Stream and Watershed Restoration: A Guide to Restoring Riverine Processes and Habitats

1

Introduction to Restoration: Key Steps for Designing Effective Programs and Projects

Philip Roni & Tim Beechie

Northwest Fisheries Science Center, National Oceanic and Atmospheric Administration, USA

1.1 Introduction

The restoration of streams, rivers, and watersheds has become a growth industry in North America and Europe in the 21st century, with an estimated $1 billion spent annually in the United States alone (Bernhardt et al. 2005). This comes with a growing appreciation from the general public of the importance of water, watersheds, and natural places not only for their wildlife and fisheries, but also for social, cultural, economic, and spiritual reasons. With this increased emphasis on restoration has come the need for new techniques and guidance for assessing stream and watershed conditions, identifying factors degrading aquatic habitats, selecting appropriate restoration actions, and monitoring and evaluating restoration actions at appropriate scales. All these require detailed consideration of not only the latest scientific information but also regulations and socioeconomic constraints at local, regional, and national levels. Thus the challenges facing watershed restoration in the 21st century are multifaceted, including both technical and non-technical issues.

As interest in aquatic restoration has increased, several texts have been produced over the last few decades to assist with various aspects of river restoration. Most have focused on habitat improvement techniques specific to trout and salmon (e.g. Hunter 1991; Mills 1991; Hunt 1993; O’Grady 2006) or design considerations for specific techniques (e.g. Brookes & Shields 1996; Slaney & Zoldakis 1997; RRC 2002). A few have provided more comprehensive regional overviews of riverine restoration planning and techniques (Ward et al. 1994 in UK; Cowx & Welcomme 1998 in Europe; FISRWG 1998 in USA; CIRF 2006 in Italy). Still others have published overviews of key concepts and principles (e.g. Brierley & Fryirs 2008; Clewell & Aronson 2008). Collectively these publications cover many of the tools, techniques, and concepts needed for restoration planning, but no single book covers the full restoration process from initial assessment to monitoring of results and adaptive management. In this book, we strive to meet the need for a comprehensive guide and educational tool that covers the key steps in this process and provide a text that links watershed assessment and problem identification to identification of appropriate restoration measures, project selection, prioritization, project implementation, and effectiveness monitoring (Figure 1.1). Each of these steps is discussed in detail in subsequent chapters. In addition, we discuss the human dimension and how one can best work with citizens, government bodies, and private companies to develop restoration projects and goals. In this introductory chapter we provide important background on the need for restoration, its relatively short history, and the major steps and considerations for planning and implementing restoration actions.

Figure 1.1 Major steps in the restoration process required to develop a comprehensive restoration program and well-designed restoration projects.

1.2 What Is Restoration?

Restoration ecology is a relatively young field with considerable confusion over its terminology (Buijse et al. 2002; Omerod 2004; Young et al. 2005). The terms restoration, rehabilitation, enhancement, improvement, mitigation, reclamation, full and partial restoration, passive and active restoration, and others have been used to describe various activities meant to restore ecological processes or improve aquatic habitats (Table 1.1). These represent a gradient of activities from creating new habitats, to mitigating for lost habitat, to full restoration of ecosystem processes and functions and even protection. In practice, the term restoration is used to refer to any of the above activities. To avoid further confusion over terminology, we therefore use the term in this sense throughout this text. Where appropriate, we distinguish between full restoration, partial restoration and habitat improvement or creation (Table 1.1).

Table 1.1 Commonly used restoration terminology and general definitions. In this book and in practice, the term restoration is used to encompass all these activities with the exception of protection and mitigation. Where appropriate, we distinguish between restoration in its strictest sense (full restoration), rehabilitation (partial restoration), and habitat improvement or creation.

Modified from Roni (2005), Roni et al. (2005), and Beechie et al. (2010).

We focus most of our discussion on ‘active restoration,’ which are restoration efforts that take on the ground action to restore or improve conditions. However, regulations, laws, land-use practices, and other forms of ‘passive restoration’ that eliminate or prevent human disturbance or impacts to allow recovery of the environment are equally important. For example, most of the improvements in water quality and habitat condition in the USA, Europe, and elsewhere would not have occurred without legislation and regulation. Similarly, habitat protection, while not typically included in definitions of restoration, is a critical watershed conservation and restoration strategy that should not be overlooked. Given the continued pressure on aquatic ecosystems, including a growing human population and climate change, habitat loss will continue and even outpace restoration efforts unless protection of high-quality functioning habitats is a high-priority component of restoration plans. In fact, habitat protection in many cases is a type of passive restoration that allows ecosystems to recover following disturbance. Ultimately, it is much more cost-effective to protect functioning habitats from degradation than it is to try to restore them once they have been damaged.

1.3 Why Is Restoration Needed?

It may seem obvious to people living in densely populated and developed areas why one might seek to restore streams or watersheds, but the level of human impact and the reasons for restoration vary widely among stream reaches, watersheds, regions, and countries. Human impacts to watersheds began well before recorded history. Archeological evidence indicates that localized deforestation and subsequent impacts to watersheds occurred in populated areas throughout the world even prior to 1000 BC (Williams 2001). For example, forest removal or conversion to agricultural lands occurred in the Mesolithic and Neolithic periods (c. 9000–3000 BC) in parts of Greece and Britain (van Andel et al. 1990; Brown 2002). Deforestation expanded during both the Bronze and Iron Age (c. 3000 BC to 500 AD) when metal tools replaced stone tools and made clearing of forests and plowing of lands easier. Extensive hillslope erosion and subsequent sedimentation and aggradation of river valleys in Greece and other areas in the eastern Mediterranean is attributed to deforestation and intensive agriculture during the Bronze Age (van Andel et al. 1990; Montgomery 2007). This was followed by diversion of rivers, draining wetlands, and harnessing waterpower in some areas of Europe and the Mediterranean with the rise of the Roman Empire (Cowx & Welcomme 1998). Deforestation, which often leads to increased silt loads, expanded rapidly during the Middle Ages not only in Europe but also in China and elsewhere, resulting in filling of coastal and low-lying areas and presumably other impacts to streams. During medieval times and through the Renaissance (c. 1000 to 1700 AD), extensive deforestation and conversion of lands to agriculture in Europe and the Mediterranean were common (Cowx & Welcomme 1998; Williams 2001). This occurred somewhat later in the New World and elsewhere following European colonization. More dramatic changes to rivers and watersheds occurred during the Industrial Revolution, as construction of dams and weirs to power industry and rapid industrialization caused the pollution of many waters. In parts of Europe, the mass production of drainage tiles and other technologies led to the drainage and conversion of vast wetlands to agricultural land (Vought & Lacoursière 2006). Increasing urban and agricultural activities resulted in some local channelization of rivers and streams. The combination of migration barriers (dams) and pollution due to industry and the rapidly growing human population led to the decline of several migratory fishes in Europe and eastern North America.

The most severe impacts to aquatic systems in North America, Europe and elsewhere arguably occurred in the late 19th and during the 20th century. Increasingly mechanized societies channelized and dredged rivers, drained wetlands, cut down entire forests, intensified agriculture, and built dams for power, irrigation, and flood control. In the UK, Ireland, Europe, the USA, and elsewhere, large river channelization and wetland drainage programs occurred from the early part of the 20th century up until the 1970s (Cowx & Welcomme 1998; O’Grady 2006). This history of land and water uses along with other human activities produced the degraded conditions we see on the landscape today. For example, it is estimated that worldwide over 50% of wetlands may have been lost (Goudie 2006). Coastal wetland loss in some US states and Europe countries exceed 80% (Dahl & Allord 1999; Airoldi & Beck 2007). Estimates suggest that globally more than 75% of riverine habitats are degraded (Benke 1990; Dynesius & Nelsson 1994; Muhar et al. 2000; Vörösmarty et al. 2010).

The above factors, coupled with an increasing human population, have led to increased air pollution, highly modified and polluted rivers, and a rapid increase in number of threatened, endangered, or extinct species (Figure 1.2; Goudie 2006). The World Water Council estimates that more than half the world’s rivers are polluted or at risk of running dry, and less than 20% of the world’s freshwaters are considered pristine (World Water Council 2000; UN Water 2009). Moreover, 80% of human water supplies are threatened by watershed disturbance, pollution, water resource development or other factors (Vörösmarty et al. 2010). As recently as 2004, 44% of the stream miles in the USA were considered too polluted to support fishing or swimming (EPA 2009). Current species extinction rates are estimated to be more than 100–1000 times background (prehistoric) rates (Baillie et al. 2004), and some studies suggest that modern rates are more than 25,000 times background rates (Wilson 1992). Extinction rates for freshwater fauna are thought to be 4–5 times that of terrestrial species (Ricciardi & Rasmussen 1999), and habitat loss and degradation are believed to be the primary cause of extinctions (Baillie et al. 2004). A suite of human activities has led to degradation of streams and watersheds and impaired their use for biota (including humans), and therefore stream and watershed restoration has become critically important worldwide.

Figure 1.2 Increase in selected human impacts during the last 300 years (percent increased compared to 10,000 BP).

From Goudie (2006). Reproduced by permission of John Wiley & Sons.

1.4 History of the Environmental Movement

The rapid modification of our natural environment was recognized centuries ago. Limited protection of forests for hunting and timber production occurred in the ancient times, middle ages (c. 500–1500 AD), and the early modern period (c. 1500–1800 AD). Ancient empires such as Assyria, Babylon, and Persia set aside hunting reserves and the Roman Empire set up a system of protected areas for wildlife (Brockington et al. 2008). The Emperor Hadrian set half of Mount Lebanon aside in the 2nd century AD to protect cedar forests (Brockington et al. 2008). As early as the 11th century in Scotland and 13th century in England, laws and fishing seasons were set to protect salmon (Montgomery 2003). However, large-scale environmental movements did not start until the late 19th and early 20th century in the UK, Europe, the USA, Australia, New Zealand and elsewhere (Hutton & Connors 1999). The late 1800s saw the establishment of some of the first national parks such as Yellowstone National Park in the USA, Rocky Mountain National Park in Canada, and Royal National Park in Australia. During the same period, the Audubon Society, the Sierra Club, the Wilderness Society in America, and the Royal Society for Protection of Birds in the UK were formed and began pushing for greater protection of wild lands and wildlife.

The modern environmental movement began in the 1960s, initially focusing on water and air quality issues. In the USA, key publications on increasing environmental problems such as Rachel Carson’s Silent Spring (Carson 1962) and a series of environmental disasters led to a large environmental movement and a series of laws to protect the environment in the 1960s and 1970s. These laws included the Wilderness Act (1964), the National Environmental Policy Act (1969), the Clean Air Act (1970), the Water Pollution Control Act (1972), and the Endangered Species Act (1973). Similar legislation was passed in the 1970s, 1980s, and 1990s in other industrialized countries (e.g. German Federal Nature Conservation Act 1976, Swiss Environmental Protection Law 1983, Canadian Fisheries Act 1985, Canadian Water Act 1985, Japanese Act on Conservation of Endangered Species of Wild Fauna and Flora 1992, Australian Endangered Species Protection Act 1992). In 2000, the European Union (EU) passed the Water Framework Directive (WFD), arguably the most sweeping legislation for the protection and restoration of watersheds and aquatic biota. The WFD combined with other EU Directives for the conservation of nature and biodiversity such as the Birds Directive (79/409/EEC) and the Habitats Directive (92/43/EEC) provide a legal basis to implement comprehensive, interdisciplinary basin-wide restoration programs.

Another key environmental aspect is the importance and economic value of ecosystem goods and services. Until recently the value of ecosystems was only based on the goods they might produce (e.g. harvestable fish, food, timber), but in recent years the services or benefits we derive directly or indirectly from ecosystem functions have also been recognized. These other services include waste processing, carbon sequestering, regulation of atmos­pheric gases, water regulation, climate regulation, genetic resources, and many others (Costanza et al. 1997; Cunningham 2002). In fact, the economic value of ecosystem services globally has been estimated to be 2–3 times that of the total global gross domestic product from world economies (Costanza et al. 1997). This realization of the importance of functioning ecosystems for our economic prosperity and our very existence has led to further emphasis on protecting and restoring natural ecosystems globally.

1.5 History of Stream and Watershed Restoration

Similar to the environmental movement, the earliest stream restoration efforts were largely undertaken by hunters and fishermen. While efforts to minimize erosion and protect water supplies and agricultural land date back thousands of years (Riley 1998), the first substantial efforts to restore streams are thought to have been made in the late 1800s by local fishing clubs in the USA and river keepers on British estates interested in improving salmon or trout fishing (Thompson & Stull 2002; White 2002). As early as 1885, Van Cleef called for the restoration and protection of trout streams in the Eastern USA (Van Cleef 1885). There is also evidence of early restoration efforts in Germany and Norway (Walter 1912; Thompson & Stull 2002). These early efforts often included stocking of fish and killing of predatory birds, fish and mammals, actions that today would be frowned upon (White 2002).

More formalized efforts to restore streams were undertaken in the USA in the early part of the 20th century (Thompson & Stull 2002). The Civilian Conservation Corps and some smaller state-sponsored stream and land restoration programs began implementing restoration projects on miles of small streams in the Midwest, Rocky Mountains and elsewhere during the Great Depression, partly to combat soil and bank erosion. These efforts tended to focus on planting trees, fencing out livestock, bank protection and stabilization, installing small log structures or weirs to create pools, and even excavation of pools. The latter three techniques were largely engineering approaches attempting to create pool habitat or a static stream channel, and often treated symptoms (lack of pools) rather than underlying problems (e.g. excess sediment, lack of riparian vegetation and woody debris) (White 1996; Riley 1998). It is however important to remember that, during this period, streams were highly degraded from decades of severe overgrazing and removal of streamside vegetation and it was not yet fully understood how quickly riparian banks and vegetation might recover once they were protected (White 2002). The 1940s and 1950s witnessed an increased emphasis on planting of vegetation to stabilize banks; however, these efforts were often not viewed as favorably as instream structures and hardening of banks, which were seen as quicker and more permanent (White 1996). Both before and after World War II in Europe there were efforts to stabilize banks using plantings and bioengineering approaches, but again these were largely to create static channels and prevent streams from moving.

Expansion of state and federal stream restoration programs in the USA continued from the 1950s through the 1980s. Following years of overgrazing and other human activities, riparian vegetation began to recover along numerous streams in the USA and Canada (White 2002). During this period, there was also an increased focus on placement of log and boulder cover structures, based largely on promising results from trout stream restoration in Wisconsin and Michigan. However, these structural techniques were largely pioneered in low-energy Eastern and Midwestern streams and met with mixed results when applied elsewhere, particularly in higher-gradient higher-energy streams of the mountainous western North America. Several of these techniques were subsequently applied in European streams in the 1980s and 1990s with varying degrees of success. Despite the emphasis on structural treatments, the key stream restoration manual (White & Brynildson 1967) recommended protecting riparian vegetation before installing instream structures. Unfortunately, this sage advice was largely ignored until recently when the importance of watershed processes became more widely accepted (Chovanec et al. 2000; Hillman & Brierely 2005; Beechie et al. 2010). Fortunately, as early as the 1960s some states were acquiring land along streams to let riparian vegetation and streams recover naturally. There was also an increasing understanding of riverine processes – partly based on Leopold et al. (1964) – which biologists were attempting to incorporate into stream restoration projects.

The late 1980s and early 1990s saw rising awareness in the importance of riparian areas, the physical and ecological importance of large wood, and a better understanding of physical and biological processes and how land use and human activities impact those processes and fish habitat (White 2002). This was initially based on extensive studies on forested streams in the Pacific Northwest of North America, but was later based on studies in a range of land uses and ecoregions. The results of these studies led to recommendations for a watershed or ecosystem approach to management and a growing call for looking beyond an individual stream reach when planning restoration (Beechie & Bolton 1999; Roni et al. 2002; Hillman & Brierely 2005). From the 1990s until today, restoration efforts have slowly been changing from a focus on localized habitat improvement actions at a site or reach scale (which often overlooked the root causes of habitat degradation) to a more holistic watershed or ecosystem approach which tries to treat the underlying problem that has led to the habitat degradation (to be discussed in great detail in the following chapters). This is not to say that certain habitat improvement techniques are not widely used or are ineffective, but rather that greater emphasis has been placed on restoring whole watersheds through improving land use, reducing sediment sources, protecting riparian areas, and other restoration efforts focused on restoring the processes that create and maintain stream habitats and health.

European river restoration efforts largely began in the 1980s and increased dramatically during the 1990s (Cowx & Welcomme 1998), focusing mostly on rehabilitation of channelized, straightened and engineered channels and floodplains. In fact, the science of floodplain restoration and remeandering of rivers was largely developed in Europe, and much of the literature on this topic comes from European case studies (e.g. Brookes 1992, 1996; Iversen et al. 1993). With the exception of some early erosion reduction efforts to reduce declining production of agricultural lands in the 1970s, restoration efforts in Australia and New Zealand and other developed countries also began in the 1980s and 1990s (Gippel & Collier 1998).

The number and scale of watershed restoration efforts, along with spending on restoration, has increased rapidly in the last few decades in North America, Europe, Australia, and elsewhere. This has been partly driven by increasing environmental awareness, stronger environmental regulations, and declines in species of fish and aquatic organisms that are of high socioeconomic and cultural value. As discussed in the Section 1.4, legal mechanisms have been developed to restore water quality, individual species, and riverine ecosystems in developed countries. Perhaps the most commonly recognized legal mandates are those requiring protection or restoration of specific species under national laws such as the Endangered Species Act in the USA, the Canadian Species at Risk Act, or the European Red List. These legislative actions are generally reactive and drive attempts to restore habitats for listed species. While the legislation behind these lists generally calls for conservation and restoration of the ecosystems upon which these species depend, restoration actions are commonly focused on restoring specific habitats deemed important for one species or another. In the USA and Canada, for example, massive efforts to restore watersheds in the Pacific Northwest of North America are almost exclusively focused on recovering threatened and endangered salmon and trout populations (Katz et al. 2007), although restoration actions such as sediment reduction and riparian restoration also benefit other species. Beyond endangered species concerns, many nations have also passed legislation aimed at more holistic attempts to restore riverine ecosystems (e.g. the Clean Water Act in the USA or the Water Framework Directive in the EU) which seek to improve more broadly defined hydromorphological, chemical, and biological conditions of rivers.

In conjunction with changing drivers of restoration and an increasingly holistic approach to restoring watersheds, the expertise needed to plan and implement projects has also evolved. Early restoration efforts were often initiated by outdoorsmen or fisheries biologists and later by engineers, and focused on structural treatments or bank stabilization. The greater emphases on watershed processes in the USA and Europe has also led to improved design of more traditional habitat improvement techniques and greater emphasis on addressing root causes of degradation. Given that streams integrate both terrestrial and aquatic processes at multiple scales, the practice of restoring processes or improving habitats of an aquatic ecosystem requires an interdisciplinary approach to be successful. This often requires the collaboration of those with expertise in fish and aquatic biology, riparian and stream ecology, geology, hydrology and water management, geomorphology, landscape architecture, and even public policy, economics, and other social sciences. That is not to say that all projects will require expertise in all these fields, but most will benefit from an interdisciplinary team; this will certainly be essential for large or comprehensive restoration projects or programs to achieve their goals. Another aim of this book is therefore to provide a common basis and level of knowledge for individuals from various backgrounds to work together on developing and implementing successful restoration programs.

1.6 Key Steps for Planning and Implementing Restoration

Despite large financial investments in what has recently been called the ‘restoration economy’ (Cunningham 2002) and increasing literature on restoration planning, numerous watershed councils, river trusts, agencies, and other restoration practitioners do not follow a systematic approach for planning restoration projects throughout a watershed or basin. As a result, a number of restoration efforts fail or fall short of their objectives. Some of the most common problems or reasons for failure of a restoration program or project include:

not addressing the root cause of habitat or water quality degradation;

not recognizing upstream processes or downstream barriers to connectivity;

inappropriate uses of common techniques (one size fits all);

an inconsistent (or complete lack of an) approach for sequencing or prioritizing projects;

poor or improper project design;

failure to get adequate support from public and private organizations; and

inadequate monitoring to determine project effectiveness.

These challenges and problems can be overcome by systematically following several logical steps that are critical to developing a successful restoration program or project (Figure 1.1). This book is designed to cover these steps in detail to assist with improving the design and evaluation of stream and watershed restoration plans and projects. We begin with a discussion of watershed processes and process-based restoration (Chapter 2), as these basic concepts underlie the restoration steps in subsequent chapters. The following chapters then explain the key steps, including: assessing watershed conditions and identifying restoration needs (Chapter 3); selecting appropriate restoration actions to address restoration needs (Chapter 5); identifying a prioritization strategy for prioritizing actions (Chapter 6); planning and implementing projects (Chapter 7); and developing a monitoring and evaluation program (Chapter 8). Goals and objectives need to be set at multiple stages of the restoration process, and there are multiple steps within each stage which we will discuss within each chapter. In addition, the human and socioeconomic aspects need to be considered throughout the planning and design process (Chapter 4). We close with a discussion of how to synthesize all these pieces to develop restoration plans and proposals (Chapter 9).

Throughout this book we emphasize the concept of process-based restoration (Chapter 2), which aims to address root causes of habitat and ecosystem degradation (Sear 1994; Roni et al. 2002; Beechie et al. 2010). Our purpose in doing so is to help guide river and watershed restoration efforts toward actions that will have long-lasting positive effects on riverine ecosystems and to ensure that, when habitat improvement is undertaken, the site potential and watershed processes are considered. We also emphasize the importance of recognizing socioeconomic and political considerations such as involving landowners and other stakeholders, permit and land-use issues, and education and outreach to the general public to build continued support for restoration (Chapter 4). Failure to consider these factors and involve stakeholders early on can prevent even the most worthwhile and feasible projects from being implemented. The following chapters go into detail on each of the steps for planning and implementing successful stream and watershed restoration programs and projects.

1.7 References

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Benke, A.C. (1990) A perspective on America’s vanishing streams. Journal of the North American Benthological Society 9, 77–88.

Bernhardt, E.S., Palmer, M.A. Allan, J.D. et al. (2005) Synthesizing U.S. River Restoration Efforts. Science 308, 636–637.

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Brookes, A. (1992) Recovery and restoration of some engineered British River Channels. In: Boon, P.J., Calow, J. & Petts, G.E. (eds) River Conservation and Management. John Wiley & Sons Ltd., Chichester, England, pp. 337–352.

Brookes, A. (1996) Floodplain restoration and rehabilitation. In: Anderson, M.G., Walling, D.E. & Bates, P.D. (eds) Floodplain Processes. John Wiley & Sons Ltd., Chichester, England, pp. 553–576.

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Brown, T. (2002) Clearances and clearings: deforestation in Mesolithic/Neolithic Britain. Oxford Journal of Archaeology 16, 133–146.

Buijse, A.D., Coops, H., Staras, M. et al. (2002) Restoration strategies for river floodplains along large lowland rivers in Europe. Freshwater Biology 47, 889–907.

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2

Watershed Processes, Human Impacts, and Process-based Restoration

Tim Beechie¹, John S. Richardson², Angela M. Gurnell³ & Junjiro Negishi⁴

¹Northwest Fisheries Science Center, National Oceanic and Atmospheric Administration, USA

²University of British Columbia, Canada

³Queen Mary, University of London, UK

⁴Hokkaido University, Japan

2.1 Introduction

Effective planning, design, and implementation of river restoration efforts each require an understanding of how watershed processes drive the structure and functions of riverine ecosystems, as well as how those processes support a wide variety of ecosystem services. In this book, the term ‘watershed process’ generally refers to movements of landscape or ecosystem components into and through river systems, which are typically measured as rates (Beechie & Bolton 1999). For example, erosion is a process that moves sediment from hillslopes to river channels, while sediment transport processes move sediment through stream and river channels to deltas and estuaries. Erosion is measured in units of mass/area/time, whereas sediment transport is commonly measured in units of mass/time. We do not restrict the term ‘processes’ to geomorphological or hydrological processes, but instead refer to a wide range of processes including erosion and sediment transport, storage and routing of water, plant growth and successional processes, delivery of nutrients and organic matter, inputs of thermal energy, trophic interactions, species interactions, and population dynamics. Understanding these processes and relationships between them is critical to the success of river restoration programs.

These driving processes influence states and dynamics of biological communities through a sequence of cause–effect linkages that connect watershed processes to habitat conditions, and habitat conditions to biota (Figure 2.1). ‘Habitat conditions’ here refers to physical, chemical, and thermal features of the river environment and ‘biota’ refers to ecological systems and functions that respond to habitat features. Humans alter watershed processes in many ways, leading to changes in habitat conditions, food webs, and biological communities (Allan 2004). Process-based restoration focuses on correcting anthropogenic disruptions to driving processes, thereby leading to recovery of habitats and biota (Sear 1994; Beechie & Bolton 1999). Restoration of critical processes also confers added resilience to river–floodplain ecosystems, as functioning processes allow the system to respond to future disturbances through natural physical and biological adjustments (Brierley et al. 2002; Beechie et al. 2010). While restoration of processes and habitats is critical to ecosystem recovery, correcting ecosystem degradation that results from lack of key species, introduction of non-native species, or poor water quality is also critical to ecosystem recovery (Karr 2006). Because these factors must also often be addressed to achieve restoration goals, we briefly address restoration of other ecosystem alterations and describe a comprehensive set of both processes and ecosystem features that may be important in restoring river ecosystems.

Figure 2.1 Illustration of process linkages between watershed processes, instream processes, and biological responses

(adapted from Beechie & Bolton 1999; Beechie et al. 2009).

In this chapter we first describe the hierarchical suite of processes that drive riverine ecosystems. We identify and describe the main processes driving riverine habitat dynamics and biota, focusing on processes that are commonly targeted by stream and watershed restoration activities. We also describe how watershed and reach-scale processes drive the expression and dynamics of habitat types in each reach of a river network, and illustrate how habitat conditions control expression and dynamics of biological communities. Throughout, our main purpose is to illustrate how these processes are hierarchically nested, and to show that higher-level controls set limits on the expression of habitat features or ecosystem attributes as influenced by lower-level controls (e.g. Beechie et al. 2010). We then describe the landscape setting, and proceed to watershed-scale processes, reach-scale processes, habitat dynamics, and instream biological processes. We also briefly describe ways in which watershed processes may be altered, thereby affecting the productivity, resilience, and functions of river ecosystems. Alterations to processes are further discussed in the chapter on watershed assessments (Chapter 3), and restoration actions that restore processes are discussed in Chapter 5. Finally, we describe process-based restoration (which is a central theme of river and watershed restoration in this book) and we present four process-based principles to help guide restoration planning and implementation.

2.2 The Hierarchical Structure of Watersheds and Riverine Ecosystems

Physical and biological features of riverine ecosystems are controlled by a hierarchy of physical, chemical, and biological processes operating across a wide range of space- and timescales (Figure 2.2). These processes control the arrangement of channel and habitat types across the riverine landscape, such as reach-scale channel types, or pool and riffle units at smaller scales (e.g. Frissell et al. 1986; Fausch et al. 2002; Allan 2004). Biological features are also controlled by a hierarchy of processes, including community composition of riparian or aquatic species and locations of suitable habitats for individual species (e.g. Beechie et al. 2008a; Naiman et al. 2010). For example, the typical sequence of channel patterns from headwaters to lower river begins with steep cascades in tributaries and progresses through step-pool, plane-bed, and pool-riffle channels (Montgomery & Buffington 1997) (Figure 2.3). As tributaries coalesce the channels become larger and more complex, and are classified as braided, straight, island-braided, or meandering (Beechie et al. 2006a). This arrangement of reach-level channel types, as well as habitat features within reaches, is determined by six main variables: channel slope, valley confinement, discharge, sediment supply and size, bank strength (including root strength), and wood supply (Figure 2.4).

Figure 2.2 Hierarchical nesting of processes controlling population and community responses of riverine biota. Higher-level controls set limits on the types of habitat features or ecosystem attributes that can be expressed at lower levels, and lower-level processes control the expression of attributes within those limits

(based on Beechie et al. 2010).

Figure 2.3 Common channel patterns in river networks,

based on Montgomery & Buffington (1997) and Beechie et al. (2006a).

Figure 2.4 Illustration of the six primary controls on channel form, including physical and biological processes that control physical habitat conditions in streams and rivers: channel slope, valley constraint, discharge, sediment supply, bank strength (including root reinforcement), and wood supply.

Two of these variables – channel slope and valley constraint (the ‘ultimate controls’) – are generally unchanging over human time frames, as the tectonic and erosional processes controlling these variables act over long time frames and across large areas (>10² years, >1 km²) (Naiman et al. 1992; Montgomery & Buffington 1997; Figure 2.5). Broad valley forms are defined by characteristics such as valley shape (V-shaped or U-shaped), the presence or absence of terraces, valley slope, and confinement of the stream (floodplain width relative to stream width). Valley forms are also relatively immutable over management time frames (<100 years) (Naiman et al. 1992), and are also controlled by past geological processes such as uplift, glaciation, and river erosion (Bishop et al. 1985; Benda et al. 1992; Montgomery 2002). While reach-scale channel slopes can change locally over shorter time frames, they can only do so within a fairly narrow range set by the valley slope, which is essentially immutable over hundreds of years. Hence, these landscape features (valley slope and constraint) control the range of channel types that can be expressed within valley segments (Naiman et al. 1992).

Figure 2.5 Spatial and temporal scales of processes that control physical features, vegetation, and aquatic biology in river ecosystems

(adapted from Naiman et al. 1992; Beechie et al. 2010).

The third and fourth controlling variables are discharge (or stream flow) and sediment supply. These two variables are controlled at the watershed scale, and the drainage boundary defines the region within which erosion and runoff processes control regimes of sediment supply and stream flow (Figure 2.5). Sediment supply and stream flow largely determine stream size and channel forms (e.g. pools, riffles, gravel or sand bars), within the limits set by valley confinement and channel slope. To some extent, patterns of stream size and sediment supply are also controlled by the structure of the river network (the arrangement of stream channels and confluences) which influence riverine habitats in two ways. First, the general downstream pattern of increasing stream flow and decreasing slope drives a corresponding shift in habitat characteristics (Vannote et al. 1980; Minshall et al. 1985; Naiman et al. 1987). Second, the network structure, arrangement of confluences, nick points, and alternating valley and canyon reaches interrupt gradual downstream trends to create local nodes or reaches with unique geomorphic features (Poole 2002; Benda et al. 2004). For example, sediment supply tends to be higher below large tributaries, creating locally complex habitat features and higher biological diversity. Together, these interacting patterns create a generally predictable arrangement of physical attributes in river systems.

At the reach level, channel forms are mutable and controlled by shorter-term variation in processes such as sediment supply regime, flow regime, and wood and plant propagule supply (Gurnell et al. 2005, 2006a; Figure 2.5). Because these processes naturally vary from year to year as a function of storms and floods, channel locations can change over relatively short time frames (years to decades) while the dominant reach type is usually relatively constant over decadal periods (Beechie et al. 2006a). Moreover, the dominant channel forms tend to exhibit a characteristic arrangement within watersheds, although geological features and the influence of river confluences can interrupt this pattern (e.g. Montgomery 1999; Benda et al. 2004; Beechie et al. 2006a). Within reaches, habitat units (e.g. pools, riffles, ponds) are even more variable in space and time (Frissell et al. 1986), often shifting positions between years as a result of wood movement. However, the relative abundance of habitats is relatively constant in the absence of major changes in driving variables (e.g. Zanoni et al. 2008).

The biological structure of riverine ecosystems is also hierarchically controlled, partly by the physical controls described above and partly by a suite of biological processes. At the landscape scale, biogeography of species controls the pool of species available to form riparian and aquatic communities (Figure 2.5). For example, native riparian species found in a watershed are limited to those that are endemic to the region and its climate, and native fishes are limited to those that can migrate to suitable habitats within their native range. Hence, in the absence of human interventions, local riparian and aquatic communities only comprise species present in the native species pool. At the watershed scale, both riparian and aquatic communities exhibit characteristic shifts in composition in the downstream direction and biological zones are often classified by the dominant species present, or by community composition and species richness (Huet 1959; Sheldon 1968; Ibarra et al. 2005). These shifts are partly due to a gradual increase in river size (e.g. the river continuum concept, Vannote et al. 1980; Minshall et al. 1985; Naiman et al. 1987; Figure 2.6). However, differences in channel slope, width, or pattern and dynamics are also influenced by changes in lithology or by network structure, which create localized discontinuities in this downstream trend. At the site scale, biological attributes are largely controlled by local habitat conditions, as well as by species interactions. For example, spawning locations of individual fish species are controlled by locations of suitably sized substrate, water depth, and velocity for the species (Beechie et al. 2008a), whereas productivity and species composition of riparian forests are controlled by local soil and moisture conditions (Naiman et al. 2010).

Figure 2.6 Illustration of the river continuum concept

(left panel, after Huet 1959; Vannote et al. 1980)

and measurements of downstream trends in channel slope, bankfull channel width, and floodplain width

(right panel, data from Skagit River, USA).

Measured trends indicate a systematic downstream trend in river size and slope as indicated by the river continuum concept, as well as local discontinuities created by changes in lithology that create alternating steep canyon reaches and low-gradient valley reaches.

2.3 The Landscape Template and Biogeography

Landscapes and watersheds are physically defined by their topography, geology, and climate, and can be classified by landscape units that stratify broad suites of habitat-forming processes and disturbance regimes (Montgomery 1999; Montgomery & Bolton 2003). This geologic, tectonic, and climatic template controls the arrangement of reach types in the network, locations of tributary junctions, and alternating canyon and floodplain reaches (Montgomery 1999; Benda et al. 2004; Brierley & Fryirs 2005; Beechie et al. 2006a). That is, the landscape template sets limits on the range of physical and biological attributes that any reach in the network is capable of expressing, and therefore sets limits on what can be achieved through restoration. For example, a steep headwater reach can only develop specific habitat types ranging from boulder cascades to short alluvial reaches upstream of wood jams, whereas a large low-gradient floodplain reach can develop a suite of habitat types ranging from mainstem pools and riffles, to bank and bar habitats, to a wide range of lentic and lotic floodplain habitats (Brierley et al. 2002; Beechie et al. 2005a).

Similarly, the landscape template influences hydrologic and sediment supply regimes that control channel forms. For example, regions with steep slopes, shallow soils, and humid climates may be characterized by high landslide rates and coarse sediment loads that tend to create braided or island-braided channels (Sidle et al. 1985; Hovius et al. 1997; Imaizumi et al. 2008). By contrast, a drier region with deep soils and gentle slopes may be characterized by high fine sediment loads and more stable anastomosing or meandering channels (Rust 1981; Beechie et al. 2006a). These differences in process regimes create diverse ranges of potential physical conditions that restoration can achieve in streams, but do not describe specific local habitat or biological conditions that are controlled by smaller-scale processes. The timescale of processes that create the landscape template also captures past climatic changes that formed regional geologic features such as glacial terraces (e.g. Benda et al. 1992), as well as large but rare disturbances such as volcanic eruptions or mega-floods that alter valley floor morphology (Beechie et al. 2001; Brown et al. 2001).

Key biological processes that drive the biogeography of species operate at similarly long space- and timescales. That is, processes such as migration, colonization, extinction, and evolution have – over tens of thousands of years – resulted in biotic assemblages that are adapted to the local geographic and climate settings of individual river systems and reaches within systems (Taberlet et al. 1998; Waples et al. 2004). Examples of these processes are migration and evolution of Pacific salmon (Oncorhynchus spp.), and colonization and establishment of plant species after glaciations (Bennett 1986; Waples et al. 2008). These processes result in species pools that limit the natural suite of riparian and aquatic communities that can be expressed in each reach. However, short-term climate variations can alter the spatial distribution or relative abundance of species over relatively short time frames (hundreds to thousands of years), yet the broad pool of species available in ecosystems generally