Ecosystem Services

For integrated assessments we have to be aware that each ecosystem service approach contains uncertainties and gener- alizations related to scale issues, methodological issues, vague classifications and definitions (Scolozzi et al., 2012). Several ideas were proposed for a better taxonomy and classification of ecosystem services, because there is still no consensus on a common strategy (Seppelt et al., 2012; Burkhard et al., 2012). Bastian et al. (2012) emphasize the need for a clear differentia- tion between landscape/ecosystem functions, ecosystem services and landscape potentials, for which van Oudenhoven et al. (2012) provide a classification and indicator scheme. The different approaches illustrate that the definition of a common classifica- tion framework remains a major challenge, because ecosystem service studies are too singular, question-dependent and context- related. However, maybe the definition of a common classification framework is neither feasible nor really necessary (Costanza, 2008).

It is also incorrect to suggest (McCauley, 2006) that conserva- tion based on protecting ecosystem services is betting against human ingenuity. Recognizing and measuring natural capital and ecosystem services in terms of stocks and flows is a prime example of enlightened human ingenuity. The study of ecosystem services has merely identified the limitations and costs of ‘hard’ engineer- ing solutions to problems that in many cases can be more efficiently solved by natural systems. Pointing out that the ‘horizontal levees’ of coastal marshes are more cost-effective protectors against hurricanes than constructed vertical levees (Costanza et al., 2008) and that they also store carbon that would otherwise be emitted into the atmosphere (Luisetti et al., 2011) implies that restoring or recreating them for this and other benefits is only using our intelligence and ingenuity, not betting against it. The ecosystem services concept makes it abundantly clear that the choice of ‘‘the environment versus the economy’’ is a false choice. If nature contributes significantly to human well-being, then it is a major contributor to the real economy (Costanza et al., 1997), and the choice becomes how to manage all our assets, including natural and human-made capital, more effectively and sustainably (Costanza et al., 2000).

Most authors agree that the term “ecosystem services” was coined in 1981. It was pushed to the background in the 1980s by the sus- tainable development debate but came back strongly in the 1990s with the mainstreaming of ES in professional literature and with an increased attention to their economic value. Over time, the definitions of the concept have evolved with a focus on either the eco- logical basis as ES being the conditions and processes through which natural ecosystems and their species sustain and fulfil human life or at the level of economic importance, where ES are the benefits humans derive, directly or indirectly, from ecosystem func- tions. As a compromise, the TEEB (The Economics of Ecosystems and Biodiversity) study (2008-2010) defined ES as the direct and indirect contributions of ecosystems to human well-being. Despite these differences, all definitions stress the link between (nat- ural) ecosystems and human wellbeing (see Figure 1) and the services are the ‘bridge’ between the human world and the natural world, with only humans being virtually sep- arated from that natural world.

mius’’—that humans act independently, rationally, and in their own self-interest. Value in this context (E-value) is based on current individual prefer- ences, which are assumed to be fixed or given (Norton and others 1998). No additional discussion or scientific input is required to form these prefer- ences (since they are assumed to already exist), and value is simply people’s revealed willingness to pay for the good or service in question. The best estimate of what people are willing to pay is thought to be what they would actually pay in a well-functioning market. For resources or services for which there is no market (like many ecosystem services) a pseudo- market can sometimes be simulated with question- naires that elicit individual’s contingent valuation.

The focus of this second edition, the ecosystem based management (EBM), is an approach that links the environmental health to the human well-being, as the environment provides valuable natural services, or “Ecosystem Services” to human communities. These services depend on the complexity of ecosystems, the connections among them, the link between land, fresh and salt water, and their integration with the human beings. The application of EBM in marine environments is new, happening mostly as an answer to the deteriorating state of its ecosystems. This necessity overlaps with the “Aichi Biodiversity Targets”, defined in the Strategic Plan for Biodiversity for the period from 2011 to 2020.

Although our ESV data came from the only global data set currently available (Sutton and Costanza 2002), these estimates are subject to a variety of uncertainties (Costanza et al. 1997). To reduce these, our study suggests that research priorities in ecosystem services should include, first, better spatial artic- ulation of services and valuations regionally and globally, and, second, more sophisticated modeling of the biophysical origin and flow of services spatially and temporally. That said, further sensitivity analysis shows that over- or under- estimation of key unit ESVs would produce little change in results. A substantial part of the ESV of biodiversity conser- vation templates is driven by the area and ESV of evergreen broadleaf forest. We thus evaluated the effect of twofold underestimation or overestimation of the ESV of evergreen broadleaf forest. The same eight templates harbored signifi- cantly more ESV than random in any case. Quantitative changes were also minor: the magnitude and ranking of

In this chapter, we compile climate change adaptation options for ecosystem services for the Wet Tropics Cluster (WTC) region, derived from the Australian literature and elsewhere. We focus particularly on water regulation, climate regulation, carbon sequestration, agricultural production, timber production, habitat provision, erosion control, and traditional values. We also discuss emerging opportunities that may become available in WTC region in the future, bring together the limitations and constraints of current payments for carbon abatement and discuss possible ways to establish payments for ecosystem services through examination of examples from across the world that may be applied to the WTC region. Finally, we discuss the barriers to climate adaptation in regard to ecosystem services. The key messages associated with each of the topics addressed in this chapter are:

We take issue with 2 of the points highlighted recently by Pomeroy et al. (2006), who state that (1) native oyster restoration or (2) the introduction of an exotic (non-native) oyster species have been widely advocated in the scientific literature as solutions to eutrophication in Chesapeake Bay. In reviewing the goals and success criteria for native oyster reef restoration, Coen & Luckenbach (2000) and others (reviewed in ASMFC 2007, Coen et al. 2007, Grab- owski & Peterson 2007) expressly noted that the sys- tem-level effects of oyster filtration have been poorly quantified, especially as they might relate to any spe- cific restoration project (but see Nelson et al. 2004, Newell 2004, Grizzle et al. 2006). The goals and suc- cess criteria emphasized by Coen & Luckenbach (2000) — and elaborated upon subsequently by Luck- enbach et al. (2005), Coen et al. (2007) and S. P. Pow- ers et al. (unpubl.) — have focused, among others, on the development of: (1) sustainable oyster popula- tions; (2) enhanced species diversity; (3) trophic com- plexity; and (4) localized material fluxes to the ben- thos. Similarly, Grabowski & Peterson (2007) point out that although effects of oyster restoration on water quality in large water bodies are difficult to quantify, localized effects of oyster filtration (e.g. reduced tur- bidity) have been observed and, together with other ecosystem services (e.g. Meyer et al. 1997, Allen et al. 2003, French McCay et al. 2003, Peterson et al. 2003) provided by oyster reefs, constitute a strong case for restoration.

Several studies from our literature review (and the empirical papers in this issue of Ecosystem Services) have sought to use deliberative approaches drawing on multiple, and mixed, methods. For example, Fagerholm et al. (2012) incorporated aerial photos into semi-struc- tured interviews with individual community members to identify and map ‘ indicators for landscape services ’ collated results were incorpo- rated into a workshop for further discussion with the wider community. In a study of sustainable forest management in Mexico, Rodriguez- Piñeros and Lewis (2013) brought deliberative discussion alongside in- depth interviews and questionnaires into a community-requested initiative to develop a new forest management plan for the commu- nity-owned forest. Haines-Young (2011) combined future scenarios with Bayesian Belief Networks to examine the latter's e ff ectiveness at integrating and visualising different types of information (qualitative, quantitative) and values across multiple stakeholders and disciplines to facilitate an analytical-deliberative approach to values identi fi cation. A number of researchers have experimented with the use of GIS to map ecological and social values of the landscape (e.g. Bryan et al., 2010; Ihse and Lindahl, 2000; Sherrouse et al., 2011; Kenter, 2016b in this issue). While some of these studies have attempted to promote participatory means to data generation, others have remained con- tained within a non-deliberative approach to public participation.

Organizations traditionally measure environmental performance in terms of inputs and outputs related to their operations. Such flows are expressed in environmental indicators that can be identified, managed, and ultimately measured by an individual organization. Measuring the quality and flow of ecosystem services is a more complex task, with shared responsibilities for those affecting and benefiting from them. Such complexity is shown, for example, when looking at nutrient cycling. The relevance of nutrient cycling is clear at a conceptual level for a company reliant on agricultural produce, but it is not obvious how to define or measure an organization’s performance with respect to this service. Measuring the financial value of ecosystem services to support corporate decision making is also challenging. To date, tools developed for organizational use have focused mainly on communicating the concepts of ecosystems and ecosystem services, and on helping organizations to understand dependencies, impacts and associated risks, while performance measurements and reporting on ecosystem services are generally focused on management activities and consumption of a few natural resources such as water.

First, tourism is a tertiary sector. For some time, at least until the construction of tourism satellite accounts, it was never really seen as a separate ‘sector’ – and in fact still suffers from numerous defi nitional debates. In terms of analysing ecosystem services it is noted that tourism receives direct inputs from many other sectors, with transport, accommodation, agriculture and viticulture important biologically-based contributors. Second, for comparative analysis, particularly when one has a growth- oriented economic focus, the concept of eco-effi ciency and twin question of sector attribution of various inputs becomes a contest- ed arena. Third, as the above analysis has shown, all of tourism’s major resource inputs (save from energy) are via indirect means. This in turn makes the exact ‘draw’ that the tourism sector makes particularly diffi cult to measure (once attribution rules have been determined).

Assessments of ecosystem services (ES) are vital for Africa’s sustainability. ES supply and demand take place in distinctive patterns in Africa due to the continent’s characteristic spatial heterogeneity, rich biodiversity, demographic developments, resource endowment, resource management conflicts, and fragile political landscapes, along with current industrialization and urbanization processes. Igno- rance of the dynamism of these parameters could diminish the capacity of the different ecosystem service providing units (SPU) to satisfy the demands in the ecosystem service benefiting areas (SBA) in Africa. The main aim of this review article is to assess the extent to which ES studies have been conducted and applied in Africa. This review analyzes those articles accessible online via the ISI Web of Science and open access journals. The online search yielded 52 ES-related studies, which were used for the review. Results indicate that most studies were conducted in South Africa, Kenya and Tanzania, and focused on services provided by watersheds and catchment ecosystems. Cru- cially, most of the studies focused on more than one ES category. Provisioning ES dominated across all the ES categories. However, ES tradeoffs and synergies were barely addressed. Economic valuation of ES and ES mapping comprised more than three-quarters of all the studies, and a quarter referred to biophysical quantification or qualification of ES. There are emerging alternative, non-monetary val- uation methods for ES, which could pave a new way of capturing value of non-monetized ES in Africa. Moreover, there is an urgent need to extend ES studies to the entire continent, in order to capture spatial and socio-economic uniqueness of various countries and focus more on local-scale assessments of multiple ES, as a means for addressing ES tradeoffs, synergies and SPU-SBA relations in Africa. Ó 2016 The Gulf Organisation for Research and Development. Production and hosting by Elsevier B.V. All rights reserved.

Floods occur when low-lying areas that are typi- cally dry become temporarily inundated with water outside of their normal confines (Rojas et al. 2013). Flooding accounts for one-third of natural disasters and affects more people than any other type of disaster (Sivakumar 2011). Flood-related impacts are expected to worsen due to global environmental change with flood risk increasing by 187% from increasing temperature in the HadCM3 climate model (Arnell and Gosling 2016). Flood magnitude is also expected to increase due to intensified water cycling resulting from as little as a 1.5 °C global average temperature increase (Alfieri et al. 2017). However, all floods are not created equal and the causes and consequences of individual floods are often unique. Floods can be seasonal as in the case of spring snowmelt or monsoon rains or they can occur randomly via several other mechanisms such as ice jams, storm surges, and heavy precipitation (Fig. 2a–c). Heavy precipitation accounts for about 65% of river floods (Douben 2006), but northern latitude areas with snow cover are also vulnerable to flooding caused by snowmelt and sometimes exacerbated by rain events (Kundzewicz et al. 2014). Flood events have been further charac- terized based on magnitude, frequency, duration, and volume (Burn and Whitfield 2016). These character- istics are important for determining the effects of floods on both aquatic ecosystems and the people who benefit from them. For example, flood magnitude can determine the amount of groundwater recharge or the extent of home and infrastructure damage during flooding. Flood magnitude is only one aspect of predicting flood impacts on aquatic ecosystems and ecosystem services. Ecosystem conditions prior to flooding are potentially equally as important as flood characteristics for determining ecosystem response to a flood event.

Conservation and Nuclear Safety; the UK government’s Department for the Environment, Food and Rural Affairs, and Department for International Development; Norway’s Ministry for Foreign Affairs; Sweden’s Ministry for the Environment; The Netherlands’ Ministry of Housing, Spatial Planning and the Environment; and Japan’s Ministry of the Environment. Interestingly, TEEB Phase II, from which the main reports emerged, was lead by Pavan Sukhdev, a career banker. The TEEB reports make the case for the systematic appraisal of the economic contribution of biodiversity and ecosystem services to human well-being, and for routine steps to prevent that contribution being lost or diminished through neglect or mismanagement (TEEB, 2010). It is an appeal to each of us, whether a citizen, policy maker, local administrator, investor, entrepreneur, or academic to reflect both on the value of nature, and on the nature of value.

and human wellbeing. As such it supports a consequentialist ethic that can be more successful than deontological approaches (see Glossary) in securing consensus and motivating action [1]. However, criticisms of the ESF as a tool for conservation raise doubts about its effectiveness and legitimacy [2, 3]. The most controversial issue is probably that of monetisation, as laid out recently by Silvertown [4]. One set of responses to such problems would continue using the ESF as a general tool for assessing habitats while recognising its multilayered structure [5], supervising it to avoid unintended consequences [6], perhaps discouraging monetisation [7], and even attempting to subjugate intrinsic value under the category of services [8]. Yet there are more profound problems with the ESF that call for a radical shift if we wish to contribute to conservation as part of a sustainable development agenda. Two outstanding issues are sufficient, in our opinion, to demand an overhaul of the ESF so radical as to require a new name. First, the definitions do not work. The fact that definitions of ‘ecosystem services’ (ES) and of specific categories are often vague, tautologous and/or at variance with the concepts actually employed is symptomatic of deep-seated problems, as we shall explain. Second, collapsing multiple human value judgments into one or a few numerical values is a form of devaluation. We unpack this claim by exploring the inescapably cultural foundation of valuation processes.

Ecosystem services contribute to human well-being directly by providing food, water, etc. and indirectly by pollination of plants, nutrient cycle, etc. These indirect services of the ecosystem are crucial for the self-sustaining of the ecosystem and have different spatial scales (Bolund & Hunhammar, 1999). Human activities in the last 50 years have severely degraded the ecosystems and hence the services they provided (Millennium Ecosystem Assessment, 2005). This degradation is brought by a number of driving factors that affect directly or indirectly. The factors are referred to as drives as they promote occurrence of ecosystem degradation that translates to degradation or decrease of ecosystem services (Anonymous, 2019). The indirect drivers do not have an effect directly on ecosystem, but rather influence or magnify the direct drivers’ effects. Examples of indirect drivers include population growth, change in economic activities, and socio-political factors. Direct drivers such as deforestation, overgrazing, irrigation, use of pesticides, affect the ecosystem directly. These drivers alone might appear insignificant but when coupled with coupled together have great effects. The degradation of ecosystems is a complex phenomenon that is spatial and temporal dependent. The ecosystem services can be categorized into four groups, namely supporting, provisional, cultural and regulating services, as shown in Figure 2 and presented below. The supporting services such as production of clean air, clean water and primary production, have an auxiliary role in sustaining other the ecosystem services.

Seasonal Fishing Ban (SFB) was introduced with the purpose of protecting the spawners during peak spawning season, reducing the fishing effort, giving respite to the sea floor and safety at sea. Since the inception of ban, the marine fisheries sector has undergone immense technological, economic and social changes. However, even after several years of implementation of SFB, there are no specific answers to the following questions: Has the natural capital asset and its value increased? Has the ban improved marine ecosystem services? What is the management cost vis-a-vis benefits? How does each maritime state perform? Answers to these questions are needed to arrive at effective management decisions to sustain this sector.

values, and health impacts from nitrogen dioxide pollution. Recreational values are provided by forest, wetlands, agricultural landscape, marine and coastal waters. Forests and wetland also act as pollutant sink by sequestration of carbon and nitrogen respectively. Given all caveats associated with finding data, the net result points at an unsustainable use of the ecosystems under the year of study, 1999. On the other hand, the correction of NDP implies an increase, which varies between 0.9 and 3.3 per cent depending on assumptions with respect to the ecosystems’ production of ecosystem services. This empirical result differs from other corrections of Swedish NDP, which instead result in a decline of NDP (Ahlroth, 2000). This is due to the difference in focus, which in Ahlroth and many other empirical studies is on pollutant emission, which enters directly into the utility function. The negative impact on utility from pollutants is then subtracted from conventional NDP. This paper also allows for a negative impact on utility from pollutant, but the main focus is on ecosystems as inputs in production of ecosystem services. This can generate a positive utility from production of non- marketed ecosystem services, which increases conventional NDP. Pollutants can reduce the ecosystems’ production capacity, but ecosystem service production must not be negative. Negative impacts on conventional NDP occur only from direct disutility of pollution and disinvestment in natural capital.

and human wellbeing. As such it supports a consequentialist ethic that can be more successful than deontological approaches (see Glossary) in securing consensus and motivating action [1]. However, criticisms of the ESF as a tool for conservation raise doubts about its effectiveness and legitimacy [2, 3]. The most controversial issue is probably that of monetisation, as laid out recently by Silvertown [4]. One set of responses to such problems would continue using the ESF as a general tool for assessing habitats while recognising its multilayered structure [5], supervising it to avoid unintended consequences [6], perhaps discouraging monetisation [7], and even attempting to subjugate intrinsic value under the category of services [8]. Yet there are more profound problems with the ESF that call for a radical shift if we wish to contribute to conservation as part of a sustainable development agenda. Two outstanding issues are sufficient, in our opinion, to demand an overhaul of the ESF so radical as to require a new name. First, the definitions do not work. The fact that definitions of ‘ecosystem services’ (ES) and of specific categories are often vague, tautologous and/or at variance with the concepts actually employed is symptomatic of deep-seated problems, as we shall explain. Second, collapsing multiple human value judgments into one or a few numerical values is a form of devaluation. We unpack this claim by exploring the inescapably cultural foundation of valuation processes.

discussed in Chapter 5, following Cowling et al (2008), we arguably assumed HWU could be considered a “learning organization” engaged in an adaptive watershed management for ecosystem services. On the other hand, interaction between HWU and the “Farmers’ community” across boundary B.2 had been far more challenging. This boundary had gone through different stages, characterized by different degrees of participation of “Farmers’ community” in decision-making processes. This included an initial stage, when the “Farmers’ community” was bound by decisions taken based on scientific knowledge, which they did not perceive as salient. Therefore, in B.2, although decisive, boundary work did not actually succeed in fully achieving the theoretical potential of interaction between knowledge users and producers. Rather, several other factors played key roles in facilitating knowledge transfer, including compensation money, new regulations, past collaborations, and peer control. In general, “contextual” (e.g. poor land productivity, a mixed farming structure) and “contingent” (e.g. increase in education level and generational change) factors as well as the relative influence of actors (Kastens and Newig 2008) were crucial for understanding knowledge to action transfer. Moreover, it is worth bearing in mind that boundary work is about managing tensions at the interface between stakeholders, whose interests and demands can sometimes be “incommensurable” (Parker and Corona, 2012).