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What a difference a map makes : including ecosystem services within systematic conservation planning Hoshizaki, Lara 2009

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WHAT A DIFFERENCE A MAP MAKES: INCLUDING ECOSYSTEM SERVICES WITHIN SYSTEMATIC CONSERVATION PLANNING by  Lara Hoshizaki  BA (Hons) Concordia University, 2005  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF  MASTER OF SCIENCE in The Faculty of Graduate Studies (Resource Management and Environmental Studies) UNIVERSITY OF BRITISH COLUMBIA (Vancouver)  June 2009  © Lara Hoshizaki, 2009  Abstract Over the past decade, conservationists have increasingly employed the concept of ecosystem services to garner support for biodiversity conservation. However, recent research has found incongruence between bio-diverse areas and areas that provide large amounts of ecosystem services. In this thesis, I investigated the spatial relationship between lands that provide ecosystem services and lands that would be prioritized for biodiversity conservation in the context of a systematic conservation planning exercise. First, I mapped economic values for carbon storage, timber production, and recreational angling using a geographical information system (GIS). These values represented the difference in the provision of services based on whether the land is conserved or subject to timber harvesting, which is the prevailing land use in the area. I integrated these values into the siteselection software Marxan using two approaches: a ‘feature’ approach and a novel ‘benefit-cost’ approach. The first approach treated ecosystem services as conservation features with targets for protection. The second approach is the incorporation of potential service values into the cost function of Marxan. I then compared the efficiency of the ‘feature’ (1) and ‘benefit-cost’ (2) approaches and found that the latter enabled Marxan to select a conservation reserve network that meets all biodiversity targets at a lower cost. I also reviewed the use of ecosystem service values within traditional cost-benefit analyses in a net benefit maximization framework and compare this with their more recent use in systematic conservation planning. With the help of concrete examples, I present a theoretical framework for the integration of ecosystem services into systematic conservation planning using the ‘feature’ and ‘benefit-cost’ approaches. I argue that before ecosystem services are integrated into conservation planning, researchers should consider particular characteristics of the services in relation to the site and purpose of the planning exercise. Conservation areas offer the opportunity to provide a haven for biodiversity, as well as essential ecosystem services for people. To ensure that they do both effectively, we must reconsider our approach to achieving these disparate goals.  ii  Table of Contents ABSTRACT .................................................................................................................................... ii TABLE OF CONTENTS ................................................................................................................ iii LIST OF TABLES ...........................................................................................................................iv LIST OF FIGURES ......................................................................................................................... v ACKNOWLEDGEMENTS .............................................................................................................. vi CO-AUTHORSHIP STATEMENT ................................................................................................. vii 1 AN INTRODUCTION TO ECOSYSTEM SERVICES AND THEIR PLACE IN CONSERVATION PLANNING ........................................................................................................ 1 1.1 ECOSYSTEM SERVICES: CONCEPT AND CONTEXT ............................................................... 1 1.1.1 What are ecosystem services? .................................................................................. 1 1.1.2 Why think about ecosystem services?....................................................................... 2 1.2 LITERATURE REVIEW ......................................................................................................... 4 1.2.1 Valuation .................................................................................................................... 4 1.2.1.1 1.2.1.2  1.2.2  Using economic values for conservation ................................................................... 7  1.2.2.1 1.2.2.2 1.2.2.3  1.2.3  Benefit transfer methods...............................................................................................10 Modeling and mapping ecosystem services for conservation .......................................11  Systematic conservation planning ........................................................................... 11  1.2.4.1  1.3 1.4  Formal markets .............................................................................................................. 8 Management programs.................................................................................................. 8 Using ecosystem services to achieve the goals of conservation ................................... 9  Spatially explicit services and values ......................................................................... 9  1.2.3.1 1.2.3.2  1.2.4  Economic valuation methods ......................................................................................... 5 Marginal values and measuring change ........................................................................ 5  Marxan ..........................................................................................................................12  RESEARCH QUESTIONS ................................................................................................... 13 REFERENCES ................................................................................................................. 15  2 ECOSYSTEM SERVICES IN THE CENTRAL INTERIOR OF BC: SPATIALLY EXPLICIT VALUES FOR CONSERVATION PLANNING ............................................................................. 20 2.1 INTRODUCTION ............................................................................................................... 20 2.1.1 Context and study area ............................................................................................ 21 2.2 METHODS ...................................................................................................................... 22 2.2.1 Economic valuation of ecosystem services and mapping of services and biodiversity features ................................................................................................................................. 22 2.2.1.1 2.2.1.2 2.2.1.3 2.2.1.4  2.2.2  Carbon storage .............................................................................................................23 Timber production .........................................................................................................24 Recreational angling .....................................................................................................25 Terrestrial biodiversity features .....................................................................................26  Inclusion of ecosystem services values within Marxan ........................................... 27  2.2.2.1 2.2.2.2 2.2.2.3  Marxan scenarios .........................................................................................................27 Targets for ecosystem services: the “feature approach” ...............................................29 Ecosystem services within the Suitability Index: the ‘benefit-cost approach’ ................30  2.3 RESULTS........................................................................................................................ 31 2.3.1 “Best” solution reserve networks ............................................................................. 31 2.3.2 Efficiency in Marxan solutions ................................................................................. 32 2.3.3 Hot spots .................................................................................................................. 35 2.3.4 Congruence of ecosystem service areas and biodiversity ...................................... 37 2.4 DISCUSSION ................................................................................................................... 38 2.4.1 Marxan results ......................................................................................................... 38 2.4.2 Unexpected similarity between solutions ................................................................. 39 2.4.3 Data uncertainties .................................................................................................... 40  iii  2.4.4 Increasing the possibility for implementation ........................................................... 41 2.4.5 Interdisciplinary communication ............................................................................... 41 2.5 CONCLUSION .................................................................................................................. 41 2.6 REFERENCES ................................................................................................................. 43 3 INCORPORATING ECOSYSTEM SERVICES THINKING INTO CONSERVATION PLANNING: A PROPOSED FRAMEWORK ................................................................................ 47 3.1 3.2 3.3  INTRODUCTION ............................................................................................................... 47 ECONOMIC VALUES OF ECOSYSTEM SERVICES.................................................................. 49 PREVAILING APPLICATIONS OF ECOSYSTEM SERVICE THINKING: VALUATION AND COST BENEFIT ANALYSES ................................................................................................................................... 49 3.3.1 Benefits of ecosystem services in CBA: internalizing externalities ......................... 50 3.3.2 Challenges of CBA: what, where and who to count ................................................ 50 3.4 RECENT APPLICATIONS IN CONSERVATION PLANNING ........................................................ 51 3.4.1 Benefits of including ecosystem services in conservation planning: for people and biodiversity ............................................................................................................................ 52 3.4.2 Challenges of including ecosystem services within conservation planning: serving disparate needs .................................................................................................................... 52 3.5 SYSTEMATIC CONSERVATION PLANNING ........................................................................... 53 3.6 A FRAMEWORK FOR DECIDING HOW AND WHEN TO INCLUDE ECOSYSTEM SERVICES WITHIN CONSERVATION PLANNING ............................................................................................................ 54 3.6.1 A feature approach .................................................................................................. 54 3.6.2 A benefit-cost approach ........................................................................................... 56 3.7 CONCLUSION .................................................................................................................. 57 3.8 REFERENCES ................................................................................................................. 59 4  CONCLUSIONS ON PRESENT WORK WITH DIRECTIONS FOR FUTURE RESEARCH 63 4.1 INTEGRATING ECOSYSTEM SERVICES WITHIN CONSERVATION PLANNING ............................ 63 4.2 STRENGTHS AND WEAKNESSES OF RESEARCH ................................................................. 64 4.2.1 Strengths .................................................................................................................. 64 4.2.1.1 4.2.1.2 4.2.1.3 4.2.1.4  4.2.2  Interdisciplinarity ...........................................................................................................64 Simplicity ......................................................................................................................65 NCC biodiversity data ...................................................................................................65 Applicability...................................................................................................................65  Weaknesses ............................................................................................................ 66  4.2.2.1 4.2.2.2 4.2.2.3 4.2.2.4  Omission of climate change considerations..................................................................66 Limited number of ecosystem services .........................................................................67 Data inconsistencies and model simplicity....................................................................67 Lack of local input .........................................................................................................68  4.3 FUTURE RESEARCH DIRECTIONS ...................................................................................... 68 4.3.1 Definitions ................................................................................................................ 68 4.3.2 Biophysical data collection and application ............................................................. 69 4.3.3 Focus on people ...................................................................................................... 69 4.3.4 Political will for implementation ................................................................................ 70 4.4 REFERENCES ................................................................................................................. 71 APPENDICES ............................................................................................................................... 73 APPENDIX A: DATA SOURCES AND METHODS FOR ECOSYSTEM SERVICE MODELING AND VALUATION ......................................................................................................................... 73 APPENDIX B: DETAILS OF MARXAN SCENARIOS.................................................................. 78 APPENDIX C: SUITABILITY INDEX TRANSFORMATION ......................................................... 80 APPENDIX D: CLIMATE CHANGE IN THE CENTRAL INTERIOR OF BC – IMPACTS ON ECOSYSTEM SERVICES AND ADAPTATION STRATEGIES ................................................... 81  iv  List of Tables TABLE 1-1: CATEGORIES OF ECOSYSTEM SERVICES............................................................................. 2 TABLE 1-2: METHODS OF ECONOMIC VALUATION.................................................................................. 5 TABLE 1-3: STEPS IN SYSTEMATIC CONSERVATION PLANNING ............................................................. 12 TABLE 2-1: MARXAN RUN SCENARIO DESCRIPTIONS AND IDENTIFYING NUMBERS ................................. 29 TABLE 2-2: COMPARISON OF COSTS AND AMOUNT OF ECOSYSTEM SERVICES CAPTURED IN SCENARIO 5 (FEATURE APPROACH) VS. SCENARIO 7 (BENEFIT-COST APPROACH WITH NO TIMBER PRODUCTION) .............................................................................................................................................. 33 TABLE 2-3: COMPARISON OF COSTS AND AMOUNT OF ECOSYSTEM SERVICES CAPTURED IN SCENARIO 5 (FEATURE APPROACH) VS. SCENARIO 6A (HYBRID APPROACH)................................................... 34 TABLE 2-4: COMPARISON OF COSTS AND AMOUNT OF ECOSYSTEM SERVICES CAPTURED IN SCENARIO 6B (HYBRID APPROACH) VS. SCENARIO 8 (BENEFIT-COST APPROACH) ............................................. 35 TABLE 2-5: CORRELATION VALUES BETWEEN SUMMED SOLUTION OUTPUTS......................................... 38 TABLE 2-6: BREAKDOWN OF COSTS AND BENEFITS IN “BEST” RESERVES FROM SCENARIOS 7 AND 8 ..... 39  v  List of Figures FIGURE 1-1: SUPPLY AND DEMAND CURVE........................................................................................... 6 FIGURE 2-1: MAP OF STUDY AREA..................................................................................................... 22 FIGURE 2-2: MARGINAL VALUES OF ECOSYSTEM SERVICES IN THE CENTRAL INTERIOR, BC BASED ON CONSERVATION AND TIMBER HARVESTING LAND MANAGEMENT SCENARIOS ................................. 27 FIGURE 2-3: “BEST” RESERVE NETWORK SOLUTIONS FOR INDIVIDUAL FEATURE SCENARIOS (1, 2, 3 AND 4) USING ROAD INDEX SCORES AS A SUITABILITY INDEX ............................................................. 32 FIGURE 2-4: TWO RESERVE NETWORK DESIGNS PRODUCED BY MARXAN – THE FIRST INCLUDING ECOSYSTEM SERVICES AS FEATURES (SCENARIO 5) AND THE SECOND INCLUDING THE SAME ECOSYSTEM SERVICES AS SIDE-BENEFITS (SCENARIO 7 ) .......................................................... 33 FIGURE 2-5:TWO RESERVE NETWORK DESIGNS PRODUCED BY MARXAN – THE FIRST INCLUDING ECOSYSTEM SERVICES AS FEATURES (SCENARIO 5) AND THE SECOND USING ECOSYSTEM SERVICES AS FEATURES AND COSTS (SCENARIO 6A ) ............................................................................... 34 FIGURE 2-6: TWO RESERVE NETWORK DESIGNS PRODUCED BY MARXAN – THE FIRST INCLUDING ECOSYSTEM SERVICES AS FEATURES AND COSTS (SCENARIO 6B) AND THE SECOND INCLUDING THE ECOSYSTEM SERVICES AS SIDE-BENEFITS AND COSTS(SCENARIO 8) .......................................... 35 FIGURE 2-7: HOT SPOT MAPS OF INDIVIDUAL ECOSYSTEM SERVICES AND BIODIVERSITY FEATURES....... 37 FIGURE 2-8: ARTIFICIAL DISCONTINUITIES AS A RESULT OF DATA LIMITATIONS IN TIMBER PRODUCTION RESULTS ................................................................................................................................. 40  vi  Acknowledgements I am very grateful for the support and encouragement of my supervisors, Drs. Kai Chan and Brian Klinkenberg. Thank you, Kai for your unwavering commitment to this work, your relentless optimism and for having an attention to detail that would make most (me included) stop and stare. Thank you, Kai. Thank you, Brian for your calming reassurances that “everything is going to be fine”, for your solutions when all I could think of were problems, and mostly, for your confidence in me and this work. Thank you also to my committee member, Dr. Kristine Ciruna. Kristy, your passion for conservation is quite simply, contagious. I would like to also acknowledge the many experts whom I consulted during the process of modeling the ecosystem services in this work. Thank you to Marcus Weiler for his insight on his dominant-runoff processes model, to Olaf Schwab, Ralph Wells, Thomas Maness and John Nelson for introducing me to the acronym-abundant world of forestry in BC, and to Eric Parkinson for his patience in teaching me the joys of fish. I gratefully acknowledge funding support from the Natural Science and Engineering Research Council of Canada and the Nature Conservancy of Canada. I would also like to thank Pierre Iachetti and Sarah Loos from the Nature Conservancy of Canada for their help and support through this project and in particular to Sarah for coaching me through some of the trials and tribulations of a little program called Marxan. Thanks as well to Ed Gregr at IRES for generously sharing his computer with me for the Marxan analyses. Finally, I truly doubt that this work could have been completed without the companionship of my fellow students at IRES. I am grateful to have begun this program at the same time as so many brilliant, funny and talented people. In particular I would like to acknowledge Jack Teng and Julia Freeman who were always there when I needed a quick statistics lesson or to assure me that yes I could, in fact, finish this. Thank you.  vii  Co-Authorship Statement Drs. Kai Chan and Brian Klinkenberg co-authored Ch.2 and Ch.3 of this thesis. In Ch.2, Dr. Chan contributed to the research design and methodology, as well as writing and reviewing. Dr. Klinkenberg contributed to the methodology, assisted with data analysis and contributed to writing and reviewing the chapter. In Ch. 3 Dr. Chan contributed to the theoretical design and writing and Dr. Klinkenberg contributed to the review. My contribution to both chapters included: •  Research design  •  Implementation of research  •  Data collection and analysis  •  Literature review  •  Writing, editing and review  viii  1 1.1 1.1.1  An introduction to ecosystem services and their place in conservation planning Ecosystem services: Concept and context What are ecosystem services? An “ecosystem service” can be a shape-shifting creature, whose definition may frequently  depend on whom you are asking, what time frame you’re inquiring about and where you are currently standing. The first widely accepted definition of this concept comes from the seminal book on the topic, Nature’s Services, edited by Gretchen Daily in 1997. In the introductory chapter Daily defines ecosystem services as “the conditions and processes through which natural ecosystems, and the species that make them up, sustain and fulfill human life” (Daily, 1997). Throughout my work, I have defined these services as the flow of direct and indirect benefits from ecosystems to people. The Millennium Ecosystem Assessment (MA) was conducted from 2001 to 2005 and involved over 1350 researchers from around the world. It was originally proposed by Dr. Walter Reid, who was then the Vice President of the World Resources Institute. The MA defined ecosystem services as “the benefits that people obtain from ecosystems” (MA, 2005). Following this definition, the MA sought to evaluate the current state of the world’s ecosystems and model how future changes in the provision of ecosystem services will impact human populations. The MA categorized services into four types: provisioning, supporting, regulating and cultural. Here, I have also included option services (Daily et al., 2000) (Table 1-1). Provisioning services are perhaps the simplest to understand - these are services that provide humans with tangible products, such as food, water, and fibers for clothing. Supporting services are those that are essential for the delivery of other ecosystem services. An example is primary production, which transforms solar energy into biomass for a multitude of uses (Farber et al., 2006). Regulating services refer to the maintenance of essential ecosystem processes for human well being. The process of plants storing carbon is a regulating service that maintains a suitable climate for human life. Cultural services are more difficult to define, in that they can vary widely within groups of people and even amongst individuals. Generally, cultural services are defined as ways in which ecosystems enhance emotional, psychological and cognitive well-being (de Groot et al. 2005; Farber et al. 2006). The serenity felt when sitting by a quiet lake, and the enjoyment of taking a walk through a forest, are the products of cultural services. Finally, an option service acts like a safety net by providing us with services that we do not currently benefit from but may in the future. Experiments like Biosphere 2 have proven that we have a relatively poor understanding of natural ecosystems and what they require to function (Heal, 2000). Therefore, it is rational to  1  assume that these systems may provide us with options, such as medicines and other helpful products, in the future. Table 1-1: Categories of ecosystem services Category Examples Provisioning services  Production of timber, food, water  Supporting services  Soil production, photosynthesis  Regulating services  Climate regulation, flood mitigation  Cultural services  Contribution to recreation, aesthetics, spiritual values  Option (preserving)  Protection of ecosystem components and genetic diversity for  services  possible future use  Regardless of the categorization of these services, this thesis assumes that ecosystem services are anthropocentric: they are the ways in which they benefit humans and are measured by the amount of benefits they provide. Section 1.2.1 will introduce the primary methods that have been used to describe, measure, and analyze ecosystem services in the current academic literature. First, however, I will briefly describe the contexts in which the concept of ecosystem services has been applied most often. 1.1.2  Why think about ecosystem services? Ecosystem services provide benefits to humans whether or not we choose to  acknowledge them. So why do we do study them? Unfortunately, although these services are essential for human well-being, we have often taken them for granted and not factored them into our decision-making processes. Early environmentalists, such as George Perkins Marsh, wrote about the work that natural ecosystems do for humans and attempted to bring awareness to the human-induced degradation of these systems. By framing natural systems in a way that was directly connected to human well-being, early ecosystem service advocates called for a greater understanding of the benefits we receive from the environment (Mooney and Ehrlich in Daily, 1997). Rachel Carson’s Silent Spring (1962) illustrated even stronger connections between environmental degradation and human well-being as she gracefully mapped the relationship between pesticides and human health. Unfortunately, decision makers have been slow to appreciate these direct links between people and their environment. The cost-benefit analyses that most developments, such as hydropower electricity and land conversion projects, are based on generally consider environmental impacts to be externalities, and thus have no real impact on the final result. Despite legal obligations to include environmental impact assessments within new  2  developments, the environment is still largely considered to be outside the economic decision making framework, even within the supposed enlightened university classroom (Shrestha, 2007). Ecosystem service research often attempts to internalize these externalities by attaching explicit values - whether they are economic, social, or cultural, relative or absolute - to the benefits humans obtain from ecosystems. In the past, it has been the work of ecologists to describe the biophysical processes behind services and the work of economists to represent these processes with dollar values. Monetary values are determined as the value of the work of an ecosystem in order to make direct comparisons in cost-benefit analyses (CBA) and costeffectiveness calculations. One of the most popular case-studies of this type of valuation took place in the Catskill watershed in upper New York State. After assigning an economic value to the water purification services of the watershed, New York City determined that it was more costeffective to conserve the watershed than to build a water purification plant for the city (Daily and Ellison, 2002). In these types of CBA the environment is explicitly factored into decision making, in a way that heavily depends on interdisciplinarity. The field of ecosystem services comprises a wide range of academics – in the natural and social sciences. Recently, conservation planning has begun to incorporate ecosystem service values to create direct links between conservation and human benefits (Egoh et al., 2007). The incorporation of ecosystem service concepts into conservation planning is relatively new, but it has spread quickly (Goldman et al., 2008). I believe that these concepts can offer an exciting new perspective on a wide variety of environmental challenges, but that we require a thorough understanding of both their strengths and limitations if governments, non-governmental organizations and other practitioners are to use them successfully to further the goals of conservation. In this thesis, I use current literature as well as my own case study to more fully explore the new and quickly growing relationship between systematic conservation planning and ecosystem services. In the present chapter, I introduce the concepts of ecosystem services and the ways in which they have typically been measured. This chapter contains a literature review that consists of three parts: the economic valuation of ecosystem services, the ways in which these values have been used, and the benefits gained from characterizing services in a spatially explicit fashion. Finally, I present my research questions. The second chapter consists of a case study in which I have estimated economic values of three services in the interior of British Columbia: carbon storage, timber production and recreational angling. These values were then included in a systematic conservation planning framework using the reserve site selection program, Marxan using two distinct approaches: the ‘feature’ approach and the ‘benefit-cost’ approach. In the third chapter I review the use of ecosystem services within cost-benefit analyses (CBA) and conservation planning and I discuss benefits and pitfalls of undertaking ecosystem service  3  research in these contexts. I also provide a theoretical framework that aims to define the appropriate application of the ‘feature’ and ‘benefit-cost’ approaches introduced in Chapter Two. In the final chapter of this thesis, I reflect upon this work as a whole and discuss its applications as well as future research possibilities given my findings. 1.2  Literature review The scope of ecosystem service literature is becoming increasingly broad as  interdisciplinarity within the field increases. Ecosystem service thinking is now being employed in conservation, resource management, public policy and community development research. Given the breadth of the field, I have chosen to review the aspects of ecosystem services that have received the greatest attention and research. This literature review will discuss the economic valuation of ecosystem services, the use of these values in policy and decision making, and finally their application in a spatially explicit framework. 1.2.1  Valuation The controversy surrounding the valuation of nature is a complex and impassioned one  that encompasses views from a wide spectrum of academia. The arguments surrounding it are outside the scope of this work however I do believe it is important to state explicitly that this work is only concerned with anthropocentric instrumental values (Turner et al., 2003). These values of ecosystem services may be further divided into economic, cultural, social, and intrinsic (NRC, 2005). The current literature has been consumed with measuring the economic values of ecosystem services. In the realm of decision making and public policy, translating ecosystem services into dollar values has offered great opportunities for real world applications (Daily et al., 2000; Postel and Thompson, 2005; Heal, 2007). However, arguments have also been made against the process of economic valuation. Some see it as an ethically questionable task – arguing that humans must be encouraged to respect nature for nature’s sake, and not for the capital gains it provides (McCauley, 2006). Others have argued against the incongruous match of linear neo-classical economics with non-linear ecosystem functioning (Chee, 2004). Although there is much validity in these opinions, it can also be argued that many development decisions have been made in the face of conservation efforts and the environment was seen as a mere externality to be accounted for elsewhere. These final economic values have also been questioned on the basis that they are socially constructed (Toman, 1998), vary across the spatial and temporal scales that they are applied to (Chan et al., 2007) as well as on who are being defined as beneficiaries (Turner et al., 2003).  4  1.2.1.1  Economic valuation methods The economic values of ecosystem services are most often derived from non-market (or  indirect market) valuation methods (Table 1-2) (de Groot et al., 2002). The task of determining the most appropriate methods for a particular service, location and situation has resulted in an ongoing discussion, most notably between ecologists and economists (Goulder and Kennedy 1997; Heal, 2000; de Groot et al., 2002; Chee, 2004; NRC, 2005). Table 1-2: Methods of economic valuation Methods of Economic  Brief Explanation  Valuation Direct market valuation  An exchange value that ecosystem services have in trade. Mostly applied to provisioning services, but also some regulating services such as through carbon credit markets.  Indirect market valuation  When no explicit markets exist, other values are used to estimate the WTP (willingness to pay) or WTA (willingness to accept compensation) for the availability or loss of these services.  1. Avoided cost 2. Replacement cost 3. Factor income 4. Travel cost 5. Hedonic pricing  1. The costs avoided due to the provision of a service. 2. The cost of replacing a service with a human-made alternative. 3. The amount by which a service increases human incomes. 4. The cost of travel incurred to benefit from a service. 5. The contribution of an ecosystem component or function to the price of a related good/service attributable to aesthetics.  Contingent valuation  Often determined from questionnaires or choice experiments, this process involves estimating value based on respondents’ stated WTP or WTA for services.  Group valuation  Similar to contingent valuation, but an emphasis is on democratic process to agree on values that reflect a group of people.  1.2.1.2  Marginal values and measuring change Economic values of ecosystem services are useful in decision making because they  allow the work of ecosystems to be compared to other relevant costs and benefits. However it is important to also consider the ways in which these values are measured. Infamously, Costanza et al. (1997) estimated these values for 17 different services on a global scale. The authors openly acknowledged many of the limitations and uncertainties within their study, but perhaps the greatest weakness was the reporting of the total value of these services at a global scale, when  5  they were first measured at a much smaller geographic scale. In addition to the problem of benefit transfer, which I will discuss further in Section 1.2.3.1, there was also the issue of assigning a total value to a service, when it is the marginal value of that service, or how much the value of a service changes with changes in its provision, that has the most meaning. How can a single total value be placed on the global supply of clean water? Arguments against total economic values are bolstered when such studies seem to put a price tag on the services necessary for all life. Therefore, it seems obvious that these large scale assessments of total value are not helpful in the policy or decision-making arena because economic decisions are often made at the margins (Daily et al., 2000; Heal, 2000). Prices are a function of the relationship between supply and demand. In the classic supply and demand curve under ideal market conditions, the price (P) is low when there is a high supply (S) and low demand (D). The price is high when the opposite is true. The market is in equilibrium when P is P* and the number of units sold (Q) is equal to Q* (Figure 1-1) (Moffatt, accessed 2008). However, it is the gradual change in prices, or the marginal change of the service’s value, that is most useful to decision makers (Daily et al., 2000). The marginal value of a good is effectively how much more someone would be willing to pay for an extra unit of the good (Heal, 2000). For example, if someone does not have a couch, they might be willing to pay $500.00 for one. However, they might only be willing to pay $200.00 for a second couch and perhaps not willing to pay anything for a third. Therefore, couches have a decreasing marginal value. Marginal prices tell us about small changes, not large ones. That is why they are the most relevant to ecosystem services – they give insight into particular decisions – such as the water purification values of a particular wetland in a particular watershed, but they don’t indicate the value of clean water all over the world (Heal, 2000). Figure 1-1: Supply and demand curve  6  A small change in the marginal value of a particular good occurs when the supply of that good changes. Therefore, when we consider ecosystem service provision in a variety of scenarios, we can use marginal values to elucidate the difference in value associated with the difference in supply. In most conversion scenarios where land transitions from one use to another, ecosystem service provision will not cease completely, so measuring the total value of a service may not be as useful as understanding the flow of benefits through time and how service values change between land uses (Balmford et al., 2002). Using these differences, tradeoff scenarios can be investigated using economic models such as production possibility frontiers to maximize the provision of individual services via different land uses with the least cost (Nalle et al., 2004; Polasky et al., 2005; Farber et al., 2006). 1.2.2  Using economic values for conservation The challenges of economic valuation go beyond attaching simple dollar figures to  services. It is the application of these values that can provide insight and advance individual research goals. The power of assigning a dollar value to an ecosystem service has most often been harnessed in three separate, but connected, ways. First, these values have been used to show the direct reliance of economies on natural systems and showcase the need for conservation purely for human benefits. Coffee production, which plays a major role in the global economy, has been shown to be dependent on the pollination services provided by neighbouring tropical forests (Ricketts et al., 2004). Similarly, maintaining habitat for diverse vertebrate communities may also decrease the risk of Lyme disease in humans (LoGiudice et al., 2003; Ostfeld and LoGiudice 2003). The direct benefits of hydrological services have also been given great attention in the literature (Brauman et al., 2007). Natural watersheds have been shown to offer many direct benefits to people such as stream flow regulation for hydropower production (Guo et al., 2000). Many organizations who have championed conservation as a means of providing clean water have made great use of the economic values associated with hydrologic ecosystem services. There is now a global trend to protect watersheds from development instead of building filtration infrastructure because of the cost savings involved (Postel and Thompson, 2005). Second, the dollar benefits that ecosystem services provide can be used indirectly to support the primary goal of biodiversity conservation. The relationship between biodiversity and ecosystem services is one that has held a significant place in the literature (Balvanera et al., 2005; Balvanera et al., 2006; Worm et al., 2006; Díaz et al., 2006). For some, biodiversity has been seen as an ecosystem service in itself as well as one which supports other services (Chapin III et al., 2000). The relationship between biodiversity and ecosystem services deserves further research for ecological management in the future (Kremen, 2005; Srivastava and Vellend, 2005). However, because of the difficulty in defining biodiversity and characterizing its functional implications for services, ecosystem services such as flood control and carbon storage have been  7  explored for their ability to finance biodiversity protection through various institutional mechanisms (Reid, 2001). Finally, using these dollar values allows researchers to explicitly describe and account for tradeoffs in conservation planning. If a certain area offers higher economic returns than another (when both are in conserved states), it may be more defensible to prioritize it for conservation (Rodriguez et al., 2006). The following sections will explore how formal markets and conservation programs have used ecosystem service values to further environmental protection goals. 1.2.2.1  Formal markets The first environmental market in the United States was a cap and trade system for  sulphur dioxide (SO2) credits. Cap and trade markets function in conjunction with government enforced policies. These policies are attempts to limit environmental degradation such as the emission of pollutants or urban developments. In the example of sulphur dioxide, companies that emit greater than the agreed upon levels of SO2 will either pay for the infrastructure improvements needed to emit less or for the credits to continue current levels of emission. These credits can be bought from other companies that do not emit the maximum allowance. The SO2 market has been a success as it has greatly reduced the amount of acid rain in the United States (Bayon, 2004). A major key to the success of this market is government supported mandatory participation. The carbon market is another example of an environmental market. Its success is also directly related to its participation levels. In Europe, where participation is mandatory, carbon credits are worth much more than in the United States, where participation in the Chicago Climate Exchange is voluntary (Daily and Ellison, 2002). Although markets are purported to efficiently allocate scarce resources, they function with many limitations. Government support is necessary to ensure equity and fairness in markets, as well establishing property rights. For public goods, such as some ecosystem services, the issue of property rights is complex and challenging. Despite these issues, markets may still offer a helpful pathway for incorporating ecosystem service benefits, such as those offered by wetlands, into real-world planning applications (Bayon, 2004). The importance of maintaining transparent information sources (Katoomba Group, accessed 2008), government and ENGO (environmental non-governmental organization) involvement as well as furthering research regarding rights allocation will be necessary for environmental market success (Bayon, 2004). 1.2.2.2  Management programs Outside of formal markets, ecosystem service values have been used to finance resource  management programs. An example of this includes municipalities that pay upstream landowners to manage their land in ways that ensure adequate water supply for downstream communities (Reid, 2001; Postel and Thompson, 2005). In other circumstances, the revenue from “ecological  8  taxes” has been used to finance local conservation efforts which return benefits like natural pest control (Reid, 2001). In Napa, California ecosystem service advocates successfully won the right to forgo the traditional hard-infrastructure flood management approach of the Army Corps of Engineers in favour of restoring the natural flood plain. This alternative flood plain management maintained the river’s sinuosity and allowed for the banks to absorb and naturally retain water without the ecological losses often incurred with Corps methods (Daily and Ellison, 2002). The BushTender program in Australia has been successful in encouraging land owners to manage biodiversity on their property and has since spawned similar programs for other services (Katoomba Group, accessed 2008). Through this program, properties are assessed for the economic benefits their ecosystem services offer. Owners are then compensated for management costs relative to the benefits their land provides via an auctioning system which ensures the greatest estimated benefit for the least cost (Stoneham et al., 2003). 1.2.2.3  Using ecosystem services to achieve the goals of conservation The examples above show the utility of assigning economic values to ecosystem  services. However, many ecosystem service advocates also caution that the concept of ecosystem services has its limitations and should not be thought of as a “silver bullet” for both human well-being and conservation (Chan et al., 2007). In some instances, maximizing service provision will conflict with biodiversity goals, and vice versa. Planting eucalyptus may increase the provision of carbon storage, but it may also decrease native plant biodiversity (Myers, 1984). This potential conflict echoes the call for a greater understanding of the relationship between biodiversity and ecosystem services, as well as for the explicit recognition of individual conservation project goals (Balvanera et al., 2001; Kremen, 2005). Funding for conservation programs is often related to their goals. In a recent study, it was found that conservation projects that explicitly list ecosystem service protection as a goal are funded four times more than conservation projects that do not mention ecosystem services (Goldman et al., 2008). It is obvious that ecosystem service thinking is a popular topic in conservation, but as a boy in a cape was once told – with great power comes great responsibility. If projects state direct human benefits, public expectations grow and leave very little, if any, room for disappointment. Failing to deliver on promises of cleaner water or increased environmental tourism revenues may result in community backlash against conservation (Chan et al., 2007). However, despite these controversies, a look at the literature suggests that ecosystem service concepts have the ability to reach a broader conservation audience and increase efforts to protect biodiversity. 1.2.3  Spatially explicit services and values Time and again, the use of ecosystem service values within a conservation framework  implies a spatially explicit representation of these services, yet few maps of multiple services exist  9  (Balvanera et al., 2001). Applying ecosystem service knowledge to management decisions often requires spatially explicit detailed knowledge of local ecosystems and the ways in which people’s communities depend on them. Indeed, Guo et al. (2001) even argue that “understanding the spatial distribution of natural capital stock is as important as the value of it.” On a large scale, ecosystem services have generally been mapped in the past using benefit transfer methods (Costanza et al., 1997; Troy and Wilson, 2006). The following sections will discuss the pros and cons of benefit transfer for mapping ecosystem service values and introduce other, more localized methods and models which I have used as a starting point in my own case study. Lastly, I will introduce the framework of systematic conservation planning and describe how some conservation organizations use the site selection tool, Marxan, within the process.  1.2.3.1  Benefit transfer methods The influential paper by Costanza et al. (1997) received criticism on many aspects of  their study – many of which were readily acknowledged by the authors within the text itself. One of the goals of the Nature paper was to facilitate and spark discussion, and this was unarguably achieved (Toman, 1998). One of the major discussions that stemmed from this research was the authors’ use of benefit transfer to extrapolate ecosystem service values across the globe (Costanza et al., 1997). Benefit transfer, or value transfer, is a technique in which the results of previous economic valuation studies are applied to new political or geographical contexts (Brouwer, 2000). This method has been used in studies where relevant local data are unavailable but information exists for what are considered to be similar systems (Troy and Wilson, 2006). Despite its utility to provide value estimates in data-deficient areas, there are severe limitations and uncertainties involved in benefit transfer. Most notably, ecosystem service values are incredibly context dependent. As mentioned in Section 1.2.1, economic values attached to these services are very likely to vary with different social, temporal, geographical political and ecological contexts, not to mention that, like snowflakes, no two ecosystems are identical. Again, economic values of ecosystem services are most meaningful when derived for the place in question. Through this lens, it is unlikely that the values of wetland water filtration for a large downstream community will be the same for a relatively isolated wetland system. Second, when these values are determined through contingent valuation methods, such as surveys and choice experiments, the ability of these methods to convey people’s perceptions of value is greatly diminished when transferred to a new place and community of beneficiaries (Bateman et al., 2006).  10  1.2.3.2  Modeling and mapping ecosystem services for conservation The limitations of benefit transfer have led researchers to develop more sophisticated,  site-specific biophysical and economic models of ecosystem services (Naidoo and Ricketts, 2006; Egoh et al., 2008; Polasky et al., 2008). These models and methods produce estimates of values that are directly relevant to individual study areas, thus decreasing uncertainty in the results. However, the majority of studies thus far have been limited by data availability and scale. Additionally, including ecosystem services in conservation planning is a relatively new challenge and as such there is no widely accepted methodology for integration between these two fields (Egoh et al., 2007). One tool that seeks to estimate ecosystem service values with relatively simple data requirements is the Integrated Valuation of Ecosystem Services and Tradeoffs tool, or InVEST. This tool has been produced by the Natural Capital Project, a partnership between Stanford University, the Nature Conservancy and the World Wildlife Fund. The main goals of this project are to provide maps of nature’s services, assess their values in economic and other terms, and incorporate those values into resource decisions (NCP, accessed 2008). Following these advancements in modeling and mapping ecosystem services and their values, research is needed to explore the most accurate way to include these maps within a systematic conservation planning framework. Incorporating ecosystem service values with biodiversity targets for conservation planning has rarely been accomplished (Chan et al., 2006). Chan et al. (2006) investigated the distribution of ecosystem services across the central coast of California in conjunction with traditional measures of biodiversity. The authors found both negative and positive spatial correlations between biodiversity and different services, proving that the relationship between biodiversity and ecosystem services is anything but straightforward. 1.2.4  Systematic conservation planning There exists a long history of protecting natural areas against the encroaching threat of  development (Margules and. Pressey, 2000). Conservation efforts not only attempt to protect known species, habitats and ecosystems, but they also attempt to protect pieces of the natural world that are currently unknown to us. The task of designing a reserve that will most efficiently conserve biodiversity and ecosystem services often seems massive in proportion to available resources; therefore, basic steps have been identified to encourage a more systematic approach (Table 1-3).  11  Table 1-3: Steps in systematic conservation planning (Margules and Pressey, 2000; Groves et al., 2002; Sarkar et al., 2006; and Wilson et al., 2007) 1) Identify the objective of the planning exercise. 2) Compile data within the planning region. Ideally, these data will be both coarse and fine filter, and describes both the biophysical and socioeconomic environment. These data will ultimately inform what the conservation targets will be. 3) Establish conservation goals. These goals are often the proportion of targets that will be represented in the conservation plan to ensure persistence of biodiversity over time. 4) Review the contribution of existing conservation areas to goals. 5) Select conservation reserve sites with an appropriate planning algorithm. 6) Review the output from the algorithms and determine the management actions necessary to implement the conservation plan.  Many environmental non-governmental organizations (ENGOs), such as the Nature Conservancy of Canada (NCC), have made it their mission to protect tracts of land and water solely for the intrinsic value of these areas and for the benefit of future generations (NCC, accessed 2008.). A novel aspect of this thesis is the consideration of ecosystem services as conservation targets, as well as costs and/or side-benefits within the NCC’s eco-regional assessment using the site selection software Marxan (Ball and Possingham, 2000). 1.2.4.1  Marxan Marxan is a site-selection software program created at the University of Queensland by  Ian Ball and Hugh Possingham (2000). It was initially intended as a computational aid in the design and site selection of marine reserves for the Great Barrier Reef Marine Authority. Marxan evolved out of another program, Spexan, which was designed solely for terrestrial environments, but both programs operate under similar assumptions and methods. Marxan is now used widely by a host of researchers, agencies and ENGO’s such as The Nature Conservancy and the Nature Conservancy of Canada, to inform decision making Chan et al., 2006). Marxan offers a range of spatially explicit options for conservation planners. It does this by selecting a group of sites which, when combined, capture a targeted amount of conservation features, such as rare species and ecosystems. At the same time, the program seeks to minimize the cost of the reserve. Marxan’s objective function is comprised of a summation of costs and penalty values; therefore, the lower the value of the objective function, the “better” the solution. In this way it seeks a minimum cost solution to capture all stated biodiversity and/or ecosystem service targets (Ball and Possingham, 2000). The objective function is mathematically described by the equation (Ball and Possingham, 2000): Objective Function Total Score = ∑Cost + (BLM*∑SitesBoundary) + (∑CFPF * Penalty) + Cost Threshold Penalty  12  The cost of the sites is based on user input and is often a summation of multiple costs that represent how much they would be to acquire and manage. These costs can also be related to a site’s suitability for inclusion within the reserve, or how likely the site would be to foster biodiversity. A suitability index could be created that assigns a cost measure to each site based on characteristics like road density and dominant land use (Chan et al., 2006). Early conservation planning attempts were based heavily on biogeographical theory. These associated principles included the creation of near-circular reserves that would maximize area and minimize perimeter. Also, connectivity between sites was prioritized for its ability to facilitate species migration between reserve planning units and ensure species persistence in an area (Sarkar et al., 2006). The BLM (boundary length modifier) is directly related to these early concepts. It is used to reduce edge effects and to acknowledge ecological and management issues related to perimeter or boundary length (McDonnell et al. 2002). This factor allows for flexibility in reserve design because the user can put a weight on the importance of spatial cohesion. If a compact reserve is highly desirable, increasing the BLM will make it more costly to create a fractured reserve. Being able to adjust the BLM allows for more realistic planning options and explicitly attempts to find a balance between cost and spatial design. Marxan uses the CFPF (conservation feature penalty factor) to assign penalties for failing to achieve conservation goals. Ideally, the value of the penalty should equal the cost of reaching the respective goal, such as the additional sites required to meet the goals. This penalty can also be proportional so that if only half of the goal is met, only half of the penalty will be charged. The objective function in Marxan seeks to minimize penalties and costs while meeting conservation targets. Finally, the Cost Threshold Penalty creates a cost ceiling for the objective function and will stop the program as it reaches the threshold. The ceiling ensures maximized efficiency within budgetary constraints. Computer algorithms have been utilized in the field of conservation planning for decades (Sarkar et al., 2006). Their utility in achieving conservation objectives, while explicitly accounting for spatial economy has made them indispensible tools for conservation planners. Spatial economy is a concept that is central to the goals of conservation planning and refers to the minimization of costs (area reserved, perimeter length, opportunity costs, etc.) while maximizing the amount of biodiversity conserved within a region (Sarkar et al., 2006). 1.3  Research questions This manuscript-based thesis consists of four chapters that consider past research in  ecosystem services, economic valuation and spatially explicit, systematic conservation planning. The work as a whole seeks to elucidate the relationship between ecosystem services and  13  systematic conservation planning by using a case study in the Central Interior of British Columbia and an analysis of both the conservation planning and ecosystem service literature to create a theoretical framework for including ecosystem service values within a program like Marxan. The second chapter describes a case study in which spatially explicit economic values were assigned to three ecosystem services in the Interior of British Columbia, Canada. The case study examines the difference in carbon storage, timber production and recreational angling values between two land use scenarios – conservation and timber harvesting. These values were then included in a biodiversity conservation plan using the reserve site selection software, Marxan using two different approaches. This chapter describes and defends the simple models used to spatially evaluate the differences in ecosystem service values between land use scenarios and posits how these values can be used to further the goals of biodiversity conservation within a systematic planning framework. The central questions this chapter seeks to answer are: •  What are the economic values of changes in ecosystem services associated with changes in land use within the study area? Are different ecosystem service values spatially correlated to each other within and across land use scenarios?  •  Do the areas that represent these greatest changes in values coincide with prioritized areas of biodiversity conservation?  •  What are the differences in Marxan solutions, when ecosystem service values are treated as targeted features vs. side-benefits or costs?  Using the lessons learned in the BC case study, the third chapter takes a step back to understand how ecosystem services can best be integrated within systematic conservation planning, and specifically within Marxan. I consider the advantages and disadvantages of including ecosystem services within cost-benefit analyses as well as the challenges and advantages of including them within a spatially explicit conservation plan. I then outline a theoretical framework for choosing between the two approaches, ‘feature’ and ‘benefit-cost’, to appropriately represent different kinds of services in different circumstances. In particular our research questions are the following: •  How can we apply ecosystem service thinking in order to most effectively attain the goals  of biodiversity conservation and ecosystem service protection? •  Under which circumstances is it appropriate to employ a ‘feature approach’ and consider  ecosystem services as features to be targeted by conservation vs. a ‘benefit-cost approach’, by which services are incorporated into the cost function of a program such as Marxan as sidebenefits or opportunity costs?  14  1.4  References  Ball, I.R. and H.P. Possingham. 2000. MARXAN (v.1.8.2): Marine reserve design using spatially explicit annealing, a manual. Balmford, Andrew, Aaron Bruner, Philip Cooper, Robert Costanza, Stephen Farber, Rhys E. Green, Martin Jenkins, Paul Jefferiss, Valma Jessamy, Joah Madden, Kat Munro, Norman Myers, 10 Shahid Naeem, Jouni Paavola, Matthew Rayment, Sergio Rosendo, Joan Roughgarden, Kate Trumper, R. Kerry Turner. 2002. 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There are many definitions of ecosystem services, but the one we employ here follows from Daily (1997:3): the flow of direct and indirect benefits from ecosystems to people. Since 2000 there has been an increasing prevalence of biodiversity conservation assessments including ecosystem service protection (Egoh et al., 2007). Including ecosystem services within conservation has been seen by many as a way to engage a community that might be unconcerned with traditional conservation goals (Daily et al., 2000; Chan et al., 2007). Indeed, many believe that ecosystem services can play a critical role in influencing environmental policies as well as garnering additional fiscal support for conservation projects (Goldman et al., 2008). However, before we assume that including ecosystem services in a conservation project satisfies the goals of biodiversity conservation and also meets the needs of human communities, it is important to consider the spatial (in)congruence of these objectives. Recent research suggests that areas with high levels of biodiversity are not necessarily the same areas that provide high levels of ecosystem services (Chan et al., 2006; Naidoo et al., 2008). Given that biodiversity and ecosystem services will not always be spatially congruent, how can we use existing planning tools to prioritize both within a particular landscape on a constrained budget? These tools include spatially explicit planning programs, such as Marxan, that create cost-minimizing solutions to satisfy biodiversity goals across a particular landscape (Ball and Possingham, 2000). The research and literature in the fields of biodiversity mapping and systematic conservation planning is expansive (Chapin III et al., 2000; Margules and Pressey, 2000; Groves et al., 2002; Margules et al., 2002; Sarkar et al., 2002; Pressey et al., 2007). Conversely, the mapping of ecosystem services and the inclusion of these services within a systematic conservation framework is a relatively new endeavor. Concrete examples are needed to bring these abstract ideas into the “mainstream” (Cowling et al., 2008). In this paper we illustrate the inclusion of ecosystem services in a conservation plan in the Central Interior region of British Columbia, Canada. First, we present a spatially explicit economic valuation of three ecosystem services: carbon storage, timber production, and the provision of recreational angling opportunities. In contrast to earlier efforts to value ecosystem services (e.g., Costanza et al., 1997), which were criticized for their calculation of total values of  1  A version of this chapter will be submitted for publication. Hoshizaki, L., Chan, K.M.A., and Klinkenberg, B. Ecosystem services in the Central Interior of BC: spatially explicit values for conservation planning.  20  services rather than the value of what might be lost due to realistic changes (Toman, 1998), we calculate the difference in the value of services across conservation and timber harvesting landuse scenarios, as the forestry industry presents the greatest direct threat to conservation in the local area (Iachetti, personal communication). By examining the differences in these values, and not the total value of the service, we represent the changes in ecosystem service provision between conservation and timber harvesting scenarios. In many land conversion scenarios it is likely that a certain level of ecosystem function will remain intact; therefore total values of the service across the landscape do not reveal which areas may have the most to lose or gain as a result of land-use change. These marginal values are more helpful for policy and decision making than total values (Daily et al., 2000). Second, we include these ecosystem service values within a systematic biodiversity conservation plan using the site selection algorithm Marxan. In the past ecosystem services have been treated as features within Marxan with minimum amounts set as targets for their protection, what we have called the ‘feature approach’ (Chan et al., 2006). Here, we introduce a novel method of integration where we include ecosystem service values within the cost function of Marxan, or the ‘benefit-cost approach’. We have also combined these two approaches by including biodiversity-congruent services (recreational angling and carbon storage) as features and incongruent services (timber production) as costs in the ‘hybrid approach’. The objectives of our work are two-fold: first, to identify areas in the Central Interior most amenable to protection from ecosystem service loss due to timber harvesting; second, to compare the previously demonstrated ‘feature approach’ with the novel ‘benefit-cost approach’ for the total costs of their solutions as well as the amount of both biodiversity and ecosystem service values that are protected using each method. 2.1.1  Context and study area This work has been done in conjunction with the Nature Conservancy of Canada (NCC)  for their eco-regional conservation assessment in the Central Interior region of British Columbia (BC). This assessment’s goal is to identify areas in the study region that should be prioritized for conservation implementation based on their high biodiversity value and their amenability to conservation (NCC, 2007). The study area consists of the Central Interior eco-province in the south and the Subboreal eco-province in the north. Eco-provinces are regions with the province of BC that share a similar climate and topography and are also at a reasonable size for policy creation and implementation (Demarchi, 1996). The study area is located in the centre of the province of British Columbia and covers roughly 46,000 km² (Fig. 2-1). The topography is relatively flat in the centre and contains the Caribou, Chilcotin and Nechako plateaus, bordered by the Coast, Skeena and Omineca mountain ranges. Vegetation in the area is dominated by the Interior Douglas-fir and Sub-Boreal Pine-  21  Spruce biogeoclimatic zones and there are large areas of Bunch grass along the path of the Fraser River, whose headwaters are located in the Sub-boreal eco-province. The region is home to diverse fauna including moose, mule deer, and over 65% of all known bird species in the province (Demarchi, 1996). Figure 2-1: Map of study area  Source: Nature Conservancy of Canada, 2007. Population density is quite low in the region; the largest cities are Prince George, Williams Lake and Quesnel with populations of 77,000, 11,000 and 10,000 respectively. The main industries within the area are forestry, with some mining and small scale ranching. 2.2 2.2.1  Methods Economic valuation of ecosystem services and mapping of services and biodiversity features We have assumed that timber harvesting is the greatest direct threat and presents the  most valuable forgone opportunity from conservation in the study area. Thus our first objective was to measure the economic values of ecosystem service provision under two land-use scenarios: conservation and timber harvesting. We have chosen to investigate three services in the study area (carbon storage, timber production and recreational angling) based on an informal survey of the NCC’s Central Interior team of experts, and in consideration of available data. In the economic valuation of these services we have assumed that the beneficiaries are largely British Columbians, with the exception of carbon storage, and that the demand for these services will remain constant over time (population growth is slow in the study area). Through this research we have investigated the difference in ecosystem service values between conservation and timber harvesting land-use scenarios. We had initially intended to  22  include as a service freshwater provision for domestic purposes. However, we discovered that the majority of residents in the study area rely on relatively impermeable groundwater aquifers, for which timber harvesting is not a relevant threat. Accordingly, we focused our efforts on the services presented here. All services were valued and mapped in 500-hectare planning units that were later included in the conservation planning exercise using software Marxan. Below we present our assumptions and briefly summarize the methods used to model and value each service. For greater detail, see Appendix A. 2.2.1.1  Carbon storage We obtained publicly available digital data on carbon storage in soil, as well as in above  and below ground vegetation from the World Resources Institute (Matthews et al., 2000). To estimate the difference in carbon storage between harvested and conservation land-use scenarios, we conducted a literature search on the topic of carbon loss in temperate forests due to logging. The Carbon Budget Model of the Canadian Forest Service has been used to model differences in carbon storage across land uses in a forested landscape similar to our study area. This model was used to determine the change in carbon storage in two hypothetical BC Interior forest landscapes that differed only by their fire disturbance and managed harvest cycles. The landscapes had fire disturbance cycles of 500 and 750 years and harvesting cycles of 100 and 120 years respectively. The study found an 18.2% and 1% loss, respectively, in carbon when the landscapes transitioned from a “natural” to a “managed” management scenario. This assumes sustainably managed forest practices, as well as regular fire and pest disturbances (Kurz et al., 1997). For our study, we took the rounded average of these findings and assumed a 10% loss of carbon when an area was logged versus when it was conserved. However, given the data we have available, we were unable to determine where harvesting is currently taking place. Therefore, we have assumed that if all land is currently being harvested, it is at 90% of its carbon storage potential and has 11% to gain if it were conserved. Conversely, if all land is currently conserved, then it may lose 10% of its carbon if it were harvested. Therefore, we have valued and mapped 10% of the carbon storage values in the study area as a conservative estimate. Carbon storage was economically valued at $8.46 (CDN) per ton of carbon dioxide using the mid-price average of three carbon trading markets: the Chicago Climate Exchange, the New South Wales and the EU Emissions Trading scheme on March 19th, 2008. We have followed the methods and assumptions outlined by Naidoo and Ricketts (2006) to justify the use of carbon credit trading prices as proxies for the value of carbon storage. First, we assume that the beneficiaries of this ecosystem service are global and that these prices reflect the amount of social damage avoided by society at large by decreasing CO² emissions. Second, we assume that protection against deforestation is a valid strategy to reduce CO² emissions and that those 23  areas outside of the study area are under imminent threat of deforestation. Given the high levels of logging activity in the area, we consider this supposition to be valid. The coarse resolution of the data hides variations in biomass within each cell, thus it represents averages across individual stands at different ages/stages, with a distribution of stages following from the historic management regime. We effectively assume that each cell has been managed according to a sustainable yield model for harvest rotations specified by its particular Timber Supply Area such that for a 40-year rotation, approximately 1/40th of the harvestable area in each cell was harvested each year for the past 40 years. To the extent that the forest stands in some cells have been heavily harvested, this will underestimate the value of the change in carbon storage associated with a change in management; our method will overestimate this value for cells where the majority of stands have yet to be harvested. Subject to these assumptions, our value represents the net present value (NPV) of the difference in carbon storage associated with timber harvesting/conservation (Fig. 2-2). 2.2.1.2  Timber production Timber production is measured here as an opportunity cost of conservation, with the  difference between the two land-use scenarios being 100% of the net value of timber harvest. The total loss of value from timber production is a valid assumption as the NCC does not intend to include reserve networks in regular timber harvesting practices. Net present values (NPV) were measured for timber production over a 1000-year time frame with a discount rate of 4%, assuming that the ratio of benefits to costs remains constant over this time period. These parameters have been supported by both expert opinion and forestry economics literature (Creedy and Wurzbacher, 2001; Nelson pers. comm.). All values were measured per 500-ha planning unit and assume uniform costs and benefits within each cell. We believe this simplification is necessary given the large number of planning units within this large study area. Costs consisted of harvesting costs, cost of transportation to the closest processing facility, and the costs of replanting (silviculture costs). These costs were based on slope, distance and biogeoclimatic (BEC) zone, respectively, and were derived from previous merchantability work in the province (Thomae, 2003). Steeper slopes and longer distances result in higher costs. BEC zones are used by the Interior Appraisal Manual to distinguish between different silviculture costs in the province (Ministry of Forests and Range, 2007). The benefits of timber production were measured as a function of leading tree species and the volume expected at its minimum harvestable age (MHA). Average timber prices were calculated from BC Interior Log Market Reports from 2003-2008 (Ministry of Forests and Range, 2003-08) (Fig. 2-2).  24  2.2.1.3  Recreational angling We have assumed that timber harvesting will have an adverse effect on recreational  angling values through an increase in sedimentation (Jordan, 2006; Lohse et al., 2008). In addition, we assume that recreational angling activities will not severely impede the goals of conservation. The value of recreational angling in the study, and how much it may be impacted by timber harvesting activities, was determined using data from an angler effort model that has assigned how much actual angler effort (measured in days spent fishing) can be supported by a particular lake given its productivity, distance from major population centers, and accessibility by roads. The model was fitted using raw data such as boat counts from aerial surveys as well as mail surveys in the region (Parkinson et al., 2004). We matched the number of angling days for each lake with economic values for the average amount of money spent per day on recreational angling in freshwater regions of BC, which include transportation as well as licenses, package deals and accommodation (Government of Canada, 2005). These values were averaged over large areas and do not take into account spatial variations, such as transportation costs. Using the Ministry of Environment’s Fisheries Sensitive Watershed (FSW) database we assigned relative sensitivity scores (from 0 to 1.0) to third-order watersheds in the study area based on six characteristics (Reese-Hansen and Parkinson, 2006). Soil type, density of alluvial streams, lake buffering capacity, amount of forest cover, annual precipitation and slope have been equally weighed in the sensitivity index scores. These data were only available for catchments that contribute to smaller lakes as they are assumed to be not artificially stocked. Given our objectives of representing the value of angling at risk due to sedimentation, the missing data are relatively unproblematic because fish populations in artificially stocked lakes are likely to be less vulnerable to an increase in sedimentation than populations spawning naturally. By this, we mean that although fish in artificially stocked lakes are still vulnerable to sedimentation, the numbers of fish in these lakes will often not decrease greatly because they will be restocked artificially. We have assumed a linear 1:1 relation between sensitivity to timber harvesting and change in economic values of recreational angling. We recognize that this is a highly simplified relation, but it seems to be the most defensible representation of our current understanding (Parkinson, pers. comm.). Therefore, we have combined the sensitivity score of each watershed with its recreational angling value, based on amount of effort, to derive a final value of the expected difference in recreational angling values between conservation and timber harvest scenarios. For example, if a watershed had a sensitivity score of 0.10 and it has a potential economic value of $10,000.00, then we would expect a difference of $1000.00 between timber harvest and conservation land-uses.  25  2.2.1.4  Terrestrial biodiversity features The NCC carried out two conservation assessments: one focused on aquatic biodiversity  and another for terrestrial biodiversity. Our work was included within the terrestrial assessment therefore we have only considered the NCC’s terrestrial biodiversity features. These include both coarse and fine filter targets such as old growth forest ecosystems and rare plant species, respectively. The fine filter data consist of over 75 plant species and 100 animal species (3 amphibians, 5 reptiles, 28 mammals and 64 birds). Animal species were selected based on their designation as IUCN red-listed species, CITES-listed species, COSEWIC and SARA-listed species, CDC red and blue-listed species, as well as other more expert-informed subjective characteristics such as whether the species is endemic, regionally important or especially vulnerable to change. Data used to represent these features came from a variety of sources including the BC Conservation Data Centre, the BC Ministry of Environment, the Canadian Wildlife Service and Ducks Unlimited. The coarse filter data represent terrestrial ecological systems, as defined by the NatureServe classification system. These systems are meant to represent groups of biological communities that are found in similar physical environments and are influenced by similar dynamic ecological processes, such as fire or flooding (Kittel, 2008). Examples of such systems include the North Pacific Interior dry grassland and the North Pacific Mountain Hemlock Forest. Coarse filter data also included particular rare or “focal” ecosystems, such as hot springs and stands of old growth forests.  26  Figure 2-2: Marginal values of ecosystem services in the Central Interior, BC based on conservation and timber harvesting land management scenarios  2.2.2 2.2.2.1  Inclusion of ecosystem services values within Marxan Marxan scenarios Our second objective was to design a conservation reserve network for the study region  that explicitly considers ecosystem services as well as the traditional coarse and fine filter biodiversity features described above. To do so, we have used the site selection program Marxan 2.0.2 (Ball and Possingham, 2000). Marxan uses a simulated annealing, stochastic optimization algorithm that seeks to minimize its objective function while capturing all targets within the reserve network. Its objective function includes a cost layer and two kinds of penalties: the Conservation Feature Penalty Factor (CFPF), for failing to achieve targets; and the boundary length multiplier (BLM), for spatially dispersed networks. As a proxy for costs—in concordance with the NCC biodiversity assessment—we used a ‘suitability index’ (SI) that assigns a non-monetary cost value to each planning unit based on its density of and/or proximity to roads. The SI was created by the Nature Conservancy of Canada and gives an indication of the (un)suitability of a particular planning unit for conservation. The index assigns a relative score to each planning unit where higher scores indicate greater density or proximity to roads, thus they are more costly and less suitable for conservation. A “flat”  27  suitability index was also used to assign a value of 500 to each planning unit as each hexagon covered an area of 500 hectares. The area of the planning unit is an often used proxy for cost in Marxan (Ardron et al., 2008). Thus in the scenario that used the flat SI, units were treated equally across the study area regardless of their suitability for conservation. We used the flat suitability index to select areas for their high timber production values, thus creating a reserve network for activities contrary to the goals of conservation (Ban, 2008)—in effect a timber reserve. All scenarios, with the exception of timber production, locked in current protected areas and parks in each solution. Timber production runs excluded these areas. Table 2-1 describes and assigns a number to each scenario that we will use to identify it in the remainder of this paper (Table 2-1). We performed Marxan runs with 500 restarts. Due to the high number of biodiversity features included in the analysis, we found through testing that the value of the BLM would have to be extremely large in order to achieve cohesion in any of the solutions. Under the direction of the NCC, a BLM of 1 was used in all runs in order to utilize the penalty, but ensure that costs to the reserve network were not astronomical. A CFPF of 10 was used for all targets as it was the lowest value we found that would still ensure all targets were met. The Marxan Best Practices Manual recommends that the CFPF should be the minimum value needed to achieve targets (Ardron et al., 2008). Scenarios were run separately for each eco-province due to the large size of the study area and the amount of data associated with it. The results were then combined for a correlation analysis. Our results focused on the “Best” and “Summed Solutions” outputs from Marxan. The “Best” solution from each scenario is defined here as the reserve network that has the lowest objective function score and meets all targets (Ball and Possingham, 2000). However, “Best” may not be the most appropriate solution given current land-use practices within the selected reserve network as the local socio-political context will ultimately dictate if and how the conservation plan will be implemented. Despite this, Marxan’s “Best” solution is regarded as a snapshot for the larger analysis, given appropriate boundary length modifiers (BLM) and conservation feature penalty factors (CFPF). The “Summed Solution” measures how many times a particular planning unit is included in a final solution, and indicates how important a particular planning unit is to the reserve network, or its irreplaceability. The details for each Marxan scenario can be found in Appendix B.  28  Table 2-1: Marxan run scenario descriptions and identifying numbers Scenario Approach Features Suitability Index 1 Feature Biodiversity Road Index 2 Feature Recreational Angling Road Index 3 Feature Carbon Storage Road Index 4 Feature Timber Production Flat Biodiversity, Recreational Angling 5 Feature Road Index and Carbon Storage Biodiversity, Recreational Angling Road Index and 6A and 6B Hybrid and Carbon Storage Timber Production BenefitRoad Index with Recreational 7 Biodiversity Cost Angling and Carbon Storage 8  2.2.2.2  BenefitCost  Biodiversity  Road Index with Recreational Angling, Carbon Storage and Timber Production  Targets for ecosystem services: the “feature approach” In the past, ecosystem services have been included in conservation assessments as  features for which particular targets are desired (Ardron et al., 2008). We ran Marxan in multiple scenarios with ecosystem services as features and assigned targets for each feature. These targets required Marxan to include at least 50% of the total available amount of ecosystem service values within each solution. There is inherent difficulty in choosing meaningful targets for services, and this challenge is discussed further in Chapter 3. Despite this, the middle value of 50% was chosen so that Marxan would have flexibility in its solutions (i.e., it would not have to choose everything) but also represent a large portion of the target. We conducted a sensitivity analysis on the targets by running Marxan with targets of 10% below and above the original values. There were relatively minor changes in the results; therefore we believe that the original values chosen are acceptable targets for the ecosystem services and the solutions are representative. The same targets were used in all of the scenarios except in 6B. In Scenario 6B we used the amount of recreational angling and carbon storage that was captured in Scenario 8 as their respective targets in 6B in order for the two scenarios to be directly comparable. Assigning targets to ecosystem services within Marxan is arguably less appropriate than assigning targets for more traditional biodiversity features (Chan et al., 2006). In the case of biodiversity features, there is a presumed ethical imperative to ensure that a particular amount of each is included in each stratification unit in the reserve network; in the case of ecosystem services, there is generally not the obligation for particular minimum levels within particular regions, in part because many services are assumed to be substitutable with human infrastructure, etc., as reflected by their valuation in dollar terms. By including ecosystem services as features within Marxan, not only are we forced to assign somewhat arbitrary targets, but we are also creating direct competition between  29  biodiversity and ecosystem service features. This situation is representative of the greater debate amongst conservationists who have identified tensions between the agendas of biodiversity conservation and ecosystem service protection (Chan et al., 2007). 2.2.2.3  Ecosystem services within the Suitability Index: the ‘benefit-cost approach’ In an attempt to bypass the aforementioned arbitrary nature of targets and direct  competition between goals, we ran separate scenarios within Marxan that included carbon storage, recreational angling and timber production ecosystem service values in the Suitability Index, or the ‘benefit-cost approach’. The angling and carbon storage values were regarded as side-benefits whereas the timber production values were seen as additional costs. We converted the Suitability Index scores into dollar values using a four-part linear transformation based on land acquisition values in the study area. We then combined the dollar values of the converted Suitability Index with our ecosystem service dollar values. We assume that higher Suitability Index scores correlate with increased urbanization, which correlate to an increase in land values. For greater detail on the Suitability Index and our transformation, please see Appendix C. We then added timber production values and/or subtracted recreational angling and carbon storage values from the transformed Suitability Index. Through the inclusion of these values, we assume that ecosystem services increase (in the case of timber production) and/or decrease (in the case of recreational angling and carbon storage) the costs or difficulty of conservation. For example, an area with high carbon storage values may be more easily conserved as some of the costs of conservation might be recouped through future fiscal returns via carbon credits. In a similar way, an area with high timber production values may be found to have opportunity costs that render conservation socially unacceptable. We measured the efficiency of this method by comparing the average cost of the “Best” solutions in both approaches, as suggested in the Marxan Best Practices Manual (Ardon et al., 2008). Firstly, we identified the areas selected to be within the “Best” reserve network for biodiversity, recreational angling and carbon storage features (our ‘feature’ approach). We then determined what the cost of this reserve network would be if the cost surface had included these ecosystem service values, and not just the road index costs by overlaying the selected planning units onto the cost layer in a GIS. We compared this summed cost to the cost of a reserve network that protected biodiversity but included recreational angling, carbon storage and/or timber production values within the cost surface (our ‘benefit-cost’ approach).  30  2.3 2.3.1  Results “Best” solution reserve networks The reserve network for biodiversity features is much patchier than the networks of the  ecosystem services (Fig. 2-3). This is due to the number and broad distribution of biodiversity features across the study area, as well as the differing scales of the coarse and fine filter features. The proposed “best” solution for timber production has much larger, contiguous areas and is spread throughout much of the study area, with the exception of the steeper terrain in the north-east corner and the protected areas in the south-west (Fig. 2-3). Areas of high slopes are not attractive for timber production because harvesting costs on steep slopes often exceed expected benefits (Thomae, 2003). The carbon storage reserve network is concentrated along the borders of the study area, away from urban areas and major highways (Fig. 2-3). This distance from population centers makes the implementation of a carbon storage reserve network in these areas more feasible. The reserve network created for recreational angling consists of small patches dispersed across the study area, with compact areas along the southern border and in the center of the study area (Fig. 2-3). These areas correspond with clusters of small lakes, the fish populations of which are thought to be most vulnerable to timber harvesting. All reserves were required to include currently protected areas and parks, with the exception of Scenario 4, which excluded them.  31  Figure 2-3: “Best” reserve network solutions for individual feature scenarios (1, 2, 3 and 4) using road index scores as a Suitability Index.  2.3.2  Efficiency in Marxan solutions The reserve network that included carbon storage and recreational angling as side-  benefits within the SI (Scenario 7) was less contiguous than the reserve network that captured the services as features (Scenario 5), although both used a BLM of 1. Despite the decreased spatial cohesion, the reserve network that used the ‘benefit-cost approach’ achieved all biodiversity targets and at a lower cost (in both road index and transformed SI terms) than the reserve network which used the traditional ‘feature approach’. In other words, when we transpose the reserve network that was achieved with the SI that included ecosystem service values, onto the road index SI, the selected reserve network was less costly than the reserve network that included all the features as targets. Although the services were not specifically targeted as features, the ‘benefit-cost approach’ reserve network achieved 78% of the recreational angling target and 72% of the carbon storage target (Fig. 2-4, Table 2-2). The ‘feature approach’ reserve network achieved over 95% of both targets.  32  Figure 2-4: Two reserve network designs produced by Marxan – the first including ecosystem services as features (Scenario 5) and the second including the same ecosystem services as side-benefits (Scenario 7 )  Table 2-2: Comparison of costs and amount of ecosystem services captured in Scenario 5 (feature approach) vs. Scenario 7 (benefit-cost approach with no timber production)  Scenario  Cost (recreational angling, carbon storage and road index dollar values)  Cost (road index values)  Amount of recreational angling values captured  Amount of carbon storage values captured  5  $93.9B  38.7M  $16.9M  $10.0B  7  $62.9B  31.7M  $13.2M  $7.18B  In Scenario 6A, our ‘hybrid’ approach, we set targets for carbon storage, recreational angling and biodiversity features and added timber production to the SI as an additional cost alongside our transformed road index values. As a result, the area of the reserve increased by 33,000 km² in order to capture targets for carbon storage and recreational angling features. Similarly to Scenario 7, the planning units that were chosen are dispersed across the study area with few compact areas outside of parks and protected areas. Each scenario achieved their feature’s targets at lower costs based on their respective SIs. Scenario 6A was 2.5M (or 16%) greater than Scenario 5 based on the SI used in Scenario 5, however the cost of Scenario 5 was 18.5B (or 7%) greater of Scenario 6A based on the SI that included timber production.  33  Figure 2-5:Two reserve network designs produced by Marxan – the first including ecosystem services as features (Scenario 5) and the second using ecosystem services as features and costs (Scenario 6A )  Table 2-3: Comparison of costs and amount of ecosystem services captured in Scenario 5 (feature approach) vs. Scenario 6A (hybrid approach) Scenario  Cost (timber production and road index dollar values)  Cost (road index values)  Amount of recreational angling values captured  Amount of carbon storage values captured  5  $123.3B  38.7M  $16.9M  $10.0B  6A  $104.8B  41.2M  $16.9M  $10.0B  Finally, in Scenario 8, we set targets for biodiversity features only and included all three ecosystem services in the transformed road index SI, either by subtracting them as side-benefits (recreational angling and carbon storage) or adding them to the cost of the planning unit (timber production). To ensure that these scenarios were comparable in terms of how much ecosystem services were we ran Scenario 8 first and recorded how much recreational angling and carbon storage values were captured. Scenario 8 captured 80% of recreational angling and 72% of carbon storage original targets, similarly to how much of each service was captured in Scenario 7. We then used the amounts captured in Scenario 8 as new targets for the services in Scenario 6B. The resulting networks were very similar, both in their spatial distributions as well as their costs. These similarities reveal the limited set of possible solutions given that both scenarios were required to represent the same amount of biodiversity, recreational angling and carbon storage values and did so with an SI that included timber production costs. However, it should also be  34  noted that the ‘benefit-cost’ approach yielded a solution that was slightly less expensive than the ‘hybrid’ approach.  Figure 2-6: Two reserve network designs produced by Marxan – the first including ecosystem services as features and costs (Scenario 6B) and the second including the ecosystem services as side-benefits and costs(Scenario 8)  Table 2-4: Comparison of costs and amount of ecosystem services captured in Scenario 6B (hybrid approach) vs. Scenario 8 (benefit-cost approach) Scenario  Cost (timber production and road index dollar values)  Cost (ecosystem service and road index values)  Amount of recreational angling values captured  Amount of carbon storage values captured  6B  $85.0B  $77.8B  $13.5M  $7.2B  8  $83.7B  $76.4B  $13.5M  $7.2B  2.3.3  Hot spots The concept of irreplaceability is also represented by proxy in Marxan’s outputs as the  summed solution (Ball and Possingham, 2000). The irreplaceability of a planning unit is proportional to its contribution to meeting the goals of the conservation network. Alternately, it is defined as the extent to which options for a conservation network are lost if the site is not included (Pressey et al., 1994). Irreplaceability can be measured by the number of times a particular planning unit was included in a solution, as is shown in Marxan’s “Summed Solution” output (Leslie et al., 2003).  35  In our study, the maximum number of times a unit could have been included in the solution is 500. The summed solution maps could also be seen as a proxy for irreplaceability (Figure 2-5). Given these definitions, we can also use these maps to identify hot spots within the study area for particular features or services. If there is a group of planning units that are consistently chosen in the solutions (and are not already protected areas), these areas could be considered priorities for conservation action. There are few defined hot spots for biodiversity features outside of protected areas. The exception to this is an area of high selection, thus greater irreplaceability, along the southern border of the Central Interior eco-province. The sites chosen for ecosystem service reserve networks were highly irreplaceable, with the exception of timber production, in which many of the planning units in the centre of the study area were chosen half of the time. The exception is along the major highways in the area where planning units were chosen in the majority of the runs. For recreational angling and carbon storage certain sites were chosen in most of the solutions and other sites were not chosen at all, offering little variability amongst the solutions. Therefore the “Best” solution maps correspond closely with the summed solution maps. Timber production represents the greatest opportunity cost to conservation in the study area. The hot spot analysis shows general incongruence between an “optimal” timber reserve network and an “optimal” carbon storage reserve network, which may increase the likelihood of carbon storage protection.  36  Figure 2-7: Hot spot maps of individual ecosystem services and biodiversity features  2.3.4  Congruence of ecosystem service areas and biodiversity In order to assess the spatial correlation between “Summed Solutions” for each scenario  we calculated Pearson’s correlation coefficients for each pair of scenarios (Table 2-5). This table gives an indication of how similar the individual networks are to each other, with a score of 1.00 being identical and a score of -1.00 being perfectly negatively correlated. These results do not show a high congruence between areas of high biodiversity and areas of high ecosystem service provision. The strongest positive correlations were between the variations of reserve networks that captured biodiversity features, and more specifically between scenarios which had similar targets and used timber production in their SI. There was a very strong correlation (0.98) between Scenarios 7 and 8. Scenario 7 included recreational angling and carbon storage as side-benefits whereas Scenario 8 used included these side-benefits as well as timber production as additional costs. This similarity is discussed further in Section 2.4.2.  37  Table 2-5: Correlation values between summed solution outputs Scenario 1 2 3 4 5 6A 1 1.00 2 0.47 1.00 3 0.37 0.43 1.00 4 -0.29 -0.39 -0.30 1.00 5 0.91 0.45 0.58 -0.32 1.00 6A 0.79 0.47 0.48 -0.40 0.82 1.00 6B 0.84 0.46 0.23 -0.28 0.70 0.86 7 0.85 0.46 0.27 -0.23 0.72 0.85 8 0.84 0.45 0.23 -0.27 0.70 0.86  2.4 2.4.1  6B  7  8  1.00 0.98 0.99  1.00 0.98  1.00  Discussion Marxan results There is an obvious difference in the spatial cohesion between reserves that used the  ‘feature’ approach vs. networks which used the ‘benefit-cost’ or ‘hybrid’ approaches. We believe that this is because of the greater range of values in any SIs that included ecosystem services. Greater variations in SI values across the study area produced fewer groupings of planning units in the same value range to be chosen together for a “Best” solution. This cohesion could be increased by increasing the BLM, however it is expected that this would also greatly increase the cost of the reserve. The weak and sometimes negative correlations between ecosystem services and biodiversity features shown here echoes past research on ecosystem services in conservation planning (Chan et al., 2006; Naidoo et al., 2008). We cannot assume that by protecting areas for biodiversity we are achieving the goals of ecosystem service provision, and vice versa. Therefore we must clearly define the priorities of the conservation assessment and identify the ways in which ecosystem services can support biodiversity goals without undermining them in the planning process. We have shown that including the marginal economic values of ecosystem services within the cost function of Marxan may be one such method of doing so. This work offers a concrete and spatially explicit example of trade-offs associated with including ecosystem services in biodiversity conservation planning (Nalle et al., 2004; Polasky et al., 2005; Rodriguez et al., 2006; Polasky et al., 2008). The reserve networks that targeted biodiversity and ecosystem services had greater correlation with biodiversity networks than with individual ecosystem service networks, however because of the large number of biodiversity features targeted and the lack of flexibility in achieving some of these targets, this correlation is not surprising. Negative correlations were also found between timber production values and all other scenarios. This is in part because timber production was run to exclude parks, whereas the others included them. Also, timber production reserve networks were most often placed along roads to ensure less costly transportation to  38  processing centers and mills. This distance is also very relevant to the results because timber scenarios were run with a flat cost surface. The spatial incongruence between carbon storage and timber production networks was unexpected. We had anticipated that areas of high timber production value would also host high carbon storage values, but this was not the case. There are two main explanations for this result: topography and data limitations. Firstly, many areas with high current carbon storage values are also areas of variable slope, which greatly increases harvesting costs thus decreasing timber production values. Secondly, data uncertainties within the timber production model also meant that some Timber Supply Areas (TSAs) reported much lower timber values than other TSAs with similar species. This limitation is discussed in further detail in Section 2.4.3. 2.4.2  Unexpected similarity between solutions There was a striking spatial similarity between Scenarios 7 and 8 despite the inclusion of  timber production as a cost in Scenario 8. We had expected that by increasing the cost of particular planning units through the inclusion of timber production values in Scenario 8, Marxan would locate the “Best” reserve in different areas than Scenario 7, in order to avoid planning units with high timber values. However, instead of locating the reserve in radically different areas to avoid timber production, the “Best” reserve in Scenario 8 was similar to that of Scenario 7, but with much higher total costs (Table 2-6). Still, this does not mean that the inclusion of timber did not have any effect on the results. Scenario 7, which did not take any timber production costs into account, did produce a reserve which included higher timber values than Scenario 8 (Table 2-6). It is clear that although Marxan was primarily guided by achieving biodiversity targets, it made small adjustments in Scenario 8 by choosing planning units with negative timber values and avoiding planning units with particularly high timber values, which was not necessarily done in Scenario 7. These small changes were possible because the timber production values were originally modeled at the scale of the individual planning unit. Therefore, in some areas, two planning units that were side by side have very different timber values, which was often a result of varying slopes between units (Fig.2-8). Table 2-6: Breakdown of costs and benefits in “best” reserves from Scenarios 7 and 8 Total Cost Total Cost Recreational Carbon Timber Road (RA, C (RA, C, TP and Scenario angling storage production index and road road index (benefit) (benefit) (cost) (cost) index values) values 7 $63.2B $76.7B $132.0M $7.19B $13.5B $70.5B 8 $64.7B $76.4B $135.0M $7.17B $11.7B $72.0B  39  2.4.3  Data uncertainties Ecosystem service research, and particularly spatially explicit work, is often restricted by  available pre-existing data. In our work we attempted to create simple models in order to measure and map three ecosystem services in the Central Interior of British Columbia. The simplicity of these models is one of their strengths, as the basic methods can be applied in broad range of areas. Our application of these models, however, is limited somewhat by inconsistencies in the original data. In particular, the model of timber production was based on Timber Supply Reviews, which are reports conducted for individual Timber Supply Areas (TSAs) in British Columbia (Ministry of Forests and Range: Forest Analysis and Inventory Branch, 2008). Individual Timber Supply Reviews use different methods to model the volume of timber at the expected minimum harvestable age. These different methods create artificial breaks between administrative boundaries in the values of timber production. These stark discontinuities in values are artifacts of the original data (Fig. 2-8).  Figure 2-8: Artificial discontinuities as a result of data limitations in timber production results  In our model of recreational angling we used an average of recreational costs across the entire study area in the absence of location-specific data. This may produce errors in the results as there may be large differences in angler costs and benefits depending on the size of the lake, the distance of the lake to the angler’s home and the available accommodation near the lake.  40  2.4.4  Increasing the possibility for implementation Including ecosystem service values within the Suitability Index is a novel approach to  Marxan analyses. By including ecosystem services within the cost function of Marxan, instead of as additional features, we were able to create a solution that met all biodiversity targets and with lower costs. However, all of the networks that used a ‘benefit-cost’ or ‘hybrid’ approach (i.e., any approach that included ecosystem services in the SI) consisted of many small patches that are not realistically implementable as conservation areas. As such, further experimentation with Marxan’s parameters is needed to fully explore the potential of this novel approach. It is also important to consider that the values associated with these ecosystem services are potential, not realized benefits. Benefits could be realized if the conservation NGO were to receive funds for carbon offsets for carbon stored on lands that would have been logged; assistance (or less opposition) from the forestry industry in exchange for bypassing conservation protection of high timber-value areas; or from recreational angling groups for conserving areas important to their sport. However, without the political will to make institutional changes, the potential for ecosystem service benefits to increase the likelihood for implementation of a conservation plan is limited. Therefore, the planning solutions discussed here are only the beginning of a process that must continue in the real world. 2.4.5  Interdisciplinary communication Ecosystem service work relies on interdisciplinary communication and research. For  scholars and practitioners who are accustomed to a particular discipline’s agreed upon patterns of communication and jargon, it is often difficult to cross disciplinary lines and understand the goals of others. It is critical that preconceived impressions of what ecosystem services are, how they should be measured, and how these concepts can be used, not limit future work in conservation. Some see ecosystem services, and particularly the economic valuation of these services, to be detrimental to the goals of biodiversity conservation (Rees, 1998; McCauley, 2006; Valiela and Fox, 2008). However, advocates of ecosystem services argue that conserving ecosystem services will also conserve biodiversity (Balvanera et al., 2001; Armsworth et al., 2007). Ecosystem service research bridges social systems and ecosystems, thus it offers an opportunity to involve disciplines within the humanities, social and natural sciences. 2.5  Conclusion We presented a concrete example of integrating ecosystem service values into a  conservation planning project using Marxan. We investigated new methods of integrating these values into the program–both as conservation features with explicit targets and as benefits within the cost layer, or Suitability Index, of the objective function.  41  Including ecosystem service values in the cost layer of Marxan enabled a biodiversity reserve network at a lower cost than if the services had been targeted as additional conservation features to biodiversity. Although these networks were not as spatially cohesive as networks that treated the services as features, we believe that this novel ‘benefit-cost’ approach offers an exciting alternative for explicitly considering ecosystem services within systematic conservation planning. By incorporating compatible ecosystem services within the cost function of Marxan, we achieved all biodiversity targets and supported the primary goals of biodiversity conservation without overshadowing them. However, despite this advance, we are also aware of the many assumptions that were necessary to complete this work. In some of these cases, further spatiallyexplicit studies could provide great value to conservation planning with ecosystem services by better elucidating the marginal changes of ecosystem services across landscapes and as a function of land use options. As well, experimentation with Marxan parameters, such as the BLM and CFPF would provide insight into how easily these reserves could be implemented. Further interdisciplinary research is necessary if we are to expect our conservation plans to satisfy seemingly disparate goals. For example, we must continue to explore how changes in one service impact others, as well as how human communities affect the supply of services. These questions could be partially answered by the development of models that investigate the spatial, ecological and economic relations between humans, the systems that provide ecosystem service provision and systems with high levels of biodiversity. It can easily feel overwhelming, “so many relationships, so little time.” However, we must remember that conservation projects have the ability to benefit both social and natural systems. 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Parkinson, Eric - Senior Scientist, Ecosystems Branch, BC Ministry of Environment, 2202 Main Mall, University of BC Vancouver BC V6T 1Z4. 2008. Personal communication. Parkinson, E., Post, J. & Cox, S., 2004. Linking the dynamics of harvest effort to recruitment dynamics in a multistock, spatially structured fishery. Canadian Journal of Fisheries and Aquatic Sciences, 61(9), 1658. Polasky, Stephen, Erik Nelson, Jeff Camm, Blair Csuti, Paul Fackler, Eric Lonsdorf, Claire Montgomery, et al. 2008. Where to put things? Spatial land management to sustain biodiversity and economic returns. Biological Conservation 141, no. 6 (June): 1505-1524. doi:10.1016/j.biocon.2008.03.022. Polasky, Stephen, Erik Nelson, Eric Lonsdorf, Paul Fackler, and Anthony Starfield. 2005. Conserving species in a working landscape: Land use with biological and economic objectives. Ecological Applications 15, no. 4 (August 1): 1387-1401. doi:10.1890/03-5423. Pressey, R.L., Johnson, I.R. and Wilson, P.D. 1994. 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Zeidler. 2002. Place prioritization for biodiversity content. Journal of Biosciences, 27(4), 339-346. Thomae, O., 2003. East Kootenay Timber Merchantabilty Analysis, Available at: http://www.for.gov.bc.ca/hcp/fia/landbase/MerchantabilityModel2003Revision.pdf [Accessed August 31, 2008]. Toman, Michael. 1998. SPECIAL SECTION: Forum on valuation of ecosystem services: Why not to calculate the value of the world's ecosystem services and natural capital. Ecological Economics 25, no. 1 (April): 57-60. Valiela, I. and S. E. Fox. 2008. Managing coastal wetlands. Science 319(5861): 290-291. http://dx.doi.org/10.1126/science.1153477  46  3 Incorporating ecosystem services thinking into conservation planning: a proposed framework2 3.1  Introduction The concept of ecosystem services has been invoked by a wide variety of academics and  practitioners as a framework for assessing the implications of environmental change for people (Bolund and Hunhammar, 1999; Daily and Ellison, 2002). Here, we define ecosystem services as the flow of direct and indirect benefits from ecosystems to people. This definition is closely aligned to the most widely accepted definition of ecosystem services from the book, Nature’s Services, which was edited by Gretchen Daily in 1997. Here, Daily describes ecosystem services as “the conditions and processes through which natural ecosystems, and the species that make them up, sustain and fulfill human life” (Daily, 1997:3). The most prevalent application of ecosystem service concepts has been to valuate them in economic terms. This valuation allows for their subsequent inclusion in cost-benefit analyses (CBA). In CBA, researchers or decision-makers typically seek to maximize net benefits through a framework that entails the assumption of substitutability between values. In other words, any values to be included, including ecosystem service benefits, are aggregated and compared with costs of different management scenarios in order to facilitate decision making. More recently there has been a call for the consideration of ecosystem services in conservation planning. By including human benefits within conservation plans, it is hoped that the opportunity costs of conservation will be perceived as lower, thus strengthening arguments for biodiversity conservation (Balvanera et al., 2001; Singh, 2002). Indeed, the number of conservation planning projects that consider ecosystem services either implicitly or explicitly has risen steadily since 2000, yet there are still large gaps in this field of knowledge (Egoh et al., 2007). Egoh et al. (2007) reviewed three ways that ecosystem services have been accounted for in conservation assessments. The first two methods involve using biodiversity pattern and ecological processes as proxies for ecosystem service provision. The third method is the explicit mapping of ecosystem services across particular landscapes (Chan et al., 2006; Naidoo and Ricketts, 2006; Egoh et al., 2008; Polasky et al., 2008). In a small number of studies, researchers have also used these maps and values of ecosystem services to create spatially explicit targets for them within conservation planning software programs like Marxan (Chan et al., 2006; Ball and Possingham, 2000). Although this software has been recently applied to ecosystem services, systematic conservation planning tools and frameworks were created to be used with traditional forms of biodiversity data such as the distribution of species occurrences across space.  2  A version of this chapter will be submitted for publication. Hoshizaki, L., and Chan, K.M.A. Incorporating ecosystem service thinking into conservation planning: a proposed framework.  47  We see two distinct approaches for including ecosystem services within a conservation planning program such as Marxan. The first, a ‘feature approach’, is to set targets for the amount of each ecosystem service one would wish to protect in addition to traditional biodiversity features such as rare plant or animal species (e.g., Chan et al., 2006). Chan et al.’s (2006) approach involved modeling or representing the values of ecosystem services in the central coast ecoregion of California and identifying targets for each service; these targets were then included within Marxan solutions alongside biodiversity targets. This approach ensures that a particular amount of the supply or value (realized or potential) of each ecosystem service is included in the reserve, treating ecosystem services as features prioritized for explicit protection alongside biodiversity. In the second ‘benefit-cost approach’, services could be considered as costs or sidebenefits to conservation. Effectively, this involves taking a CBA approach and assuming that the services are fungible with each other and other costs, such that one unit of a service can be substituted by a certain amount of other resources. This second approach considers how the provision of ecosystem services makes areas more suitable for biodiversity conservation, but does not attempt to protect them explicitly. We seek to answer Egoh et al.’s (2007) call for an appropriate framework to integrate ecosystem services within conservation planning, using Marxan as a site selection tool. Although Egoh et al. (2007) do not distinguish between mapping the relative supply of ecosystem services and mapping their economic values, we have chosen to focus on economic values. Using economic values eases direct comparisons with opportunity costs to conservation, as well as for possible inclusion in cost-benefit analyses outside of the systematic conservation planning framework. We begin by providing a brief rationale for assigning economic values to ecosystem services within conservation planning. We then examine the past applications of ecosystem services within cost-benefit analyses and the theoretical differences between their use in CBA and their inclusion in conservation planning as features. We use this background to explore the integration of ecosystem services within Marxan, presenting a framework for choosing between the two approaches to appropriately represent different kinds of services in different circumstances. In particular, we ask the following questions: •  How can we apply ecosystem service thinking in order to most effectively attain the goals  of biodiversity conservation and ecosystem service protection? •  Under which circumstances is it appropriate to employ a ‘feature approach’ and consider  ecosystem services as features to be targeted by conservation vs. a ‘benefit-cost approach’, by which services are incorporated into the cost function of a program such as Marxan as sidebenefits or opportunity costs?  48  3.2  Economic values of ecosystem services Both approaches we discuss for including ecosystem services in conservation planning  may include the economic valuation of these services, a practice that has been critiqued by many (Rees, 1998; Ludwig, 2000; Gatto and De Leo, 2000; McCauley, 2006). However, a large proportion of ecosystem service research has been devoted to improving valuation methodologies, and improvements are always on-going (Costanza et al., 1997; Wilson and Carpenter, 1999; de Groot et al., 2002; Bateman et al., 2006; Turner et al., 2003; Brouwer, 2000). By estimating economic values of ecosystem services, we are able to incorporate services into a systematic conservation planning process by comparing and integrating them with other costs or benefits. Alternatively we can also use these values as an indication of a service’s relative importance to stakeholders in the planning process and use them to valorize conservation and so enhance its political feasibility (Satterfield and Kalof, 2005). 3.3  Prevailing applications of ecosystem service thinking: valuation and cost benefit analyses Economic values for ecosystem services have often been intended to internalize  traditional environmental externalities in cost-benefit analyses (CBAs). These CBAs have been used by decision-makers for a wide variety of policy questions, further proving that in some cases it may be more beneficial to attach a monetary value to nature and consider it in decision-making rather than ignore the impacts on nature altogether (Costanza et al., 1997; Heal, 2000b). Economic values for ecosystem services have been applied by a wide variety of stakeholders in order to respond to the needs of environmental management decisions (Daily & Ellison, 2002). In some situations, such as managing the Catskills watershed to provide clean drinking water for New York City, the decision was simple – improving the health of the ecosystem cost less than a new filtration plant (Heal 2000a). Similarly, it has also been shown that there are major economic benefits provided by natural pollination services to the production of coffee, a valuable export commodity in Costa Rica (Ricketts et al., 2004). However, in other circumstances, whether or not ecosystem service valuation will support a decision that is more costly in economic values, is less clear. In Napa, California residents opted to restore their floodplain to its natural condition and preserve the sinuosity of the river rather than straighten it and build levees and dykes to guard against flooding. The cost of natural restoration was greater than the traditional methods proposed by the Army Corps of Engineers, however the perceived additional benefits that included recreational opportunities, tourism, enhanced property values, and fish and wildlife habitat tipped the scale, demonstrating that local residents attributed nonmonetary values and anticipated many long-lasting benefits of a “living river” (Turner and Daily, 2008).  49  These case studies are examples of how ecosystem service thinking has been applied in the past. By attaching economic values, we connect market prices to the perceived “gratis” work of nature (Ricardo in Heal, 2000b) and can factor these values into decision making. As with any tool, CBAs have their share of benefits and challenges. 3.3.1  Benefits of ecosystem services in CBA: internalizing externalities Although some scholars have stated ethical problems with attaching monetary values to  ecosystem services, many argue that unless we do so, the environment will continue to be ignored in most decisions and policy making arenas (Mooney and Ehrlich, 1997). Not all values can be measured quantitatively for appropriate inclusion in a CBA (Chan et al., in press). However, at times, certain (acknowledged) underestimates are more helpful than no estimate, and it must be appreciated that many, if not most, of the decisions that affect our environment are often based largely on economic measures (Armsworth et al., 2007). This language of dollars and cents (but not necessarily sense) is globally understood and allows for a direct comparison of trade-offs between different land-use scenarios (Rodriguez et al., 2006). By applying ecosystem service thinking to this model, we gain a more realistic appreciation of the challenges inherent in balancing the needs of humans and natural systems and can offer possible solutions. Finally, a major benefit of including ecosystem services within a cost-benefit analysis is the opportunity to engage and gain support from stakeholders who might otherwise not be interested in what may be “environmentally friendly” solutions. By demonstrating the economic value of these services and the ways in which their protection can increase human welfare, a wide variety of stakeholders may become advocates for other conservation initiatives (Chan et al., 2007). CBA offers a unique opportunity to include ecosystem service thinking in conservation planning by simultaneously acknowledging both market and non-market benefits. 3.3.2  Challenges of CBA: what, where and who to count Despite a strong belief that some economic measure of ecosystem services is better than  none, it is obvious that not every service can be included within a CBA. Furthermore, some services—such as the provision of subsistence harvest experiences—are associated with multiple dimensions of value such that monetary values are likely to bear little relation to the service’s true value. Those conducting CBAs necessarily make decisions based solely on the information that is available to them in a common unit of measure, typically dollar values. If a service has been over/underestimated this may lead to an inappropriate CBA and a subsequently costly outcome for both economic and environmental interests. For this reason it is imperative that ecosystem services are clearly defined and measured if they are to be included with other measures of well-being in a formal economic analysis (Banzhaf and Boyd, 2005). An incorrect valuation of services can also occur when economic values from one location are assigned to another, seemingly similar location. This method of valuation, called  50  benefit transfer, is often employed because there is a lack of site-specific data (Costanza et al., 1997; Troy and Wilson, 2006; Plummer, 2009). This dearth of data is arguably one of the single biggest issues in ecosystem service research today and greatly impedes the accuracy of most CBAs (MA, 2005; Turner and Daily 2008). The impacts of this data deficiency become exacerbated when we attempt to ensure that economic values foster equitable decision-making across social groups and generations as we and others would prefer (Costanza and Folke, 1997; Howarth and Farber, 2002). Unfortunately, inter-generational equity is simply ignored by CBA, which focuses only on net benefits and not the distribution of these benefits (Goulder and Stavins, 2002). Finally, one of the largest ethical pitfalls of cost-benefit analysis is its necessary assumption of fungibility. By aggregating all values it is assumed that they are all equally tradable and thus substitutable. This is clearly not the case with all ecosystem services. For example, recreational services can be provided by both natural (parks) and artificial (gymnasiums) environments. However it is unlikely that all beneficiaries of this service would place the same value on walking on a treadmill as they would to hiking to a mountaintop. 3.4  Recent applications in conservation planning More recently, ecosystem services have played an important role in the field of  biodiversity conservation planning after a call was made by the Millennium Ecosystem Assessment to increase conservation actions that could provide benefits to human well-being (MA, 2005). In most cases these services have been included as economic values; however, some studies have also chosen to map the provision of services in biophysical terms, such as amount of freshwater provided by a watershed (Egoh et al., 2008). Although this shift in application appears as a novel way to increase stakeholder engagement in biodiversity conservation implementation, explicitly linking biodiversity conservation with ecosystem service concepts is not new (Mooney and Ehrlich, 1997). Indeed, the relation between biodiversity and ecosystem service provision has often been used to give further weight to arguments for biodiversity conservation (Chapin III et al., 2000; Balvanera et al., 2001). Given this connection, explicitly accounting for ecosystem services in conservation planning is a natural progression. Much recent literature on ecosystem services and biodiversity conservation begins with a focus on the latter and uses the former to demonstrate win-win scenarios (Chan et al., 2007). In this way, ecosystem service thinking is being applied to make the environmental and economic benefits of conservation explicit through spatial trade-off scenarios and correlation analyses (Nalle et al., 2004; Polasky et al., 2005; Rodriguez et al., 2006; Nelson et al., 2009). This new direction in the field of ecosystem services introduces the need for spatially explicit modeling and mapping. In some cases, these maps have been used to show a positive relationship between areas of high biodiversity and ecosystem service provision (Naidoo and Ricketts, 2006; Egoh et  51  al., 2008) and in other cases there is a greater tension between biodiversity and ecosystem service goals (Chan et al., 2006; Naidoo et al., 2008). These relations can be further explored with computer algorithms, such as Marxan, which can consider both ecosystem services and biodiversity goals (Chan et al., 2006; Ball and Possingham, 2000). This synthesis of traditional biodiversity conservation planning and ecosystem service thinking is an exciting frontier for conservation, but we should take pause and consider the various factors that affect the appropriateness and effectiveness of including ecosystem services within conservation plans. 3.4.1  Benefits of including ecosystem services in conservation planning: for people and biodiversity Generally, the application of ecosystem service thinking to conservation planning differs  from its application in cost-benefit analyses because of the differences in the primary goal of each exercise. In cost-benefit analyses, the goal is to maximize net benefits associated with different management decisions, land-use scenarios or policy options. However, in most conservation planning exercises, the primary goal is to achieve a certain minimum level of biodiversity conservation. In this way, studies seek to protect biodiversity while offering side benefits of possible economic and human well-being benefits associated with ecosystem services. In the traditional conservation planning framework, certain ecosystem services, such as carbon storage, as well as existence- and option values, are retained just by virtue of being provided by the ecosystems in protected areas. By being explicit about the provision of these kinds of values as side-benefits of conservation, the number of beneficiaries from conservation planning with ecosystem services grows beyond the borders of the protected area. Finally, a major benefit in applying ecosystem service thinking to conservation planning is the possibility of alleviating tensions between conservation efforts and local communities. In some cases, conservation areas have locked local residents out of their own land, inevitably leading to an unsustainable system where the goals of conservation are undermined by the people closest to the land (Chan et al., 2007). However, by engaging local residents in ecosystem service valuation assessments and by planning for and communicating the tangible benefits that ecosystem service protection offers, conservation may be more successful in the long term (Turner and Daily, 2008). 3.4.2  Challenges of including ecosystem services within conservation planning: serving disparate needs Applying ecosystem service thinking to conservation planning assumes compatible goals.  This is not always the case, which sets up a possible collision of agendas (Chan et al., 2007). For example, if we wish to maximize carbon sequestration, a regulating service, it may be most efficient to replace diverse natural land cover with eucalyptus trees that are able to sequester large amounts of carbon in a relatively small period of time (Arroja et al., 2006). Not only may the  52  goals of biodiversity protection and ecosystem services provision require different species compositions, they may also have different spatial scales for habitat protection. Natural pollination of crops is an example of an ecosystem service that requires small patches of natural habitat in proximity to agricultural landscapes. However, this small scale is counter to that of reserve networks, which seek to protect large areas of contiguous natural land cover (Kremen et al., 2004; Chan et al., 2006). In many circumstances, protecting an area for biodiversity will also entail the provision of some ecosystem services, but what do we do when prioritizing one feature/service incurs a cost to another? Some conservationists would argue that the needs of biodiversity must be met first and that ecosystem services should be included in conservation only when they lend support to its primary goals (Salafsky, 2008). This problem becomes more complicated, however, when funding for conservation projects is based more on the promise of ecosystem service delivery than biodiversity protection. As the concept of ecosystem services becomes more popular amongst policy makers and funding agencies, a pattern has emerged that biases support for conservation projects that include ecosystem services as goals. For example, conservation projects of The Nature Conservancy that explicitly list ecosystem services as goals received almost four times the funding of projects that list only biodiversity goals (Goldman et al., 2008). Since ecosystem services and biodiversity goals do not align perfectly, what should be prioritized, and will the ecosystem-service goals distract conservationists from biodiversity goals? This tension requires reflection on the goals of individual organizations in order to maintain a clear and commonly understood vision. 3.5  Systematic conservation planning How can conservation planners improve on systematic conservation planning methods in  order to increase the prevalence of win-win scenarios and serve the goals of both biodiversity conservation and ecosystem service provision? The steps involved in planning begin with determining the primary objective(s) of the project, then compiling a library of data for the study region. Based on the compiled data, conservation targets are set to define the desired amount of protection for each feature, such as a rare species, habitat type, or ecosystem services such as carbon storage. After the appropriate targets are agreed upon, existing reserves should be reviewed to determine if, and to what extent, they contribute to the current targets (Margules and Pressey, 2000). Data can then be input into a program, such as Marxan, that offers possible solutions at the lowest possible cost while still meeting all targets (Ball and Possingham, 2000). The cost of the sites is based on user input and is often (at best) a proxy for the multiple possible costs of conservation such as land acquisition and management (Naidoo et al., 2006). The costs also reflect the spatial design of the reserve network, because the cost of a reserve with many small patches will generally be more expensive to manage than a contiguous reserve. 53  3.6  A framework for deciding how and when to include ecosystem services within conservation planning The call has been made for a greater integration of ecosystem service values in  conservation planning (Egoh et al., 2007). With few exceptions (see Chan et al., 2006) there have been few contributions to the literature that investigate concrete integrations of ecosystem services within systematic conservation planning software such as Marxan. The majority of the discussion has been somewhat abstract, although it is generally assumed that the inclusion of ecosystem services in conservation plans will reinforce the goals of biodiversity conservation. But whether and by how much such inclusion reinforces conservation will depend in part on how these services are included. Their inclusion must also depend on the purpose of the plan, whether it is education and research or if it is intended for actual implementation. This final distinction is crucial as we investigate the ability of ecosystem services to increase the likelihood of conservation implementation. Below, we propose two approaches for including services in conservation planning, and we illustrate the framework using concrete examples. How ecosystem services are included in conservation plans should depend largely on the stakeholders and beneficiaries of the project. Here, we define beneficiaries as anyone who will benefit from the conservation action. Stakeholders are defined as those that are affected, either positively or negatively, by the conservation action. The conservation organization will often identify who their stakeholders are and try to address their needs to ensure the highest amount of co-operation. The identities and characteristics of these stakeholders should inform the selection of services, their representation in Marxan, and most importantly, the goals for each in individual conservation planning exercises and the priority of these goals. For example, in an area with a large community of recreational anglers, efforts could be made to ensure that favorite fishing spots are maintained in exchange for angler’s support of conservation actions. As well, if the stakeholders in a conservation project rely on clean water flowing from the conserved area, ensuring clean drinking water should be prioritized within Marxan with sufficient targets set to meet demand. These decisions have important implications for how we should apply a program such as Marxan to represent ecosystem service goals in systematic conservation planning. Our two approaches outline the circumstances in which we should be considering ecosystem services as features or as costs within Marxan and conservation planning generally. 3.6.1  A feature approach The choice to represent ecosystem services in conservation planning as features or as  costs (or side-benefits) is a crucial one. One key element of the difference is the issue of substitutability, which we have discussed above. Including services as costs or side-benefits is the same in this respect: it treats services as substitutable. To include a service as a feature in Marxan generally means meeting the target specified for that feature, even if it requires considerable additional habitat and effort. Accordingly, this ‘feature approach’ entails treating  54  features as independently valuable, not as substitutable for one another. Doing so sets up a potential conflict: if including services as features does not result in additional resources for conservation, services and biodiversity features compete for protection unless their spatial distributions overlap substantially, which seems rare (Chan et al., 2006; Naidoo et al., 2008). When we include ecosystem services as features alongside biodiversity features in Marxan 2.0.2, we are required to use the same cost surface for all. This effectively assumes that the same management costs are equally applicable for biodiversity features and ecosystem services yet, with most services, this assumption is not appropriate. For example, Marxan uses the amount of patchiness in a reserve as a proxy for cost because it assumes that many species require a contiguous area to persist. However, in the case of carbon storage the level of storage will not depend on the contiguity of the network. Marxan was designed to create reserves for biodiversity features, driven by ethical imperatives; how often this applies appropriately for particular ecosystem services remains to be seen. Ecosystem services that are subject to legal requirements, such as the provision of opportunities for subsistence and ceremonial harvest for First Nations people in Canada, are one example of services driven by clear ethical imperatives and are good candidates for being treated as features. The spatial and temporal scale of both the supply and demand for a service will influence how it is best integrated within Marxan. Recreational angling is a service whose supply is limited to few, particular areas across a landscape. As well, the demand for this service is often relatively local and therefore it seems more appropriate to include this service as a feature to be targeted for a specific amount in each sub-region to fulfill the demands of local beneficiaries. Furthermore, the ‘feature approach’ is justifiable when addressing equity issues and the particular distribution of a service across the land or seascape. For example, we would be justified in treating food production as a feature to be ensured at given levels in various geographical areas if beneficiaries were unable to secure food through inter-regional trade because of poverty and marginalization (van Jaarsveld et al., 2005). As well, the current (in)ability of valuation methods to accurately represent many cultural services means that these services cannot be included within a cost layer amongst dollar values (Chan et al., in press). Therefore they may be best represented as features within a conservation plan in order to ensure their provision in the landscape. Finally, whether the service can be classified as a public or private good should also be considered before including it as a feature or as a cost in Marxan. Ecosystem services that are public goods are by definition non-excludable (i.e., it is impossible to prevent people from consuming the good) and, because they offer little opportunity for private profit in provision, are very likely to be underprovided (Heal, 2000a). Therefore, they may be more appropriately included as features so that they can be guaranteed a certain level of provision within a reserve  55  system. An example of this could be aesthetic values on public lands. Therefore, individual targets should be set for stratification units in order to ensure the provision of these beneficial public goods (Chan et al., 2006). 3.6.2  A benefit-cost approach In some circumstances it will be more appropriate to represent ecosystem services as  costs or side-benefits in the cost layer of Marxan. In our previous study we included timber production in the cost function of Marxan as an opportunity cost to conservation because areas in which high values of timber production are realized (i.e., harvested) are often less probable to be reserved for conservation (Chapter 2). Conversely, carbon storage and recreation values were included in the cost layer as benefits (and subtracted from the other costs), thereby decreasing the final cost layer by introducing the possibility of conservation-friendly revenue schemes and making a particular area more desirable for protection. Including ecosystem services in the cost function of Marxan corresponds somewhat to cost-benefit analysis in that all the values are aggregated; it also means that services may support the protection of biodiversity features and not compete with them. Given this, services should only be included within the cost layer if they can be and have been appropriately valued in economic terms. As mentioned above, this quantification is often not possible for cultural services (Chan et al., in press). By including services in the cost function, we also avoid the task of choosing monetary targets for certain ecosystem services, which can be difficult, ethically questionable, and possibly arbitrary. For example, it is hard to justify requiring Marxan to meet a specific target when there may be a high possibility of substitution for the service provided in a particular location. An example of a highly substitutable service is carbon storage for climate regulation. This service should be included within the cost function of Marxan because a loss of carbon stores in the study area could be mitigated by reduced fossil fuel emissions or carbon sequestration elsewhere. In the BC case-study (Chapter 2), the provincial government has set emissions reduction targets, with a preference for local projects, so despite global atmospheric circulation, there is a pertinent component of demand at the scale of the province (Ministry of Environment news release, 2008). In this case, it is less meaningful to designate particular local targets within a relatively small conservation region when the demand could be met in other areas of the province. Including ecosystem services within the cost function internalizes their values to the conservation planning process in the sense that they are automatically considered in the resulting plan. However, doing so (for a plan that is intended to be implemented as opposed to one for educational purposes) is only appropriate if the values can realistically be internalized to the conservation implementation process. Thus, if high values for the ecosystem services in areas conserved actually translate into lower costs, increased revenues, or enhanced feasibility of  56  conservation, their inclusion in the planning process may increase the feasibility of implementation. Services that may be good candidates for this criterion are recreational angling, carbon storage and timber production. For example, recreational anglers are frequent allies of conservation organizations and would be more likely to donate to projects if they believed conservation would benefit them. As a second example, as the province attempts to reduce its greenhouse gas emissions, including carbon storage in a conservation plan may be financially supported by the government due to reduced GHG emissions by means of avoided deforestation. Finally, the timber industry is a potent political player that prevents conservation actions in areas of high timber value (Karmona, 2007). Therefore it may be appropriate to consider the opportunity costs of timber production within the planning process. However, when we include ecosystem service values as costs and/or benefits we must consider the likelihood of these values actually coming to fruition. The probability of actually obtaining benefits from ecosystem services will be increased if there are institutional measures already in place, such as carbon markets. If this is not possible, for example if carbon storage benefits are included for an area that does not participate in a carbon trading market, Marxan analyses could also be used as educational exercises. In closing, when it is inappropriate to include ecosystem services as features within Marxan, it may still be appropriate to include them within Marxan’s cost layer in order to envision how the services might help an organization meet the goals of biodiversity at a lower cost. 3.7  Conclusion Conservation planning offers a venue in which ecosystem services can be incorporated  in decision-making in diverse ways, and according to characteristics of each. Unlike cost-benefit analyses, conservation planning assessments (such as using Marxan software) offer the opportunity to treat ecosystem services as independently valuable (i.e., non-substitutable). The field of systematic conservation planning has been often criticized for not explicitly considering socio-economic concerns early enough in the process (Margules and Pressey, 2000; Richardson et al., 2006; Wilson et al., 2007). From this perspective, the inclusion of ecosystem service values that represent benefits to humans, promises to advance the field of conservation planning by explicitly considering the needs of people. Considering ecosystem services may also enhance local buy-in to a conservation project and possibly offer new ways to finance conservation, increasing the probability of implementation (Wilson et al., 2007; Goldman et al., 2008). The first step in including ecosystem services in a systematic conservation planning framework is to determine the goals of the project and the needs of the stakeholders. These needs and goals should greatly influence the Marxan parameters and frame the analysis. Asking particular questions at the beginning of the exercise will determine whether a feature approach or a cost approach is appropriate. For example, are biodiversity features the priority? Then the 57  ‘benefit-cost approach’ is appropriate in order to include ecosystem services as opportunity costs, or added benefits within the cost function, if at all. Is a particular service required in an area? If so, it should be targeted as a separate feature, using the ‘feature approach’. These are ultimately internal decisions to be made by an informed conservation organization. External stakeholders may also factor in to the framing of the exercise as different ecosystem services will have unique implications for particular agencies or partners. For example, a government’s forestry office may be highly interested in preserving areas of high timber production value, and they may have the political power to ensure this area is prioritized for timber production, not conservation. Therefore, conservation agencies may use Marxan to identify reserves that host high timber values in order to gain an idea of where the greatest barriers to conservation exist (Ban, 2008). In most cases these insights will not heavily influence the final plan, as conservation assessments prioritize biodiversity conservation first, arguing that we need to conserve biodiversity regardless of human costs and benefits. In this paper we have presented two approaches for including ecosystem service values within Marxan. We also recognize that ecosystem services and biodiversity may not be spatially congruent and the tools for biodiversity conservation planning may not always be appropriate for ecosystem service planning. In particular, biodiversity conservation planning generally involves designing compact reserves to minimize costs. Accordingly, there is a risk that ecosystem services will be insufficiently represented because services differ in several critical ways from biodiversity features: ecosystem service values depend upon the demands of beneficiaries, and these social characteristics will vary spatially and temporally, in addition to the variation of biophysical characteristics of both biodiversity features and services. For example, conservation planning does not currently offer straightforward ways to account for how the value of a service will change based on the proximity of service-production to beneficiaries. To use an example, a park that is close to an urban area but equal in other respects should be of higher value than one which is more inaccessible to recreationists (Chan et al., 2006). Finally, we cannot assume that all ecosystem service values can be internalized to assist conservation implementation. The inclusion of ecosystem services to a conservation plan may not assist in implementation at all, and may be beneficial from an educational perspective only. Nevertheless, the accurate portrayal of ecosystem services and their values within conservation planning offers a means of enhancing biodiversity protection while also improving human wellbeing in a crowded world.  58  3.8  References  Armsworth, P. R., K. M. A. Chan, G. C. Daily, P. R. Ehrlich, C. Kremen, T. H. Ricketts, and M. A. Sanjayan. 2007. Ecosystem-Service Science and the Way Forward for Conservation. Conservation Biology 21, no. 6: 1383-1384. doi:10.1111/j.1523-1739.2007.00821.x. Arroja, L., A. Dias, and I. Capela. 2006. The Role of Eucalyptus Globulus Forest and Products in Carbon Sequestration. Climatic Change 74, no. 1 (January 12): 123-140. doi:10.1007/s10584006-3461-1. Ball, I.R. and H.P. Possingham , 2000. MARXAN (v.1.8.2): Marine reserve design using spatially explicit annealing, a manual. Balvanera, P., G.C. 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Environmental and Resource Economics 39, no. 1 (January 12): 25-35. doi:10.1007/s10640-007-9176-6. Turner, R. Kerry, Jouni Paavola, Philip Cooper, Stephen Farber, Valma Jessamy, and Stavros Georgiou. 2003. Valuing nature: lessons learned and future research directions. Ecological Economics 46, no. 3 (October): 493-510. doi:10.1016/S0921-8009(03)00189-7. Westman, Walter E. 1977. How Much Are Nature's Services Worth? Science 197, no. 4307. New Series (September 2): 960-964. Wilson, Kerrie A., Emma C. Underwood, Scott A. Morrison, Kirk R. Klausmeyer, William W. Murdoch, Belinda Reyers, Grant Wardell-Johnson, et al. 2007. Conserving Biodiversity Efficiently: What to Do, Where, and When. PLoS Biology 5, no. 9 (September 1): e223 EP -. doi:10.1371/journal.pbio.0050223. Wilson, Matthew A., and Stephen R. Carpenter. 1999. Economic Valuation of Freshwater Ecosystem Services in the United States: 1971-1997. Ecological Applications 9, no. 3 (August): 772-783.  62  4 Conclusions on present work with directions for future research 4.1  Integrating ecosystem services within conservation planning This research builds upon previous work in systematic conservation planning and the  more recently emerging field of ecosystem services, specifically the economic valuation and mapping of ecosystem services. In doing so, this work contributes to the small but growing set of case studies, methods and theoretical frameworks for creating spatially explicit maps of ecosystem services for conservation planning (Chan et al., 2006; Naidoo and Ricketts, 2006; Egoh et al., 2008). Ecosystem service literature is becoming broader as a wide range of interdisciplinary scholars begin to apply these concepts to their particular field of study. With this growth, definitions of ecosystem services and ideas regarding how these concepts are most effectively applied have become the focus of some debate. Indeed, one only has to read the introductions of a few ecosystem service papers to gather a few contrasting definitions of ecosystem services. In this work, we adopt a definition and conceptual framework of ecosystem services similar to that of Daily (1997). In this work, we have defined ecosystem services as the provision of both direct and indirect benefits to people from ecosystems. In Chapter 2, following this definition, we mapped and valued (in economic terms) the marginal values of three ecosystem services in the Central Interior region of British Columbia. This work was done in conjunction with the Nature Conservancy of Canada, which is in the midst of creating a conservation plan for the study area and they will be considering our ecosystem service values in their plan. We used our ecosystem service maps in conjunction with economic models to determine the value of ecosystem services that would be lost or gained between two land-use scenarios: conservation and timber harvesting. We were then able to identify areas in the study region that would benefit the most from conservation. In a novel ‘benefit-cost’ approach, we also included ecosystem service values in the cost layer of the conservation planning software, Marxan, in order to create a reserve that conserved biodiversity targets, but also explicitly considered ecosystem services in a way theoretically similar to their traditional application in cost-benefit analyses. The theoretical differences between using ecosystem services as side-benefits or costs, rather than as targeted features, in Marxan’s objective function was explored further in Chapter 3. There, we answered a call by Egoh et al. (2007) for a theoretical framework to include ecosystem services in conservation planning projects. We continued to use the Marxan case study from Chapter 2 to illustrate this framework and characterize two separate approaches: net benefit maximization (‘benefit-cost approach’) and cost minimization (‘feature approach’). Net benefit maximization as an approach aligns with the theories of cost-benefit analysis and includes ecosystem services in Marxan as costs or benefits. A cost minimization approach considers  63  ecosystem services as conservation features alongside traditional biodiversity features targeted for conservation. In Chapter 3 we outline the circumstances in which we believe one approach may be more appropriate than the other, depending on the characteristics of the ecosystem service and its beneficiaries. We also consider which approach will ultimately create the most effective conservation reserve for protecting biodiversity and or ecosystem services under various circumstances. The results in Chapter 2 demonstrate that the most effective reserve for both biodiversity and ecosystem service protection was found using the ’benefit-cost approach’, although the solution was less spatially cohesive than the reserve created using the ‘feature approach’. In Chapter 3, we qualify these results by stressing the importance of the individual situation. The spatial and temporal dimensions of the supply and demand for an ecosystem service, as well as the population defined as beneficiaries of a particular service, will ultimately dictate the most appropriate method of including ecosystem service values into a systematic conservation plan. Ultimately, these two chapters nest together by using a case study in Chapter 2 as a concrete example that illustrates the different approaches outlined in Chapter 3’s proposed theoretical framework. 4.2  Strengths and weaknesses of research  4.2.1 4.2.1.1  Strengths Interdisciplinarity This research was done in consultation with a wide range of experts from the fields of  forestry, fisheries, and carbon storage modeling. We also consulted with hydrologists in the beginning phases of the research when we intended to measure water provision and flood mitigation services. We believe that the breadth of expert advice we received offers substantial strength to this work. This wide range of input, as well as theories and methods borrowed from both the natural and social sciences make this work truly interdisciplinary. A considerable amount of effort was expended in communicating across disciplines to ensure that an accurate portrayal of ecosystem service concepts was understood by the different experts. This understanding guaranteed that the final work would accurately portray the concepts of ecosystem services without betraying the biophysical and/or economical underpinnings of the individual services. For example, through consultations with experts from the Forestry Department at the University of British Columbia we ensured that all major costs of timber harvesting were included in the timber production model. Similarly we also communicated extensively with experts from the Canadian Forest Sector regarding their Carbon Budget Model in order to understand the biophysical dynamism inherent in carbon cycling and to better inform our understanding of how carbon storage in the study area may change with land-use alteration.  64  4.2.1.2  Simplicity The models used to measure the ecosystem services are intentionally simple. Their  simplicity allows for greater ease of use in other locations and situations as the data they require is commonly available. The carbon storage data used have global coverage and is publicly available from the World Resources Institute (Matthews et al., 2000). Recreational angling was modeled using available indices of fisher effort and watershed sensitivity. This information is only available for the province of British Columbia, but could be replicable for other areas using relatively simple data sets such as slope, soil and density of alluvial channels. Recreational angling economic statistics are also available publicly and have been compiled for the entire country. The timber production model was the most data- and time-intensive to develop due to the majority of information being only available in print format. However, our study, along with the merchantability index created by Thomae (2003), could be used as a measure of possible costs and benefits for timber production as our methodology relies on publicly available data. 4.2.1.3  NCC biodiversity data The data collected by the Nature Conservancy of Canada was used as a measure of  biodiversity in Chapter 2 and gives great strength to this work. The primary goal of the NCC conservation assessment of the Central Interior was to map areas of high biodiversity value for conservation planning. Therefore the coarse and fine filter biodiversity data was collected and analyzed by a wide collection of experts working on the conservation assessment. These regional experts identified which rare species and terrestrial systems to be included as well as their appropriate conservation targets within Marxan runs. Because of provincial government participation in this NCC project, we were also given access to data that would ordinarily not be publicly available and are considered to be the most detailed and accurate information available for the region. 4.2.1.4  Applicability Finally, a major strength of this work, as demonstrated above, lies in its direct connection  to the Nature Conservancy of Canada’s eco-regional assessment of the Central Interior. The inclusion of our research findings in the NCC’s final report is a testament to its utility to this organization and their conservation planning process. However, although this two-way relationship provided data on ecosystem services to the NCC and on biodiversity for our Marxan analyses, there were also challenges encountered regarding data sharing and timelines. These difficulties are to be expected in such a massive research undertaking that includes groups from government, academic and NGO sectors, but are often less problematic in solely academic research. These connections across institutions are necessary in order to effectively implement ecosystem service research.  65  Thus some of the insights we have offered in Chapters 2 and 3 have come about via the first hand witnessing and participating in a real-world conservation planning exercise, which lends them greater legitimacy. We believe that this experience has offered us a realistic understanding of the obstacles faced by conservation organizations in both the planning and implementation phase of biodiversity conservation assessments. 4.2.2 4.2.2.1  Weaknesses Omission of climate change considerations The case study presented in Chapter 2 was meant to identify areas in the Central Interior  where ecosystem services are the most threatened by forestry, or alternatively where conservation could have the greatest impact on the provision of these services. Timber harvesting was chosen as the alternative land-use scenario because it is believed to be major local threat to the goals and implementation of conservation actions (Iachetti, 2008). However, another leading threat to biodiversity and the provision of ecosystem services in the Central Interior of British Columbia is climate change (Boon, 2007; Hamann and Wang, 2006; Loukas et al., 2002). The complexity and dynamism of climate change has and its impact on ecosystem services is beyond the scope of this research. It is nevertheless necessary to identify this omission from our research, especially in light of the mountain pine beetle (MPB) epidemic, which is exacerbated by climate change. The extent of damage caused by the mountain pine beetle in the study area and the subsequent salvage logging that has taken place has major implications for biodiversity and ecosystem services in the area. Our work however does not account for these impacts in our models and subsequent maps. An example of the impact of climate change on ecosystem service provision in the study area is demonstrated by recent research from Kurz et al. (2008). This study shows much of the forests in the Central Interior becoming a source for carbon, rather than a sink, which would alter our results in Chapter 2. However, the MPB-affected areas and the associated prescribed salvage logging plans are in constant flux, making modeling their impacts on ecosystem services a task that was outside the scope of this research. Our main intentions for this work were to explore the marginal values of ecosystem services in the study area and introduce a conceptual framework for including them into conservation planning in a broadly applicable way. However, the problems of climate change and, in the Central Interior, mountain pine beetle, are great and have major local ramifications that deserve further exploration. We explored the impacts of climate change on ecosystem services briefly in a report that was submitted to the NCC’s Climate Change team (Appendix 4.1).  66  4.2.2.2  Limited number of ecosystem services We were only able to model and map three ecosystem services in the study area.  Although we had intended to include freshwater provision, we subsequently deemed it to be inappropriate given local reliance on groundwater, which is not influenced substantially by timber harvesting in this region. However, flood mitigation is an important ecosystem service for this particular area and is expected to be directly impacted by timber harvesting. Unfortunately, due to time constraints, we were unable to investigate this service despite its local significance. Only modeling three ecosystem services limited the analyses that we could perform regarding the correlation between services and biodiversity. Therefore, we are unable to say a great deal about the interactions between ecosystem services in the study area. However, this is a common problem as each service is complex and time-consuming to map and model. As well, services are constantly interacting with each other and producing a wide range of direct and indirect benefits that we may not be aware of or have the capacity to measure. 4.2.2.3  Data inconsistencies and model simplicity Chapter 2 relies on multiple secondary data sets, all of which contain uncertainty. The  carbon storage data used have a coarse resolution of 83 km² in comparison to the 500 hectare, or 5 km², planning units. Also, the estimation of a 10% loss of carbon in the transition from a conserved to a harvested landscape was taken from the literature but could have been more accurately described had we been able to directly apply a model such as the Carbon Budget Model of the Canadian Forest Sector (CFS, 2006) to individual planning units. The timber production model relied on information gathered from individual Timber Supply Reviews (TSR) for each Timber Supply Area (TSA) in the study area. These reviews provided information on the minimum harvestable ages for a particular species and gave an indication of the projected timber volume at this age. However, these ages and volumes are modeled differently across TSAs and the results reflect these administrative boundaries to such a degree as to introduce artifacts. In retrospect we may have obtained more meaningful results had we applied the minimum harvestable ages and expected volumes reported by the largest TSA across the entire study area. This treatment may have been insightful as we could more easily compare the relative changes in timber production values in the study area. However, by using the individual TSR data for each corresponding TSA we were able to produce results which were consistent with the other information available for each area and could be seen as more useful to individual TSA managers. As was stated in Chapter 2, the recreational angling model used to measure the economic losses in the face of timber harvesting was very simple. The relationship between watershed sensitivity and economic losses is undoubtedly more complicated than we have  67  portrayed it (as a negative linear one-to-one relationship) and could be refined to more accurately describe the impacts of forestry on recreational angling values in the study area. Although we believe that the simplicity of our models allows for their broad application, some may view this aspect of the work to be a weakness. For example, our carbon storage model is extremely simple and does not explicitly consider fluxes in the carbon cycle. Therefore, although the simplicity of our models means that they could be applied anywhere on the globe, the results may be less helpful locally, as they are rough estimates, and not guaranteed economic values. 4.2.2.4  Lack of local input We identified the ecosystem services of interest in the study area by surveying the NCC  Central Interior team of experts who had prior experience working in the region. This research would have produced a richer picture had we traveled to the study region to speak with local stakeholders. Local communities are an important source of data for ecosystem service projects. Their participation could have enriched this research by helping to identify which ecosystem services are of greatest interest in the area as well as given an indication of how these services are valued by the community. Local people may also be a resource of detailed, spatially explicit knowledge as well as traditional ecological knowledge which could be used to enrich non-site specific biophysical models. Ultimately, we believe that ecosystem service research should be centered on the people who benefit from the work that ecosystems provide and our present work would be more informed by including the opinions of those living amongst the ecosystems we studied at a distance. 4.3 4.3.1  Future research directions Definitions Without a clear, commonly shared definition of what ecosystem services are, research in  this field has become splintered in many directions. The concepts of ecosystem services are broad and applicable to many disciplines and this is why, in part, we believe they are so powerful. Yet this array of definitions also creates confusion amongst those less familiar with the ideas and history behind them. Research in the field of ecosystem services, especially in connection to conservation, has proliferated in the past decade. However, this growth has been primarily in the methodological aspects of modeling, valuing and mapping services. We must also begin to ask more theoretical questions, like those posed in our third chapter. How do different groups of people define ecosystem services and how do these differences affect their research? Do people performing ecosystem service research have the same fundamental goals? Should they?  68  4.3.2  Biophysical data collection and application In the field of ecosystem services in conservation planning, many theoretical and  methodological questions still remain. Despite many calls in the ecosystem service literature, rigorously collected and detailed biophysical data are still needed in order to accurately portray the supply side of ecosystem services. These data could be collected through basic environmental monitoring and surveying via mean such as weather stations, river gauges, and topographical surveys. Sophisticated models must also be developed in conjunction with data collection in order to play out possible scenarios and offer greater understanding of impacts such as climate change. Especially in areas where demand for services such as freshwater and flood mitigation is high, empirical research and data collection are needed to ensure greater certainty in models that predict the provision of services. If we are to argue that conservation will provide such necessary life-sustaining services we must be confident in our predictions. 4.3.3  Focus on people Future ecosystem service research should also have a greater focus on local  beneficiaries, especially in the case of conservation planning. In many areas, conservation is seen as a threat to local livelihoods and without a say in the process many people will feel alienated from the research that says it is explicitly considering their needs (Chan et al., 2007). Research and conservation actions that include ecosystem service concepts should explicitly account for greater spatial and temporal equity amongst beneficiaries. The framework that we present in Chapter 3 offers the opportunity to consider the demand side of ecosystem services as well as the provision, or supply, of them. This means recognizing more than just the needs of current generations but the needs of future generations as well. It also includes attempts at equitably distributing services across space, such as providing freshwater downstream and/or across national borders. Because of gaps in ecosystem service research, cultural ecosystem services such as aesthetic enjoyment, intellectual development and spiritual fulfillment have also often been overlooked (de Groot et al., 2005). The methods for valuing these services are still new and under scrutiny by many (Chiesura and de Groot, 2003; Chan et al., in press). However, despite these inevitable growing pains, more case studies that value and map cultural ecosystem services are needed to fill this space in the literature and to be integrated with natural science ecosystem service research. Just as we require more accurate biophysical datasets, we must also keep asking the difficult questions of how ecosystems contribute to aesthetic pleasure and spiritual fulfillment of people.  69  4.3.4  Political will for implementation Perhaps the greatest challenge in the field of ecosystem services and conservation  planning revolves around the collection of the required political will for the implementation of conservation plans. Despite the importance of local people’s values and opinions, they rarely have the opportunity to influence policy and decision making. Ecosystem service and conservation planning researchers must make every effort to connect with local policy makers in order for their work to be relevant to society. As academics, we cannot rely on others to seek out and interpret our findings but instead raise our own voices in the communities that we live in.  70  4.4  References  Boon, Sarah. 2007. Snow accumulation and ablation in a beetle-killed pine stand in Northern Interior British Columbia. BC Journal of Ecosystems and Management 8, no. 3: 1-13. CFS, (Canadian Forest Sector). 2006. Carbon Budget Model of the Canadian Forest Sector (CBM-CFS2). http://carbon.cfs.nrcan.gc.ca/cbm/index_e.html. Chan, K. M. A., R.M. Pringle, J. Ranganathan, C.L. Boggs, Y.L. Chan, P.R. Ehrlich, P. K. Haff, N.E. Heller, K. Al-Khafaji, and D.P. Macmynowski. 2007. When Agendas Collide: Human Welfare and Biological Conservation. Conservation Biology 21, no. 1: 59-68. doi:10.1111/j.15231739.2006.00570.x. Chan, Kai M. A., M. Rebecca Shaw, David R. Cameron, Emma C. Underwood, and Gretchen C. Daily. 2006. Conservation Planning for Ecosystem Services. PLoS Biology 4, no. 11 (November 1): e379 EP -. doi:10.1371/journal.pbio.0040379. Chan, Kai M. A. Joshua Goldstein, Terre Satterfield, Neil Hannahs, Kekuewa Kikiloi, Robin Naidoo, Nathan Vadeboncoeur, and Ulalia Woodside, in review, "Cultural Services and Non-Use Values", in Peter Kareiva, Gretchen Daily, Taylor Ricketts, Heather Tallis and Steve Polasky, eds., The Theory & Practice of Ecosystem Service Valuation in Conservation, Oxford University Press. Chiesura, Anna, and Rudolf de Groot. 2003. Critical natural capital: a socio-cultural perspective. Ecological Economics 44, no. 2-3 (March): 219-231. doi:10.1016/S0921-8009(02)00275-6. Daily, Gretchen C., ed. 1997. Nature's Services: Societal Dependence on Natural Ecosystems. Washington DC: Island Press. Egoh, Benis, Belinda Reyers, Mathieu Rouget, David M. Richardson, David C. Le Maitre, and Albert S. van Jaarsveld. 2008. Mapping ecosystem services for planning and management. Agriculture, Ecosystems & Environment 127, no. 1-2 (August): 135-140. doi:10.1016/j.agee.2008.03.013. Egoh, Benis, Mathieu Rouget, Belinda Reyers, Andrew T. Knight, Richard M. Cowling, Albert S. van Jaarsveld, and Adam Welz. 2007. Integrating ecosystem services into conservation assessments: A review. Ecological Economics 63, no. 4 (September 15): 714-721. doi:10.1016/j.ecolecon.2007.04.007. de Groot, R., P. Ramakrishnan, and A. Berg. 2005. Cultural and amenity services. Ecosystems and Well-being: Current Status and Trends. Milliennium Ecosystem Assessment. Washington DC: Island Press. Hamann, A., and T. Wang. 2006. Potential effects of climate change on ecosystem and tree species distribution in British Columbia. Ecology 87, no. 11: 2773-2786. Iachetti, Pierre - Director of Conservation Science & Planning, Nature Conservancy of Canada, BC Region, 300 -1205 Broad Street Victoria, BC V8W 2A4. 2008. Personal communication. Kurz, W. A., C. C. Dymond, G. Stinson, G. J. Rampley, E. T. Neilson, A. L. Carroll, T. Ebata, and L. Safranyik. 2008. Mountain pine beetle and forest carbon feedback to climate change. Nature 452, no. 7190 (April 24): 987-990. doi:10.1038/nature06777. Loukas, Athanasios, Lampros Vasiliades, and Nicolas R. Dalezios. 2002. Climatic impacts on the runoff generation processes in British Columbia, Canada. Hydrology and Earth System Sciences 6, no. 2: 211-227.  71  Matthews E, Payne R, Rohweder M, Murray S . 2000. Pilot analysis of global ecosystems: Forest ecosystems. Washington, DC: World Resources Institute. Naidoo, Robin, and Taylor H. Ricketts. 2006. Mapping the Economic Costs and Benefits of Conservation. PLoS Biology 4, no. 11 (November 1): e360 EP -. doi:10.1371/journal.pbio.0040360. Thomae, Oliver. 2003. East Kootenay Timber Merchantabilty Analysis. June. http://www.for.gov.bc.ca/hcp/fia/landbase/MerchantabilityModel2003Revision.pdf.  72  APPENDICES Appendix A: Data sources and methods for ecosystem service modeling and valuation Carbon storage The World Resources Institute provided a dataset for carbon storage in above and below ground vegetation and soil, in an 82.81km² grid (Matthews et al., 2000; Loveland, Reed & Brown, 2000). The values of this grid, which measured carbon in tons/ha, were disaggregated to 500-ha hexagons using area weighted averages. To determine the difference in carbon storage between an actively forested landscape and a conserved landscape, we conducted a literature review of carbon storage changes in landscape transitions. Kurz et al. (1998) provided a relevant case study as a basis for our work as they modeled carbon changes between a managed and natural forest in the Interior region of British Columbia. This study reported a 10% decrease in stored carbon when an area transitioned from a natural to sustainably harvested state. Their model also accounts for losses in carbon due to natural disturbances, such as fire and pests. Similar losses were also found by Sanscrainte et al. (2007), and Leighty et al. (2006). Therefore, we assumed that the difference in stored carbon between a conserved and harvested landscape was 10%. However there was no available data to tell us what areas of the landscape were currently being harvested and which were in a natural state. Therefore, if we assume that the entire landscape is currently in a natural state, i.e. 100% of possible carbon storage potential, the difference between conservation and forestry will be a 10% loss. However, if we assume that the entire landscape is being harvested, i.e. 90% of possible carbon potential, the potential gain in carbon storage would be 11%. To determine the average difference in carbon storage between the two scenarios, we used the more conservative percentage difference, 10%. Therefore the benefit function to describe the change in carbon storage between a conserved and forested landscape was as follows: ∆Carbon Storage = 0.1 x Current carbon storage CO2 storage was valued at $8.46 (CDN) per ton using the mid-price average of three carbon trading markets: the Chicago Climate Exchange, the New South Wales and the EU Emissions th Trading scheme on March 19 , 2008. This value was then multiplied by 500 and assigned to individual hexagons in the study area. Timber production Timber production was considered as an opportunity cost of conservation, therefore the difference in production values were effectively 100%. These values were measured in a way that considered both the merchantability of the timber within a particular hexagon as well as the net present value of that timber. To measure timber production value across the landscape, a spatially explicit database was created using a variety of sources (Table A1). Hexagons were removed from the database if they were considered to be outside of any possible timber harvesting land base (THLB). Because the THLB changes over time, we assumed hexagons will not be harvested if their centroid was within a land use that would not be suitable for timber harvesting (Figure A1). For each remaining hexagon within the THLB, an average leading species was determined using Vegetation Resource Inventory (VRI) raster data and zonal statistics in ArcGIS 9. The economic values  73  assigned to the hexagon were based on this species and assume that the species which covers the majority of the hexagon is a representation of the range of values present within the hexagon. Table A1: Spatial data sources for timber production Data Leading species  Resolution 1 ha  Site index  1 ha  Leading age  1 ha  Biogeoclimatic zones Slope  1:20, 000 25m  Mill locations  unknown  Source Vegetation Resources Inventory via HectaresBC Vegetation Resources Inventory via HectaresBC Vegetation Resources Inventory via HectaresBC Ministry of Forests and Range Nature Conservancy of Canada Ministry of Forests and Range  Figure A1: Land uses assumed to be outside timber harvesting land base • • • • • • •  Alpine and Tundra BEC zone Bunch Grass BEC zone Urban Lakes Shrub – from Baseline Thematic Mapping Range – from Baseline Thematic Mapping Parks and Protected Areas  Each hexagon was assigned a leading species, an average site index of that species and the Timber Supply Area (TSA) number that it was in. Using this information we reviewed the Timber Supply Review (TSR) Analysis report for each hexagon and determined the minimum harvestable age (MHA) and Volume of harvestable timber present at that age. The MHA is the earliest age at which the timber could be harvested and using this value offers a conservative estimate of volume and market value. The leading species and volume at MHA determined the net benefit of each hexagon based on average prices from the 2003 – 2008 BC Interior Log Market reports (Ministry of Forests and Range, 2003). Costs were calculated as a summation of harvesting, transportation and silviculture costs. Harvesting costs were based on the average slope of the hexagon and transportation costs were based on the distance of each hexagon to the nearest primary processing facility using the spatial join function in Arc. Both costs were measured using an index from a similar merchantability assessment done in the Interior of BC (Thomae, 2003) and were assumed and confirmed to be appropriate by experts. A loop function was written in Visual Basic to determine the NPV of each hexagon given its total value as well as the number of harvesting rotations expected within a given time period (Figure A2). The costs and benefits of each hexagon were assumed to be constant throughout the time frame. The number of harvests, or rotations, within the time period was determined by the minimum harvestable age of the stand, as well as its current age. If the average current age of the hexagon was greater or equal to its MHA, it was assumed that the hexagon would be harvested immediately. If the stand’s age was less than the MHA, the function used the difference in years between the current age and the MHA to account for the time before the initial harvest. After determining the number of rotations, the NPV was calculated using a time frame of 1000 years  74  and a discount rate of 4%. Both of these figures are commonly used in forestry value assessments (Alaouze 2004; Heal 2000; Creedy & Wurzbacher 2001).  Figure A2: VBA script for NPV of timber values 'B is benefits--assumed to be constant 'C is costs--assumed to be constant Function NPV_tree(B, C, r, Tp, MHA, age) If (age >= MHA) Then init_harv = 0 'If the current age of the stand is greater or equal to the MHA then there will be a harvest immediately. 'Rot is the # of rotations Rot = Round((Tp / MHA) + 0.5) ' If the current age is greater or equal to the MHA, then the number of rotations will be the time period (1000)/MHA plus 1 (for the immediate harvest). Adding .5 and rounding effectively adds one harvest rotation. Else Rot = Round((Tp - (MHA - age)) / MHA) + 0.5 ' If the current age is less than the MHA then the number of rotations will be the time period minus the difference between the MHA and current age divided by the MHA. This will be rounded up to account for the time spent reaching the initial harvest. init_harv = MHA - age 'If the stand's age is less than the MHA, the initial harvest will occur once the stand has reached the MHA End If ti = init_harv ' Time when harvesting begins Sum = 0 Dim i As Integer 'declares i as an integer For i = 0 To (Rot - 1) 'for loop function Sum = Sum + (B - C) / ((1 + r) ^ ti) ' Summation of NPV function ti = ti + MHA Next i NPV_tree = Sum Recreational angling The expected decrease in the economic values of recreational angling in the study area due to timber harvesting was measured using two separate models. Angler effort, or the amount of angling days that would be supported by a particular lake, was modeled previously by Eric Parkinson (Parkinson, Post & Cox, 2004). These data were only available for non-stocked lakes however this was not seen as a limitation as the amount of angling available in artificially stocked lakes is assumed to be independent of surrounding timber harvesting. Effort was modeled as a function of lake productivity, distance from major population centers and proximity to roads. The amount of effort per lake was converted into dollar values by dividing the total amount of days fished by the amount of revenue generated in that year. This figure includes, but is not limited to, money spent on transportation, licenses, equipment and package tours. These revenue statistics are only relevant for freshwater regions in British Columbia and are derived from the Survey of Recreational Fishing in Canada (Government of Canada, 2005). rd  The second set of modeled data was used to classify each 3 order watershed in the study area with a relative sensitivity to timber harvesting score. In general, when we discuss sensitivity we are referring to the likelihood of increased sedimentation within streams. To rank this sensitivity, we used data from the Fisheries Sensitive Watershed (FSW) database that evaluates watersheds in British Columbia that may be particularly susceptible to a decrease in fisheries due to logging (Reese-Hansen & Parkinson, 2006). The variables included in the FSW index are: soil type,  75  annual precipitation, lake buffering capacity, amount of area with gradients over 60%, amount of forest land cover, and the density of alluvial channels. For example, a watershed with highly erodible soil, heavy annual precipitation, low lake buffering capacity, high slopes, large amount of available timber and a high density of alluvial channels would be considered highly sensitive to logging. Using a GIS, these sensitivity variables were combined with equal weighting within a multi-criteria evaluation. There is current debate regarding the combination of these different variables however at the time of writing equal weighting is common practice. All watersheds were then given a relative ranking from 0 -1.0. The amount of effort was assigned to the entire watershed that the lake was present in to create a continuous surface and scale of both effort and sensitivity. rd  To disaggregate the individual 3 order watersheds into 500ha planning units for a Marxan analysis, we assumed that steep planning units will contribute to economic losses more than flat ones. However, all planning units will contribute to an area’s overall sensitivity. Given this, we assigned individual values to planning units using the equation: l = (L/N)*((s+x)/(S+x)) Where: l = economic loss attributed to individual planning unit L = total economic losses in watershed N = number of planning units in watershed s = average slope of individual planning unit x = 49% (largest planning unit slope) S = average slope of watershed References: Alaouze, C.M., 2004. The Effect of Conservation Value on the Optimal Forest Rotation. Land Economics, 80(2), 209-223. Creedy, J. & Wurzbacher, A.D., 2001. The economic value of a forested catchment with timber, water and carbon sequestration benefits. Ecological Economics, 38(1), 71-83. Government of Canada, F.A.O.S.S., 2005. 2005 Survey of Recreational Fishing in Canada. Available at: http://www.dfo-mpo.gc.ca/communic/statistics/recreational/Canada/2005/toc_e.htm [Accessed May 5, 2008]. Heal, G., 2000. Nature and the Marketplace: Capturing the Value of Ecosystem Services, Washington DC: Island Press. Leighty, W., Hamburg, S. & Caouette, J., 2006. Effects of Management on Carbon Sequestration in Forest Biomass in Southeast Alaska. Ecosystems, 9(7), 1051-1065. Loveland, T., Reed, B. & Brown, J., 2000. Development of a global land cover characteristics database and IGBP DISCover from 1 km AVHRR data. International Journal of Remote Sensing, 21(6-7), 1303-1330. Matthews E, Payne R, Rohweder M, Murray S , 2000. Pilot analysis of global ecosystems: Forest ecosystems, Washington, DC: World Resources Institute. Ministry of Forests and Range, 2003. Log Market Reports - Revenue Branch Publications - MFR Province of B.C., Available at: http://www.for.gov.bc.ca/hva/logreports.htm [Accessed August 31, 2008].  76  Parkinson, E., Post, J. & Cox, S., 2004. Linking the dynamics of harvest effort to recruitment dynamics in a multistock, spatially structured fishery. Canadian Journal of Fisheries and Aquatic Sciences, 61(9), 1658. Reese-Hansen, L. & Parkinson, E., 2006. Evaluating and Designating Fisheries Sensitive Watersheds (FSW): An Overview of B.C.'s New FSW Procedure, BC Ministry of Environment. Available at: http://www.env.gov.bc.ca/wld/documents/fsw/FSW%202006%20Information%20Paper%20v1.1.p df [Accessed May 2, 2008]. Sanscrainte, C.L., Peterson, D.L. & McKay, S., 2007. Carbon storage and soil properties in latesuccessional and second-growth subalpine forests in the North Cascade Range, Washington. Northwest Science 77(4):297-307 Thomae, O., 2003. East Kootenay Timber Merchantabilty Analysis, Available at: http://www.for.gov.bc.ca/hcp/fia/landbase/MerchantabilityModel2003Revision.pdf [Accessed August 31, 2008].  77  Appendix B: Details of Marxan scenarios Table B1: Marxan scenarios for study region  Suitability Index  Features  Road Index Road index including carbon storage and recreational angling Flat SI = 500ha (area of planning unit)  Biodiversity x  Recreation x  Carbon storage x  Timber production  Rec., Carbon, & Bio. x  x x  Table B2: Targets for individual ecosystem services Feature  Conservation Goal Unit  Standard Targets  Percentage of total value in study area  Timber production  CDN Dollar  $35.6 B  75%  Carbon storage  CDN Dollar  $10.0 B  50%  Recreational Angling  CDN Dollar  $169.9 M  50%  Rationale Goals relate to the creation of a timber reserve. This target is not congruent with biodiversity goals and should not be run with biodiversity features, or other ecosystem services. Precise targets for market-based ecosystem services are difficult to assign because of the unpredictability of markets. Also, there is no ethical imperative to ensure a minimum representation of market-based services through the use of goals. In the face of this uncertainty, we have suggested general target scenarios. See above: Carbon storage.  78  Table B3a: Details of Marxan scenarios for Sub-boreal eco-province  3  Scenario  Features  SI  BLM  CFPF  Avg. Score  1 2 3 4 5 6A 6B 7 8  Biodiversity Recreational Angling (RA) Carbon Storage (C ) Timber Production (TP) Biodiversity, RA and C Biodiversity, RA and C Biodiversity, RA and C Biodiversity Biodiversity  Road index Road index Road index Flat (500 per hexagon) Road index Road index and TP Road index and TP Road index, RA and C Road index, RA, C and TP  1 1 1 1 1 1 1 1 1  10 10 10 10 10 10 10 10 10  3.40E+07 9.11E+06 2.35E+07 1.50E+07 3.62E+07 5.32E+10 4.16E+10 2.97E+10 3.7715E+10  Avg. Cost 1.53E+07 3.48E+06 1.06E+07 6.63E+06 1.72E+07 5.07E+10 3.90E+10 2.67E+10 3.46E+10  Avg. Boundary Length 1.85E+07 5.64E+06 1.29E+07 8.34E+06 1.86E+07 4.68E+07 3.87E+07 3.75E+07 3.82E+07  Table B3b: Details of Marxan scenarios for Central Interior eco-province Scenario  Features  SI  BLM  CFPF  Avg. Score  1 2 3 4 5 6A 6B 7 8  Biodiversity Recreational Angling (RA) Carbon Storage (C ) Timber Production (TP) Biodiversity, RA and C Biodiversity, RA and C Biodiversity, RA and C Biodiversity Biodiversity  Road index Road index Road index Flat (500 per hexagon) Road index Road index and TP Road index and TP Road index, RA and C Road index, RA, C and TP  1 1 1 1 1 1 1 1 1  10 10 10 10 10 10 10 10 10  3.53E+07 1.25E+07 2.91E+07 1.13E+07 3.77E+07 5.63E+10 4.78E+10 3.79E+10 4.3519E+10  Avg. Cost 1.92E+07 5.57E+06 1.72E+07 5.42E+06 2.22E+07 5.53E+10 4.67E+10 3.70E+10 4.25E+10  Avg. Boundary Length 1.60E+07 6.90E+06 1.18E+07 5.91E+06 1.54E+07 3.29E+07 2.70E+07 2.63E+07 2.67E+07  3  The values for the boundary length modifier (BLM) and conservation feature penalty factor (CFPF) were determined by the Nature Conservancy of Canada.  79  Appendix C: Suitability index transformation The road index, or suitability index, was used as a cost layer in some Marxan runs. This layer was used to overlay the road index scores, used as a proxy for cost, between 0.0086 and 0.73 to all planning units in the study area. It represents the assumption that the cost of conservation in a planning unit is high if that unit has high road density and/or is in close proximity to roads. In particular Marxan runs, we also included ecosystem service values within the Suitability Index. To do this, we transformed the road index value to a dollar value that we could then add or subtract our ecosystem service dollar values to and create a cost layer that considers both roads as well as ecosystem services.  To transform the road index into dollar values, we used land market acquisition values in the study region as a substitute, effectively assuming (1) that with increased road density or proximity to roads, we can expect greater urbanization and an increase in land acquisition costs; (2) that land acquisition costs are an important contributor to, and an appropriate proxy for, total costs; and (3) that land acquisition costs are over-represented by land prices to the same degree that total costs are greater than land acquisition costs. Land costs for rural areas were provided by the Nature Conservancy of Canada and costs for urban areas were found on the Multiple Listing Service (MLS) of the Canadian Real Estate Association.  We used a four part linear transformation to assign dollar values to each planning unit, based on four land values, which were placed at frequency peaks in the distribution of road index values. We used the equation y = mx + b where: y = Land acquisition cost x = Road index score m = (y² - y¹) / (x² - x¹) b = y¹ - m¹*x¹ Table C1: Values used in linear transformation of Road Index (SI) to dollar values using local land acquisition costs ($) $/Planning unit 617750 1482600 4942000 7536550 1606150000  SI 0.0086 0.017 0.35 0.5 0.7  y1 617750 1482600 4942000 7536550  y2 1482600 4942000 7536550 1606150000  x1 0.0086 0.017 0.35 0.5  x2 0.017 0.35 0.5 0.7  m 102958333.33 10388588.59 17297000.00 7993067250.00  b -267691.66 1305993.99 -1111950.00 -3988997075  80  Appendix D: Climate change in the Central Interior of BC – Impacts on ecosystem services and adaptation strategies The impacts of climate change on ecosystem services in the Central Interior region of British Columbia demands separate consideration from impacts on biodiversity, in part because we are not necessarily concerned with differences between fine and coarse filter targets. Instead, we are concerned with the impact of climate change on both the provision and demand for ecosystem services in the study area. The following brief report attempts to outline possible impacts on the provision of and the demand for our chosen ecosystem services: carbon storage, timber production, freshwater provision, flood mitigation and recreation. We discuss each service in its own section and conclude with possible strategies for adapting to future changes in services attributable to climate change. We conclude by discussing the benefits and limitations associated with the integration of a suite of Global Climate Models (GCMs) within the planning process: they add a spatially explicit dimension to possible impacts, but their interpretation must be respectful of the range of uncertainties inherent in these models.  Introduction Ecosystem services are “the conditions and processes through which natural ecosystems, and the species that make them up, sustain and fulfill human life” (Daily et al., 1997). They have also been more simply defined as the benefits that people obtain from ecosystems (MA, 2005). The concept of ecosystem services is just that – a concept. In a planning exercise they may be most usefully considered in terms of their biophysical supply, as well as human demand for them. It is less appropriate to separate these services into fine and coarse filter targets... When we begin to consider the impacts of climate change on the supply and demand of these services, it may be difficult to delineate the appropriate scope of analysis of the interactions between different ecological processes. For this reason, we are limiting the impacts that we consider to those directly related to changes in the following: •  Precipitation  •  Temperature  •  Threats directly associated with climate change (as mediated by precipitation and temperature) such as mountain pine beetle infestation, forest fires, and extreme weather events such as storms.  •  Demographics as they relate to a change in demand for services.  This report is divided into subsections for each ecosystem service; it outlines several key ways that ecosystem service supply and demand may be impacted by climate change. We then offer ways in which the Central Interior project could influence the mitigation of these impacts through various management options. Much of this thinking has been based on the expected changes to historical baseline climate scenarios as described via ClimateBC (Spittlehouse, 2006). We feel that the incorporation of these GCMs along with others, will ultimately allow for a richer understanding of climate change impacts in the study area. 81  ClimateBC We feel that ClimateBC would offer beneficial spatially explicit information when estimating impacts of climate change on ecosystem services. The computer program uses different global climate models (GCMs) to inform their predictions (Hamann and Wang, 2006). The historical data are at a fair grid resolution of 2.5 arcmins, however the predictive scenario results are on a 1° grid. We would appreciate a fuller understanding of how these different data resolutions have been combined. Applications of ClimateBC have predicted warmer and wetter winters with less precipitation in the late winter and summer. It also predicts a general trend of higher temperatures in the spring and summer, which would result in an earlier snowpack melt (Spittlehouse, 2006). Keeping these general thoughts in mind we can more specifically discuss the effects of climate change on ecosystem services and what we could do to adapt to these changes.  Carbon storage Impacts In general, we expect that the carbon sequestration rates would change with a change in temperature and precipitation in BC. Decomposition rates may increase with higher temperatures resulting in a decrease of carbon storage, however, tree growth may also increase, which would mitigate losses of stored carbon. High tree mortality rates due to MPB may also negatively affect carbon storage in the study area. Although new trees will likely grow and store carbon eventually, these secondary (or tertiary) forests generally store less carbon than their old growth predecessors. However, despite these changes, it is important to note that carbon storage is a service that is highly substitutable with other measures to mitigate climate change.  Adaptation We might mitigate some of these possible changes in carbon storage by investigating the relatively high carbon content in woody debris and coarser soils. Management scenarios that were informed by spatial knowledge of high carbon areas could then consider their actions to conserve carbon. These could take the form of not clearing away debris after logging and encouraging the preservation of coarse soil types. As stated previously, climate change will affect this ecosystem service, but there is such a high possibility for substitution available that it may be more appropriate to adapt to these impacts on carbon storage through actions that decrease greenhouse gas emissions.  Timber production Impacts on growth rates It is generally predicted that there will be an increase in temperature due to climate change. In the temperate climate of BC, we can therefore expect growth rates to increase and an increase in possible 82  2  timber production in the long term future. The increase in CO concentrations in the atmosphere can also be expected to increase growth rates. 2  The likely benefits of increased temperature and CO concentrations may be limited or negated by changes in precipitation. If a location experiences an increase in temperature but maintains constant precipitation, this will likely increase evapotranspiration and decrease water availability and summer tree growth. In the summer, ClimateBC predicts less precipitation; we therefore expect an overall decrease in growth in many water-limited areas. Threats and vulnerabilities The impacts of Mountain Pine Beetle (MPB) on timber production will differ depending on the time scale we choose to examine. In the short term, timber production will increase because of the immediate need to salvage dead trees before the timber is no longer valuable. In the medium term future, we expect a decrease in timber production because of the time lag between replanting and a stand reaching maturity. With climate change we would also expect a change in the range of local harvestable tree species. These changes could leave the trees vulnerable to attack from other, currently unknown pests, which would also create a flux in timber production. Not only are pests an issue, but invasive plants could displace harvestable species and threaten timber production. Planted monocultures exacerbate the threat of widespread disease and pest damage and may therefore cause subsequent decreases in timber production. Monoculture regimes would also be contrary to biodiversity goals. A decrease in precipitation means an even greater vulnerability to forest fires, and subsequent losses in timber production. Higher temperatures could also increase fire risk due to extreme weather causing lightning strikes, as well as increase the potential fuel load because of increased biomass. This contributes to the need for replanting and possible problems related to that process. Adaptation To adapt to these possible issues, we must carefully consider which trees are replanted to ensure adequate species diversity to mitigate risks associated with poor climate suitability and to safeguard against threats like pests and invasive species, but also choose species which are still economically profitable. The choice of species would ideally be based on which species are more likely to be productive in the future considering the range of future climate change scenarios. We should consider not only the species of stands, but also their age. Large stands that include same-age trees may also be more vulnerable to infestation and wide spread disease. Finally, we are considering timber production in our assessment because it is an important ecosystem service at it has a high relative contribution to our provincial economy. However, given the major changes that are predicted in the future due to climate change, it may be necessary to discuss ways to diversify the BC economy and become less dependent on the timber sector.  83  Freshwater provision Availability of water ClimateBC predicts generally higher precipitation rates in the winter, so we can expect an increase of freshwater provision in these months. However, this increase in precipitation may also be accompanied by an increase in temperature, which would increase evapotranspiration rates and might maintain available freshwater at baseline levels. There may be more precipitation in the winter, however because of warmer temperatures this precipitation may fall as rain instead of snow, decreasing snow pack reservoirs. What snow pack exists is also expected to thaw earlier in the spring. This early thaw coupled with less precipitation in the summers is expected to decrease freshwater provision in the mid and late summer months, when water scarcity is already an issue. These extreme peaks in annual water availability may also mean that communities become more dependent on constructed reservoirs and dams for their regulation abilities. Quality of water The current increase in MPB infestation, which is linked to climate change, brings an increased need for salvage logging, and therefore an increased probability of soil erosion. This increased sediment load due to logging will probably decrease water quality. We will not only suffer water quality disturbance due to the sediments themselves but also from pathogens that attach easily to sediments and are more difficult to filter and treat. It is the relationship between the pathogens and sediments that cause the greatest threats to human health. Demand for water Higher temperatures in the summer could also result in an increased demand for water provision services, which are likely to increase anyway as a result of population growth and immigration. This demand would not necessarily be met due to decreased stream flow in the summer, which may create tension in some communities. For example, agricultural communities may need to switch to more conservative irrigation technologies, like drip watering. This increased demand for water could also result in higher water taxes and water metering programs to encourage household water conservation, which may have detrimental impacts on certain industries. Demand may also increase for freshwater with a growth in population. At this point, however, we are uncertain what predictions exist for local demographic trends. Adaptation A greater dependency may on regulating infrastructure such as reservoirs and dams may be required to adapt to changes in annual water availability. Also, smaller reservoirs should be maintained and important filtration areas like wetlands should be protected for their natural filtration abilities. Areas that experience particularly high levels of precipitation should also be considered for protection from urban development. Impervious surfaces in urban areas greatly contribute to water quality degradation by contributing pollutants such as oil and grease, as well as inhibiting any natural filtration  84  through soils and vegetation. These impenetrable surfaces also decrease filtration of water to the water table and encourage greater peaks and lows in stream systems. Community watershed programs could be initiated to connect communities with their water systems and encourage household conservation. With an expected increase in water demand, we need to encourage water conservation from the individual household up to the municipal, industrial and commercial levels.  Flood mitigation Decreased ability to control flooding: As stated previously in reference to carbon storage, increased temperatures could also increase the decomposition of biomass. With a decrease in biomass, we would also see a decrease in the levels of standing soil available, and with it a lessened ability to store water.  Seasonality of demand Increased precipitation in the winter could mean an increased risk of winter floods, thus a greater demand or need for natural flood mitigation services. Also, with higher temperatures we can expect melting snowpack earlier in the season, increasing further still the threat of floods in that season. In the summer, however, there is a decreased chance of flooding due to the decrease in precipitation. Generally speaking, there is a greater chance for higher peaks and lower lows because of dramatically variable annual precipitation which may increase annual demand for mitigation and flood regulation. Ice jam floods Flooding due to ice jams may be more prevalent with climate change as well. Ice jams are caused when temperatures fluctuate closely around the freezing point of water. This causes ice crystals to coalesce and layer on top of each other to create an unusually thick cover of ice in streams. When a thaw is sudden, as may result from climate change, these thick layers of ice break apart and create instream jams resulting in floods. Land use change The threat of pests such as MPB and subsequent salvage logging decreases a watershed’s ability to mitigate floods due to the basin’s reduced ability to store water naturally. Also, we can consider that possible future land conversion from forest to grassland or agriculture could decrease flood mitigation capabilities (Naef, Scherrer and Weiler, 2002). An increase in urban development and population may also contribute to greater density of impervious surfaces, like roads. These changes would result in greater peaks in storm water discharge and increase the demand for flood mitigation.  Adaptation Salvage logging plans should be attendant to their impacts on flood mitigation services. These plans could take into consideration logging patterns and amount of coarse woody debris that is left on site. 85  Human-made infrastructure, such as dams and levees may be used to mitigate flood risk and increased seasonal hydrograph peaks. However, we should also target natural flood plains for protection against development, as these areas naturally mitigate risk by allowing seasonal stream fluctuations. Discouraging development in the natural flood plain reduces damage to human infrastructure and nurtures non-human ecological floodplain communities. Adaptation strategies to mitigate increased risk of flooding in the future should also discourage surface impermeability (i.e., paved surfaces) whenever possible. This could mean requiring a certain depth of top soil for all new developments and or protecting areas which have high levels of soil permeability.  Recreation Angling An increase in air temperatures would likely result in an increase in stream temperatures, which would make borderline habitat unsuitable to certain cold-water species of fish, such as salmonids. This degradation of habitat would mean that recreational angling opportunities would also decrease. Stream habitat for desirable angling species may also be degraded by MPB infestation and salvage logging, and by more frequent fires, which would contribute higher amounts of sediment to streams, further diminishing recreational services. In the summer the problem of fish habitat degradation may be exacerbated by a decrease in precipitation and subsequent lower flow volumes in streams and lakes, which will likely result in higher water temperatures. In lakes especially, cold-water refuge areas for certain species may disappear all together. A decrease in the flow volume could also eliminate the necessary pool riffle sequence for spawning if riffles areas completely dry up. These general changes to fish habitat may be undesirable for certain popular angling species, however it may be more conducive for invasive species. These invasive species could then also contribute to the disappearance of current fish through interspecific competition. Outdoor recreation Increased fire risk due to decreased summer precipitation, increased temperature and more frequent storm events may also decrease the supply of outdoor recreational services because of habitat destruction, public safety and aesthetics. However, demand for recreation may also shift because of habitat destruction. If certain recreational areas are not considered favorable any longer, people may choose to go elsewhere. In an angling example, this may increase pressure on fish stocks in lakes previously thought undesirable. This logic suggests that planning needs could be adapted to include areas which may be underused for recreation now, but that may become popular in the future. Adaptation Similarly to carbon storage, there are many substitutes for ecosystem-provided recreation. However, this service is also directly linked to cultural and spiritual services which there are, arguably, very few substitutes for. 86  Points of integration There are multiple points of integration across the range of ecosystem services with the other teams involved in the assessment. Ecosystem services are complex functions of many features across the landscape, therefore it is expected that just as climate change affects terrestrial and freshwater targets, it will also affect the supply of ecosystem services. Examples of this integration include: •  Carbon storage targets influenced by coarse filter terrestrial team input about systems that are rich in carbon content.  •  Timber production adaptation strategies formed with terrestrial teams and forestry experts to identify appropriate species for re-plantation.  •  Coordinate information with the freshwater team regarding expected changes in stream flow.  •  Integrate knowledge regarding the role of wetlands in flood mitigation with terrestrial and freshwater teams, and how this role may be altered with climate change.  •  Coordinate with coarse filter terrestrial teams to determine how climate change may alter watersheds and their ability to mitigate floods on the landscape.  •  Coordinate with freshwater teams to predict the increased likelihood of higher peaks in hydrographs and increased risk of flooding due to climate change.  •  Integrate knowledge from the animals team regarding the vulnerability of certain recreational angling species to climate change.  Further work and considerations We have previously mentioned the utility of using a spatially explicit tool to inform our thinking about climate change and its impacts on ecosystem services. Although we agree that there is a great deal of uncertainty involved in these types of tools, we also believe that they are never the less useful in guiding our thinking. Indeed, many of the assumptions we have used to underline the issues in this report are based on the predictions of ClimateBC. Therefore, we feel it necessary openly communicate how we have used these results and that we are aware of the large probability of error. Secondly, if we are to more appropriately estimate impacts on ecosystem services for the study area it is important to consider the demand for services as well as supply. This demand would most likely be predicted in conjunction with relevant population growth projections for the area, similar to the work of the Millennium Ecosystem Assessment (MA, 2005; Tallis and Karieva, 2006). Finally, future work that considers the impacts of climate change on ecosystem services must explicitly consider the interactions involved between services; does a change in service enhance or degrade the functionality of another? For example, we have anticipated a decrease in timber production for particular areas, but this may be followed by an increase in production in other areas, creating a need for increased accessibility. This increase in the road network may also increase the opportunities for recreation in the area, depending on access rights. Not only could we consider the tradeoffs involved in 87  the supply of services, but we may also consider the shifting of employment opportunities from one sector (timber) to another (ecotourism).  Conclusion In contrast to the other teams involved in this assessment, we have not divided the impacts of climate change on ecosystem services into coarse or fine filter targets. Instead, these impacts are more appropriately described as how they relate to the supply and/or demand of services. In this brief report we have generally described what some impacts of climate change may be on ecosystem services for the central interior region of BC. Adapting to these changes will require cooperative integration across a wide range of stakeholder agencies and should include not only humanmade substitutions, but also the protection of particularly vulnerable and important features on the landscape, such as wetlands and undeveloped flood plains. Finally, many of these services require a spatially explicit investigative approach. An example of this need is found when discussing flood mitigation services. The flood mitigation services of a particular landscape have a unique relationship with areas downstream, and the understanding and accounting of this spatial relationship is crucial for appropriate planning that would also consider future global climate change. Tools such as ClimateBC, which integrates global climate models, coupled with demographic projections regarding demand for services, may aid the process by offering a spatially explicit dimension to the issues that we have outlined in this report.  References Daily, G. C., Ed. 1997. Nature's Services: Societal Dependence on Natural Ecosystems. Washington, DC, Island Press. Hamann, A. and T. Wang. 2006. Potential effects of climate change on ecosystem and tree species distribution in British Columbia. Ecology 87(11):2773-2786. Millennium Ecosystem Assessment (MA) 2005. Ecosystems and Human Well-being: Synthesis. Washington, DC, Island Press. Naef, F., S. Scherrer and M. Weiler. 2002. A process based assessment of the potential to reduce flood runoff by land use change. Journal of Hydrology 267: 74-79. Spittlehouse, D. 2006. ClimateBC: Your access to interpolated climate data for BC. Streamline: Watershed Management Bulletin. 9(2): 16-21. Tallis, H. and P.Karieva. 2006. Shaping global environmental decisions using socio-ecological models. Trends in Ecology and Evolution. 21(10): 562-568.  88  

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