Research

INTRODUCTION

Living within the ecological boundaries of our biosphere (Rockström et al. 2009) while ensuring the sustainable and equitable use of its natural resources is one of the biggest challenges facing humanity in the present century. Rivers serve as the chief source of renewable water supply for humans and freshwater ecosystems (Vörösmarty et al. 2010), and water scarcity is a global threat to both society and freshwater biodiversity (Ruhí et al. 2016). The effects of water scarcity accentuate the incident threat to human water security and biodiversity, especially in drylands and desert belt transition zones across continents (Vörösmarty et al. 2010).

This scenario is particularly accentuated for the western United States, where “Manifest Destiny” and a favorable hydroclimate led to the establishment of a significant agricultural economy, especially in the southwest, despite the warnings of early naturalists such as John Wesley Powell (Sabo et al. 2010). Concern over water quality and quantity, biodiversity, and land preservation along rivers has led to a boom in restoration activity across the United States at an annual cost of roughly $1 billion USD (Bernhardt et al. 2005). River restoration in the U.S. Southwest has followed national trends to a large degree, but has also been shaped by influences unique to the region (Follstad Shah et al. 2007).

Previous system-scale studies of river basins have used panarchy theory or Ostrom’s social-ecological systems (SES) framework to assess the resilience of river basin systems in various ways. One of the premier examples of the application of an SES framework in river basin resilience assessment was that by Cosens and Fremier (2015) in which the Columbia River basin’s resilience was traced through historical timelines divided into pre-contact, post-contact, dam-building, and civil and environmental justice eras. Cosens and Fremier (2015) presented “eyeballed” estimates of ecosystem services present in the Columbia River and then used expert elicitation to quantify resilience based on defined resilience metrics. This has been an often-used method in other case studies using panarchy theory and other resilience related assessments in adaptive governance of river basins (Cumming 2011, Nemec et al. 2014, Allen et al. 2018). Although it is a proven technique, it draws attention to the underlying problem of insufficiently incorporating widely available ecological metrics into SES analyses, which may obviate the need for estimating resilience and bring more exactitude in resilience measurements.

The common tendencies of SES analyses to be all-encompassing and include integrated analyses of SES systems often results in an overemphasis on institutional aspects and governance regimes and very little emphasis on ecology. To the extent that biophysical attributes are described at all in commonly used frameworks such as the institutional analysis and development, SES, and robustness frameworks, among others, attribute descriptions tend to be limited to resource unit mobility, resource system productivity, clarity of system boundaries, and size of the resource system. Furthermore, these variables are considered only as they relate to the action situation, ignoring the potential contribution of biophysical processes to the system (Epstein et al. 2013, Vogt et al. 2015). While there is a call from the social sciences to incorporate more “ecology” into SES analyses, there is simultaneously a call from the ecological and biophysical sciences to incorporate more social science and human aspects into management of and decision-making for river systems (Poff et al. 2003, Martin et al. 2015).

Here, our aim is to bridge the gaps in system-level river basin resilience assessments and offer an alternative approach that brings ecology into the forefront. We build on the principles of SES and complex adaptive systems science principles to assess adaptive governance in river restoration programs in the arid Southwest, in particular, the Colorado River basin, to answer three questions: (1) To what extent do basin-scale restoration programs contribute to system resilience of the Colorado River? (2) To what extent are the social and ecological parts aligned in these restoration programs? (3) What are the opportunities for a transformative pathway going forward? We analyze the evolution and performance of two basin-scale mitigation and restoration programs: the Upper Colorado River Endangered Fish Recovery Program (UCR-EFRP) and the Lower Colorado River Multi-Species Conservation Program (LCR-MSCP).

Adaptive river governance

Resilience is defined as the capacity of a system to absorb disturbance and reorganize while undergoing change so as to retain the same structure, function, identity, and feedbacks (Walker et al. 2004). In the SES context, management of ecosystem resilience requires the ability to observe and interpret essential processes and variables in ecosystem dynamics to develop the social capacity to respond to environmental feedback and change (Carpenter et al. 2001). Because the self-organizing properties of complex ecosystems and associated management systems seem to cause uncertainty to grow over time, understanding should be continuously updated and adjusted, and each management action should be viewed as an opportunity to learn further how to adapt to changing circumstances (Carpenter and Gunderson 2001).

A leading approach to successfully meet the challenges of SES changes is adaptive governance (Koontz et al. 2015). Adaptive governance is defined as “changing rules and norms from a static, rule-based, formal and fixed organization with clear boundaries” to a view of institutions as “more dynamic, adaptive and flexible for coping with future climatic conditions” (International Institute for Sustainable Development 2006). Governing complex adaptive ecosystems requires adaptive managers that are supported by flexible and problem-oriented organization, networks of collaboration at all levels, and leadership. Institutions governing multispecies resource commons should avoid compromising the functional aspects of the ecosystem by implementation of rule systems that maintain the diversity and sustain a multitude of species (Becker and Ostrom 1995).

In the context of river basin organizations, a diagnostic framework to analyze complex policy situations and their coevolution with ecosystem effects calls for assessing the interactions within and between the social-institutional and biophysical systems in the basin (Bouckaert et al. 2018). River basin organizations traditionally focus on holistic demand management based on integrated water resources management principles, and their institutions are clearly defined and conceptualized based on specific criteria, including the presence of international water treaties, institutionalization of cooperation, specific governance mechanisms, and a list of other factors (Schmeier et al. 2016).

The dynamic interplay of the institutional and ecological processes and their coevolution characterize the system’s state and trajectory (Fig. 1). Bouckaert et al.’s (2018) framework is a reconstitution, within a specific river basin context, of a broader framework for assessing fit between ecosystem characteristics and regime variables in terms of stocks, flows, controls, and resilience, among other factors (Young 2002). Analyses of two separate components in a complex adaptive system such as a river basin can yield insights into the factors that independently affect institutional resilience as well as ecological resilience. However, to assess system-level social-ecological resilience, the two separate components and the feedbacks between them must be looked at as a coevolving system. Bouckaert et al.’s (2018) framework (Fig. 1) breaks down each system into irreducible, complementary, and codependent components that can influence each other in nonlinear ways.

Bouckaert et al.’s (2018) framework makes parallels between corresponding social and biophysical elements by, for example, categorizing: collaboration as a social connector and water flows as a biophysical connector; structuring as social assemblage and species diversity as biophysical assemblage; leadership as social capital and material cycling as natural capital; and finally, learning as social renewal and species recruitment as biophysical renewal. They then use a rating system for social and biophysical elements to chart the trajectory of river basin governance over time, much like how the panarchy theory is used. We argue that the strength of the framework is not simply in scoring the system properties but in using variables (either qualitative or quantitative metrics, as appropriate; see Table 1 and Appendix 1) to measure social and biophysical characteristics and then rescaling to indicators to describe a state of the system across time with a clear indication of trade-offs (defined as the distance between social and biophysical indicators for that system property).

METHODS

We next describe the system architecture or system dynamics properties and start with a conceptual foundation for the specific metrics being discussed (Table 1). An assessment of each of variable is carried out by drawing from available and indicated peer-reviewed and grey literature, both historical and current, on social and ecological indicators. The system properties are broadly categorized into system architecture and system dynamics. Building on Bouckaert et al.’s (2018) study, the following system properties and constituent elements were identified:

1. System architecture: connectivity and distribution, with collaboration as a social indicator and water flows as a biophysical indicator;
2. System architecture: assemblage of elements, with structuring as a social indicator and species diversity as a biophysical indicator;
3. System dynamics: social and natural capital, with leadership as a social indicator and material cycling as a biophysical indicator;
4. System dynamics: renewal and continuation, with learning as a social indicator and species recruitment as a biophysical indicator.

System architecture

Changes in hydrological connectivity patterns will affect the water cycle and, consequently, the regulatory capacity of the river (Gao et al. 2018), making both social and hydrological connectivity significant metrics of assessment for system architecture. Dams have significantly altered natural flow dynamics, with changes in natural flood pulse dynamics having significantly altered assemblage structures of aquatic communities (Ngor et al. 2018), making both governance and species assemblage also important assessment metrics for system architecture.

Social indicators

“Connectivity” is defined as the manner by which and extent to which resources, species, or social actors disperse, migrate, or interact across ecological and social landscapes (Biggs et al. 2012). Collaboration is actualized institutionally through the degree of connectivity of all relevant stakeholders and their capacity to participate in governance processes (Bouckaert et al. 2018). Wantzen et al. (2016) proposed a notion of river culture that recognizes the intersection of hydrological, biological, and cultural uses and values of the river as a basis for preserving ecological and cultural diversity along rivers, much of which is tied to seasonal pulses in flow. The scale of the river strongly influences the river’s social role (Kondolf and Pinto 2017). In this context, the collaborative initiatives developed to balance water development with ecological concerns embody social connectivity around the use of the Colorado River.

In terms of “assemblage”, the sustainable management of freshwater resources requires a shift from conventional hierarchical models of water governance focusing on regulatory controls to hybrid governance models in which collaborative, market-based, and regulatory elements all play a role. The structuring of stakeholders’ different value positions influences the kinds of decisions that are made on various governance issues and is important in how a wide range of decision problems can be framed in terms of choices between alternative options and the development, adaptation, and refinement of such options (Lennox et al. 2011). Stakeholder participation is essential for system design (Ackoff 1974), and there are three levels at which stakeholder analysis could be conducted: rational level (who are the stakeholders and what are their perceived stakes), process level (how the organization manages stakeholder relationships), and transactional level (the set of transactions or bargains among the organization and stakeholders; A. A. Elias and R. Y. Cavana, Stakeholder Analysis for Systems Thinking and Modeling, unpublished manuscript).

Biophysical indicators

Connectivity, in a biophysical sense, is the degree to which components of a watershed are joined and interact by transport mechanisms that function across multiple spatial and temporal scales; it is determined by the characteristics of both the physical landscape and the biota of the specific system (Alexander et al. 2018). This definition reflects a systems perspective of watersheds as heterogeneous mosaics of interacting ecosystems in which variations in the duration, magnitude, frequency, timing, and stability of flows form dynamic, spatiotemporal continua of connectivity (Alexander et al. 2018).

In terms of species assemblage, ecological indicators can be used to assess the condition of the environment, to provide an early warning signal of changes in the environment, or to diagnose the cause of an environmental problem. The use of ecological indicators relies on the assumption that the presence or absence of, and fluctuations in, these indicators reflect changes taking place at various levels in an ecological hierarchy (Dale and Beyeler 2001). Among the range of hypotheses that summarize possible general responses of ecosystem processes to reductions in species richness, there is considerable variation in what minimal diversity is needed for proper ecosystem functioning, which species make significant contributions, and what the effects of changes in diversity are on ecosystem function (Lawton 1994). Indicator selection is scale dependent, and, for our purposes, assemblage indicators at the ecosystem level include species abundance, richness, evenness, diversity, and distributions as compositional elements in the ecosystem (Dale and Beyeler 2001).

System dynamics

Although it is accepted that humans are part of the environment, it is not always recognized that they perform multiple roles as coproducers of ecosystem services, as beneficiaries of those services, and through the addition of capital to realize those services (Jones et al. 2016). Ecosystem accounting approaches have tended to separate out the natural capital and human capital elements (Jones et al. 2016). Here, we look at them conjointly as key elements of system dynamics. Furthermore, actions such as constructing environmental flows that mimic pre-development river flows to conserve selected biodiversity, reserving aquatic refugia, constructing fish passages, restoring riparian vegetation to cool rivers, and so on, have been categorized as “renewal ecology” initiatives, distinct from conservation ecology or restoration ecology (Bowman et al. 2017). The coupled human-natural system that embodies aspects of renewal and transformation is, therefore, a key metric of assessment in system dynamics.

Social indicators

Collective action can often involve leaders, who have a larger role than other group members in goal establishment, logistics, coordination, effort monitoring, dispute resolution, and so on. “Leadership” is multidimensional and can vary from: passive influence to active motivation of group members; distributed across multiple individuals (polycentric) to concentrated in a single individual; based on persuasive reasoning to coercion; situational to institutional; and achieved due to past actions to ascribed based on kinship or social identity (Glowacki and von Rueden 2015). How power distribution occurs when leadership is concentrated based on an individual’s social value orientation, defined as a relatively stable preference for how valuable outcomes are distributed between oneself and others (Harrell and Simpson 2016). Social value orientations can range among: “individualists” seeking to maximize their own outcomes with little regard for the outcomes of others, “competitors” who seek to maximize the difference between their own and others’ outcomes, and “prosocials” who tend to maximize joint outcomes and to minimize differences between their own and others’ outcomes (Van Lange et al. 1997).

A “learning organization” is “an organization that is continually expanding its capacity to create its future” (Senge 1990); in other words, it is an organization that is striving for excellence through continual renewal (Hitt 1995). Learning at different scales manifests differently with the examination of operational paradigms at the system level, strategic planning exercises at organizational and program levels, and changes in behavior, attitudes, relationships, and activities at the individual level (Watts et al. 2007). Lant et al. (1992) offer a more complex explanation of organizational learning by stating that “renewal hinges not so much on noticing new conditions, but on being able to link environmental change to corporate strategy and to modify that linkage over time”. Furthermore, another shortcoming is the failure to address the fundamental tension of strategic renewal: the tension between exploration and exploitation (Crossan and Berdrow 2003). Organizational learning theory does not address new competency development while concurrently exploiting existing ones (Watts et al. 2007).

Biophysical indicators

There are several compelling reasons to consider “material or nutrient cycling” in streams. First, to the extent that nutrients are limiting in streams, they regulate rates at which important ecological processes take place, such as primary production or decomposition. Changes in these processes alter stream community structure. Second, elemental dynamics in streams link aquatic and terrestrial or riparian ecosystems, and in-stream processes are sensitive to watershed alterations (Meyer et al. 1988). Third, within-stream processes can alter the timing, magnitude, and form of elemental fluxes to downstream ecosystems, altering downstream community structures. Fourth, dissolved organic carbon dynamics have an important role to play in stream energy budgets. Fifth, many anthropogenic assaults on streams have been nutrient additions that led to major alterations of stream communities (Meyer et al. 1988). Furthermore, the intermixing of surface water and groundwater occurs at different spatial scales connected by two possible vectors that may be either groundwater flow from uplands through riparian zones to the active channel, or surface water recharging groundwater along an upstream-downstream gradient (Dahm et al. 1998).

“Species recruitment” occurs either through connectivity provision or, more predominantly, through stocking from hatcheries. Fisheries management and conservation biology have similar agendas because both seek the long-term viability of fish stocks, albeit for different reasons. Conservationists are interested in maintaining biodiversity, whereas fisheries managers are interested in maximizing productivity. Conservation biology has long emphasized the importance of practices such as environmental enrichment, pre-release training programs, and soft release to improve post-release survivorship of captive-bred animals. In contrast, the production of ecologically viable individuals is not part of the hatchery equation because the production of large quantities of fish, rather than natural history, behavior, and ecology, largely guides hatchery practices. The level of success and funding of hatchery programs is often determined by the number of fish released rather than by survival rates of those fish (Brown and Day 2002).

ASSESSMENT OF SYSTEM STATE

The assessments below are obtained from available peer-reviewed and grey literature on Colorado River management and governance structures, as well as hydrological and ecological studies to date. For each indicator, the results are described first for the UCR-EFRP and then for the LCR-MSCP. The first four subsections discuss how the social indicators, and the subsequent four subsections discuss the biophysical indicators. A map of the two cases is illustrated in Fig. 2.

Social connectivity and distribution: collaboration

The UCR-EFRP was established in 1987 after three years of discussion, data analysis, and negotiations by representatives of the U.S. Fish and Wildlife Service (USFWS); the Bureau of Reclamation; the States of Colorado, Utah, and Wyoming; and environmental and water development interests (U.S. Fish and Wildlife Service 1987). Three species, the Colorado squawfish, humpback chub, and bonytail chub, had been listed as endangered by the Secretary of the Interior under the Endangered Species Act (ESA) of 1973. A fourth species, the razorback sucker, was a candidate for Federal listing under the ESA. The recovery program was developed as part of a cooperative effort involving multiple agencies and organizations that had an interest in how the upper Colorado River basin and its resource are managed. The upper basin States have a development interest in the river’s resources, while the Bureau of Reclamation operates a number of small to large Federal reservoirs. The USFWS is responsible for administering the ESA. Water resource organizations also have a development interest that balances States’ water rights systems, interstate compacts, and fish recovery goals. A number of national and statewide conservation organizations are interested in realistic and effective fish recovery and habitat preservation (U.S. Fish and Wildlife Service 1987).

The April 2005 Record of Decision for the LCR-MSCP (U.S. Department of the Interior 2005) created an equivalent program for the lower basin, with a similar institutional structure in which the Bureau of Reclamation and USFWS are designated to act on behalf of the Secretary of the Interior to ensure compliance with the ESA, the National Environmental Policy Act of 1969, and state environmental regulations for the three lower basin States of California, Nevada, and Arizona. The program is a cooperative effort between Federal and non-Federal entities, over a 50-year period, for the purpose of conserving habitat and working toward recovery of threatened and endangered species; accommodating present water diversions and power production and optimizing opportunities for future water and power development; and providing the basis for incidental take authorizations. Other prominent Federal agencies include the Bureau of Indian Affairs, National Park Service, Bureau of Land Management, and the Western Area Power Administration. Covered actions and activities for participants occur in La Paz, Mohave, and Yuma counties in Arizona; Imperial, Riverside, and San Bernadino counties in California; and Clark County in Nevada (U.S. Department of the Interior 2005).

Social assemblage of elements: structuring

Both the UCR-EFRP and the LCR-MSCP fall under the broad purview of the Secretary of the Interior. The Implementation Committee of the UCR-EFRP was created in 1987 and was charged with overseeing the development and implementation of specific recommendations for each of the recovery elements. The committee comprises representatives from Federal agencies, the three upper basin States, water development interests, and conservation organizations. The Secretary’s ultimate responsibility is in administering the ESA without impeding States’ abilities to manage and administer their water and wildlife resources. The committee has more responsibility with management than recovery teams, which are generally biological and research-oriented groups. It provides an oversight forum for major participants. The Secretarial Observer is the liaison between the Secretary and the Implementation Committee, the Program Director provides staff assistance to the USFWS and Implementation Committee, and the management and technical groups provide assistance to the committee (U.S. Fish and Wildlife Service 1987).

The Secretary is authorized to manage and implement the LCR-MSCP in agreement with the lower basin States for providing for the use of water for habitat creation and maintenance in compliance with the ESA. Given their legal entitlements to the Colorado River water and hydropower resources, the three lower basin States, Indian Tribes, and other non-Federal interests have a vested interest in the outcome of any consultations between the Bureau of Reclamation and USFWS. The LCR-MSCP is governed by a 35-seat Steering Committee, which comprises five members: the U.S. Department of the Interior (Department of the Interior, Bureau of Reclamation, USFWS, National Park Service, Bureau of Land Management, Bureau of Indian Affairs); the lower basin States Water Resources Departments; agricultural and drainage districts; urban interests; power-generation interests; and wildlife, game, and fish departments. The Steering Committee appointed a working group to oversee the technical development of the LCR-MSCP with Steering Committee oversight.

Social capital: leadership

The Implementation Committee of the UCR-EFRP consists of representatives of major participants, including the Regional Director for Region 6 of the USFWS, the Regional Director of the Upper Colorado Region from the Bureau of Reclamation, and representatives (one each) appointed by the Governors of Colorado, Utah, and Wyoming. The Area Manager of the Western Area Power Administration is also a member because of its relationship with the Bureau of Reclamation and program revenues. Additionally, founding documents recommended including a representative of water development interests and a representative of conservation organizations. The Implementation Committee selects its own chairperson and also includes two non-voting members: one is appointed by the Secretary as an observer to provide a direct liaison between the Implementation Committee and the Secretary; the other, a Program Director, is appointed by the USFWS Regional Director, to serve as a staff person (U.S. Fish and Wildlife Service 1987). Initial funding costs were split between Federal and State governments, with the Bureau of Reclamation bearing the lion’s share of costs, followed by the USFWS. Annual base funding to date has been provided from Colorado River Storage Project hydropower revenues (Upper Colorado River Endangered Fish Recovery Program, public laws authorizing the program: https://coloradoriverrecovery.org/uc/documents/foundational-documents/).

A Memorandum of Understanding, signed in 1995 among the three lower basin States (including wildlife resource agencies) and the Department of the Interior, led to a Memorandum of Clarification in July 1996 for the development of the LCR-MSCP. The Bureau of Reclamation and USFWS are co-leads for ensuring compliance with the National Environment Policy Act of 1969. The LCR-MSCP permit applicants have applied to the USFWS for an incidental take permit, pursuant to section 10(a) of the ESA. The total cost of the program is estimated at $626,180,000 (in 2003 USD) over the 50-year period, with 50% of the costs borne by the permit applicants and 50% borne by the U.S. government. The Steering Committee meets at least once a year to work on program implementation, work plan, and budget. The Program Manager is under the supervision of the Regional Director for the Bureau of Reclamation’s Lower Colorado Region (U.S. Department of the Interior 2005).

Social renewal and continuation: learning

The Department of the Interior has clear cut procedures and documentation for the implementation of adaptive management in river restoration programs. Both the UCR-EFRP and the LCR-MSCP incorporate adaptive management in program activities and decisions. The basis for the programs rested on a “biological opinion” issued by the USFWS, and a habitat management plan was accordingly drawn. The plan has been revised annually based on available scientific research and monitoring efforts. The programs incorporate a trial-and-error approach and constantly improve and innovate based on changing conditions.

The UCR-EFRP program considered two alternatives for in-depth evaluation, including a “no action” alternative that would involve continued Section 7 consultations with basic and applied research and monitoring, over an initial 15-year period and a “proposed action” alternative that would involve habitat management, development, and maintenance; rare fish stocking, non-native fish management, and sportfishing; and research, monitoring, and data management (U.S. Fish and Wildlife Service 1987). The program has come up for extended authorization twice since 1987 and is undergoing a reevaluation of conservation priorities following a recent decision to down-list two species that have achieved sufficient levels and aim for continued, long-term management.

The LCR-MSCP program was conceptualized as a 50-year program in 2005 and has so far not undergone significant reframing of program policies and procedures. The Steering Committee commissioned two separate scientific reviews of interim conservation strategy documents during program development in 1999 and 2002. The first review was conducted by the American Institute of Biological Sciences, and the second Science Review Team comprised 6 members selected from a list of 18 active interdisciplinary scientists with a working knowledge of Southwest ecosystems. Of the three action alternatives considered, including a “no action” alternative, a combination of the two other plans was selected as the preferred alternative on the basis of it realizing the full range of environmental goals to conserve species effectively while allowing water use under existing entitlements. This alternative has been the basis of yearly plans and progress to date.

Biophysical connectivity and distribution: water flows

The upper Colorado River and its principal upper basin tributary, the Gunnison River, have their headwaters in the Rocky Mountains in central Colorado. The Yampa River and White River, major tributaries of the Green River, likewise have their sources in the Rocky Mountains. The annual hydrographs of these rivers are dominated by snowmelt runoff, which usually begins in late April, reaches a peak in May or early June, and recedes through July. Summer thunderstorms are common and can cause localized flooding on tributaries and increased turbidity on the larger rivers for days, but they do not have a significant effect on main stem discharges (Van Steeter and Pitlick 1998).

Natural streamflows of the Colorado and Gunnison rivers are affected by many diversions and dams. Collectively, the reservoirs upstream of the Flaming Gorge and Glen Canyon dams store only approximately 10% of the total volume of water in Lake Powell. However, these reservoirs are near the source of the runoff and thus alter the annual hydrograph significantly. Composite records indicate that in the post-development period (1950–1995), annual peak discharges of the Colorado River at Glenwood Springs have averaged 286 m³/s, which represents a 43% decrease relative to the pre-development period (1900–1949) average of 504 m³/s. The effects of reservoirs and transbasin diversions in the upper Colorado River basin diminish downstream because of added flow from unregulated tributaries (Van Steeter and Pitlick 1998).

Since the Glen Canyon dam first began to store water in 1963, creating Lake Powell, some 430 km (270 miles) of the Colorado River, including Grand Canyon National Park, have been virtually bereft of seasonal floods. Before 1963, melting snow in the upper basin produced an average peak discharge exceeding 2400 m³/s. After the dam was constructed, releases were maintained at < 500 m³/s. The dam has also trapped > 95% of the sediment moving down the Colorado River in Lake Powell (Poff et al. 1997). The resultant changes in flow regime and withholding of sediment have induced drastic changes in the downstream Colorado River.

Flows of the lower Colorado River historically displayed tremendous annual variability. Prior to major flow regulation imposed by construction of the Hoover Dam in 1936, instantaneous peak discharges as high as 8500 m³/s and as low as 0 m³/s were recorded below Yuma, Arizona (Sykes 1937). Pre-regulation photographs of the Colorado River on its delta and historical accounts depict a highly sinuous channel with a broad flood plain harboring a diverse assemblage of lotic and lentic habitats, with fine sand and silt dominating sediments (Sykes 1937). The most notable flood events in the lower Colorado River occurred in the mid-1980s and the early and late 1990s (Tiegs and Pohl 2005). These floods rehabilitated much of the riparian vegetation in the delta that was lost as a consequence of flow regulation. The flood regime of the contemporary Colorado River at its delta is event-based, and floods are often associated with the El Niño phenomenon (Glenn et al. 1996). Observations based on these events culminated in the implementation of Minute 319 in 2014, which released 130 million m³ of water in a pulse flow from Lake Mead to rejuvenate riparian ecosystems and support species in the delta region (Flessa et al. 2013).

Biophysical assemblage of elements: species diversity (aquatic and riparian)

The Colorado River mainstem fish community historically comprised ten freshwater species, of which seven are currently federally listed as endangered and one is of special concern. Of these latter species, Colorado pikeminnow (Ptychocheilus lucius), bonytail (Gila elegans), and razorback sucker (Xyrauchen texanus) were widely distributed throughout the mainstem of the river and have been the subject of various management actions for more than three decades (Mueller 2005). European settlement brought dramatic biological and physical changes through the introduction of channel catfish (Ictalurus punctatus), carp (Cyprinus carpio), largemouth bass (Micropterus salmoides), bluegill (Lepomis macrochirus), and several other species (Dill 1944). Hoover Dam construction in 1935 greatly altered the physical conditions, which benefitted invasive species. The reservoirs and their tailwaters were stocked with recreational species, and after World War II, an estimated > 80 fish species, the majority of which were aggressive predators, were stocked (Mueller and Marsh 2002).

Martinez et al. (1994) investigated the effects of the completion of Taylor Draw Dam in 1984 on the White River, the last significant free-flowing tributary in the upper Colorado River, which formed Kenney Reservoir. Fishes were sampled above and below the dam axis prior to closure of the dam and in the reservoir and river downstream following impoundment. They found that while the immediate effects of the dam to ichthyofauna included blockage of upstream migration to 80 km of documented range for the endangered Colorado pikeminnow, the reservoir also proved to have profound delayed effects on the river’s species composition (Martinez et al. 1994). Pre-impoundment investigations in 1983–1984 showed strong domination by native species above, within, and below the reservoir basin. By 1989–1990, non-native species comprised roughly 90% of fishes collected in the reservoir and 80% of fishes collected in the river below the dam. The key take-away from the study was that smaller scale mainstem impoundments that do not radically alter hydrological or thermal regimes can still have profound effects on native ichthyofauna by facilitating the establishment and proliferation of nonnative species.

Native fish in the lower mainstem of the river had become rare by the mid-1930s due to a combination of predation and habitat destruction (Dill 1944). Numbers of razorback sucker and, to a lesser extent, bonytail rebounded when lakes Mead, Roosevelt, and Mohave formed (Minckley 1983). Colorado pikeminnow were extirpated from the lower basin by 1975, but small populations persist in the upper basin. Bonytail and razorback sucker have experienced recruitment failure for nearly four decades. Wild bonytail are believed to be gone, with the last one captured from Lake Mohave during the late 1990s. Estimates of wild razorback sucker dropped to < 1000 individuals: approximately 100 in Green River, 300 in Lake Mead, and 500 in Lake Mohave. The USFWS attempted a stocking of > 12 million razorback sucker fry from 1981–1991 in an attempt to reestablish the species in Arizona and avoid federal listing (Johnson 1981). However, survival was extremely poor, as < 200 of these fish were ever captured (Minckley et al. 1991). It was found that following initial releases, razorback suckers were lost to resident catfish within a matter of hours (Marsh and Brooks 1989). This discovery led to the growing realization that predator and invasive species control needed to be adopted as basin-wide strategies.

Tamarisk (salt cedar) is a shrub that was introduced into the American Southwest in the late 1800s and has spread throughout the Colorado Plateau by occupying islands, sand bars, and beaches along streams. Historical photographs show that tamarisk spread from northern Arizona to the upper reaches of the Colorado and Green rivers at a rate of approximately 20 km/yr (Graf 1978). Tamarisk has a reputation for having negative effects on riparian ecosystem structure and processes, including high water use compared to native plants (Taghvaeian et al. 2014), increased soil salinization (Ohrtman et al. 2012), displacement of native vegetation (Glenn and Nagler 2005), changing erosion and sedimentation regimes (Vincent et al. 2009), increased fire frequency (Busch and Smith 1993), reduced biodiversity (Harms and Hiebert 2006), and reduced habitat quality for wildlife (Bailey et al. 2001, Hinojosa-Huerta et al. 2013). Zavaleta (2000) reported that the negative effects of tamarisk water consumption on agricultural and municipal water supplies, hydropower generation, and flood control reach an annual value of $285 million USD.

Tamarisk control attempts have had varied success, with control strategies such as mechanical removal, fire, and herbicide treatments, although these have proved costly and had negative effects on the native plant and soil communities (Hultine et al. 2010). A biological control program led to the selection and approval of two insects in 1994 for tamarisk control: the saltcedar leaf beetle from central Asia, and a mealy bug from the Middle East. In 1999, populations of tamarisk beetles were imported into the United States (Dudley and Deloach 2004). However, in the early 1990s, it was determined that the endangered Southwestern Willow Flycatcher (Empidonax traillii extimus) was nesting in tamarisk habitat and showed a distinct preference for disturbed habitat (Hultine et al. 2010). Given its current preference for tamarisk habitat, replacement of tamarisk by native vegetation may negatively affect Southwestern Willow Flycatcher conservation efforts. Furthermore, the rate of defloration of tamarisk has far exceeded the rate at which native vegetation is regenerating (Dudley and Bean 2012), resulting in large areas showing lower total foliar cover, which has implications for vital ecosystems processes such as nutrient cycling (Hultine et al. 2010).

Natural capital: material cycling

Floodplain nutrient cycling is often measured by following leaf litter production and decomposition. Although floodplain nutrient cycling is relatively well understood in mesic regions, decomposition patterns on dryland floodplains are complex and not well studied. Before the construction of Hoover and Glen Canyon dams in 1935 and 1964, respectively, discharges to the delta reached an estimated 6000 m³/s, and the delta occupied 780,000 ha (Glenn et al. 1996). Post-dam construction, practically no water flowed into the Gulf of California, and sediment supply to the basin has been held in the upstream reservoirs. High precipitation-induced water releases in the early 1980s and early and late 1990s allowed sufficient water to reach the delta, but the water was polluted by agricultural and municipal water returns (Glenn et al. 2001). High concentrations of selenium and organochlorine pesticides have been observed in the biota of the estuary, exceeding the toxic threshold in 23% and 30% of biota, respectively (García-Hernández et al. 2001). Part of the estuarine basin of the Colorado River is now regarded as the agricultural “sewer” of the Wellton-Mohawk irrigational system of Imperial Valley (Carriquiry et al. 2011).

Miller (2012) studied longitudinal patterns in dissolved organic carbon loads and chemical quality in the Colorado River from the headwaters in the Rocky Mountains to the United States–Mexico border from 1994–2011. His findings reveal that a shift from the historically snowmelt-driven Colorado River to a heavily regulated system dominated by storage levels in Lake Powell has coincided with a shift from a net increase to a net decrease in dissolved organic carbon loads (Miller 2012). This hydrological shift has also resulted in a geopolitical shift in defining the Colorado River reaches 1 through 5 as being in the upper basin and reaches 6 through 10 as being in the lower basin. The study revealed that net dissolved organic carbon input in the upper basin was greater than the net loss, whereas the reverse was true for the lower basin. The average annual discharge and dissolved organic carbon loads in the upper basin increased from reach 1 to reach 5 by 6.6 × 10⁹ m³/yr and 2.7 × 10⁷ kg/yr, respectively. Increased dam storage in the lower basin resulted in average basin-scale loss of 4.4 × 10⁹ m³/yr of water, in part, to evaporation, and a corresponding decrease in average dissolved organic carbon load to 2.2 × 10⁷ kg/yr.

No reliable data are available for nutrient fluxes in the Colorado River across the international border prior to dam construction. However, nitrogen cycling and phosphate precipitation in U.S. reservoirs resulted in removal of nutrients (Daesslé et al. 2017). Despite this removal, the upper Gulf of California is still considered an important area for marine primary productivity, with peak chlorophyll-a concentrations of 18.2 mg/m³ (Millán-Nuñez et al. 1999). The upper Gulf of California hosts important fisheries of shrimp, shark, and sea bass (Galindo-Bect et al. 2000), and the rich coastal productivity is assumed to remain sustained by the addition of nutrients via sediment resuspension and surface–groundwater input from agricultural runoff, associated wetlands, recycling of nutrients in the water column, and input from the Gulf of California (Millán-Nuñez et al. 1999). In the Mexicali Valley, fresh surface water is limited to irrigation and drainage channels. A few wetlands are supported by drainage and wastewater flows, including the Cienega de Santa Clara. However, the main riverbed remains mostly dry in its course along the estuary (Daesslé et al. 2017).

Biophysical renewal and continuation: species recruitment

The sheer amount of grey infrastructure in the mainstem of the Colorado River has created insurmountable barriers for migratory fish such as Colorado pikeminnow to migrate upstream and spawn. The Hoover and Glen Canyon dams have cut off the possibility of integrated management of upper and lower basin species. The upper basin offers more opportunities for species recruitment because of the presence of unregulated tributaries that serve as refugia for endangered native fish species. The lower basin has seen the sharpest declines and near extermination of Colorado pikeminnow, as well as razorback sucker and bonytail, both of which saw some resurgence with the creation of lower basin lakes and reservoirs.

The current restoration strategy is to limit and prevent movement of invasive game fishes out of impoundments and to curtail future stocking by enacting public education programs. Increasing harvests of carp and channel catfish are also promoted. Because of the recreational values put on sport fishing, these measures of non-native fish control have faced challenges in implementation. Non-native sport fishes continue to proliferate by reproducing in river channels and invading from off-channel habitat (Tyus and Saunders 2000). Additionally, small, nongame fishes such as the red shiner and fathead minnow, which were unintentionally introduced, have proven to be aggressive, abundant, and widely distributed, constituting > 90% of the standing crop of fishes in backwater habitat used as nursery areas by the listed fishes (McAda et al. 1994).

Two resource philosophies evolved in the Colorado River basin in the late 1980s: (1) establishment of the Upper Colorado River Basin Recovery Implementation Plan in 1987, and (2) a conservation movement to actively manage two endangered species in the lower basin, which began in 1989 (Mueller 1995) and later became the LCR-MSCP. In the upper basin, a consortium of resource agencies and water users came together to establish a recovery regime that would occur in 15 years in conjunction with continued water development. Initial recovery centered on habitat restoration, including the restoration of historical flow regimes that had been disrupted by reservoir storage. Since 1990, emphasis has shifted toward restoring floodplain wetlands and predator removal and control (Lentsch et al. 1996, Wydoski and Wick 1998).

Both razorback sucker and bonytail established impressive communities when several reservoirs filled in the lower basin, with the razorback population in Lake Mohave swelling to > 100,000, while bonytail were less numerous. However, these increases occurred before the introduction of non-native species. Bonytail became extremely rare by the early 1980s. Stocking of bonytail in Lake Mohave began in 1980 (Minckley and Thorson 2004), and a similar stocking effort for the razorback sucker began in 1989 (Mueller 1995). The approach involved capturing wild larvae and rearing them to a size large enough to avoid predation. The goal was to capture genetic variability that would have been lost in hatchery production (Mueller 2005). Because of the short supply of hatchery rearing space, fish were reared in municipal ponds, isolated reservoir coves, and backwaters blocked by nets (Mueller 1995). The concept expanded to other reaches of the lower river under the LCR-MSCP.

RESULTS

Based on the detailed analysis of various social and biophysical indicators in the previous section, we scored the indicators using ratings on a scale of 1–5. The scoring was undertaken for both the upper and lower basin restoration programs (Table 2; see also Appendix 1). The indicators are rated on a scale of 1–5, with the ratings representing a gradient or degree of variation of the indicator based on available literature reviews.

DISCUSSION

The discussion is organized into two subsections that discuss the UCR-EFRP and LCR-MSCP, respectively, and link back to the research questions asked pertaining to the extent to which these programs contribute to system resilience as well as how the social and ecological parts are aligned in these programs. The subsequent conclusion discusses transformative pathways forward in the context of the current scenario of a Federal water shortage declaration on the Colorado River.

The Upper Colorado River Endangered Fish Recovery Program

The UCR-EFRP has performed better in terms of contributing to social and institutional resilience than biophysical resilience (Fig. 3). The collaboration as well as system structuring of the program around the goals of habitat restoration and native species conservation is strong, though the biophysical system shows moderate amounts of connectivity owing to the presence of significant dams and reservoirs, including Flaming Gorge Dam, the Aspinall Unit, Navajo Reservoir, and Glen Canyon Dam. These structures have affected species diversity by resulting in the development of a two-state system: recreational values have resulted in invasive species dominating reservoir systems, whereas downstream areas are dominated by native species at risk of invasive species encroachment, thereby requiring long-term management to ensure native species preservation.

Furthermore, a number of tributaries in the upper basin are relatively undeveloped and more free-flowing and provide crucial refugia for migratory species spawning. This situation is a principal reason why natural capital cycling and ecosystem processes flourish better further downstream of the big reservoirs. The social capital is also strong, and because of the implementation of strong adaptive management processes, the feedback between the social and biophysical elements appears to be strong. Although the initial “biological opinion” appeared to have limited options scientifically for restoration scenarios, there has been significant improvement in scientific studies over the decades since program inception and a strong public awareness campaign to recruit public support for achieving program goals through invasive species control.

System renewal in terms of institutional learning is lower rated because the predominant paradigm of the program is mitigation rather than full-scale restoration. So, although significant financial capital has been invested in building infrastructure to facilitate migratory fish passages and stocking programs to maintain native fish populations, the overdependence on technological solutions in a highly engineered system is less conducive to incorporating actual biophysical feedbacks and responding to them. Despite the high levels of infrastructure, the system still retains some biophysical resilience because of flow control and recruitment facilitation as well as large stretches of undeveloped tributary systems.

The Lower Colorado River Multi-Species Conservation Program

The LCR-MSCP is less resilient overall and shows significant systemic vulnerabilities (Fig. 4), partly because of the way the program is structured and partly because of the context in which the program is situated. The lower basin mostly comprises the main stem of the river, with the tributaries also being in a greater state of development as compared to the upper basin. The upper and lower basins are divided by the insurmountable barriers of the Hoover and Glen Canyon dams, and there are large and small dams and diversions at various reaches all the way to the Mexican border. The amount of flow regulation is significant, which creates a highly disconnected system at a biophysical level. Therefore, even though the collaboration based on economic and regulatory incentives is strong (though not all-inclusive, with Indian Tribe involvement being minimal to insignificant), the prevalence of a feedback between social and ecological elements is broken.

The way adaptive management is structured in the lower basin is geared toward establishing isolated backwaters and aquatic landscapes for native species to flourish and managing invasive species. Therefore, although species diversity may exist in these isolated pockets, it is being propped up by technological interventions such as hatchery breeding and stocking programs. Institutional structuring in the lower basin is also heavily geared toward mitigating the effects of take by senior water-rights holders at various scales from state to local. Therefore, both social and biophysical structuring elements are a lot weaker here than in the upper basin.

The institutional structuring is also the reason for social capital being more concentrated in a hierarchical structure to allow for water allocation goals to be met. This situation creates a significant disruption in ecosystem processes because over-allocations, pollution from overuse, and impending drought have combined to create a heavily degraded system at a biophysical level. The adaptive capacity of the institutions is also more toward meeting compliance requirements. Because of the high barriers to species recruitment and the creation of a two-state system as in the upper basin, the capacity for the system to renew itself is heavily compromised.

Transformational pathways

The ESA has been the principal driver of actions for aquatic species protections in the upper and lower basins. The main mechanism of ESA implementation is the placement of individual species on an official list as either threatened or endangered, triggering species-specific actions. From the perspective of resilience theory, however, the ESA has several limitations, most of which point to a lack of system focus, very little emphasis on overall functionality of ecological systems, as well as being reactive rather than proactive in species protection (Benson 2012). Furthermore, Federal statues pertaining to restoration are vague and inadequate and do not provide any guidelines or practices to approach restoration from a systems perspective (Palmer and Ruhl 2015).

A systems-based approach to transformative stewardship of the Colorado River (and other heavily developed arid rivers) would include engineering both natural and built infrastructure for optimal river and riparian ecological health (Poff et al. 2016). There are increasing calls for supplementing or even replacing gray infrastructure with green infrastructure and nature-based solutions that can not only preserve biophysical processes but also ensure water security (Palmer et al. 2015). These solutions require the extensive engagement of urban, agricultural, and industrial actors to strategically apply green infrastructure solutions that contribute to river health and sustainability.

CONCLUSION

We have presented the case of two sub-basin-scale river recovery programs in the Colorado River basin using a diagnostic framework for river basin organization governance and stewardship. Our findings indicate that the upper basin has performed comparatively better than the lower basin because the ecological connectivity is higher, allowing for healthier levels of ecological dynamics. Socially, the system structuring is strong in the upper basin, though the social system dynamics in both basins are relatively weak, which might be because of infrastructural lock-in that limits the actors and solutions that could be used. The transformative pathways to improve river health need a systems-based approach to legal, policy, and social action that incorporate the principles of resilience and adaptive management. Srinivasan et al. (2021) point to pathways to governance for anticipatory and adaptive resilience, and Srinivasan et al. (2022) also highlight the need for planning for altered river futures because the biophysical capacity of the system to respond meaningfully has been reached, and social system coevolution to states of uncertainty and risk spreading (Folke 2003) must occur.

This is an intensive place-based study of river stewardship at sub-basin levels for a heavily developed arid river system. As such, it represents a first step toward developing archetypes of river systems. While the generalizability and transferability of such place-based research is limited, recurrent but nonuniversal patterns can hold well for geographically distant systems (Václavík et al. 2016, Eisenack et al. 2019, Oberlack et al. 2019). We contend that the Colorado River is archetypally similar to the Murray-Darling River in Australia, for instance, as both have been the subject of multiple comparative studies (Ladson and Argent 2002, Grafton et al. 2012, Wheeler et al. 2018), as well as to the Sacramento and San Joaquin rivers (Barnett et al. 2004, Foti et al. 2014). In a scenario in which global rivers are drying up, in the words of Richter and Postel (2004), the preservation of ecosystem health must become the explicit goal of water development and management.

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