability of scarce species. Even when regularly updated, the behaviors of common species may not indicate the individualistic behaviors of rare species.
Facing climate change directly, Groves et al. (2012) doubted the utility of any biological approach to management units and promoted a more complete reliance on geological indictors
U.S. Army Corps of Engineers 99 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence of regional biodiversity. They argued that biodiversity is often correlated with the diversity of geological features, which are more likely to persist through climate change than community- based indicators of biodiversity. Whatever the means of selection, reserves require appropriate connecting habitat for species dispersal or, in many island-like situations, actual translocation of species and that typically requires a larger regional perspective. Many aquatic ecosystems fall into this latter category.
Contemporary resource managers and conservation biologists tend to differ on the question of ecosystem management boundary placement although most see some value in it. Resource managers tend to align more with the ecosystem managers of earlier years, accepting property and political boundaries as well as ecologically defined boundaries, depending on circumstances (Krausman and Cain 2013, Fogarty and McCarthy 2014). Conservation biologists are more leery of political determinations. They have developed numerous ecological principles for guiding management unit location, size, shape, and connectivity (Van Dyke 2008, Primack 2014) based largely on the needs of targeted species. They also recognize the difficulties. Van Dyke for example, believed that definitions of naturally bounded ecosystems “can be elusive” and “notoriously leaky”. In light of subsequent environmental changes, Groves et al. (2012) highlighted the uncertainty associated with past community-based approaches to identifying conservation-based management units.
The trends have been toward delineating larger units, but uncertainties exist at landscape scales as well, since landscape boundaries are often determined by matrix ecosystem boundaries. Agency interest in landscapes as a focus of management analysis has grown since global climate change has been accepted as a problem for fish and wildlife management. In 2009, The U. S. Department of the Interior created the LCCs, which are regional, stake-holder driven partnerships designed to redirect the disparate approaches of conservation organizations into a more focused but comprehensive attention to environmental change and ecosystem management at landscape scale (Cole et al. 2018). A primary goal is to identify landscapes of high management priority in a rapidly changing environment.
The relationship of landscape units to ecosystem units is not always clear. That appears to be the case for nature reserve planning based on either biological or geological indicators of management units. The hierarchical approach of Comer et al. (2003) may be appropriate if a landscape is defined by the pattern of ecosystem patches in a larger ecosystem matrix (such as cold water lakes in a boreal forest matrix). But the approach used must be able to regularly update management-unit maps and community-change projection models. Even so, existing paradigms suggest that the behaviors of common community members can be misleading and should be complemented with information on individual species of exceptional concern.
Contemporary eco-managers continue to believe that delineating ecosystem/landscape management units is useful when they are defined by the specific goals accepted by management stakeholders and updated as needed. However, no standard approach has been proposed or accepted for all purposes. Most of the various approaches are spatially fixed in management-units often defined by inflexible property boundaries. Eco-managers now generally agree that these diverse approaches need to be well coordinated to achieve sustainability goals, but how that will play out is unclear.
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6.2.4 Ecosystem Management Should Complement Species-Based Management Paradigm: Ecosystem management should complement species-based management to achieve species-based objectives and sustainable ecosystem diversity and services. 6.2.4.1 Before 1996 Early observations following pollution control and restoration actions confirmed that simply removing the cause of ecosystem alteration could not be counted on to restore species composition (Cairns and Dickson1977, Cottam 1987). Botkin (1977) addressed this issue of species failure to establish in apparently suitable habitats. He provided a list of 18 species- specific risk-management questions that read like the attributes of populations described under the section on population paradigms. In sum, managers needed a lot of species-specific population information to significantly improve the probability of success when the targeted services are specifically linked to species recovery and protection from destructive use. Once management actions are taken, the populations need to be carefully monitored and adaptively managed. With this brief description, Botkin (1977) produced a framework for the species-based population approach to conservation biology applied in the 1980s and 1990s (Primack 1993).
In addition to species needs, Noss (1983), Hutto et al. (1987) and others emphasized the importance of also considering the integrity of ecological patterns and processes at landscape scales, especially where species data were limited. Use of this coarse-filter approach at ecosystem and landscape scales by itself assumed that when landscape patterns and process are maintained (or restored), the full complement of species will persist and biodiversity will be maintained (NRC 1992). In contrast, total reliance on the fine-filter approach assumed that the appropriate landscape patterns and process could be determined entirely through the needs of targeted species.
Integration of both approaches, it was argued, would more effectively manage the risks associated with each approach. The species-based approach was risky because the data were rarely complete. Yet the assumption that the coarse-filter approach alone would protect all species in a landscape was even riskier for rare species. Bourgeron and Jensen (1994) noted that the coarse-filter approach had been used most successfully to maintain common species over large areas while the fine-filter approach had been used to maintain rare species and communities “that would fall through the cracks of the coarse-filter”. Bourgeron and Jensen (1994) believed that ecosystems can change at variable rates and into very different stable states that were not “perfectly predictable” and needed to be managed adaptively to accommodate “surprise” (Kay 1991). But they accepted the need for both fine- and coarse-filter approaches, consistent with the recommendations of Hutto et al. (1987). Primack (1993) overwhelmingly emphasized the fine-filter approach while recognizing the potential for an ecosystem and landscape approach in settings where little fine-filter information exists. 6.2.4.2 After 1995 Acceptance of an integrated species and ecosystem approach to management voiced by Bourgeron and Jensen (1994) continued to grow (Vogt et al. 1997) as doubts about the effectiveness of ecosystem restoration spread. Recognition of the individualistic response of many species to climate and other environmental change reinforced these trends. Most contemporary eco-managers now believe ecosystem management provides a framework for integrating the fine-filter approach into a more holistic approach to biodiversity conservation.
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Noss and Scott (1997) split the ecological filtration concept into a species-based, “fine-filter” approach and an ecosystem-based “coarse-filter” approach, with the idea that ecosystem management could complement species-based management, but not necessarily replace it. Meffe et al. (2002) believed that a blend of the coarse-filter and fine-filter approach to nature- reserve selection would be more likely to protect all species present under the umbrella of the ecosystem-landscape approach while reducing the risks of not covering species protection priorities. A common theme throughout discussions of novel ecosystem management for biodiversity maintenance by Hobbs et al. (2013) was the need to consider species requirements as well as ecosystem-level functions and structure. Krausman and Cain (2013) believed that the coarse-filter approach was bound to fail if applied without the species-based, fine-filter approach.
With certain caveats, Cain et al. (2011) confirmed the importance of integrating species and ecosystem approaches through habitat assessment and management. They believed protection priorities for “the rarest and most rapidly declining species” need to be identified first. For the purposes of conservation biology, habitat selection needs to be based on the needs of imperiled species; not the scarcity of habitat types independent of any identified species. A scientifically objective database is needed for this assessment. The IUCN red list is useful globally, but NatureServe Explorer is generally considered the most complete and up to date database for the United States (NatureServe 2015).
The concept of biodiversity hot spots (Myers et al. 2000) also integrated species and ecosystem-level considerations. Hotspots are often indicated by the numbers of species in a well-known group, such as birds, based on the assumption that the diversity within groups is correlated (Primack 2014). The assumption may not always hold true, but the fractional loss of hotspot area is generally accepted as a better indicator of biodiversity conservation needs than the fractional loss of an ecosystem type (Primack 2014). The biodiversity hotspot approach has been refined in indices of rarity weighted species richness (Williams et al. 1996, Cole 2016), which may be useful where the requisite information is available. 6.2.5 Ecosystem Goods and Services Inform Ecosystem Management Paradigm: Ecosystem structure and function provide human goods and services that inform eco-managers about ecosystem values, management problems and management objectives. 6.2.5.1 Before 1996 The concepts of goods and services have a long history in eco-management thinking, primarily with respect to human enhancements of natural process. The idea of unenhanced “natural” services has circulated for more than a century. Marsh (1864), for example, spoke of the service rendered by insect consumption of decaying matter and by forest regulation of soil erosion and local climate. Many ecologists since then referred to ecological functions that had positive effects on human well-being, implying the concept of service (Mooney and Ehrlich 1997). But the concept of ecosystem services did not become an explicit topic of communication between ecologists and economists until the 1990s. During that era, environmental protection and restoration in federal government was typically justified based on resource significance, consistent with NEPA. That relatively crude approach to determining ecosystem value began to change rapidly when ecological economics emerged as a cross-disciplinary approach to valuation of ecosystem outputs and ecosystem management principles emphasized the importance of ecosystem services.
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Ecological economics—the economics of ecological sustainability (e.g., Costanza 1991)— provided an analytical bridge between ecosystem management and the people it served. It drew from traditional economics to develop the concepts of ecological-economic models, “natural” capital, ecological goods and services, and investment discounting more favorable to future generations. Most ecologists and eco-managers accepted ecological economics slowly. Of the major text writers, only Primack (1993) included them. The NRC (1992) report on aquatic ecosystem restoration mentioned the importance of integrating “ecological values” and “human values” in consideration of plan evaluation, but did not refer to ecological goods and services. The reader also was not told how ecological values differ from human values, nor has any other author reviewed here been able to explain this cogently. This separation of ecological and human values was, perhaps, typical of a more general confusion over emerging concepts. 6.2.5.2 After 1995 The emphasis of ecological economics on natural capital (Goodland and Daly 1996) and ecological goods and services (Daly 1996) grew stronger. In a book edited by Daily (1997) on “natural services”, Goulder and Kennedy (1997) clarified some confusing economic issues. They made clear that both consumptive and non-consumptive use of goods and services could be valued directly and that the ecosystem production systems for those goods and services was indirectly valued through the value of the produced goods and services. Totaling direct and indirect values would be, incorrectly, “double counting” the value of ecosystem goods and services.
Goulder and Kennedy (1997) also distinguished nonuse value from use value. Nonuse value is typically revealed by the desire to set aside land, water, and other resource use for the possible benefit of others. In federal eco-management, nonuse value is typically revealed in land, water, and species protections from use for the benefit of future generations. A report of the National Academy of Sciences on ecosystem service valuation (NRC 2005) enlarged upon the concepts described by Goulder and Kennedy and generally confirmed their assessment. It also described the economic controversy associated with measuring nonuse value in monetary units. Numerous economists objected to monetization based on theoretical and practical grounds, as do the Corps of Engineers (USACE 2000). When destructive use is set aside out of respect for the rights of future generations to inherit the diversity of ecological opportunities left to the existing generation, the value is intrinsic and entirely outside the scope of economic valuation. This includes the existence value of all species (defined by their security from extinction) and areas set aside from destructive use in perpetuity for cross-generation benefit.
The importance of ecological goods and services grew in conservation investment planning (Groves et al. 2003). Jax (2005) emphasized the connections between human service and ecosystem functions. The Millennium Ecosystem Assessment (2005) of the United Nations used ecological goods and services as means for identifying the benefits of sustainable development, but limited their concept of service recipients to present generations. They did not consider biodiversity maintenance as an option-maintenance service that benefits future generations and sustains human well-being. In addition, biodiversity conservation organizations began shifting from an investment decision process based on the existence value of nature (a nonuse value) to the use value of services (Mace 2014). They apparently assumed that conserving ecosystems for their use value would also sustain biodiversity more effectively than focusing on its nonuse and intrinsic value (“nature for its sake” or for future generations), which are difficult concepts for the public. The assumption is controversial (Soulé 2013), however, largely because of a fear by some conservation biologists that the scientific underpinnings of ecological sustainability will be displaced by decisions based too much on use value (Mace 2014).
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Shackleford et al. (2013) stressed the need to clearly link ecological restoration objectives to the human service value that justifies project investment. Restoration ecologists increasingly believe that restoration ecologists should turn away from restoring previous conditions to focus more on desired future conditions determined in large part by ecosystem service provision (Hobbs et al. 2013). The concepts of ecosystem goods and services are now quite widely accepted as important in federal environmental and natural resource management agencies, and are likely to be integrated into future management planning.
The term “natural services” continues to be used in eco-management while being over- shadowed by a largely rejected idea of nature as separate from humanity, because of the complications it causes (e.g., Uggla 2010; refer to the first paradigm under ecosystem management for details). The difficulty inherent in identifying services free of all human effect can unnecessarily complicate planning. Focusing on “natural” attributes can divert from careful identification of the human services desired of ecosystems regardless of how free of human effect they may be. Most contemporary ecologists (as opposed perhaps to advocacy environmentalists) accept humans as integral and inseparable parts of nature. 6.2.6 External and Internal Forcing Functions Often Need Joint Management Paradigm: Achievement of a sustainability goal frequently requires joint management of external forcing functions originating in the abiotic environment as well as internal forcing functions originating in the biotic community. 6.2.6.1 Before 1996 The “bottom-up” effects of energy input (mainly light), nutrients and other external forcing functions that contribute to ecosystem regulation were recognized and managed long before the “top-down” effects of predation, competition, and other internal forcing functions. Before its importance was recognized, regulation by internal forcing functions may have been ascribed to external forcing functions. For example, Ryther (1969) and Russell-Hunter (1970) used ecological efficiency theory and measurements of oceanic primary productivity to estimate a maximum ocean fishery yield based on a bottom-up process, which proved to be reasonably accurate based on later data. The apparent accuracy of their model may have been fortuitous, however, because the possible effects of internal forcing functions were not considered. Disproportionately reduced consumption by heavily fished top carnivores may have boosted the total yield available to fisheries to an unsustainable level.
The potential effectiveness of managing trophic cascades to improve environmental quality was experimentally demonstrated by Shapiro et al. (1975), who revealed the effect of predator manipulation on water clarity. By removing top-level predators, the abundance of first-level predators increased and the biomass of herbivorous zooplankton decreased. That allowed phytoplankton biomass to increase, reducing light transmission. The implications for fishery management were important because changes in light transmission could affect the balance of phytoplankton and rooted aquatic plants in many ponds and shallow lakes that support popular sport fisheries. Biomanipulation for water clarity was not discussed much in the major fishery management text of the time (Kohler and Hubert 1993), but grass carp introductions for control of aquatic plants (Pípálova 2006) and benthic fish removal to decrease turbidity caused by sediment disturbance were well established management tactics by the 1990s (Kohler and Hubert 1993).
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A big concern in terrestrial environments was the effect of overgrazing on habitat quality, including reduced food production for grazers and aquatic habitat degradation by soil erosion (Bolen and Robinson 1995, Scalet et al. 1996, Kohler and Hubert 1993). The connections between wild herbivore overabundance and depressed predation on the herbivores were widely recognized and generally managed through compensatory hunting regulations. Primack (1993) demonstrated the importance of trophic-cascade management by describing the stabilizing effects of sea otters on kelp-forest communities. Sea otter eat various invertebrates, including herbivorous sea urchins. When sea otters are depleted, sea urchins increase in number and consume most of the kelp. 6.2.6.2 After 1995 The paradigm shifted toward more emphasis on managing internal forcing factors associated with top-predator and other keystone species in concert with external forcing functions, but with respect for the complexity of interactions and uncertainty of results. Based on their review, Chipps and Graeb (2010) believed that aquatic food webs could be managed top-down by managing trophic cascades, as well as bottom-up. They emphasized, however, that food web management is complex and can produce inconsistent results because food webs may be structured by keystone species from the middle trophic levels as well as the upper levels. Knowledge of the system is critically important.
Conservation biologists in general recognize the importance of predation and competition in ecosystem dynamics and believe their effects, while complex, are essential to consider in ecosystem management for sustainability (Van Dyke 2008, Primack 2014). Predators are often the most far-ranging species for which corridors are designed and managed. Krausman and Cain (2013) described the application of the “top-down” principle of ecosystem regulation in Yellowstone National Park where wolves were reestablished in the park after decades of extirpation. The elk population expanded following the loss of wolves and probably reduced the abundances of young aspen, willows, and cottonwoods. Wolves were expected to reduce elk populations; increase survival of aspen, willows and cottonwoods; and stabilize species abundances. The community responded about as expected, although other changes in consumer relationships and climate complicated conclusions about the degree to which wolves alone were responsible (Marris 2014, Morell 2015).
The importance of the paradigm for invasive species management became clearer. One dramatic effect of zebra mussel invasion of Lake Erie was its impact on water clarity and subsequent changes in nutrient cycling and community composition (Karatayev et al. 2002). Accurately predicting which nonnative species are likely to have such extensive effects on species competition and survival is considered a critical need for invasive species management (Krausman and Cain 2013).
The relationships among keystone species and the production, biomass, and gross structure of communities have major management implications for the maintenance of desired species. This is particularly well recognized for apex predators (Terbough and Estes 2010, Ripple et al. 2014), but keystone species at all levels can have profound effects in many ecosystems. Terbough et al. (2010) emphasized that decades of research revealed that internal forcing functions were not, as some had suggested, acting independently of external functions and limited to certain ecosystems, but “are complementary countercurrent flows, inextricably bound together” in ecosystems of many terrestrial and aquatic types. Apex predators are believed to contribute exceptionally to the stabilizing effect of increased biodiversity on ecosystem structure and function. At the extremes, virtual elimination of apex predators by fishing or hunting may result in a “catastrophic” shift to an ecosystem state with very different composition and functionality
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(Scheffer et al. 2001, Scheffer 2010, Essington 2010, Terbough and Estes 2010). Sophisticated determination of harvest regulations considers these possibilities.
Eco-management has not entirely caught up with the science despite recognition by leading eco-managers of the importance of managing predation and competition for desired results (Krauseman and Cain 2013, Primack 2014). Estes et al. (2011) argued that the continued inability to predict and moderate human impacts on the environment comes from misunderstanding of the causes, which many managers continue to assume are largely external forcing functions. Others believe that human predators can take the place of other predators, but Estes et al. (2011) argued that humans are not nearly as effective at stabilizing communities. They are thought to be less flexible predators, which are more likely to reduce prey species to exceptionally low abundances and possible species “collapse” (Terbough and Estes 2010, Darimont et al. 2015). Estes et al. (2011) believed that human replacement of and impacts on apex consumers may be “humankind’s most pervasive influence on nature”. Ripple et al. (2014) thought its impacts are on par with global climate change. Unlike other predators, humans have the unique ability to analyze and modify their behavior—a point that is essential for continued coexistence of other species with humanity (Worm 2015). 6.2.7 Restoration of Ecosystem Elements is a Legitimate Complement to Protection Paradigm: Restoration of ecosystem elements (but not entire ecosystems) is an essential strategy when protection alone cannot sustain biodiversity. 6.2.7.1 Before 1996 Public protection of natural landscapes has a much longer history (Bean 1983) than ecosystem restoration. The landscape concept so important in setting aside national parks in 19th century America was more aesthetic than ecological (Novak 1995), but the parks protected many species from habitat destruction and human predation. When conservation turned toward recovery and protection of imperiled species, it continued to emphasize a “nature reserve” strategy (Primack 1993). The earliest manifestations of ecological restoration were not nearly as well regarded by conservation biologists. Restoration was a very loosely interpreted concept, which often varied from a strict historical standard for ecological condition. Ecosystem retention of erosive soils was often “restored” by planting exotic species. The Dingle-Johnson Sport Fish Restoration Act provided the states with funds to manage sport fisheries in various ways, including construction of artificial lakes and introductions of nonnative species. “Restoration” was nearly always intent on establishing some desirable type and level of ecological service that often differed from “natural” conditions. Conservation biologists trusted more in the protection of “natural” areas and generally regarded ecological restoration as a much riskier venture.
The restoration and protection of ecosystems supporting imperiled species populations became more attractive as more conservation biologists and other eco-managers realized that protection alone was insufficient (Curry 1977, Noss 1983). The generally positive response of aquatic ecosystems to point-source pollution control made a strong case for restoring gross aspects of ecosystem function and structure, if not the more nuanced aspects of ecosystem composition. Because the usual interest was focused on gross function, the completeness of recovery after restoration activity was rarely documented (Welch and Cooke 1990, Makarewicz and Bertram 1991). When it was documented, species compositions had changed. Monitoring the results of point-source pollution removal frequently revealed the return of water chemistry and dominant sportfish to conditions more like unpolluted reference conditions, but the species composition
U.S. Army Corps of Engineers 106 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence generally differed from the original condition (Cairns and Dickson1977). Similarly close monitoring of the species richness at two restored prairies documented restoration of gross functions and many native species, but the composition differed from undisturbed sites (Cottam 1987).
The uncertainty observed in complete ecosystem restoration to a historically “natural condition” led Jordan et al. (1987) to promote restoration as an ecological research tool rather than a reliable way to manage ecosystems for specific conservation goals. Restoration remained little more than a footnote in the conservation biology of the 1990s. Primack (1993) recognized its potential importance for conserving biodiversity, but considered it to be insufficiently researched and unreliable. Bolen and Robinson (1995) and Scalet et al. (1996) accepted restoration of specific ecosystem elements, such as bison and turkey, to abundances more like previous times (more like some aspects of the rewilding concept already described), but not the restoration of entire ecosystems.
The NRC (1992) report on federal ecosystem restoration adopted an ecosystem-scale approach to management. Yet it only superficially explored the applicability of important ecological concepts including ecosystem succession and resilience, energy flow, biogeochemical cycling, top-down and bottom-up regulatory mechanisms, cross-boundary ecosystem interactions, or many of the population concepts necessary to carry out a complementary fine-scale approach to achieving restoration objectives. The authors emphasized the importance of a “holistic process”, such as restoring wetland hydrology, geomorphology and the gross structure of vegetation. The importance of biodiversity and species roles was noted, but not developed, and providing for the needs of keystone, dominant, or other essential support species was hardly addressed.
The case studies described in the NRC report primarily reflected federal interests in mitigating impacts on gross ecosystem functions that were economically valued, such as flood moderation, potable water supply, and generic habitat support for common recreational fish and wildlife. Ecosystem restoration was considered a more reliable alternative to artificially created functional states. Federal policies were not averse to creating a functional state with a value that exceeded the historical condition. Probably because government conservation biologists were skeptical of its effectiveness, little interest had been shown in restoring ecosystems to sustain imperiled species. 6.2.7.2 After 1995 Doubts about the value of ecosystem restoration for conservation purposes continued. Simberloff et al. (1999) criticized the many “restoration” projects with vague goals and objectives as well as the wide acceptance of a “goal” to reproduce a past ecosystem condition (essentially the NRC [1992] goal). In their view, objectives should scientifically specify the time and composition of the desired community or ecosystem condition and should “not let the process become the goal of restoration efforts at the expense of species.” They promoted careful ecosystem restoration where preservation alone was insufficient, but also recognized the risks. They also emphasized that much could be learned by treating restoration projects as scientifically structured and tested hypotheses—a reflection of the rigorous approach to adaptive management described by Walters (1986). More recently, Suding (2011) documented numerous restoration challenges and failures. But ecological restoration, if not precisely ecosystem restoration, gradually gained broad acceptance (Cain et al. 2011, Primack 2014, Krausman and Cain 2013).
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The goal for ecosystem restoration chosen by the Society of Ecological Restoration (SER) was to restore ecosystem “health and integrity” (SER 2004). In this context, ecosystem integrity generally means the ecosystem is complete in structure and function as indicated by some reference condition presumed to be structurally and functionally complete. A healthy ecosystem is determined by functional sustainability, self-regulation, and serviceable (Rapport et al. 1998). The SER acknowledged the challenges involved in achieving this goal, but failed to resolve them clearly or carefully consider the implications of global climate change.
The old paradigms of ecosystem restoration presented in NRC (1992) and SER (2004) were beginning to give way to new concepts more compatible with the realities of climate and other environmental change (Choi et al. 2008). Shackelford et al. (2013) pointed out the failure of SER (2004) to allow for differential shifts in species ranges in response to environmental changes, including global warming. When biodiversity maintenance is the issue, success may depend on targeting management needs for individual species based on possible future conditions rather than on historical or existing conditions. Because of the changes taking place, Shackelford et al. (2013) and Harris et al. (2013) believed that reference ecosystems may serve to inform specific understanding of relationships, but not as general models for holistic ecosystem restoration because reference sites and project sites are increasingly diverging from historical conditions as the climate and other environmental change occurs. They accepted some nonnative species in community assemblages as proxies for native species to sustain supportive biodiversity. The pollinating effectiveness of the nonnative Eurasian honeybee is a common example (Kennedy et al. 2013). Hulvey et al. (2013) accepted that premise as long as the inevitable differences among species in the same guild were carefully considered (Eviner 2004).
Additional criticisms of past practices were directed at the unevaluated use of indirect indicators of conditions. Kennedy et al. (2013) demonstrated how much plant ecologists had dominated restoration ecology and how often assumptions that successful animal restoration can be reliably indicated by abiotic features and plants proved false. Hulvey et al. (2013) recognized that these assumptions are “an extension of the “Field of Dreams” hypothesis” described by Palmer et al. (1997) and Hilderbrand et al. (2005). The hypothesis—if you build it, they will come—is commonly proven false by failure to meet specialized habitat and community needs or connecting restored sites to essential source materials and populations (Kanowski et al. 2006). Based on the work of Morrison (2009), Shackelford et al. (2013) also knew that, regardless of other indicators of success, “the actual size of a site to be restored may constrain restoration success; for instance, by impeding the establishment of viable populations of a key desired species or trophic level”.
Jackson (2013) questioned whether any ecosystem is restorable for any long-term duration. Quaternary paleoecologists have repeatedly documented unprecedented intermingling of species in a long sequence of new assemblages without modern counterparts (Graham 1986, Overpeck et al. 1992, Williams et al. 2001, Jackson and Williams 2004, Jackson and Overpeck 2000, Jackson 2006, Williams and Jackson 2007). Jackson (2013) strongly implied that the types of community changes associated with projected climate change would most likely be irreversible sooner or later. Krausman and Cain (2013) accepted the conclusion that rather than attempting restoration of a historical state, a more achievable objective is to manage a damaged system for a supportive condition similar to reference-site conditions. They believed the focus should be more on developing the desired conditions for targeted species than on restoring some previous condition.
In the new paradigm that is emerging, virtually all ecosystems have changed relatively recently in geological time in response to human effects, including the effects of anthropogenic climate
U.S. Army Corps of Engineers 108 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence change. Old concepts of ecosystem protection and ecosystem restoration based in assumptions about stability and stationarity are not feasible in the long run. Many existing ecosystems are novel ecosystems very resistant to restoration of some previous condition even in the near term (Hobbs et al. 2013). Wohl et al. (2015) documented practitioner realization that the results of river restoration are typically unavoidable hybrids of humanly modified and unaffected conditions. While some field practitioners remain unconvinced about the need for any management change (Marris et al. 2013), the criticism of old paradigms has become more credible as more and more evidence of environmental and ecosystem change accumulates.
Ecosystem elements may be restored to a quantity and quality more like some previous conditions somewhere on the continent, but restoration of ecosystems to a more natural previous condition is becoming increasingly infeasible. With this observation in mind, the concept of ecosystem integrity needs to be elevated to a much larger geographical scale of consideration and based fundamentally on preserving and restoring the diversity of ecosystem elements. The most threatened of those elements are species. 6.2.8 A Landscape Approach Is an Essential Aspect of Ecosystem Management Paradigm: Management of ecosystem patterns in the regional landscape plays an essential role in the protection and establishment of desired species and materials support systems. 6.2.8.1 Before 1996 Leopold (1933) had prepared later eco-managers for accepting the wildlife management utility of landscape ecology and management at landscape scales. Many wildlife species addressed by Leopold and later wildlife ecologists are large and highly mobile animals that rely on certain landscape patterns for survival. Managers were well aware of these needs and planned accordingly. But it was many years before Noss (1983) called attention to the importance of including landscape perspectives in conservation planning and landscape ecology was formally recognized as a discrete discipline in the United States.
Responding to federal agency needs, Bourgeron and Jensen (1994) accepted Noss’s call for a broader ecosystem approach to conservation planning and the growing belief that stability increased with increased scale of ecological management from population to landscape and larger regional scale. They believed landscape boundaries should incorporate “the full regime of disturbances and processes” that occurred historically (such as fires and floods), but also recognized that extreme events can overwhelm landscape stability and management plans. They believed that the effects of global climate change could be destabilizing and recommended active manipulation of landscape patterns and process to stay within the range of historical variability. Their approach assumed that climate change effects on ecosystems and landscapes could be resisted through proper management and that historical variability could be simulated effectively.
A book edited by Naiman (1992) approached the concept of landscape and ecosystem management through the watershed and emphasized the need to assess and manage problems at much larger watershed scale. The ecological concepts described in the book linked terrestrial ecology with aquatic ecology more explicitly than the landscape approach. Several principles were developed that were commonly recast in subsequent concept development for ecosystem restoration. Consistent with hierarchy theory, management for watershed outputs needed to account for the effects of change at regional and global scales as well as local scales. The scale of the problems required much more information, better information
U.S. Army Corps of Engineers 109 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence management, and management decisions based more on data and less on “conceptual solutions”. Management also needed to more fully integrate satisfaction of human needs while sustaining ecosystem vitality and long-term resource supply. It required much more cooperation across the administrative boundaries of agencies, universities, industry, and other nongovernment organizations.
By the mid-1990s, leading eco-managers were familiar with basic concepts of landscape ecology and the utility for wildlife management in particular (Bolen and Robinson 1995, Scalet et al.1996). While they were well aware of the importance of increasing ecosystem edge to increase abundance of many game species, Scalet et al. (1996) also believed that a high edge to ecosystem ratio was a major source of adverse effects on numerous species vulnerable to extinction. They were particularly convinced that the concept of corridors had applications for both fisheries and wildlife management, even though demonstration of their importance was questioned by some basic ecologists (Simberloff et al. 1992). They were particularly concerned about the loss of wetland connectivity because of its possible role in the decline of amphibians. While these generalities were accepted, however, there was not much guidance about how to simultaneously manage landscapes for the protection of species at risk and provision of more game in response to demand.
Primack (1993) recognized the potential importance of landscape ecology in management for biodiversity, but devoted little of his book to explicit application of landscape ecology concepts. Some conservation biologists strongly advocated maximizing geospatial diversity to favor greater species diversity (Noss 1983), but Primack believed that not enough research had been completed to confirm the value of a landscape approach for conservation biology based on certain premises. He questioned for example, the idea that landscapes with high geospatial diversity will support greater biodiversity because they may be too patchy and edge dominated to favor many species at risk of extinction. Reflecting conventional thinking in conservation biology, Primack (1993) focused on the need to manage landscapes for the numerous imperiled species threatened by high amounts of fragmentation and ecosystem edge.
The authors of the NRC (1992) report on ecosystem restoration were aware of the importance of geographical scale considered in ecosystem restoration planning and the need to consider all aspects of restoring different elements of ecosystems at various scales. The report mentions the importance of addressing aquatic ecosystem restoration at watershed scale, but also mentioned the need to consider species needs on appropriate scales, such as the scale of habitat use by migratory birds. 6.2.8.2 After 1995 Recognition of the importance of management at landscape scale has continued to grow. Vogt et al. (1997) recognized the value of the landscape approach for terrestrial animal management and summation of different property uses and conditions. But they also emphasized how difficult it was placing boundaries on terrestrial ecosystems. They found watershed boundaries much easier to determine and useful for managing watershed process for desired habitat qualities in aquatic ecosystems. They also recognized that watershed boundaries are not very useful for terrestrial wildlife management.
Primack (2014), Cain et al. (2011) and Ricklefs and Relyea (2014) all agreed on the importance of the ecosystem-landscape approach to conservation biology when integrated with species- based approaches. They accepted much of the theory of reserve design based on landscape and ecosystem-level approaches to conservation that was just becoming established in the 1990s. The reserve-design paradigm that they accept has several elements based in landscape
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ecology. Ordinarily, for species-oriented conservation purposes, larger reserves are better than smaller reserves, one large reserve is better than several small ones, closely spaced reserves are better than reserves spaced far apart, reserves connected by corridors or “stepping stone” habitats are better than unconnected reserves, compact reserve shapes with low edge exposure are best, and any reserve is improved by establishing a buffer zone around it (Primack 2014). Fine-filter information frequently reveals important exceptions to these rules. The most important principle for reserve design is to closely examine the design closely for its fit to the needs of targeted species.
The corridor rule has been among the least settled. While protection of existing corridors was shown to preserve use by targeted species, corridor restoration often failed to be used by the intended species (Turner et al. 2001). Yet the corridor concept has become widely accepted for conservation purposes (Hilty et al. 2006). In more recent studies, corridors have been shown to facilitate movement of some species while other studies showed no clear effects (Cain et al. 2011). Some provide avenues for invasion by undesirable species (Resasco et al. 2014). The emerging management paradigm assumes corridor connectivity is carefully considered with respect to the needs of targeted species. In a meta-analysis of over 1000 species, Prugh et al. (2008) found that the distances between and sizes of habitat patches were generally poor predictors of species occupancy, but were better predictors where the matrix ecosystem conditions intolerable barriers to species dispersal.
7. Species Richness Regulation Paradigms
7.1 Science Paradigms Overarching Paradigm: The species richness of a local ecosystem is regulated largely by the species richness of and immigration from the ecoregional species pool and the local extinction of species resulting from insufficient habitat heterogeneity, enough habitat to support minimum viable population sizes, insufficiently moderate seasonal and other periodic disturbances, insufficient refuge from and regulation of predators (including herbivores) and competitors, and stochastic events.
The species richness of an area is determined by the balance between species immigration into the area from the surrounding ecoregion and species extinction within the area (e.g., Cain et al. 2011, Ricklefs and Relyea 2014). Thus the species richness of a local ecosystem is determined indirectly by the factors determining the species richness of the region as well as local variables. Numerous ecological processes influence that balance at ecoregional and at local ecosystem scales typical of the dimensions of many management areas. In many locations that balance has been tipped in favor of extinction by human actions. Recent recognition of rapidly changing global climate and associated acceleration of ecosystem change has elevated interest in how the regulation of species richness in an area may be better managed. Understanding of species-richness regulation is, perhaps, the most fundamental knowledge needed to generally guide restoration and maintenance of species numbers in rapidly changing ecosystems. This section demonstrates the complexity of that process and the challenges it implies for management.
The importance of scale, extending from the global arrangements of continents and oceans to populations in local ecosystems, is revealed in the hypotheses, theories, and paradigms pertaining to species-richness regulation. What happens globally affects regional species richness, which, in turn, affects local species richness and vice versa. For the practical aspects
U.S. Army Corps of Engineers 111 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence of species conservation, the most relevant interactions among species and their environment occur at local to regional scales. Coordinated management for sustaining biodiversity at all scales is counted on among conservation biologists.
Species diversity and species richness are often used interchangeably, but differ technically. Species richness is a simple indicator of species diversity. More complex indices include relative abundance as well as species richness (e.g., Pielou 1975). Species richness information was and continues to be much more widely available than abundance information and has been more useful for assessing the regulation of species diversity over a wide range of global, regional, and local conditions.
While species richness is easier to measure than other indicators of diversity, it is complicated by uncertainties associated with the definition and identification of species. The biological definition of species is grounded in evolutionary theory and based largely on biological isolation of species by anatomical, physiological, behavioral or genetic impediments to successful reproduction (Mayr 1970, Primack 1993). This definition is widely accepted among biologists, but is difficult to apply in a timely way. Most species taxonomy has been based instead on variation in morphological traits. The “morphological species” is a practical but imperfect indicator of the “biological species”. Morphological distinctiveness is statistical rather than absolute—variation in trait expressions can overlap—except when structural differences clearly preclude successful mating. Under close observation, some morphological species prove not to be reproductively isolated and hybridization is more common than once thought (Pennisi 2016). Taxonomies are often revised as more information is gained about the completeness of reproductive isolation. This source of uncertainty could contribute to the uncertainty associated with correlations based on geographical variation in species richness, but is generally assumed to be a minor source of error.
A potentially more problematic source of error is incomplete species classification. Most ecologists agree that a small fraction of all species have been described. Ecologists rely largely on taxonomic groups that have been most completely classified, such as birds and mammals, to identify species richness of an area. Relying on one or two taxonomic groups to indicate variation in the geographical richness of all species may be biased by unique adaptations of groups. Birds and mammals, in particular, may be biased by their endothermic adaptation to environments and comprise a small fraction of the total species richness. Nearly all other species are ectotherms. Whether the taxa are predominantly aquatic or terrestrial is important too because of the way environmental variables behave in and species adapt to these very different environments.
During the 1970s, the interest in diversity broadened beyond species richness to biological diversity at genetic, ecosystem, and landscape levels as well as the species level. The U. S. Congress Office of Technology Assessment (1987) defined biological diversity as “the variety and variability among living organisms and the ecological complexes in which they occur” (Scalet et al. 1996). Wilson (1988) is credited with integrating “biological diversity” into one word, “biodiversity”, which is now widely accepted. Despite biodiversity occurring at all levels of biological organization, the center of analytical attention over the years has remained focused largely on species populations. Biodiversity at larger levels than species is of interest to conservation biologists primarily because of its influence on biodiversity at the species population and genetics levels (Cain et al. 2011, Primack 2014, Ricklefs 2014). Genetic diversity within and among populations provide the biological information necessary to develop structure and function at higher levels of community, ecosystem and landscape organization. From a management standpoint, both traditional resource management and conservation biology have
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focused largely on individual species, but increasingly with an eye to the genetic basis of resource diversity and the interactions with other species and environments at larger scales. 7.1.1 Regional Species Richness Decreases as Latitude Increases Paradigm: Regional species richness, energy availability, production, and environmental extremes vary together along a latitudinal gradient from highest at the equator to lowest at the poles. 7.1.1.1 Before 1996 The generally accepted correlation of species richness with latitude is the springboard for other paradigms explaining the correlation. As naturalists began to more thoroughly explore the globe, they observed much higher species numbers in the tropics than elsewhere and least in the Polar Regions starting late in the 18th century. They included a young Charles Darwin as a ship’s naturalist. The experience peaked Darwin’s (1859) interest in explaining the origin of species and species richness. While the judgments of the earliest naturalists were qualitative, subsequent documentation of increasingly more quantitative species counts consistently indicated decreasing species richness with increasing latitude (Pianka 1967).
Ecological interest in what regulates species richness increased rapidly during the 1960s following publication of a paper called Homage to Santa Rosalia or why are there so many kinds of animals? (Hutchinson 1959). This work, like some other hypotheses and theories for explaining variation in species richness, was based largely on integrating observations of regional species richness variation from low to high latitudes in terrestrial environments (much less was then known about marine gradients) with other ecological knowledge. Ecologists have generated numerous specific hypotheses and theories for ecological regulation of species richness, much of it based on correlations between the global distributions of species numbers and the distributions of possible regulators. The hypotheses drew upon many of the paradigms of population, community, and ecosystem ecology.
The more commonly encountered of these hypotheses and theories are described below, starting with Hutchinson’s (1959) theory of energy limitation. His theory and several others are based on the numbers of species discovered along the latitude gradient and some understanding of ecological and evolutionary mechanisms. Ecologists during the 1990s did not accept any one hypothesis or theory as completely explanatory and believed that most of them applied in various circumstances (Colinvaux 1993, Brewer 1994). By that time, however, there had been some attempt at synthesis into a more coherent theory linking regional species richness to the determination of local species richness (Ricklefs 1993). 7.1.1.2 After 1995 With additional data to draw on, further research has generally confirmed that species richness generally increases from the poles to the tropics, among continents even at the same latitude, and among biomes in different locations. The effects of climate change on the latitudinal distribution of species richness have yet to be disclosed, but there appears to be no obvious reason to expect that the general relationship between latitude and species richness will collapse. Hillebrand (2004) documented consistent general patterns in terrestrial, freshwater and marine environments. A study by Jablonski et al. (2006) of marine fossils indicated that greater rates of species origin and emigration poleward contributed to explaining the species- richness gradient, but it did not identify causes.
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Cain et al. (2011), Primack (2014), and Ricklefs and Relyea (2014) voiced general agreement that the latitudinal gradient of species richness has multiple interacting causes that are not fully understood, but a few variables are more widely accepted than others. Mittelbach et al. (2007) categorized these causes as being based on one of three assumptions: 1) productivity is greater in the tropics, providing a higher total carrying capacity for more species; 2) diversification rates are similar globally, but evolutionary time under stable conditions has been greater in the tropics; and 3) rates of terrestrial species diversification are greater in the tropics because of larger land mass and thermal stability. Several of these and other hypotheses that are increasingly accepted as contributing to the determination of large scale regional species richness are described below. 7.1.2 Total Energy Availability Limits the Maximum Species Richness of Regions Paradigm: The maximum number of species niches and richness that have evolved in a large region (e.g., Asian arctic; South American tropics) is ultimately limited by the total energy available for species partitioning and the minimum size of species niches. 7.1.2.1 Before 1996 Hutchinson (1959) was interested in the factors that determine the maximum number of animal species that ecosystems could support. His perspective was largely global to regional climate zones (tropical, semi-tropical, temperate, boreal, polar) as he tried to explain why so many kinds of animals had evolved over time and why the number of species varied geographically among climate zones. Hutchinson proposed that, at the large regional scales he examined, enough evolutionary time had been available for species to fully diversify and fill the maximum number of possible niches in the region, resulting in regional species saturation at the maximum possible level. Based on that assumption, he also assumed regional patterns of animal species richness were useful indicators of what limited the evolutionary diversification of species. Hutchinson did not address regulation of species richness at local ecosystem scales, which could be substantially lower because of locally limiting factors (this is addressed in later paradigms).
Drawing on ecosystem energetics theory, Hutchinson believed that global variation in species richness could be explained largely by variation in the energy input into each region, which is highest in the tropics and lowest in Polar Regions. He believed that energy input was stable over long periods of evolutionary time and reasoned that only so much partitioning of energy into species production was possible before further evolutionary diversification reached a limit defined by the average energy needed to support each niche. By assuming that the ecological efficiency of consumers was constant across regions, as first proposed by Lindeman (1942), he dismissed it as a source of regional variation in species richness. In this context, regions are broadly defined by latitude (e.g., tropical, subtropical, temperate, subpolar, polar). Hutchinson concluded that energy alone could not fully explain what regulated species numbers in each local ecosystem, but it limited the animal species richness possible in any local ecosystem to the number that existed in the entire region.
Through correlation analysis, Currie (1991) examined 21 indicators of the different species- richness regulation hypotheses developed and found that an evapotranspiration index explained most of the regional variation in the species richness of mammals, birds, reptiles, amphibians, and trees. Currie concluded that evapotranspiration supported the energy input hypothesis of Hutchinson based on the response of evaporation and transpiration to energy input and
U.S. Army Corps of Engineers 114 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence associated temperature. Evapotranspiration is also highly correlated with solar radiation and latitude.
Hutchinson’s (1959) theory explains a lot that has been claimed under the hypothesis that differences in primary production largely determine variation in species richness. Solar energy drives primary production and chemical energy used by bacteria adds a small fraction to the total. On a global scale, solar energy input, primary production and species richness all decrease along a latitudinal gradient from equator to poles. However, Currie (1991) found substantially less correlation of vertebrate species richness with primary production than with the evapotranspiration index, suggesting that temperature, moisture, or other factors correlated with evapotranspiration and energy input, are also contributing separately to regional limitations. Cooler temperatures, for example, may decrease the efficiency of energy transfers in community trophic dynamics. The possible effects of climate change in modifying thermal limits on species diversification was not considered, although it had been implicated with certain mass extinctions.
Water availability may also play a role in terrestrial environments. Whitaker (1977) observed a change in species richness along an Arizona mountain slope that appeared to be related both to temperature and precipitation. Species richness was lowest at the exceedingly dry lower elevations, then increased to peak species richness with increasing precipitation and more moderate temperatures, followed by a decrease at the highest elevations as precipitation continued to increase and temperatures cooled dramatically. Temperature and precipitation may have separate influences, but combinations of temperature and precipitation appeared to contribute significantly to regional differences in terrestrial primary productivity.
Connell and Orias (1964) generally accepted Hutchinson’s theory that energy input most fundamentally limits regional species numbers along a latitudinal gradient, but also proposed that increased environmental instability associated with increasing latitude contributed to the decrease in total energy available for species productivity. They reasoned that environmental instability originating largely from seasonal variation in solar radiation required species to divert more energy to respiration, thereby reducing ecological efficiency and limiting the number of niches possible. Colinvaux and Barnett (1979) added evidence for this hypothesis when he discovered a much lower ecological efficiency in an ecosystem dominated by endotherms, which generally increase in relative importance in the terrestrial ecosystems of colder climates.
Based on the work of MacArthur (1955), Connell and Orias also hypothesized that community stability increased as increased species numbers led to greater food-web complexity. They believed that community stability decreased with increasing latitude, which further depressed ecological efficiency and energy available for diversification. The fossil record, however, is not entirely consistent with this hypothesis. It reveals a general trend upward in the number of families (e.g. Wilson 1989). However, diversification at that taxonomic level is not necessarily indicative of diversity at the species level.
By the 1990s, increased productivity (and by inference, energy input) was generally accepted as a major factor contributing to increased regional species richness along the latitudinal gradient, but not exclusively so (Colinvaux 1993), Ricklefs 1993, Brewer 1994). Even if it were influential regionally, variation in production does not explain much about the variation in species richness at local scales. Researchers had identified numerous examples of local ecosystems with high productivity and low species richness and vice versa. The efficiency of energy conversion by primary producers varies significantly because of local nutrient and water limitations, and it may influence the maximum possible species richness of local ecosystems. Connell’s and Orias’s (1964) hypothesized contribution of environmental instability to observed differences in species
U.S. Army Corps of Engineers 115 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence richness may be indirectly supported by acceptance of real variation among ecological efficiencies within ecosystems, including variation in the proportion of exotherms in the species pool, by Colinvaux , Brewer, and Ricklefs. 7.1.2.2 After 1995 For decades, acceptance of the energy-flow paradigm of ecosystem ecology supported the belief that maximum regional species richness is limited by the amount of energy available for production. While the idea is theoretically logical, it may be impossible to test empirically. Whether regional species richness is at a maximum is testable. Ricklefs (2008) believed close examination of the fossil record of diverse groups consistently indicates that taxonomic diversity is generally stable for millions of years (when extinction about equals species origin) consistent with Hutchinson’s hypothesis. But others believed that other variables may contribute to determining the maximum possible regional species richness. For example, Allen et al. (2002) made a case for temperature influencing regeneration and mutation rates, which in turn influence speciation rates, as previously proposed by Rohde (1992). Then Currie et al. (2004) found a greater correlation of energy input with species richness than with any other variable, suggesting that energy is the dominant factor limiting the existing regional species richness.
In describing latitudinal gradients, Ricklefs and Relyea (2014) accepted the attractiveness of the energy limitation hypothesis, based on its theoretical logic, but were less convinced that other factors did not reduce regional species richness to numbers below the maximum. They were aware that numerous ecologists believed other variables contributed to determining regional equilibriums in species richness resulting from species evolution and extinction. Cain et al. (2011) and Primack (2014) did not comment on the maximum possible species richness, but believed closely related productivity is one of several contributors to latitudinal variation.
In theory, global climate change has little effect on the solar and chemical energy entering the Earth’s ecosphere, but it could have some influence on the quality and quantity of solar energy reaching the Earth’s surface as a consequence of atmospheric changes. 7.1.3 Species Richness Increases as Competitive Exclusion Is Reduced Paradigm. Species niches and species richness increase as competitive exclusion is reduced by predation and other sources of mortality. 7.1.3.1 Before 1996 Hutchinson (1959) proposed that competitive exclusion was a major impediment to species niche diversification that ultimately reaches a regional maximum determined by energy availability. He first proposed that predation played a key role in reducing competitive exclusion and then added abiotic disturbance two years later (Hutchinson 1961). Hutchinson reasoned that new resources were made available at higher levels as new niches were established in the lower trophic levels, allowing further “niche packing” until energy limited any further niche diversification. In his thinking, the cumulative effects of predation and disturbances had occurred over long periods of evolutionary time to establish saturated regional species richness. Later ecologists also proposed that predation and disturbances of a historical type played a role in reestablishing and sustaining species niche occupancy in local ecosystems by reducing the advantages of superior competitors.
Connell and Orias (1964) accepted Hutchinson’s basic energy limitation hypothesis, but placed more emphasis on spatial separation of subpopulations as a determinant of new species origin.
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They theorized that greater energy input resulted in greater productivity, which allowed species populations in an area to concentrate, increasing the likelihood of spatial isolation and genetic differentiation. This hypothesis was consistent with Mayr’s (1970) theory of species origin, which was widely accepted at the time. There is nothing inconsistent with reduction of competitive exclusion, however, which could also contribute to population isolation and differentiation. Isolation alone did not explain how most species became more specialized, which Connell and Orias accepted as the typical result of increased species richness.
Empirical research during the 1960s supported Hutchinson’s competitive-exclusion hypothesis Paine (1966) clearly demonstrated that predation is a major player in the determination of local species diversity. He experimentally removed predatory starfish from intertidal communities and observed a reduction in the number of prey species. Evaluating the role of disturbance, Connell (1978) examined data from the tropics and hypothesized that an intermediate level of disturbance enhanced diversification while extremes reduced it. Experimental studies subsequently supported Connell’s hypothesis (Sousa 1979).
Hutchinson (1959) also proposed that top-level predation capped the potential number of consumer trophic levels based on energy availability alone. It was assumed that evolutionary selection favors the most efficient top carnivores which may be less vulnerable to predation than other levels (being in the top position). Consequently, they are more likely to consume too much of the underlying trophic level, causing them to shift to lower trophic levels or cannibalism and limiting the total number of trophic levels. In sum, Hutchinson anticipated the cascade theory that top-down forces operating through predation complement the effects of bottom up energy flow to determine the total number of species.
By the 1990s, ecologists generally accepted the proposition that species must specialize to avoid extinction from competitive exclusion (Colinvaux 1993, Ricklefs 1993, Brewer 1994). They also accepted predation as an important agent for reducing competitive exclusion, but not the only one. By the 1990s, it was well recognized that many forms of disturbance and predation could influence species coactions and competitive exclusion (Colinvaux 1993, Brewer 1994). Because keystone and dominant species have disproportionate community effects, their demise can significantly disturb the community, reducing species richness. However, Hutchinson’s (1959) theory that competition was not important in structuring most communities was not backed up by empirical studies (Shoener 1983, Connell 1983). Some niche overlap was apparently common in diverse settings, indicating that competitive exclusion was operative to some degree. By the 1990s, competition was generally accepted as a force in shaping community structure and species diversification along with predation, other biotic interactions and abiotic variables (Colinvaux 1993, Ricklefs 1993, Brewer 1994). 7.1.3.2 After 1995 The competitive exclusion principle has held up well and evidence that predation and disturbances interfere with competitive superiority is now generally accepted (Cain et al. 2011, Ricklefs and Relyea 2014), reinforcing the idea that predation and environmental disturbances reduce competitive exclusion, which allows more species to occupy a region until a maximum number is limited by energy availability. Cain et al. (2011) also mention speciation as the ultimate mechanism for reducing competitive exclusion, but did not explicitly discuss the effects of predation and disturbances on speciation at the regional scale.
Humans play a complex role in the processes of resource partitioning and its regulation by predation. Humans are omnivores that are competitively superior to other animal species in all consumer levels of many food-webs, including the top carnivore level, and also compete
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indirectly with plants through management of plant species compositions. Humanity profoundly limits the energy that flows from net primary production through other consumers by diverting production to its own needs and to microbial decomposition (largely in sewage treatment plants in North America). More recent estimates range between 20 and 32 percent (Rojstaczer et al. 2001, Imhoff et al. 2004, Haberl et al. 2007). The effect is primarily terrestrial. Assuming this paradigm and the estimates of effects are generally accurate, humans could be limiting the maximum possible diversity of terrestrial plants and animals proportionally worldwide and even more so in those regions where the appropriation of net primary production is greatest. The paradigm suggests that significantly greater plant and animal species extinction may be only a matter of time in areas like temperate grasslands, where the effects of agriculture are concentrated. 7.1.4 Evolutionary Time May Contribute to Limiting Regional Species Richness Paradigm: The realized number of species niches and species richness in a region may be determined in part by the degree to which environmental instability limits the availability of evolutionary time for speciation. 7.1.4.1 Before 1996 The evolutionary time hypothesis is based in the assumption that environmental stability favors evolutionary diversification and decreases with distance from the tropics. Over a century ago, Wallace (1878) first proposed that limited amounts of time available for evolution was a major determinant of the latitudinal gradient of species richness. Based on the evidence at the time, he proposed that the source of instability disrupting evolutionary time was decreasing climatic stability at higher latitudes. Yet the fossil record revealed that mass extinctions of marine organisms were particularly large in the tropics (Jablonski 1986). More recent evidence indicates that significant environmental changes also occurred in the tropics during periods of past glaciations, making greater long-term stability in the tropics a less convincing argument for the high species richness observed there (Colinvaux 1993). However, there is no evidence that the tropics were exposed to the same instability extremes as areas closer to the poles, such as extended exposure to freezing temperatures.
Rohde (1992) believed that the ecological evidence indicated there were more “empty niches” outside the tropics. He accepted the high correlation of species diversity with energy supply along latitudinal and altitudinal gradients, but believed there was little evidence to support the assumption that species richness was saturated along latitudinal gradients as Hutchinson (1959) and others had proposed. Rhode (1992) proposed instead that a better explanation for the relationship between energy input, latitude, and species richness could be found in different evolutionary rates. He proposed that “greater species diversity is due to greater ‘effective’ evolutionary time (evolutionary speed) in the tropics, probably as the result of shorter generation times, faster mutation rates, and faster selection at greater temperatures.” He admitted an “urgent need” for more research, however.
In the 1990s, ecologists generally believed that natural immigration from other continents, resulting to some extent from climate instability, was a significant source of continental species richness and that secondary evolution of new species from the founding species further diversified into new species, filled most of the possible niches, and determined continental differences in species composition (Colinvaux 1993, Ricklefs 1993). This was most evident in the connection of the Americas to Eurasia as a consequence of lowered sea levels during periods of Pleistocene glaciation. Some ecologists minimized the importance of environmental
U.S. Army Corps of Engineers 118 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence instability as a regulator of differences in contemporary regional species richness, however. Ricklefs (1993) dismissed differences in regional environmental stability during the Pleistocene because there was evidence that all terrestrial regions underwent substantial disturbance and were well connected, allowing relatively rapid equilibration of invading species with the existing species composition of each region. In contrast, Brewer (1994) thought climatic instability may have contributed to reduction of species richness below saturation numbers in terrestrial regions outside the tropics.
While often related to environmental instability, environmental extremes may resist speciation to some extent independently of instability. The degree to which environments vary from those most hospitable for life process in general has been suggested to be a primary cause for the variation in species richness, the most benign environments having the greatest number of species (Paulson and Culver 1969). Environments with stable but extreme temperature, acidity, oxygen concentration, aridity, and other variables generally have low species richness (but they also tend to be small and may be limited by spatial factors described later). Decreases in species richness and productivity with increased latitude and elevation could be a response to more extreme temperatures. The metabolism of primary producers and ectothermic animals is definitely a function of temperature and the species richness of ectotherms also decreases as latitudes approach the poles. As temperature decreases, the efficiency of energy conversion decreases even among endothermic animals (colder temperatures increase metabolic costs).
However, an important exception to the proposed effect of extremes is found in marine benthic environments where temperature decreases with greater depth as species richness increases. Therefore one or more factors other than environmental instability and extremes appear to be contributing to the high diversity of marine benthos in perpetually dark, cold, and unproductive waters. Sanders and Hessler (1969) had hypothesized that the long-term environmental stability of deep ocean bottom communities explained the high species richness found there. Both temperature and energy input in the form of organic sediments are relatively low, but the environment is believed to be among the most stable on earth. Sanders and Hessler (1969) suggested that long-term environmental stability allowed greater diversity in deep environments than in coastal waters where there was higher productivity, but greater exposure to climate variation. However, consistent with Rhode’s (1992) thinking, Colinvaux (1993) considered the generally high species richness of marine environments more a consequence of the huge expanse of the marine ecosystem than its stability over evolutionary time (more is said about this later). While Colinvaux (1993) and Brewer (1994) mentioned environmental extremes as a hypothesized factor, Ricklefs (1993) did not. 7.1.4.2 After 1995 Taylor (2004) supported Jablonsky’s contention that extinction rates of marine molluscs were greatest in the tropics. Ricklef’s (2008) study of the fossil record generally revealed a relatively stable past extinction and speciation rate punctuated by five exceptional mass extinctions over the past 450 million years (Raup and Sepkoski 1982). The last of these mass extinctions occurred about 66 million years ago. The fossil record also indicated that the recovery to a generally stable level of diversity took 10 to 20 million years following a mass extinction.
There is no indication that a mass extinction had already occurred in association with Pleistocene glaciation, but a more selective extinction of large mammals and birds, starting about 30,000 years ago, may have been the beginning of a sixth mass extinction, now in progress (Primack 2014). The ecological impacts of these extinctions are likely to have been substantial because of the outsized role that large species tend to play in their communities. As a consequence of those effects, there may have been extinctions of other species that do not
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show up well in the fossil record. The extinction of large animals in North America during the late Pleistocene is increasingly linked to human impacts interacting with abrupt climate change about 15,000 to 11,000 years ago (Levy 2011, Roberts 2014, Cooper et al. 2015). Present rates of climate change now appear to be even more abrupt and have greatly raised the level of concern over a sixth mass extinction. Humans are also implicated with earlier Australian extinctions (Rule 2012)
Ricklefs and Relyea (2014) mentioned that there continues to be a school of thought that proposes the instability of global cooling and extended glaciations caused extinctions differentially along the latitude gradient. In general, Cain et al (2011), Ricklefs and Relyea (2014), and Primack (2014) accepted latitudinal variation in environmental stability and evolutionary time as one of the hypotheses that most likely contributes to the global distribution of species richness. The issue has significant implications for the effects of global climate change on species richness and how they may be distributed along a latitudinal gradient. Climate data indicate that warming is greater at the polar end of the gradient than in the tropics (IPCC 2007), suggesting a disproportionate effect there. However many of the terrestrial species located there tolerate significantly more climate variation than species in the tropics, possibly making many terrestrial tropical species more vulnerable to the effects of global climate change (Perez et al. 2016). 7.1.5 Area Size and Isolation Contribute Largely to Species Richness Regulation Paradigm: Species richness is regulated largely by areal size and, locally, by isolation, which determine the balance between the rates of immigration into local ecosystems and local extinction. 7.1.5.1 Before 1996 A very early advance in understanding variation in species richness was the development of a quantified approach to comparing species numbers in different areas directly using the slope of the relationship between the number of species and the geographical area sampled (Arrhenius 1921, Wilson 1943). Because different researchers typically sampled areas of different size, it was not possible to compare the species richness of areas directly without some sampling bias entering into the comparison. The species-area curves usually indicated an increase in species richness at a decreasing rate as more area of the region was sampled. As more data were gathered, it became clear that the slope of the relationship determined for different areas varied widely, significantly reflecting differences in species richness (Rosenzweig 1995). The relationships graphically demonstrate that the species richness of a region is always equal to or more than the richness in any part of it.
Cody (1975) examined bird species-area curves along similar latitudes in the temperate zones of three different continents (to control for the latitude effect). The surveyed area ranged from a single point survey to an area approaching a million square miles. He found that, while the composition differed, the species richness increased with area sampled and was very similar on each of the three continents. Species richness differed slightly more as geographical scale increased, but nowhere was the difference large, suggesting that regulating processes were operating in the same general way at the continental scale. This result also supported the Hutchinson theory that most regions have reached a similar equilibrium number of species at continental scale where total energy input is about the same within a limited latitude range.
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Species-area relationship models were well accepted by ecologists in the 1990s. Colinvaux (1993) believed that the high species richness of marine benthos observed by Sanders and Hessler (1969) was more likely due to the large area over which the total number of species was estimated than to environmental stability over evolutionary time. Grassle and Maciolek (1992) had estimated that the number of benthic species in the oceans was between 1 and 10 million based on extrapolations of a species-area curve to full ocean size. This estimate rivaled a contemporary median estimate of 6.1 arthropod species, which dominate the richness of tropical rainforests (Hamilton et al. 2013). Colinvaux’s conclusion seemed reasonable, considering the wide range in marine species richness estimates and the much larger size oceans compared to the terrestrial tropics.
MacArthur and Wilson (1967) enlarged upon the concept of species-area with their equilibrium theory of island biogeography, including an empirical test (Simberloff and Wilson 1969). It would prove to become a fundamental paradigm for explaining local differences in species richness. They assumed, as Hutchinson did, that the species richness on continental land masses is largely saturated and at equilibrium with the total energy resources available, sources of Figure 19. Island biogeography theory relates the size of islands and distances of islands from mainland to the mortality that reduced competitive exclusion, balance between immigration and extinction that and evolutionary process. But they were more determines the species richness on the islands. interested in the processes that determined the equilibrium number of species in local THEORY OF ISLAND BIOGEOGRAPHY areas and tested their concepts using species Small, near or richness data from oceanic islands. They large, far island proposed that successful immigration of new species on islands is a decreasing function of immigration extinction the number of species already present on the islands. They also proposed that extinction is Rate near small an increasing function of the number of far species on the island (this could result from large competition for limited resources). An equilibrium number of species occurs where large, close island the two functions intersect (Figure 19). The small, far island distance of the islands from regional land masses and the size of the islands contribute to determining the equilibrium number. The s1 s2 s3 extinction rate decreases with increased Number of species island size because a larger island size can support larger species populations and more numerous populations per species.
The distance to the island determines the probability of immigrant survival as it moves through inhospitable habitat. The theory was influenced by the general concept of a population dispersal survival curves described by Kettle (1951). Many of the first species to reach islands, and repeatedly reach them thereafter, are r-selected species (Diamond 1975). The size of the island influences the probability that a dispersing organism will intercept the island. The equilibrium number also varies among species groups, depending on their mobility and tolerance for conditions in the isolated environment. Birds are much more successful colonizers of remote ocean islands than amphibians or freshwater fish and invertebrates. In addition to immigration, evolutionary diversification also contributes to species enrichment on islands. The composition of terrestrial and freshwater species becomes increasingly different from mainland species with distance from the mainland, indicating that evolution on the most remote islands is the major source of species diversification and richness.
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MacArthur and Wilson (1967) recognized that isolated ecosystems occur in many forms other than ocean islands. The equilibrium between immigration and extinction could also explain changes that take place in a disturbed patch during community succession (Colinvaux 1993). Many freshwater ecosystems are isolated by land or by land and ocean in the case of rivers that run to the sea. Patches of terrestrial ecosystems also may vary enough from their surroundings to be likened to islands—the most obvious example being mountain-top “sky islands”. Their equilibrium number of species was also expected to be limited by the same variables as species on ocean islands. Colinvaux (1993), for example, believed that small ecosystem size explained a lot of the low species richness observed in extreme environments like thermal springs.
Terbough (1973) proposed that diversity in a particular ecosystem condition is regulated in large part by speciation at the regional “core” and dispersal into peripheral ecosystems where different environments selected for more species. He also proposed that the total number of species in a terrestrial region generally defined by latitude, such as the tropics, is largely a function of the land area, consistent with species-area observations. Rosenzwieg (1992) proposed that the tropics were as diverse as they are because they include the largest land areas among the major zones defined by latitude of similar range and they are thermally homogenous. He believed that homogenous temperatures and large land area allowed larger numbers of individuals in each species to survive, thereby resisting extinction by spreading risk over larger area. He also suggested that the two factors favored larger species ranges and the probability that some parts may become reproductively isolated allowing for speciation to occur. That argument could also apply to the high species richness found in deep marine benthic environments (Sanders and Hessler 1969).
Food chain length is a crude indicator of species richness and Shoener (1989) found that it increased with increased size of water bodies (aquatic islands). He showed that the rate of primary productivity alone could not explain this relationship. By revisiting elements of Hutchinson’s theory and the theory of island biogeography, Shoener proposed the “productive space hypothesis” to explain how ecosystem size helps to determine food chain length (and species richness by implication). Productive space is the product of ecosystem area and primary production rate per unit area. As the productive space increases, the species richness and food chain length increase.
The productive space hypothesis integrated the energy and productivity limitation theories with the theory of island biogeography. As islands selected for study become smaller, the productive space decreases, prey productivity becomes incapable of supporting viable predator populations, and the number of species decreases. The total energy available to support full- time resident niches approaches zero on very small islands and no populations can persist if they depend exclusively on the islands’ productivities. Many small oceanic islands are used only seasonally or less frequently by vertebrate species that depend on the sea for sustenance. This concept applies broadly to any island ecosystem including lakes, rivers, and terrestrial ecosystems with clearly defined ecological boundaries.
By the 1990s, Ricklefs (1993) and Brewer (1994) reflected the general acceptance of the equilibrium theory of island biogeography and the importance of land size and isolation as factors in determining the species richness of different ecosystems. However, Colinvaux (1993) noted that while there have been numerous tests of the equilibrium theory that appear to support it for some groups of organisms, plants returning to Krakatoa Island following a major volcanic eruption in 1880 did not follow the expected pattern. This led Colinvaux (1993) to question the completeness of the equilibrium theory in all situations, even though he recognized the importance of island size in determining the number of species. He thought that other, unstated properties of island size may play a role.
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7.1.5.2 After 1995 More recent ecologists generally accept the equilibrium theory of island biogeography as evidence for its validity has continued to accumulate (Cain et al 2011, Ricklefs and Relyea 2014). Cain et al. (2011) believed that the primary explanation for variation in species richness was the size and isolation of the area while other factors may also play some role. Ecologists also widely accept the applicability of the equilibrium theory to any geographical area that differs from surrounding areas. Based on observed relationships between species richness and the size and isolation of forest fragments following human impact, Cain et al. (2011) concluded that the species-area and equilibrium theory of island biogeography are basic paradigms that are “timely and relevant to the issues of conservation that we deal with today.”
However, Cain et al (2011) also believed, after reviewing empirical studies of terrestrial ecosystems defined by vegetation types, that the porosity of matrix ecosystems separating patches of other terrestrial ecosystems varies for a species as vegetation type varies and may be much greater than that of highly isolated land islands and lakes. Therefore, the relationship between distance across the matrix ecosystem and species success crossing the matrix ecosystem is more complex than the models developed from island biogeography theory. Highest permeability exists when habitat differences meaningful to the species are small in the matrix or a sufficiently wide enough corridor of connecting habitat exists between the patches. Models can adjust for the differences in permeability with the appropriate information. Global climate change will most likely affect model variables, but is not likely to affect acceptance of the paradigm. Based on correlative and experimental studies, Ricklefs and Relyea (2014) accepted the general premise of the equilibrium theory of island biogeography for any type of island-like habitat condition. 7.1.6 Habitat Heterogeneity and Moderate Disturbance Increase Species Richness Paradigm: Habitat heterogeneity and moderate disturbance increase the number of species that can be supported by an area. 7.1.6.1 Before 1996 Additional hypotheses have been proposed to “fine-tune” the species equilibrium model of island biogeography at local geographic scales. Up to the limits generally imposed by the equilibrium theory of island biogeography, the habitat heterogeneity hypothesis postulates an increasing number of niches with increasing habitat heterogeneity as a consequence of spatial partitioning of resources (Ricklefs 1993, Colinvaux 1993, Brewer 1994). An island of uniformly flat topography would be expected to support fewer species than a topographically diverse island of similar size and remoteness. The moderate disturbance hypothesis postulates an increased number of niches in moderately disturbed ecosystems (Connell 1980). Moderate disturbance may interact with homogenous environments to develop more spatial heterogeneity or, if regular enough, such as diurnal and seasonal changes, competition for the same resources may be reduced through temporal partitioning of the resources. The diversity of conditions that result from spatial and temporal variation creates more local conditions like those that occur within the climatic region that is the source of immigrants into the local area.
Spatial partitioning of resources is well documented where habitat structure is diverse. Brewer (1994) referred to this as resource diversity. Pianka (1966) reasoned that more spatially complex habitats should support more species because they can inhabit different parts of the “environmental matrix”. Pianka identified differences in the effects of macro- and micro-spatial
U.S. Army Corps of Engineers 123 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence heterogeneity. Macro-heterogeneity is associated with topographic diversity, which interacts with climate to create quite different abiotic conditions regionally. Micro-habitat diversity is more localized, such as the structural diversity that results from plant species stratification at different elevations above the ground. The positive effect of habitat diversification on species richness can only go so far before small habitat sizes limit establishment of viable populations.
The effect of spatial heterogeneity is often witnessed in bumpy species-area curves (Rosensweig 1995). As a species-area transect crosses topographically diverse areas, the rate that new species are encountered bumps up. At the micro-heterogeneity level, the species richness of birds is at least partly correlated with increased number of discrete strata in vegetation (Cody 1975). Pianka (1966) concluded that lizards in the southwestern United States have specific micro-habitat preferences that influence the total species richness. Gorman and Karr (1978) found a correlation between fish species richness and habitat structural diversity in streams. Huston (1979) highlighted the role of low to intermediate disturbance in reducing competitive displacement of species and increasing species diversity.
Sale (1977) recognized the importance of spatial heterogeneity in determining the species richness of coral reefs, but also emphasized the role chance plays in species establishment in the diverse microhabitats. He proposed that suitable locations for the survival of many different coral species within the spatially diverse features of coral reefs depended largely on where the larvae of the species were carried fortuitously by currents; the first to occupy a site had an advantage. Since the early life stages of many coral reef species are dispersed passively by water currents and have an equal chance of colonizing a particular site, his “lottery model” could contribute to the maintenance of species that are not at a competitive advantage. By the 1990s, however, the role of random chance had not been widely recognized as an important contributor to maintenance of species richness, but it could play an important role in determining the species composition of an area following a major disturbance or area restoration activities.
Moderate flooding, fire, storms, disease, and other disturbances have been hypothesized to increase species richness locally at intermediate intensities and frequencies (Connell 1980). Ricklefs (1993) and Brewer (1994) cited intermediate disturbance as a separate factor regulating species richness at local scales. Colinvaux (1993) implied its contribution by including weather and other disturbances with predation as a source of differential mortality that reduces competitive exclusion.
Temporal partitioning of resources in the same location is enabled by generally predictable change in environmental conditions. For example, Lawler and Morin (1993) found that frog egg laying at different times in the spring reduced tadpole competition for the same types of food. Kronfeld-Schor and Dayan (1999) described how different adaptations for diurnal and nocturnal feeding avoided competitive exclusion between two mouse species preying on the same insect species. 7.1.6.2 After 1995 Contemporary ecologists continue to accept the importance of habitat heterogeneity and moderate disturbances among the local conditions shaping ecosystem species richness. Based on much empirical evidence, both Cain et al. (2011) and Ricklefs and Relyea (2014) accepted the role of habitat complexity in species resource partitioning and emphasized the correlations between habitat diversity and plant and animal diversity. They also accepted the importance of moderate disturbances in promoting species coexistence and diversity by mediating food and other resource availability. Groves et al. (2012) promoted the importance of geological and topographic heterogeneity in selecting future preserves for biodiversity protection in rapidly
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changing environments. The importance of topography in maintaining climate change refugia also pertains (Morelli et al. 2016).
7.2 Management Paradigms Overarching Paradigm: The many variables that filter and regulate the total species richness and composition of communities are difficult to predict, but can be adaptively managed when projects are planned and implemented at well-integrated spatial and temporal scales for clearly specified objectives.
In the early days of fish and wildlife management for recreational and commercial use, management for species richness was of little concern. Managers were intently focused on providing for the needs of specific resource species, often at the expense of other members of the managed community. That approach to management changed quickly after the ESA was passed, at least in policy if not immediately in practice. As federal and state agencies incorporated endangered species programs, state game and fish agencies widened their management perspectives and often created their own lists of state protected species. Conservation biologists became more interested in understanding and working with the factors that regulate the species richness of conserved areas as it became clearer that species richness often contributed to the relative stability of community composition (Primack 2014). A few important management paradigms began to emerge and continue to be refined. 7.2.1 Species Richness Management Requires an Integrated Approach Paradigm: Specie-richness management requires an integrated approach to managing variables that influence species richness regulation at regional to local scales. 7.2.1.1 Before 1996 Eco-managers began to seek useful information about the processes that regulate species- richness as they assumed a more holistic approach to managing for resource use compatible with sustaining native biodiversity. Most of the existing science was not motivated by the practical needs of management. They found many hypotheses and little coherent theory. Most ecologists of the 1990s had not attempted to integrate the hypothetical processes of species- richness regulation into theory of potential use to eco-managers. While Colinvaux (1993) and Brewer (1994), for example, believed that more than one hypothesis was likely to apply, they did not attempt to integrate them into coherent theory.
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Ricklefs (1993) was an exception. He integrated some of the most accepted concepts of species richness regulation into a conceptual model of species richness flow from large regions to local ecosystems within the regions (Figure 20). He assumed that at the regional scale, the main source of species richness is organic evolution (species origination). Species immigration from other regions also plays a secondary but increasingly Figure 20. Integrated theory of species richness regulation in continental and oceanic ecosystems (modified from Ricklefs 1993). important role, as a consequence of human Stochastic activities. Using island Regional Disturbance biogeography theory, he Immigration (Extinction) Competitive proposed that the regional Exclusion species richness is determined Regional Local by the balance between Local Species Species Species Habitat regional species extinction and Immigration Origination Richness Richness Diversity the sum of evolutionary and immigrant species origins. The Natural regional equilibrium condition Selection Extinction Predator Exclusion & forms a species “pool” from Poor Habitat Quality which species disperse and (Extinction) immigrate into local environments. Because dispersing species are ecologically filtered (blocked or die) from the number of successful immigrants, the maximum species richness in local areas is usually less and never more than the richness of the regional species pool. Once in a local area, the number of species equilibrates as immigration is balanced by local extinction caused by unsuitable habitat qualities, excessive predation, and destructive stochastic events (e.g., extreme floods, fires, and droughts). But local extinction is moderated by habitat diversity and probably by moderate disturbances, although he did not identify disturbance as a separate effect.
Ricklef’s model does not distinguish between native and non-native species richness or human and non-human impacts. Human-influenced immigration of non-native species could increase pool sizes while anthropogenic environmental change could contribute to reductions of regional and local species pools. The key to restoring and protecting targeted species is appropriate management of both the regional and the local ecological filters. At regional scale, eco- managers are typically more interested in controlling introductions of exceptionally destabilizing species and managing ecological filtering for the desired immigration flow and conditions in the managed area than in managing for long-term species origination in the region. Yet coordination of the collective ecosystem management can influence future rates of species origin as well as extinction. 7.2.1.2 After 1995 Post (2002) developed a theoretical conceptual model of species richness regulation (indicated by food-chain length), which included most of the elements in the Ricklefs model with additional considerations. Like Ricklefs, he assumed that the regional species-richness pool was a reliable resource for species immigration into local ecosystems and emphasized the importance of island biogeography, evolutionary time hypotheses, and regional barriers (ecological filters) to local ecosystem immigration. Within the local ecosystem, Post also identified extreme disturbances as a cause for local extinction and specified a number of conditions that fit into the more general category of unsuitable habitat in the Ricklefs model. These conditions included evolutionarily young areas created by lava flows, retreating glaciers, and areas limited by low energy input, such as nutrient-limited plant production and depressed detritus input from outside the ecosystem. He included habitat space as a potentially limiting resource combining it with
U.S. Army Corps of Engineers 126 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence productivity based on the productive-space concept of Vander Zanden et al. (1999). He also recognized the importance of predators in controlling food chain lengths and hypothesized that variation in predator and prey sizes influenced the food chain length. Neither Ricklefs nor Post identified the stabilizing effect of increased species richness as a variable that could result in even greater species richness
Cain et al. (2011) and Ricklefs and Relyea (2014) basically accepted many of the elements of the Ricklefs and Post models for explaining variation in species richness at the local community- ecosystem level without specifically mentioning the earlier models. Reflecting development of conceptual models for ecological filtration (e.g., Belyea 2004), they also highlighted the importance of moderate disturbance and predation in reducing competitive exclusion within the local community-ecosystem. Certain interactions can facilitate survival of a species exposed to intermediate levels of stress indirectly through food web and other interactions that moderate the effects of competition (e.g., Hacker and Gains 1997). Cain et al (2011) also admitted that the role of chance in determining species persistence in diverse communities “has intuitive appeal”, but stopped short of full acceptance.
In sum, eco-managers generally recognize the importance of an integrated approach to species richness management while recognizing that the existing theories and models for species richness regulation are works in progress, which should be applied cautiously and adaptively. Many of the variables that influence the species richness and composition of a protected or restored area are to some degree manageable. While Ricklefs and Post demonstrated that general theories of species-richness regulation could be developed from the integration of various hypotheses, no theory is now widely accepted as totally complete. The hypotheses and theories are useful for eco-management conceptualization, including restoration actions, but the unique specifics of most management areas need to be defined and considered as well. 7.2.2 Managing for Species Composition Requires a Species Approach Paradigm: When planning and managing for a specific species composition, eco-managers need to assess and manage risks faced by the desired species. 7.2.2.1 Before 1996 The ecologists who proposed species richness paradigms did not claim that the equilibrium number of species established in an area would be continuously composed of the same species (Hutchinson 1959, McArthur and Wilson 1967, Ricklefs 1993). Changes in species composition became increasingly evident as more and more exotic species from other continents took their place in continental ecoregions and as more nonnative species from distant locations in the region became established in local ecosystems (Elton 1958, U. S. Congress Office of Technology Assessment 1993). The numerous introductions of highly invasive exotic species were among the critical complications for managers intent on the sustainability of native species (Primack 1993). During the equilibration of species numbers between immigration and extinction, populations of the rarest species were often observed to be most vulnerable to extinction.
Highly invasive species are typically broadly adapted, r-selected species. They are frequently among the first to reach and become established in areas following major disturbances and are often among the species that persist in an area during environmental change (Kolar and Lodge 2001). Broadly adapted species often decline when more specialized species immigrate and compete with, consume, and parasitize them as environments become more hospitable. Yet,
U.S. Army Corps of Engineers 127 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence because they are more vulnerable to random variables, until populations of specialists are large enough, they remain highly vulnerable to destructive random events. Exceptionally aggressive invasive species may evade biological regulation and remain dominant in ecosystems where they can contribute to the local extinction of rarer species. This frequently happens when nonnative species invade an area unaccompanied by their native predators and parasites.
Local extinctions of rare species may be caused by deterministic or stochastic processes (Gilpin and Soulé 1986, Brewer 1994) that typically act in species-specific ways and require species- specific management. Deterministic processes are generally manageable once they are identified for desired species. They operate in a fixed direction, such as habitat destruction, unregulated hunting, or highly invasive predators. Stochastic processes vary randomly and usually cause local extinction once a population declines to low numbers as a consequence of deterministic processes. Stochastic processes include weather changes and related events (e.g., floods, fires, droughts), random variation in population dynamics (e.g., birth rates), or random genetic changes that result in the loss of alleles. Stochastic process is more difficult to manage and frequently requires some form of species-specific captive breeding program and carefully monitored reintroduction into suitable habitat. By the 1990s, eco-managers concerned about extinction were generally committed to species-specific, fine-filter approaches to extinction management (Primack 1993) 7.2.2.2 After 1995 Contemporary managers continue to accept the uniqueness of species and management needs as a fundamental management paradigm when objectives target individual species or groups of species. Much of conservation biology assesses species status and management needs at the population and metapopulation level. In addressing management for conservation purposes, Van Dyke (2008) stated: “Populations are the fundamental unit of conservation and the primary target of management…”. Recognition of rapidly changing climate elevated the importance of this paradigm. Van Dyke (2008) and Primack (2014) recognized that each species is genetically unique and responds differently to climate and other habitat change. Vulnerability to climate change needs to be assessed on a species basis (Glick et al. 2011). Specific bioclimatic models are required to project range shifts and various management requirements (Van Dyke 2008, Primack 2014). Cain et al. (2011) emphasized the importance of individualistic adjustments of many species to climate and other environmental change. Hobbs et al. (2013) generally stressed that, because species composition is always changing, ecosystem restoration needed to consider the unique needs of targeted species. Underlying much of the contemporary assessment of wildlife resource management in Krausman and Cain (2013) is the assumption that targeted species need to be managed based on their specific needs. 7.2.3 Managing for High Biodiversity Requires a Regional Approach Paradigm: The limits imposed by the richness of the regional species pool, project-area size, and barriers to project-area immigration need to be considered and managed to achieve local protection and restoration objectives for biodiversity maintenance. 7.2.3.1 Before 1996 After the ESA was passed, federal eco-managers were expected to manage for biodiversity maintenance as they managed for achievement of specific agency objectives. Conservation biologists now recognize the importance of protecting the regional species pool, managing regional as well as local ecological filters, and sizing protection and restoration actions to support viable populations in any managed area. In the 1990s, these were new and
U.S. Army Corps of Engineers 128 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence incompletely assimilated management concepts. Managing barriers by establishing species travel corridors was just beginning to be seriously considered in conservation biology. The Ricklefs (1993) model for species richness regulation promoted consideration of species kinds, numbers and locations in the regional pool and barriers to their dispersal and immigration into the management area. However, the model originated too late for general acceptance by eco- managers by the mid-1990s.
Eco-management was still based largely on the assumption that species richness and composition were generally stable at large regional scales. This assumption was based in widespread acceptance of the community-unit and stability paradigms as well as widespread use of community and ecosystem typologies and maps. That eco-management paradigm was eroding, however, as the number of native species extinctions and exotic species increased. Conservation biologists identified habitat fragmentation resulting from human land use as a major reason for the decline of many species. Managing fragmentation drew questions about how big the habitat should be and how many were needed. Conservation biologists addressed the issue by promoting preservation of the largest “natural” areas feasible and as many as possible (Soulé and Simberloff 1986), Primack 1993). The boundaries based on that type of vision depended less on the ecological elements to be preserved and more on the economic constraints exerted by land ownership and its management. Total protection was obviously an unrealistic goal.
In the usual situation where reserve size was constrained by costs, the travel corridor concept was proposed as the primary solution to overcoming barriers to species dispersal and immigration into suitable habitat elsewhere. But it had yet to be thoroughly evaluated for effectiveness in the 1990s (Primack 1993 and Brewer 1994). Despite some potential problems, the merits of corridors appeared obvious (Simberloff et al. 1992). Federal government began considering networks of interconnected federal lands to maintain populations of large and scarce wildlife (Salwasser 1987). The importance of connectivity had long been recognized in river ecology because of the needs of fish that migrated through rivers between different habitats required to complete their life cycles. But concerns were raised over corridors not always working or, worse, opening avenues for invasive species and other threats (Simberloff et al. 1992). Further research was endorsed.
Perhaps the most important point to be made by the debates over reserve size and corridors was the importance of considering species-member dispersal through fragmented landscapes and regions and the need to plan management at far larger regional scales (Primack 1993). Scalet et al. (1996) and Bolen and Robinson (1995) also recognized the need for a regional perspective and had the same concerns for those far ranging resource species that were becoming increasingly scarce. 7.2.3.2 After 1995 Acceptance of evidence that global climate is changing rapidly accelerated acceptance of a much larger ecoregional perspective in planning and management. A volume edited by Soulé and Terbough (1999) focused conservation planning on large eco-regional to continental scales. They emphasized reconnection of habitat fragments to restore and sustain apex predators that range over large areas. This was a form of the coarse-filter approach to species conservation guided by the needs of “umbrella” and “keystone” species. The approach would presumably facilitate local ecosystem immigration from the regional species pool and restructure the various ecosystems as the effects of apex predators cascaded downward through community trophic levels. One of their main points is that management activities at local scales need to be highly
U.S. Army Corps of Engineers 129 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence coordinated across large regional scales. Climate change is not mentioned at all, however, even though it is now viewed as a critical consideration (Primack 2014).
Ecologists have made some progress in developing species assembly “rules” for managed areas (Temperton et al. 2004). They advanced conceptual models for considering ecological filters (Belyea 2004) that operate between regional species pools and the local management areas. They also began to accept as paradigm the frequent futility of protecting and restoring local species elements of ecosystems in the absence of a regional perspective on the distributions of existing populations, their dispersal capacities, and their ecological filters.
Van Dyke (2008) described advances in conservation planning that considered species flows from regional to landscape and smaller ecosystem scales when designing regional networks of reserves and connecting corridors. He and Primack (2014) also emphasized the importance of regional to local accounting of species distributions and reserve network gaps to assess species vulnerability. More recently, more emphasis is being placed on inter-organizational cooperation for climate change adaptation in the consideration of needs and design for biodiversity sustainability. One example is the LCCs set up by the U. S. Department of Interior (Cole et al. 2018). 7.2.4 Manage For Historically Moderate Disturbances and Against Extremes Paradigm: To restore and protect species richness, eco-managers need to promote the effects of historically moderate disturbances and mitigate the effects of extreme disturbances at the source and by increasing the number, size, connectivity and heterogeneity of project areas.
7.2.4.1 Before 1996 By the 1990s, moderate disturbances were recognized as a source of environmental heterogeneity that favored niche diversity and greater species richness. Primack (1993) knew the importance of maintaining all stages of plant succession in terrestrial ecosystems to support the full array of species. He implied the importance of small reserve size in limiting a full array of successional conditions and the possible need for management to replace historical disturbances. He specifically noted the importance of maintaining historical fire regimes in fire- maintained terrestrial systems using controlled burns. For aquatic ecosystems, managing for moderate hydrological fluctuations starts with watershed management. Primack 1993) also recognized the importance of water management outside protected areas for maintaining historical hydrology, but was more concerned about water diversions than watershed condition.
Extreme floods, fires, drought, and other extreme disturbances can cause extinctions both locally and sometimes more broadly in ecoregions (Primack 1993). In the 1990s, human-caused climate change was still viewed as worrisome, but uncertain, and there was little emphasis on managing it. The more frequent approach to moderating the effects of extreme conditions or events was to create or restore refuge habitats within the managed area. Refuge habitats provide retreats from exposure to the worst conditions associated with extreme events. Fisheries managers, for example, increasingly emphasized more restoration of geophysical variation in channelized streams based on the assumption that a physically less diverse river has fewer spatial niches for species diversity and fewer hydraulic retreats during high-velocity flood events (e.g., Rabeni and Jacobson 1993).
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Extreme floods, fires, droughts, storms and other climate-related events commonly occur over large geographical areas. Their effects may be moderated by either increasing the size of projects to a geographical area larger than expected from a major event or by increasing the total number of projects in the region and dispersing them to reduce aggregate exposure to single events (Primack 1993). Ordinarily, larger protected areas were preferred in keeping with island biogeography theory, but, where that was not possible, establishing a number of scattered smaller projects was a secondary consideration. Regardless of the approach, a regional perspective was increasingly believed to be essential for coordinating preserve planning by the same or different organizations, whether for protecting intact areas or restoring and protecting damaged areas.
7.2.4.2 After 1995 Since the 1990s, climate change has dominated a lot of the thinking about regional eco- management. Lawler et al. (2010), for example, recommended prioritization of eco-management actions based on the uncertainties associated with climate change and emphasized the need for adaptive management regardless of the measures chosen. Because they believed them to be least influenced by climate change uncertainties, Lawler and his coauthors recommended expanding reserve networks, enhancing connectivity among reserves, and removing other threats within management areas (e.g., aggressively invasive species, poaching, contamination). They considered population translocations especially risky and habitat restoration only slightly less so when the justification depended on climate change forecasts. Fewer measures are available to protect cave, freshwater-spring, and other species limited to very small ranges from extreme events. They can be protected to some extent by minimizing other threats (e.g. Lawler et al. 2010). For example, better protection of aquatic cave and spring species would typically entail watershed improvements and tighter regulation of groundwater use.
Aquatic ecologists have long recognized the importance of watershed condition for conserving freshwater species, but many terrestrially-oriented conservation biologists tended to treat watershed boundaries as a secondary consideration in reserve design. Recognizing the importance of watershed condition for sustaining moderate hydrologic variability required by vulnerable aquatic species and support systems, Primack (2014) recently suggested that reserves be designed to include entire watersheds. He did not, however, describe the challenges associated with efficiently designing reserve networks for both aquatic and terrestrial conservation. Conservation biologists now prefer management using methods that mimic unmanaged processes. Examples include use of controlled burns that mimic historic fire history to reduce the probability of more destructive fires (Van Dyke 2008) and increasing the spatial heterogeneity of rivers below dams by simulating flood releases from reservoirs. 7.2.5 Manage Local Areas for Appropriate Habitat heterogeneity Paradigm: Management for habitat heterogeneity appropriate for the species desired in local areas can contribute significantly to the sustainability of high species richness and the selection of future reserves. 7.2.5.1 Before 1996 Emphasis on habitat heterogeneity (habitat diversity) has a long history in wildlife resource management. Leopold (1933) was very aware of the importance of habitat heterogeneity appropriate for the needs of numerous game animal populations. He advocated forest cutting and rotations that would mimic the “interspersion and juxtaposition” of different habitat succession stages suitable for some species. Population-specific habitat heterogeneity tailored
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for the needs of desired fish and wildlife species became a staple of conventional fish and wildlife resource management (Neilsen 1993, Bolin and Robinson 1994, Scalet et al. 1996).
Conservation biologists also considered the species-specific importance of habitat heterogeneity for supporting populations of many species vulnerable to extinction (Primack 1993). They were also aware of the role that habitat heterogeneity could play in determining the species richness of an area, but relied much more on knowledge of species richness when it was available. As remote sensing technology advanced, they became more interested in habitat heterogeneity as an indicator of relative species richness where information on species richness was limited, typically in locations outside of North America.
Those eco-managers who turned toward ecosystem management to sustain biodiversity, gave up on tailoring habitat heterogeneity for particular species. Ecosystem restoration and protection relied heavily on mimicking the predisturbance condition of physical environments as closely as possible, using prescribed burns, simulated flooding, and other simulations of predisturbance processes (Bourgeron and Jensen 1994). But most conservation biologists regarded ecosystem management as too coarse-scaled for specific population needs (Primack 1993). They remained more concerned about the specific habitat heterogeneity and other qualities needed by imperiled species and accepted management intervention when simply setting aside reserves was insufficient. Conservation biologists were well aware of the specific population effects of habitat fragmentation and habitat quality degradation, and were rapidly learning the importance of considering the species-specific habitat sizes required to support viable populations and metapopulations (Primack 1993). Those managing for sustained use of fish and wildlife resources had long been aware of habitat needs, including size and qualities, to sustain harvestable populations. They were also increasingly aware of the habitat protection needs of rarer species during resource management, and the complexity that consideration added to management responsibilities (Bolen and Robinson 1995, Scalet et al. 1996).
The relevance of habitat heterogeneity was still being considered under the assumptions of ecological stability and stationarity in the early 1990s. Eco-managers in general limited their considerations of habitat heterogeneity effects to fixed locations determined largely by property, watershed, and vegetation-type boundaries. They generally assumed that the interactions between species and habitats would remain about as they had been in the past (Primack 1993, Bolen and Robinson 1995, Scalet et al. 1996).
7.2.5.2 After 1995 As eco-managers accepted the high probability that global climate was changing, they began to consider the effects of spatial heterogeneity in a different light. While eco-managers continue to focus local habitat management on the habitat needs of targeted species (Krausman and Cain 2013, Van Dyke 2008, Primack 2014), most are emphasizing the importance of considering larger landscape and ecoregional scales for maintenance of species diversity, with much more emphasis on maintaining suitable habitat connectivity. These approaches usually start with identifying areas with a “full range of environmental conditions” needed to support numerous targeted species (Krausman and Cain 2013). This approach typically relies on a set of focal species to guide the site selection process and any management that may be needed. It is basically the approach embraced by the LCCs to set landscape conservation priorities (Cole et al. 2018).
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Because of the uncertainty of future ecological changes and their effects on species survival and redistribution, some conservation biologists are suggesting that future nature preserves and connectivity should be based more on physical aspects of habitat heterogeneity than on existing species richness or projected changes in the suitability of habitat for desired species (Groves et al. 2012). This proposal has not yet reached a paradigm status in eco-management thinking (Primack 2014), but it adds another consideration among many considerations needed as eco- managers attempt to adapt to climate change.
Like other approaches proposed in the past, eco-management paradigms are likely to become a more complex mix of considerations that result from integrating coarse-and fine-scaled ecosystem- and species-based approaches. Increased habitat heterogeneity resulting from an ecosystem approach to management may provide more fundamental niches, which may increase the probability of greater species richness, but a coarse-scaled approach used alone will not necessarily provide for the needs of the most desired species. The area of each future habitat type must not only provide species-specific qualities, but also be large enough to support viable species populations. In those situations where more hands-on management is possible, the goal is to optimize, rather than maximize, spatial heterogeneity by tailoring it as much as possible to the needs of target species while retaining much of the stabilizing effect of higher species richness.
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8. Implications of Paradigm Shifts for the Corps
8.1 The Relevance of the Paradigm Shifts to the Corps Contemporary ecological paradigms no longer support the explicit and implicit assumptions of stability, stationarity, community-unit coherency, management independence, and a holistic ecosystem approach to protection and restoration of a “more natural” ecosystem condition that is implied by much of Corps environmental guidance. The paradigm shifts require deep consideration of species sustainability needs in ever changing ecosystem and landscape contexts, many of which have undergone extensive habitat fragmentation and quality degradation. However, Corps guidance often allows experienced Corps practitioners to incorporate the new ecological thinking into guidance interpretation. Corps practitioners are left much room for interpretations of guidance that are reasonably consistent with contemporary paradigms. Ultimately though, future Corps personnel, partners and other cooperators are likely to benefit from revisions of the policy guidance and development of new technical guidance that more clearly reflects the paradigm shifts.
The new paradigms require different approaches to ecological protection and restoration within the constraints of property-based management authorities, with much more emphasis on cooperation, coordination, and collaboration among diverse organizations and authorities. To be consistent with the paradigms, Corps planners need to not only refine its predominant emphasis on management of the geophysical environment by considering more pointedly the needs of the scarcest and most vulnerable ecosystem elements in the nation, its threatened species, but also to greatly increase their emphasis on protecting and restoring essential biological functions, including the top down functions of large predators and other keystone species.
The paradigm shifts have profound implications for the Nation’s sustainability goals, which Corps practitioners need to more carefully consider to set proper priorities and specific management objectives. A consistently held goal of national and international governmental policies is to improve and maintain human welfare from environmental use while sustaining the full diversity of options for future welfare within and across human generations. This goal is clear in the goals of NEPA, which establishes Federal environmental policy in the United States. The diversity of ecological options is typically maintained in the Nation’s biodiversity, and the Nation’s commitment to maintaining that diversity in a national heritage sustained for future human choice and benefits. This national commitment to maintenance of ecological options is most clearly expressed in the goals of the ESA and related federal laws.
The Corps expresses these national goals through its broadly stated sustainability goals, which are generally expected to be fundamental considerations in all Corps activities. These goals are integral to the Corps Environmental Operating Principles and to impact mitigation considerations for all civil works projects, regulatory functions under the Clean Water Act, stewardship of Corps lands, and the ecosystem restoration program. Yet the importance of restoring and maintaining biodiversity tends to get lost in the broad generalities of its sustainability goals and the complexity of Corps policy guidance. It is hardly addressed in technical guidance.
To be more consistent with recent shifts in eco-management paradigms, Corp practitioners will need to shift certain emphases within their approaches to achieving ecosystem restoration and environmental protection objectives. Because biodiversity maintenance is essential for maintaining national ecosystem diversity, Corps practitioners should more explicitly emphasize a focus on maintenance of biodiversity at local to national scales. They also need to focus more
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explicitly and specifically on the needs of those living ecosystem elements most vulnerable to decline and eventual extinction, including management of threats to species survival from local to global scales. Fully recognizing that paradigms can continue to shift, Corps practitioners should carefully consider the new eco-management paradigms to accommodate the instability, non-stationarity, and changing species compositions of ecosystems brought about by climate and other environmental change. They also need to more fully accept the probability of constant change and working more thoroughly with its nonfederal sponsors to adaptively manage for objective achievement over the entire project planning period and beyond. Perhaps even more so than under previous eco-management paradigms, Corps effectiveness will depend on more fully communicating, coordinating, and otherwise cooperating more comprehensively with other organizations in pursuit of sustainability and conservation objectives at local to national scales.
These needs are discussed in detail in the next five sections.
8.2 Strategic Adaptation of the Corps to Paradigm Shifts 8.2.1 Focus on Sustaining Local and National Biodiversity The goals and objectives of Corps policy guidance emphasize environmental and ecosystem sustainability over a more specific focus on maintaining biological diversity at local to national scales. Despite policy guidance identification of high native biodiversity and “more biologically desirable species” as indicators of success, the existing emphasis on Corps actions has been translated largely into restoration and/or protection of the existing and past habitat. The usual Corps concept of “habitat” is closely aligned with geophysical condition (getting the hydrology and geomorphology “right”), including vegetation structure, but does not typically include getting the biology right, which is also part of the habitat of any one or more species. To be consistent with existing paradigms and more effective in the future, technical guidance needs to approach ecological restoration and/or protection under changing environmental conditions with more specific biological protection and restoration objectives in mind. “Habitat” improvement and protection needs to be more consistently viewed as a means toward preserving national biodiversity over restoration and/or protection in any specific location. While implementation of the Corps ecosystem restoration program must focus its activities on geophysical restoration, consistent with its authorities, project and program planning needs to be more comprehensive and engage cooperators as needed to completely achieve ecological objectives.
Most Corps environmental activities are localized in civil work projects and mitigation actions, including actions required under Corps regulation of the discharge of materials into the Nation’s waters and wetlands. Given the paradigm shifts that have occurred, the possibility of restoring and maintaining communities of the past is increasingly unlikely. Consistent with its commitment to long term sustainability of project outputs, the Corps needs to plan with a greater focus on sustaining the most globally threatened elements of ecosystems somewhere in the Nation, if not precisely where they once occurred in the Nation. The elements recognized nationally as most threatened with total loss are biological species and the legal focus of the Nation, through legislation such as the ESA, is on sustaining their unique biological attributes in perpetuity. Since many of these species may need to redistribute over tens to hundreds of kilometers to adapt to environmental change during a 50 year planning period, the dimensions of project areas will often need to be substantially greater than in the past and significantly more concerned with connectivity to other suitable habitat. Since areas that support the highest historical levels of species diversity are often most likely to support more globally vulnerable species, Corps personnel also need to manage for historical levels of species richness on Corps properties regardless of anticipated changes in the composition of both geophysical and
U.S. Army Corps of Engineers 135 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence biological variables. One place to start approaching the problem is with our limited understanding of the factors that appear to regulate species richness. They may serve as a first step in managing for more sustainable species and associated biodiversity at National scale (Figure 21).
Several major geophysical factors underlying determination of total regional species richness are not anticipated to change much during the next century and, fortunately, are not significantly manageable in any practical way. Figure 21. A focus on sustaining biodiversity starts with an They include total solar and understanding of species richness regulation. chemical energy entering the ecosphere, total land and water Regional Regional Regional Sufficient areas of the continents and Energy Geophysical geographical Evolutionary oceans (even allowing for some Input Diversity area Time change due to sea-level rise), and Maximum geophysical diversity at regional Regional Species to continental scale. We may Richness Rapid Greatly Reduced Balanced Anthropogenic assume that those factors will Evolutionary Regional environmental Time Immigration, Reduced change continue to contribute largely to evolution & Regional Species setting a maximum on the total Extinction Richness Landscape Fragmentation species richness of large regions (e.g., North American temperate Local Reduced Altered Habitat Geophysical Dispersal/Local Quality zone) in the future. The maximum, Diversity Immigration Altered however, is eroding as the rate of Disturbances global extinction exceeds the rate Local Reduced Local Geographical Species Human of species origin. While the Area Richness Predation species richness of large regions sets a limit on local species richness, local processes, including Corps and other human activities, contribute to both reduced and restored species richness at local and regional scales (as indicated by two- directional arrows in Figure 21). Ecological filtration by landscape and ecosystem fragmentation sets a lower limit on local species richness and impedes recovery of imperiled species richness. Fragmentation reduction is one of the main management strategies for recovering species.
Since high species richness may contribute to greater community stability as climate changes, eco-managers also need to minimize local impact on habitat suitability for diverse species (the feedback indicated from “Local Species Richness” in Figure 21 reflects this belief). Of particular relevance to the Corps, local differences in geophysical diversity (spatial heterogeneity) and geographical area contribute to differences in local species richness. Geophysically diverse areas generally provide a greater diversity of tolerable conditions for a greater number of species than more uniform areas as long as the spatial diversity is not so great it reduces habitat sizes below minimum requirements for desired species. Even as species composition changes, areas that were species rich in the past because of geophysical diversity are likely to remain species rich in the future as long as landscape connectivity and local habitat qualities are maintained and unsustainable human predation is minimized. But biological considerations, including species introductions, are also important to consider. From a functional standpoint, including contributions to ecological resilience, individual species contribute uniquely and often quite subtly. The functions of numerous species overlap closely, providing a functional redundancy that collectively contributes to resilience and stability but often does not stand out at the individual species level. The importance of dominant species becomes obvious where there are few substitute species and the dominant species are decimated. Oysters are one example and cottonwoods in many western riparian areas are another. Regardless of their physical
U.S. Army Corps of Engineers 136 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence dominance, keystone species perform functional roles with major effects on community characteristics that often cannot be assumed by other species. Eco-engineers and top predators are important examples. Historically common top predators have been so greatly reduced in so many ecosystems for so long, that managers are inclined to forget their keystone importance, which research has often shown to be dramatic.
The present species extinction trend could lead to a sixth mass extinction worldwide unless the imbalance between extinction and species origin is managed better at global to local scales. How well Corps practitioners optimize resource use and biodiversity maintenance for the greatest public benefit depends on how effectively it and other organizations plan specifically for protecting vulnerable populations and restoring them to abundances that are generally secure from a need to list them under the protections of the ESA. The Corps has significant resources that can be used for that purpose. One way to encourage that use is to focus protection and restoration policy more clearly on biodiversity maintenance.
Corps practitioners also need to carefully consider how geophysical diversity in the landscape interacts with new and changing climates. Species richness could actually increase in the temperate zone, depending on how much temperature and landscape connectivity are limiting factors. Careful assessment of habitat fragmentation on vulnerable species and its management needs should be a primary focus of Corps practitioners. Challenges are likely to exist everywhere the Corps is authorized to manage resources through engineering and other resource management measures. Based on recent observations, future climate change is likely to be more dramatic in the northern temperate and polar zones, but because of higher species richness, at least as many if not more species may be threatened in southern temperate and semitropical zones, especially in isolated habitats. Montane patterns of species response to climate change could change similarly as temperature zones move up slope. Species adapted to mountain top conditions, whether aquatic or terrestrial, could be driven to extinction. Both continental and montane patterns are also likely to be influenced by precipitation changes. Acidity is likely to increase in most aquatic environments, possibly limiting the potential for increased species richness as numerous species requiring stable alkalinity and calcium supply fail to adapt.
Obviously, the complexity of these possible interactions demands much more care in planning and adaptive management than the old paradigms misleadingly required. 8.2.2 Emphasize Species-Based Planning at All Ecological Scales Planning for high species richness alone is not enough to assure that the most vulnerable species are among them. To be consistent with national goals that target biodiversity maintenance at national scale, Corps practitioners need to consider the importance of more specifically guiding ecosystem- and landscape-scale restoration and/or protection through the needs of the threatened species elements in the Nation. While it may be less costly to rely heavily on restoring scarce habitat to sustain biodiversity, its assumed effectiveness in maintaining national biodiversity is based on a questionable assumption that habitat scarcity and species scarcity are proportional and a generally rejected assumption that previous communities will reestablish as a complete unit in an improved habitat without much species- specific consideration. This approach is a less reliable means of restoring species diversity than focusing on the habitat and other needs of threatened species more directly, especially since habitat characteristics are rapidly shifting in response to environmental change. While restoring the previous hydrology and geomorphology is still an important consideration, it should not be the only consideration in numerous situations because of new interactions associated with climate and other environmental change beyond reach of local management. Species will cope
U.S. Army Corps of Engineers 137 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence with environmental change in different ways and redistribute at different rates even with optimal connectivity. Without more focus on them, the threatened species, which are typically among the scarcest and least adaptable species, are more likely to be filtered out of the project area while the more adaptable and common species become established and persist.
Corps policy guidance tends to emphasize a holistic habitat or ecosystem approach to environmental management. That coarse-filter approach is likely to be less effective at sustaining national biodiversity in changing environments than a fine-filter focus on the habitat and other needs of the biologically desirable species and high native biodiversity identified as indicators of success in Corps project planning policy. This coarse-filter approach may restore gross function and structure, when carefully considered, which is necessary for support of desired species, but fails to emphasize consideration of the more specific risks impeding the establishment and long-term viability of the least sustainable species. Making this fine-filter approach work in the context of Corps project planning policy probably requires a shift in emphasis from restoring a more natural ecosystem condition to restoring a more natural level of abundance and productivity of the specifically desired species somewhere in the nation if not necessarily in the area where they occurred in the past.
Clearly, species are not equally vulnerable to extinction caused by rapid environmental changes. The unique biological attributes of narrowly adapted specialists usually face the greatest risk of global loss. The Corps has commonly relied on more broadly adapted generalists to indicate restoration and protection needs, which fails to indicate many residual risks faced by the narrowly adapted scarce species that are the appropriate target of the ecosystem restoration program and sustainability. The paradigm shifts indicate that planning for species-informed habitat connectivity at ecosystem, landscape and ecoregion scales is as important as planning for supportive qualities within residential habitats. For plans spanning 50 years or more, the scale of consideration needs to include possible population redistribution needs of 100 to 200 km or more in flat landscapes and elevation changes of 0.2 to 0.3 km or more.
Especially in aquatic and shoreline environments, Corps eco-managers may need to shift even more of their attention toward connectivity for the desired species. The environments where the Corps is most active are typically characterized by continuums of ecological change where material fluxes are often high. Rivers, floodplains, coastlines, coral reefs, and many types of wetlands form highly dynamic interactions at their interfaces between land and water. Many aquatic species have adapted to these continuums and suffer from their fragmentation. The interruption of species movements is widely recognized by many Corps practitioners. Perhaps less widely recognized are species requirements for an optimal range of material inputs, from both proximal and remote sources far from the location of particular management concerns. Significantly increased or decreased material inputs, including suspended sediment and nutrients, typically causes major changes in riverine, wetland, and coral ecosystems. Understanding how the quality of connecting habitats and material input-output dynamics interact with climate change is essential for maintaining optimal connectivity for biodiversity maintenance.
Because of the individualistic responses of ecosystem elements to climate and other environmental change, future eco-management guidance should clarify the importance of species-based population assessment and management planned from local to national ecosystem scales. Managing for ecosystem sustainability must allow for major geographical redistributions of habitat variables and species, and concentrate on maintaining the diversity of possible species compositions and interactions with their environments at a national ecosystem scale rather than local scales. In that context, a more geographically and organizationally
U.S. Army Corps of Engineers 138 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence comprehensive approach to ecosystem management for species sustainability is more effective than a localized and organizationally fragmented approach. The comprehensive approach targets the sustainability of national ecosystem biodiversity based on integrated, species-based assessments of needs that are pursued through inter-organization coordination and cooperation.
The use of surrogate species may be the only practical way to guide conservation planning, but careful selection of a focal group is generally much more effective than single-species indicators. Planning for the needs of a carefully selected focal group should go a long way toward protecting or restoring historically high species richness as well as species targeted for special attention because of their vulnerability to global extinction. Since national biodiversity protection or restoration should always be a driving concern, comprehensive selection of focus species should include globally threatened species and supportive dominant species, keystone species, and other essential species. Remaining uncertainties require planning for adaptive management and monitoring population response to improve the chances for objective achievement.
PVA is an essential aspect of species-based protection and restoration in a changing ecosystem context, including the concept of MVPS, the habitat size needed to support it, and vulnerability analysis. When done using a focal group approach, PVA assumes a broader community and ecosystem perspective. While characterizing the fundamental niches of the focal group is important, the biological interactions that shape the realized niches are equally important. Population viability planning and management often requires knowledge of the metapopulation dynamics of the species. The use of the habitat evaluation procedure and habitat suitability indices for a single species or a closely related group of species (e.g., dabbling ducks) can guide environmental impact mitigation planning, it’s original purpose, but is less likely to succeed for ecosystem restoration than PVA using the focus group approach. Most Corps planners and other eco-mangers cannot be expected to be expert in the population ecology and techniques, but they should be aware of their essential importance and able to team with Corps ecologists to select and work with outside experts.
Because of the relationship between spatial and temporal scale, future eco-management will generally require a longer-term strategic approach to regional assessment to inform the project investment decision process. Improvements will likely require more strategic considerations of investment priorities in ecosystem restoration and protection actions. For example, the habitat suitable for many Caribbean coastal marine species, now limited to southern Florida or southern Texas, could substantially expand along the Gulf coast and up the Atlantic coast in response to climate change without much need for human help. Range expansion could significantly relieve threats to those species. In contrast, many American rivers run generally east and west to waters that are uninhabitable for threatened freshwater species. Their options for redistribution along a North-South axis as climate changes are limited by land barriers and redistribution to higher elevations are often limited by dams, pollution, and other anthropogenic barriers. 8.2.3 Accept New Restoration and Protection Paradigms The paradigm shift in eco-management perspective from local ecosystem restoration and sustainability to biodiversity maintenance at national scale requires quite different ecological restoration and protection perspectives. Most leading eco-managers now doubt the practicality of restoring a more natural ecosystem condition like that which would have existed in the project area if human impact had not occurred. They now believe that neither biotic communities nor their physical environments behave as cohesive units (species composition and geophysical variables are always changing). Nor do they believe they are geographically stationary or
U.S. Army Corps of Engineers 139 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence sustainable in any one place for an indefinitely long time. Consequently, the practicality of achieving sustained restoration of a more natural ecosystem is rapidly losing appeal. The pervasive influence of human-caused climate change and other locally unmanageable environmental change on “reference” and “restored” conditions argues against the possibility of locally restoring an area to a condition like that which would have existed without human effect and then sustaining that condition for many decades into the future.
Setting aside areas to protect biodiversity from local human impacts continues to be an important strategy, but one that must be complemented with management that allows for species redistribution and other adaptation to environmental and community change. While maintaining a natural condition has been part of the strategy in the past when ecological stability was widely assumed, it is less likely to remain so in the future. Many applied ecologists realize the increasing impossibility of sorting out and preventing all but the most obvious and threatening environmental impacts in project areas and reference sites as the pervasive effects of anthropogenic climate change, nutrient cycling, synthetic chemicals, decimation of keystone consumers, invasive species, and other environmental effects have been more fully recognized.
Because of its past geophysical orientation toward environmental management, the Corps has emphasized getting the ecosystem hydrology and geomorphology “right” and has assumed that the biotic community would recover as a unit in the improved habitat and then remain stable and self-sustaining without much consideration of biological risks. Unfortunately, this assumption is now regarded as management myth and “holistic” short cuts to restoring and/or protecting sustainable abundances of targeted species vulnerable to extinction rarely work reliably. At best, such efforts are statistical gambles with considerable risk of failure elevated by uncertain interactions between climate and the geophysical conditions of the landscape. Preserving biodiversity requires management objectives to be more precisely defined by the needs of the most threatened species, including their supporting species and habitat conditions.
The old eco-management focus on place-based protection and restoration of “more natural”, self-regulating, and sustainable ecosystems is fading rapidly in favor of more specific objectives that can be achieved in ever-changing “novel” ecosystems significantly influenced by the effect of human activities. Maintaining “more natural ecosystems” has much less meaning for biodiversity maintenance than sustaining national species richness and other biodiversity at regional to global scales, regardless of how free the ecosystems are of human effect. This does not mean that management measures with the least need for human intervention should not be considered for their cost-effectiveness. But the risks of failure also need to be carefully weighed and counterbalanced by human intervention when necessary to contribute meaningfully to maintenance of the diversity of species and ecosystems at a national scale.
The concept of targeting the restoration of plant and animal diversity native to a particularly location needs to be rethought. As anthropogenic changes in climate and other environmental variables occur, many geographical areas will become “unnaturally” inhabited by “nonnative” species. Many native species may ultimately become “invasive” and “nonnative” at particular project locations. The likely alternative to successful redistribution is the extinction of many narrowly adapted species. While ecosystems occupying specific geographical areas will not be the same as past systems, or free of human effect, most will continue to function as nonnative species from elsewhere in the nation functionally replace many of the previous inhabitants. In many situations, eco-managers need to foster these changes, which help to sustain functional resilience as well as the maintenance of native biodiversity at national and international scales.
Leading eco-managers are increasingly treating nonnative species of foreign origin as valued contributors to local biodiversity and the functional stability that biodiversity tends to impart. For
U.S. Army Corps of Engineers 140 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence the sake of practicality and the increased community stability associated with high biodiversity, leading eco-managers are more likely than in the past to accept species of foreign origin into novel community compositions as long as they do not threaten stability. Thus a focus on native species is losing ground among conservation and restoration practitioners to a focus on threatened species and functionally supportive species of most any origin.
To become more consistent with ecological paradigm shifts, Corp practitioners need to reorient their interpretation of the restoration planning objective, as stated in restoration guidance, from one apparently focused on a more natural and self-regulating ecosystem to one clearly focused on biodiversity sustainability at a global scale. A similar reorientation is appropriate for its project management objective, which targets ecosystem sustainability. The Corps also needs to explain how that objective should be pursued at program and project scales in more clearly written and complete policy and technical guidance. 8.2.4 Adapt Management to Constant Environmental Change The paradigm shifts in ecological science and management are not the best news for federal agencies increasingly encouraged to be more effective with less funding. Although there were many signs of growing skepticism in ecological science, the old eco-management paradigms misled many into believing that environments and ecosystems could be quite certainly and cheaply improved as desired because their responses to management were deterministic and predictable. The certainty so securely held in past scientific paradigms pertaining to ecological stability, stationarity, holism and coherent integrity at local community and ecosystem levels are still attractive for their management convenience, but the evidence against these beliefs and assumptions can no longer be ignored. Perhaps a degree of certainty may be expected at the level of gross functional and structural scale. Industrial brown fields, for example, can be more or less reliably restored to aesthetically more pleasing green space where stressors have not been too severe and intractable. But that is not the aim of the Corps ecosystem restoration program, which is to restore the sustainability of nationally important ecosystem attributes, including, most clearly, imperiled species. Reversing the decline of numerous species is much less predictable and more challenging than restoring gross structure and function.
Collapse of the community-unit and stability paradigms caused a chain of ecological paradigm shifts that will affect Corps mitigation, restoration and management decisions far into the future. Further ecological change appears inevitable for decades to come, but much uncertainty is associated with projections of change and adaptation to that uncertainty is necessary. Most informed professionals now know the results of eco-management are much less certain than early practitioners appreciated, requiring monitoring and adaptive management to stay on course. They cannot rely on deterministic “nature knows best” management and restoration strategies. Many species are expected to behave individualistically as they attempt to adapt to climate and other environmental change, rearranging in ways that are difficult to predict. The gross functions and structure of ecosystems resist major changes up to a point and then shift, often without much warning, to very different ecosystem conditions. Between these major shifts, the fine-scale structure and functions of ecosystems are continuously changing while gross structure and function remain stable.
The fine-scaled uncertainty usually involves the most vulnerable elements of national biodiversity that are among the greatest environmental concerns of federal agencies. As scientific and technical knowledge increases, reducing some uncertainty, other uncertainty is often revealed; especially about the needs of vulnerable species. Whether for restoration or protection, Corps eco-managers must manage uncertainty by forming closer relationships with research scientists and other eco-managers, participating in more spatially and temporally
U.S. Army Corps of Engineers 141 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence comprehensive planning, including many more species-based considerations in planning, adopting scenario-based planning to increase management flexibility, monitoring objective achievement following plan implementation, and modifying management as needed to achieve objectives. The uncertainty of climate change effects generally requires greater emphasis on planning with management flexibility in mind over the entire period for which a project is planned. For the sustainability focus of Corps restoration and protection activities, that period has no clearly determined end. When nonfederal sponsors are expected to protect the sustainability of the managed ecosystem, monitoring and adaptive management expectations need to be more clearly defined than in past guidance. Adaptive planning may require planning for several scenario conditions to anticipate flexibility needs. This of course implies that planning costs would need to increase if the Corps is to become more effective.
Even with fine-tuned planning and management, outcome uncertainty is unavoidable. Because the types and rates of future ecological change are uncertain, ecological planning must include room for adaptive management at scales dependent on possible rates of change. But the details often require more information about targeted species and their support systems during the planning process and following implementation of management measures than is appreciated. Carried out in its most effective form, adaptive management requires a close marriage of research scientists, information technologists, and field-managers. Models of the managed system are developed and continuously improved by pilot studies and monitoring of project outcomes. Adaptive management extends beyond the project level to program levels as more is learned about project effectiveness. It often incorporates a scenario-based approach to planning for alternative future conditions, which forces a check on management flexibility and suppresses the frequency of surprises. Careful monitoring of objective achievement is essential, which typically means monitoring populations of the species desired in and around the managed area. Monitoring of gross functional indicators, such as nutrient concentrations and water flow, is unreliable.
Whether through protection, restoration or other avenues to environmental improvement, the uncertainty about future environmental conditions requires Corps practitioners to adopt strategies that more effectively manage or adapt to uncertainty. Objective achievement needs to be based on variable outcomes rather than a particular composition of species. While there may be a dozen candidate species targeted for population restoration, success may be defined by the actual restoration of two or three of those populations. This need may be particular true for organizations that approach problem “solutions” largely through environmental engineering, as the Corps does. But the finality implied by the word “solution” is not applicable to biodiversity maintenance problems. They need to accept a range of possible outcomes as successful and continuously adaptive management to sustain success. For those concerned about the sustainability of biodiversity, successful species recovery and protection in changing ecosystems will require much more knowledge, inter-organizational cooperation, comprehensive and intricate planning, and money to manage risks than the old paradigms implied.
While much of the responsibility for long-term adaptive management of Corps restoration projects rests in the hands of the Corps non-federal project sponsor or sponsors. The partnership demands much of all partners in the process, particularly in the care taken to clearly lay out project objectives, to plan completely for contingencies, and to clearly spell out the obligations of all partners during the very long post-construction phase of project management.
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8.2.5 Cooperate More Effectively With Others Close cooperation among project sponsors is, of course, essential, but even broader cooperation among agencies and other organizations with diverse authorities and skills is needed. The paradigm shifts toward much broader management consideration at landscape to continental scales indicate a need to better coordinate inter-organizational eco-management planning and implementation activities more completely at regional and national scales. The LCCs exemplify the cooperative approach to sustaining resource use and biodiversity in a commonly understood framework of conservation goals (Cole et al. 2018). They were formed to facilitate the transformation of species-based conservation from local management units to a landscape approach coordinated by an ecoregional and national visions of management needs.
Contemporary tenets of ecosystem management are now widely accepted as complementary to population-based management and are no longer promoted as a replacement for species-based approaches. Outside of its emphases on cooperative, objective- and service-driven adaptive management, the most important contribution of ecosystem management may be its emphasis on assessing and managing the needs of targeted species from population to landscape and regional scales. A new management reality requires much more specific information about the needs of species most vulnerable to extinction as well as those desired in greater or lesser abundances for economic reasons, such as pest control and recreational wildlife.
When ecological stability and stationarity was assumed, most organizations were likely to manage their assigned properties more or less in isolation from other management jurisdictions. Since those assumptions were dismissed, most eco-managers have been more willing to adopt the more cooperative, coordinated, and collaborative inter-organizational approach to planning that is needed to adapt to projected rates of ecological change within and across property boundaries. The technical challenges to cooperation are daunting, but the many social-political barriers impeding a more integrated approach to management may be the greater impediment to sustaining biodiversity and associated diversity of resource options for future generations (Gunderson et al. 1995). Many of those barriers originate in a property-based approach to resource management. Transcending those barriers will require the full engagement of agencies in multi-organizational cooperatives despite limited budgets and uncertain futures.
The fragmented nature of government and private enterprise in the United States and elsewhere is an impediment to progress. Despite willingness, it is difficult for eco-managers, up to the top executives in each organization, to expand their highly limited view of the issues at hand and the roles they can most effectively play. The day to day demands placed on eco- managers tend to overwhelm the more strategic needs of inter-organizational cooperation at program levels of management. Assuring the long-term sustainability of ecological diversity will require the highest level of interest in and commitment to such cooperative efforts. But, because there is such a piecemeal demand for cooperation in government agencies, various cooperatives compete for attention. The Corps needs to evaluate the likely effectiveness of different cooperative interactions and set priorities for engagement that are most consistent with national goals and the Corps environmental and restoration missions. In the United States, inter-organizational cooperatives are the only viable means for integrating numerous public and private conservation organizations into a nationally comprehensive adaptive approach to climate and other environmental change.
Among conservation cooperatives, the LCCs are uniquely comprehensive and adaptive to climate and other environmental change. The LCCs are defining landscape-level conservation priorities based on the ecological needs of a species focus group, which may include resource species, species vulnerable to extinction, dominant and keystone species, and problematic
U.S. Army Corps of Engineers 143 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence invasive species. This species-based approach to eco-management is reinforced by the paradigm shift from community-unit assembly behavior of species to one based in the belief that small groups of tightly associated species assemble individualistically. The effectiveness of ecosystem management is greatest when it takes individualistic behavior into consideration and its effectiveness is evaluated based on the sustainability of all of the ecosystem elements somewhere in the ecoregion. The approach of the LCCs reflects growing acceptance of the importance of top-level, “apex” consumers and other keystone species in regulating the composition and functions of multiple communities at landscape scale. It is moving eco- management toward a more thoroughly integrated population, community, ecosystem, landscape and larger regional approach to management.
Many contemporary restoration ecologists would agree that the ecosystems that now exist in most American rivers and coastal areas are likely to resist restoration to any previous condition, let alone a pristine state. But many also believe that restoration of species to a sustainable state is possible in other ecosystem arrangements and locations when careful, cooperative attention is paid to their individual needs, including acceptable dispersal corridors, at local to regional scales over long periods of time. In this way, the diversity of species may be sustained nationally in a wide variety of changing ecosystem assemblages without depending on the improbable restoration and sustainability of past community and ecosystem conditions.
A better alternative to outmoded management concepts based on each organization independently sustaining local ecosystem compositions and processes is to concentrate instead on maintaining the diversity of biotic and abiotic attributes at regional to national scale though the concerted efforts of cooperatives. Partner organizations in the LCC need to be prepared to commit sufficient time and resources to coordination; clearly articulate their project and program objectives and constraints; identify species and support-system management priorities consistent with their organization missions; contribute meaningfully to spatially and temporally comprehensive conservation planning; consider improved implementation methods, smarter monitoring, and more insightful adaptive management; continuously update ecological thinking; and, of course, use their funds more effectively once informed by the cooperatives desired future conditions.
The most effective approach to making research and management investments in conservation activities is to coordinate needs and pool resources across organizations. Integrated ecosystem, landscape, and ecoregional approaches to species population management are essential because of the numerous direct and indirect environmental influences on populations, many of which originate far from population locations. In the Corps, project size and location cannot be limited to cost-effectiveness and partnership considerations addressed in an ecological vacuum and still expect success. Because the long-term sustainability of targeted populations is necessary for success, new projects need to manage areas large enough and well enough connected to support viable populations, including apex predators and other keystone and dominant species that sustain much of the character of a desirable support system for the species populations targeted by management. That requires more information about the abiotic and biotic needs of targeted species and their support systems at a wide range of ecological scales.
A landscape approach to species-based conservation planning typically emphasizes both habitat quality and corridor connectivity. Cooperative approaches at regional to national scales can inform eco-managers about whether restored and protected areas will best serve corridor or residential habitats for targeted species. Corridors may compensate somewhat for the isolation of management areas when the corridors specifically serve the needs of the targeted populations and their support systems while also managing access by undesirable species.
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Sustaining the populations of many species faced with climate change probably requires extensive cooperatively planned corridor connectivity along with assessments of risks associated with the influx of undesirable materials and species, including diseases vectors.
The LCCs are making progress while faced with many challenges, which often arise from within government. Their success or failure largely depends on the commitment of cooperating organizations and involvement of organizational personnel with appropriate understanding of contemporary ecological and eco-management paradigms. The Corps has been unevenly involved at field levels with little sustained support at the level of the Corps national headquarters. While Corps policy embraces cooperation, project level activities are emphasized. Corps participation in the LCCs has been tepid, in part because the relevancy of the LCC approach to landscape and ecoregion planning is not widely considered a priority (Cole et. al. 2016). 8.2.6 Make Use of Policy Guidance Flexibility and Consider Future Revisions The paradigm shifts in ecological science and management may imply to some that major changes in policy guidance are needed before planners can adapt to those paradigm shifts, but much can be done within the limits of existing policy guidance. The project planning policy regulations for ecosystem restoration (USACE 2000) allows broad flexibility in its restoration approach to achieving project success, including the long-term sustainability of desired ecosystem resources once they are restored. Stewardship and project impact mitigation policies are also broadly stated while also focusing on sustaining ecosystem resources in their protection approaches to environmental sustainability. Particularly important aspects of policy flexibility are described in the following paragraphs.
Policy acceptance of partial ecosystem restoration allows Corps project planners to refocus from holistic approaches, which are largely beyond reach, to specific parts of the ecosystem. It also implies that not all parts of ecosystems are equally significant resources or equally important contributors to objective achievement success. It allows the more natural condition identified as part of the objective of ecosystem restoration to apply to certain publicly desired parts of ecosystems, such as the more natural abundance of species that are now much scarcer as a consequence of human impacts. In fact, when searching for evidence of a broad national public desire for a more natural previous condition of certain resources in laws and public opinion, it is difficult to find evidence of an interest in the restoration of entire ecosystems or of geophysical parts of ecosystems (the emphasis on certain ecosystem types, such as wetlands, is on protection rather than restoration). The focus of national public interest is on species as expressed in laws such as the Endangered Species Act or in public surveys continuing to support the goal of the Act, which is to sustain an intact fish and wildlife heritage (including plants). When the focus is on restoring publicly desired parts of ecosystems, whether or not the entire ecosystem context is more natural is less important than it is in a holistic approach to ecosystem restoration. In fact, by allowing mimicry of natural conditions and naturalistic outcomes, project planning policy implies that ecosystem naturalness is not essential.
Unlike the Corps ecosystem restoration policy regulations (USACE 1999), which interpret its project implementation authorities as limited largely to geophysical measures for managing risk and uncertainty, project planning policy regulations (USACE 2000) directs consideration of all acceptable restoration measures in search of the most complete and cost effective plans. This includes a wide variety of biological measures as well as managing the risks and uncertainties associated with the possible effects of anthropogenic climate change and all other
U.S. Army Corps of Engineers 145 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence environmental change. While the Corps typically does not engage in planning projects that are highly unlikely to include geophysical improvements, after detailed study a plan may ultimately be recommended that does not involve the Corps in project implementation. Such a recommendation would have to be implemented under another organizations authority. Policy guidance generally leaves the specifics about risk and uncertainty identification and management to technical guidance. Having focused primarily on hydrology and geomorphology, Corps technical guidance has advanced relatively little for the many biological aspects of risk and uncertainty management, More complete technical guidance, including the biological aspects, would encourage a more complete approach to risk and uncertainty management for the Corps environmental and ecosystem restoration missions.
While a common interpretation of all policy guidance assumes a past and future ecosystem stationarity that is no longer scientifically accepted as valid, policy guidance flexibility allows the national diversity of ecosystem elements to be sustained at a national scale as local ecosystems change both physically and biologically in response to global drivers of change. Policy guidance also allows considerable flexibility in determining the appropriate dimensions of the project study area and where the desired ecosystem resources may ultimately reside. Even though local ecosystems cannot be restored to a compositionally stable state for very long, the national composition of species and habitat conditions is much more likely to be sustained when landscape connectivity and other aspects of landscape ecology are properly considered and managed. To do that, practitioners may need to focus on the specific habitat and other needs of individual species elements not only in the local project area, but also in a larger geographical area that considers future connectivity and residential needs elsewhere. It requires consideration of likely geographical redistributions of both geophysical and biological attributes of ecosystems as well as a larger geographical consideration of risk and uncertainty management needs. While desired ecosystem resources may not be sustainably restored to a more natural level of productivity and viability in the local ecosystem context they may be restored somewhere in the larger geographical context needed to accommodate global environmental changes.
Less flexibility is allowed for the types of outputs allowed among the desired ecosystem resources. The planning guidance indicates that the desired ecosystem resources are species desired in greater quantity and/or quality for continued existence of their unique biological attributes. This is first indicated by the National Ecosystem Restoration goal, which indicates that success is to be measured by an increase in ecological resource quality as a function of habitat improvement. Since habitat improvements are management inputs that may or may not be restoration measures, the desired outputs (determined by evidence of resource significance) are species inhabitants in a supportive habitat setting (but not necessarily a more natural setting). The indicators of ecosystem restoration objective achievement provided in project planning guidance confirm this emphasis on species as significant resources desired by the national public in greater quantity and/or quality. The indicators of success are the most explicitly stated outcomes expected from ecosystem restoration projects. They make clear that the project-area ecosystem is expected to sustain support for desired outputs, which are more biologically desirable species and high native plant and animal diversity. Based on these indicators, restoring degraded native biodiversity and sustaining it are key aspects of objective achievement. The emphasis on high native biodiversity may have been focused on conditions in the local project area in the past, but enough policy flexibility exists to allow restoration of indicators to extend to a larger geographical area and to the nation as a whole.
Where landscapes are suitable, the redistribution of species populations is a critical means for species survival in globally changing environments. Many species native to specific locations
U.S. Army Corps of Engineers 146 Institute for Water Resources Advances in Conservation Ecology: Paradigm Shifts of Consequence are likely to invade previously uninhabited areas in response to environmental changes as their original habitat becomes unsuitable. These species would be regarded as nonnative when the concept of native is defined by local inhabitation of an area. They would, however, remain native to the United States. This ability of species to redistribute as conditions change is a means for sustaining ecosystem resilience despite different combinations of habitat and species elements than previously existed in any one area. While many planners may have assumed that native plant and animals meant native to a locally inhabited area, enough flexibility in policy interpretation exists to allow native species to be defined as native to the United States. This means that for either ecosystem restoration, mitigation, or stewardship purposes, consideration of what desired species are likely to move into and out of the area in response to global changes is critical for selecting the best plans and projects for implementation any single location. Planning for appropriate connectivity between past, present, and future habitats in support of desired species is a critical consideration in any complete plan.
While existing policy guidance allows interpretation flexibility, future Corps practitioners, cooperators, and local project sponsors would most likely benefit from policy guidance revision and new technical guidance that more clearly reflects the paradigm shifts described in this report. Perhaps most important for the Corps is to more explicitly recognize the importance of improving habitats needed to restore the Nation’s threatened biodiversity and sustaining it nationally while deemphasizing restoration to some previous condition of biodiversity in particular project locations. Ideally, policy guidance would increase emphasis on restoring and sustaining nationally threatened elements of ecosystems and deemphasize any suggestion that holistic restoration and/or protection of past and existing ecosystems in a specified location is usually practical. The uncertainty surrounding what is natural or not would be clearly explained or, perhaps better still, the naturalness of the result would be deemphasized in favor of ecological self-regulation (a “wild” state) wherever it contributes positively to maintenance of national biodiversity.
The occurrence of rapid shifts in species distributions and the much larger geographical dimensions and connectivity needed to sustain those species would also be emphasized. Briefly explaining how local biodiversity contributes to maintenance of national biodiversity would reinforce the need for much broader geographical perspectives and the cooperation of other organizations in achieving national goals pertaining to biodiversity. The indicators of objective achievement for restoration planning would be further clarified with more explicit explanation of what “biologically desirable species” are and how they relate to high native plant and animal diversity and sustained ecosystem support for the desired outputs. Other possible indicators would be identified, if they exist, and if none can be identified the guidance would explicitly say so. Some of the outstanding anthropogenic sources of risk and uncertainty facing ecological sustainability would also be identified, including the wide range of environmental changes that have occurred and continue to occur. The details would be left to technical guidance.
Technical guidance would also be developed to describe in much more detail than policy guidance the uncertainties and potential risks associated with objective achievements and how to manage them. It would expand well beyond past emphasis on geophysical management measures (including vegetation management) to the wider array of ecosystem variables that need to be included in any complete planning process. Population ecology, community ecology, and the principles of conservation biology would play a much larger role in technical guidance than in the past. The contemporary paradigms described in this report can contribute to development of improved technical guidance.
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9. Summary and Conclusions
• The paradigms of ecological science and management have shifted dramatically over the past few decades. Among the more general conclusions pertaining to these paradigms shifts are the following widely accepted beliefs of ecological scientists and managers: • Global climate is changing rapidly in response to human activities and the cause has already had widespread ecological impacts as a consequence of changing temperature, hydrology, water acidity, geomorphology, and other primary and secondary effects. • While the extent of future climate change and impact are uncertain, it is virtually unavoidable and requires cooperative approaches to adaptively manage at small to large geographic scales of consideration. • Most contemporary managers now consider species populations to be the basic units of eco-management, which require more specific management planning than holistic ecosystem management, which is generally thought to be less reliable. • Many species populations adapted to more stable environments are less stable now because of environmental change and more vulnerable to the effects of future climate and other environmental change. • Many specific population variables need to be considered for the most effective eco- management; including species niche widths, dispersal capabilities, distributions, population growth characteristics, reproductive potentials, life-history strategies, metapopulation connectivity needs, minimum viable population sizes, and minimum habitat sizes for viable populations. • Since the responses of different species populations to environmental change vary widely, management designed to reduce the impacts of environmental change on particular species must be, to practical extent, tailored to both the abiotic and biotic needs of those species and monitored for adaptive management. • The compositions of biotic communities are continuously changing in ways that often stabilize changes in gross community structure and function (until biotic or abiotic extremes cause dramatic shifts), but also frustrate attempts to holistically restore and protect community biodiversity. • To contend with uncertainty in future community composition, eco-managers now generally believe that biodiversity maintenance at local to global scales requires more of a species-based approach to management at local to regional scales using species focal groups to indicate broader needs if, for practicality, indicators are required. Use of single-species or guild indicators of community needs is generally thought to be unreliable. • With some exceptions (e.g., some aggressively invasive species), high species diversity is now believed to contribute importantly to the stability of community support for populations specifically targeted for management. • Ecosystems are now believed to be constantly changing, forming indistinct and changing spatial and temporal boundaries around functional units having limited long-term management value only when carefully defined, used and frequently redefined. • Species-guided ecosystem management is a useful complement to species-based management for biodiversity maintenance because of its focus on long-term sustainability and emphases on clearly specifying objectives based on desired ecosystem services, local-to-regional approaches to assessment and planning, and
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adaptive flexibility, but holistic ecosystem management, when done independently of desired population needs, is largely regarded as unreliable in changing environments. • The belief that ecosystems can be sustained in or restored to a more natural condition is now widely doubted and eco-managers increasingly emphasize management targeted on desired ecological services and biodiversity maintenance. • Ecological restoration for biodiversity maintenance is generally regarded as a legitimate complement to biodiversity protection whenever former abundances of threatened species elements are the targets of restoration and the specific locations of restored populations are anywhere appropriate for future biodiversity maintenance. • In addition to external forcing functions, such as hydrology and geomorphology, eco- managers now believe that reestablishing apex predators and other keystone species contribute importantly to ecosystem regulation and are generally essential for other biodiversity protection and restoration. • Once overlooked or dismissed, many of the movements of materials and populations among ecosystems at project and mitigation-bank scales are now widely recognized for their importance in restoring and protecting the long-term sustainability of local to global biodiversity. • Once dismissed or deemphasized, managing for species composition and diversity in typical project areas requires a species approach with a regional planning perspective and comprehensively integrative inter-organizational management, maintenance of historically moderate disturbances, moderation of extreme environmental conditions, and appropriate habitat heterogeneity and connectivity within and outside the project area. • Despite emphases inconsistent with the paradigm shifts, Corps policy guidance allows broad enough interpretation by Corps practitioners to reasonably accommodate the shifts, but future Corps personnel, partners, and cooperators would likely benefit from revision of policy guidance and development of technical guidance more reflective of the changes.
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