Contract n° 006538 (SSPI)
BR B RI ID D GE G E
B B ac a ck kg gr r ou o un nd d c cR Ri it te er ri ia a f fo or r t th he e I ID De en nt ti if fi ic ca at ti io on n o of f G G ro r ou un nd dw w at a te er r th t hr rE Es sh ho ol ld ds s
Specific targeted Research Project Scientific Support to Policies (SSP)
Deliverable number: Methodological Framework WP5 (D24)
Due date of deliverable: 1 April 2005 Actual submission date: 1 July 2005
The deliverable authors are responsible for the content
Start date of the project : 1 January 2005 Duration : 24 months
AUTHOR: Roy Brouwer
AFFILIATION: Institute for Environmental Studies (IVM) ADDRESS: De Boelelaan 1087
TEL.: +31 20 598 5608 EMAIL: Roy.brouwer@ivm.vu.nl FURTHER AUTHORS: -
Revision [draft, 1, 2,…]
Project co-funded by the European Commission within the Sixth Framework Programme (2002-2006) Dissemination Level
PU Public
PP Restricted to other programme participants (including the Commission Services) X RE Restricted to a group specified by the consortium (including the Commission Services) CO Confidential, only for members of the consortium (including the Commission Services)
Contents
1. Introduction
Aquatic ecosystems are adaptive, but ecologically sensitive systems, which provide many important services to human society. This explains why in recent years much attention has been directed towards the formulation and operation of sustainable management strategies, the recent adoption of the European Water Framework Directive (2000/60/EC) being a good case in point. Both natural and social sciences can contribute to an increased understanding of relevant processes and problems associated with such strategies. The key to a better understanding of aquatic ecosystem problems and their mitigation through more sustainable management, lies in the recognition of the importance of the diversity of functions and values supplied to society at different spatial and time scales. This includes a better scientific
understanding of aquatic ecosystem structure and processes and the significance of the associated socio-economic and cultural values.
The Water Framework Directive (WFD), which provides a much more integrated catchment approach to water policy, is the first European Directive to explicitly recognize the importance of this interdependency between aquatic ecosystems and their socio-economic values.
Investments and water resource allocations in river basin management plans will be guided by cost recovery, cost-effectiveness criteria and the polluter pays principle. The plan
formulation and assessment process must furthermore include a meaningful consultative dialogue with relevant stakeholders. Such a dialogue will inevitably raise socio-political equity issues across the range of interest groups and therefore affect the management strategies.
Although groundwater resources are an integral part of catchment wide aquatic ecosystems, their position and role are not well defined in the WFD. For the definition of good
groundwater chemical status, it was not considered appropriate to list new quality standards that would be applied uniformly to all groundwater bodies throughout Europe, because of the natural variability of groundwater chemical composition and the present lack of monitoring data and knowledge. Article 17 was introduced stipulating that the European Parliament and the Council shall adopt specific measures to prevent and control groundwater pollution on the basis of a proposal for a new Groundwater Directive. The new Groundwater Directive complements the provisions already in place in the WFD and in the existing Groundwater Directive 80/68/EEC, which will be repealed in 2013 under the WFD.
In its communication COM(2003)550, the European Commission states that groundwater bodies shall be considered as having good groundwater chemical status when the measured or predicted concentration of nitrates, pesticides and biocides do not exceed standards laid down in existing legislation (Directives 91/676/EEC, 91/414/EEC and 98/8/EC respectively).
For other pollutants good groundwater chemical status is reached when it can be
demonstrated that the concentrations of substances do not undermine the achievement of the environmental objectives (good ecological and chemical status) for associated surface waters or result in any significant deterioration of the ecological or chemical status of these surface water bodies, nor should concentrations result in any significant damage to terrestrial ecosystems which depend directly on the groundwater body. For these other pollutants, groundwater quality threshold values have to be established by Member States in all bodies of groundwater that were characterized in the recent first WFD reporting obligations as being at risk.
In order to support this process of determining future groundwater quality threshold values for all European groundwater bodies, the Environment Directorate-General of the European Commission commissioned a 2-year project to develop a general methodology for
establishing groundwater threshold values called BRIDGE (Background cRiteria for the IDentification of Groundwater thresholds, contract n° SSPI-2004-006538). The methodology has to apply to substances from both natural and anthropogenic sources and threshold
values defined at the level of national river basin districts or groundwater body levels should be representative for the groundwater bodies at risk in accordance with the analysis of pressures and impacts carried out under the WFD. In the proposal for the new GWD these threshold values will be used for defining good groundwater chemical status.
As part of BRIDGE, a socio-economic assessment is carried out in support of setting thresholds values for specific groundwater pollutants and the evaluation of the social and economic consequences of specific threshold values. The assessment procedure will follow the economic analysis outlined in the Water Framework Directive (WFD) and more
specifically in the WATECO guidance for the economic analysis. The main objective of this paper is to present the general methodology for the socio-economic assessment procedure.
This procedure will be tested in several practical case studies in BRIDGE and is to a large extent based upon the interdisciplinary and integrated research concepts and methods developed in recent years in other European research projects, such as ECOWET
(Ecological-Economic Analysis of Wetlands: Functions, Values and Dynamics, contract n°
ENV4-CT96-0273) and EUROCAT (European Catchments, contract n° EVK1-2000-00510).
The remainder of this paper is organized as follows. Chapter 2 briefly introduces the WFD and more specifically the required socio-economic analysis in the WFD. In chapter 3 the basic concepts of integrated groundwater assessment are introduced and the integrated assessment methods that will be used and tested in the practical case studies in BRIDGE.
Finally, chapter 4 presents the economics of groundwater protection and remediation, including an overview of the costs and benefits of groundwater protection and remediation.
2. The European Water Framework Directive
2.1 Environmental objectives
The WFD’s main objective is to improve the chemical and ecological status of European water bodies by reducing or eliminating the emission of polluting substances into these water bodies and protect them from further deterioration. By 2015 water bodies have to be in a so- called ‘good’ chemical or ecological state. This is to be achieved through a (international) river basin approach and the identification of a cost-effective program of measures laid down in river basin management plans. The program of measures has to be defined by 2009 and implemented by 2012.
The environmental objectives associated with good chemical and ecological status are not defined in the WFD. It is up to the Member States to define these environmental objectives.
They may differ depending upon the nature of the water body (surface water, groundwater or protected area), while in some cases environmental objectives can be lowered for water bodies, which are characterized as artificial or heavily modified through irreversible hydro- morphological changes (e.g. dams). In the latter case, the WFD requires that Good
Ecological Potential (GEP) is reached instead of Good Ecological Status (GES). Both GES and GEP have to be defined by Member States.
If GES or GEP cannot be achieved, Member States can apply for derogation as detailed in Article 4 of the WFD. This may be for technical reasons such as high natural background levels of arsenic in ground water bodies or for economic reasons when the costs of additional
measures to reach good status are considered disproportionately high. In any case, a
deterioration of surface water and groundwater quality has to be avoided at any time. This is the so-called ‘stand still’ principle laid down in Article 4, 1, sub a (i) for surface water and Article 4, 1, sub b (i) for groundwater. Furthermore, a number of conditions hold when applying for
derogation. Environmental objectives in other water bodies cannot be jeopardized and exemptions have to be in accordance with other existing EU legislation.
2.2 Environmental objectives
The WFD is one of the first European Directives in the domain of water, which explicitly recognizes the role of economics in reaching environmental and ecological objectives. The Directive calls for the application of economic principles (e.g. polluter pays principle), approaches and tools (e.g. cost-effectiveness analysis) and for the consideration of
economic instruments (e.g. water pricing methods) for achieving good water status for water bodies in the most effective manner. The Guidance Document on the Economic Analysis prepared in 2002 by the European Water and Economics Working Group (WATECO)
advises that the various elements of the economic analysis should be integrated in the policy and management cycle in order to aid decision-making when preparing the river basin management plans. The integration of economics throughout the WFD policy and decision- making cycle is presented in Figure 1.
Source: WATECO Guidance Document
Figure 1: The role of economics throughout the WFD implementation process The main elements of the economic analysis are found in Articles 5 and 9 and Annex III in the WFD. Economic arguments also play an important role in the political decision-making process surrounding the preparation of RBMP in Article 4 where derogation can be
supported by the strength of economic arguments when setting environmental objectives.
The economic analysis can be summarised as follows:
1) Economic characterisation of the river basin (Article 5)
• Assessment of the economic significance of water use in the river basin.
• Forecast of supply and demand of water in the river basin up to 2015.
• Assessment of current cost recovery by estimating the volume, prices, investments and costs associated with water services in order to be able to assess cost recovery of these water services, including environmental and resource costs.
2) Cost-effectiveness analysis (Article 11 and Annex III)
• Evaluation of the costs and effectiveness of the proposed programme of measures to reach the environmental objectives.
3) Disproportional costs (Article 4)
• Evaluation whether costs are disproportionate.
4) Cost recovery and incentive pricing (Article 9)
• Assessment of the distribution of costs and benefits and the potential impact on cost recovery and incentive pricing.
The three main steps in the economic analysis identified by Wateco (2002) and the
associated time path are illustrated in Figure 2. The first step, the economic characterisation of river basins, has recently been completed. In the next two years (2005-2006) the
preliminary risk analysis carried out so far for the different European river basins will be further elaborated (including the more detailed definition of environmental objectives) and a start will be made with the identification of additional measures needed to reach good water status in a second step.
Source: modified from the WATECO Guidance Document
Figure 2: Steps in the economic analysis in the WFD and corresponding time path
By the end of 2007 each EU Member State has to produce an overview of its basic and additional measures according to Article 11, from which the most cost-effective programme of measures will be selected in step 3 by the end of 2008. Based on a cost-effectiveness analysis of programmes of measures, the question whether the total costs of additional measures to reach good water status are disproportionate will be addressed by the end of 2009. Finally, the financial implications of the basic and additional measures for different groups in society has to be evaluated by 2010, including the level cost recovery, changes in the use and level of
economic instruments (e.g. levies, taxes, water prices) and their role in achieving a more efficient and sustainable water use.
3. Integrated Groundwater Assessment
3.1 Integrated assessment
In the light of an increasingly emancipated society, decision-makers are held responsible and are made accountable for their decisions sometimes years after decisions were taken. Public scrutiny, accountability and trust have become of paramount importance. Decisions have to be explained and justified, especially to those who are directly affected by them. In complex decision-making situations, various, often competing, interests may be at stake. Demand for information, which reflects this plurality and diversity of interests and the way these interests are affected by decisions is increasing. Ideally information has to encompass the various relevant dimensions of the problem a decision-maker tries to solve, i.e. information has to be comprehensive and complete, even though in practice decision-making takes place in contexts in which uncertainties and incomplete information are inevitably present to different degrees.
At the same time information has to be communicated in a meaningful and persuasive way to both decision-maker and those affected by decisions. As societal-environmental change becomes more complex and decisions, interests and value systems more inextricably connected, there is growing interest in integrated approaches to inform policy and decision- making (see, for instance, Janssen, 1996 and Rotmans and van Asselt, 1996).
Information provided to support decisions is determined to a large extent by the political characteristics of the decision-making system and the phase in the decision-making cycle (Figure 3). Figure 3 provides a general characterisation of the decision-making cycle. A similar general model of decision-making processes is given, for instance, by Mintzberg et al.
(1976), who distinguish three phases in a decision process:
1) Identification of opportunities, problems and crises, activating the decision process (recognition and diagnosis).
2) Development of solutions (search and design).
3) Selection from available solutions (screening, judging, bargaining) and obtaining approval for the selected solution (authorisation).
In Figure 3, two axes are included, depicting the demand and need for horizontal integration across scientific disciplines and the demand and need for vertical integration of information as decision-making contexts become ever more complex. An important issue linking knowledge and information across different sciences and between sciences and decision- making processes is the temporal and spatial consistency. The same problem may play at significantly different geographical and time scales from a natural scientific perspective, a social scientific perspective and an institutional-political perspective (Boesch, 1999).
Source: Brouwer et al. (2003)
Figure 3: General framework linking information supplied by multidisciplinary science and decision-maker demand for information.
Socially and politically sensitised forms of integrated assessment are an important step towards:
i) increasing awareness about the complex nature of the interdependency between our physical and socially constructed environment; (i.e. co-evolutionary processes of change);
ii) greater recognition that uncertainties and risk of irreversible change require careful consideration in decision-making (precautionary principles), which may be facilitated by prior agreement on a sensible, preferably social learning based, evaluation process;
iii) recognition that costs and benefits in complex decision-making circumstances are dynamic, as knowledge and experiences progress;
iv) increasing public support for and trust in decisions because of greater transparency in the ex ante evaluation phase.
Various assessment forms exist, including technical, health, environmental, economic and social appraisals. In general, the need for these appraisals results from problems of choice, where decision-makers (e.g. policy makers, producers, consumers) face one or more options (e.g. policies, projects, measures, products etc.) in a given policy context. Integrated assessment procedures have been developed in order to avoid as many unforeseen consequences of policy decisions as possible (de Vries, 1999). The literature contains various definitions of integrated assessment. According to Weyant et al. (1996), an assessment is integrated when it draws on a broader set of knowledge domains than are represented in the research product of a single discipline. Parson (1995) argues that integrated assessment is a policy relevant whole, which is greater than the sum of the disciplinary parts. An appealing definition is given by Rotmans et al. (1996), who define integrated assessment as ‘an interdisciplinary process of combining, interpreting and communicating knowledge from diverse scientific disciplines in such a way that the whole cause-effect chain of a problem can be evaluated from a synoptic perspective with two characteristics (emphasis added):
1) integrated assessment should have value added compared to single disciplinary oriented assessment;
2) integrated assessment should provide useful information to decision makers.’
This definition contains three important components, which are considered essential in such an approach. First of all, integrated assessment not only provides a conceptual framework, it is an interdisciplinary learning process, for experts and decision makers, and in its most inclusionary form other types of stakeholders. Setting up a collaborative framework between experts with different scientific backgrounds and experiences is often a time consuming procedure. Participants have to get used to and acquainted with each other, overcoming different uses of language, and their often fundamentally different ways of thinking, before their work can actually be put together in a meaningful and coherent way (Turner, 2000).
Secondly, in order for this collaborative process to be successful, communication is essential, i.e. communication between experts (scientists) and communication between experts and (lay) policy or decision-makers, especially in the case of complex decision- making contexts. Policy and decision-makers should be involved in this process right from the start for a number of reasons.
First, the inclusion of their policy or decision objectives determines the scope of the
assessment. Although assessments may be set up in such a way as to minimise subjectivity, judgements are inevitable in any evaluation. If these judgements influence outcomes in a major way, they should be made explicitly clear to the users of the evaluation. Discussing this sooner rather than later with the users of the information generated by the assessment will minimise the risk of producing overly controversial results.
Secondly, given the inevitability of scientific uncertainties and risks, the involvement of lay decision makers in discussions about these uncertainties and risks may (a) improve the decision maker’s understanding of the complexity of the problem, (b) modify accordingly expectations regarding ‘the deterministic truth’ behind the outcome of the integrated assessment should these uncertainties persist throughout the research, and (c) encourage the adoption of policy principles such as the Precautionary Principle or use of Safe Minimum Standards.
Involving policy or decision makers in the integrated assessment is more likely to ensure that the assessment will deliver useful information to decision makers (the third component emphasised). According to Rotmans et al. (1996), integrated assessment is policy motivated
research to develop an understanding of the issue, not based on disciplinary boundaries, but on boundaries defined by the problem. It offers insight to the research community for
prioritization of their efforts and to the decision-making community on the design of their policies. Also Janssen (1996) argues that integrated assessment is an iterative process where, on the one hand, integrated insights from the scientific community are communicated to the decision-making community and, on the other hand, experiences and learning effects from decision-makers form the input for scientific assessment. This complex and value- loaded process cannot be captured by one single approach. Depending on the problem it tries to address, it usually consists of a variety of approaches and methods, including formal, explorative, experimental and expert judgement methods.
In addition, by including also other stakeholders, an interactive, participatory and inclusionary bottom-up approach results, which may be beneficial for a number of reasons:
(1) It will help to elicit public perception of problems and possible solutions in addition to expert judgement and therefore ensure that decisions focus on the right problems (as perceived by all parties involved). In the field of risk assessment, Cvetkovich and Earle (1992) argue, for instance, that effective management of environmental hazards requires knowledge of both physical environmental systems and the social-
psychological processes affecting human responses to environmental conditions.
Integrated assessment can also be seen as a (communication) process bringing together the knowledge and experiences of policy or decision makers, experts and lay public.
(2) Early involvement of stakeholders can increase broader social support for decisions, whatever the outcome. In sociology, positive relationships are found between the intensity of interaction and the degree of commitment (e.g. Broom et al., 1981).
(3) Early involvement of other stakeholders will facilitate the identification of the distribution of costs and benefits to different groups of people.
(4) It may also help decision-makers and experts to identify relevant criteria to evaluate policy outcomes, planning and implementation procedures.
The exchange of information between (representatives of) various stakeholder groups and communication of different perspectives on perceived problems will inevitably result in some sort of social learning process, which may change perceptions, attitudes and behavioural patterns underlying these initially perceived problems.
3.2 Procedural rationality in integrated assessment
A number of steps can be identified underlying decision support systems in general, including integrated assessment. The logic behind these steps is closely related to ideas, which have been referred to in the literature as procedural rationality (Simon, 1964; 1972 and 1982). The stepwise procedure encompasses an iterative communication and learning process, including various feedbacks to previous levels of analysis and evaluation. The following steps will be discussed in more detail below:
1) Problem definition and identification main objective(s)
2) Identification of potential solutions and program of measures 3) Identification of effects
4) Measurement of effects
5) Evaluation of effects
6) Risk and uncertainty analysis
3.2.1 Problem definition and identification main objective(s)
Any assessment procedure starts with an analysis which scopes the present situation, resulting in the diagnosis of (potential) problems, their nature and scale, the definition of objectives and the identification of possible solutions to perceived problems. Although not necessarily yet commonly understood, a (shared) belief in and a sense of (shared) responsibility about the problem is essential in this first stage to mandate an integrated assessment.
A number of conceptual frameworks have been developed, which help to map cause-effect relationships underlying environmental problems, including problems related to water resources management. These frameworks are especially helpful when facing complex problems surrounded by uncertainties. One widely applied framework is the Driving Forces- Pressure-State-Impact-Response (DPSIR) framework.
The DPSIR framework provides a conceptual and organising backdrop for the contributions of different disciplines to the description and analysis of environmental problems. The socio- economic aspects of environmental problems are an integral part of the framework.
Environmental problems can be described, analysed and evaluated vis-à-vis the economic, social and cultural context in which they arise (Figure 4). The framework provides an important tool for achieving a common level of understanding and consensus between researchers, natural resource managers, policy makers and other stakeholders. It provides the link between the various driving forces (endogeneous and exogeneous such as
urbanisation and climate change), which pose threats to ecosystems (pressures), their impact and possible policy responses (Turner et al., 1998; Turner, 2000).
Another internationally applied mapping tool to scope problems is the Source-Pathway- Receptor (SPR) framework. The SPR framework forms the basis of Ecological Risk Assessment (ERA) methods that have been developed to help manage the impacts of contaminated sites on plants, animals, and ecosystems (e.g. Banwart, 2000). The role of ERA within a broader risk management framework is illustrated in Figure 5.
ERA is a method by which the actual or potential adverse effects of contaminants on plants, animals, or ecosystems can be assessed in a systematic fashion. Essentially ERA asks the question: how likely is it that damage will be or has been done by contaminants? The
objective of carrying out an ERA is to determine whether any identified ecological values are likely to be adversely affected by soil, water or air contamination. This will enable land managers to make decisions about managing contaminant risks on sites of concern.
Source: Adapted from Weber and Crouzet (1999)
Figure 4: The DPSIR-framework in the context of water policy and management
ERA is based on a causal stress-response model in which a contaminant is transported from a source through a known pathway to a receptor (e.g. a plant or animal). The basic components of an ERA include (NIWAR and MWLR, 2001):
1) Problem identification 2) Receptor characterisation 3) Exposure assessment 4) Toxicity assessment 5) Risk characterisation.
ERA can be undertaken at three distinct levels of detail. At each level, the five key tasks mentioned above are undertaken to provide information and data in order to make a risk management decision or to decide whether it is necessary to proceed to the next level of detail. Broadly, the degree of detail and quality of the data at each level can be described as:
- Level 1: qualitative
- Level 2: semi-quantitative - Level 3: quantitative
Source: National Institute of Water & Atmospheric Research and Manaaki Whenua Landcare Research, New Zealand
Figure 5: The role of ERA in a broader risk management framework
An important difference between the DPSIR and the SPR framework is that the former focuses explicitly on the interaction between pressures exerted by human (including economic) activities and their effects on environmental and social systems. The SPR-framework is usually somewhat more restricted to the geophysical modelling of ecological risks. However, the framework can be expanded relatively easily to also include risks to humans, for instance when water resources are also used as drinking water. In that case the SPR approach can be seen as a component of the DPSIR approach.
3.2.2 Identification potential solutions and program of measures
Solutions are usually worked out in more or less detail progressively, usually stepwise, sometimes over several years. The process can evolve through policy scenarios to concrete management measures and instruments to implement solutions. This process occurs in a given institutional setting in which several decision-makers play a role, at different
institutional and administrative levels. Through interactions with their decision-making environment, including stakeholder groups and other groups of experts, deliberation occurs within existing formal and informal power structures, which may shift as a result of the decision-making process.
3.2.3 Identification of effects
Based on individual experience, expert knowledge and judgement, a preliminary assessment is usually made of the nature of likely impacts related to different policy options, even though quantitative data and information about the exact magnitude and extent of the impacts may not yet be available. It is usually at this stage when potential effects are discussed, that previously implicit criteria become explicit and the grounds are laid on which policy options or management actions ultimately will be evaluated. Ideally, evaluation criteria are made explicit first, based on the objectives set out in the previous stage. Subsequently the relevant effects associated with these criteria would be identified. However, in practice the identification of evaluation criteria often follows after discussion of the relevant effects. Including (other) stakeholders at this stage in the process may help to identify the distribution of positive and negative effects (costs and benefits in physical and monetary terms), across environmental assets and different institutionally categorised groups of people.
3.2.4 Measurement of effects
Models and indicators are two important ways to assess the effects of alternative policy options or management actions. However, the procedures adopted can be quantitative and/or qualitative. They may be based on highly advanced models in which dose-effect relationships are formalised in mathematical terms, or simply involve the use of expert judgement. The choice of any method obviously depends upon the specific nature and scale of the problem and decision-maker demand for specific types of information. Natural and social scientists have developed their own tools, methods and procedures to measure the impacts of human intervention on aquatic ecosystems (environmental, social and economic impact assessment methods and procedures).
Groundwater policy and management will directly or indirectly affect both the environment and society (depending among others on the scale of implementation). Environment and society are usually inextricably linked. For example, groundwater management affects the economic activities, such as agriculture in an area, but it also affects the biological diversity present. Wastewater treatment and water purification are essential for good-quality drinking water. However, their provision not only affects public health, but also the ecological quality of aquatic ecosystems. The presence of water in a landscape often gives it its characteristic feature(s), and may play an important role in people’s perception of its beauty, or be an important determinant in people’s motivation to live or visit a particular location.
Understanding the interactions between aquatic ecosystems and society, including the economy, cannot be achieved by observational studies alone. Modelling of key
environmental processes is a vital tool that must be used if water and wetland management
is to achieve its overall sustainability goals and objectives, including a further quantification of the uncertainties in existing ecological process based models. For any group of researchers wishing to investigate and model a particular local groundwater system, or aspects of that system, to provide upscaling data for larger models, there are initially two types of
information required (Turner, 2000):
1) Estimations of biogeochemical fluxes in the system as it is now and dynamic simulations of processes in aquatic ecosystems which can be used to explore the consequences of environmental change and produce forecasts of future fluxes.
2) Measures of the forces of socio-economic changes (e.g. population growth, urbanisation etc.) on fluxes of toxics, nutrients and sediment (pressures), and
assessment of the human welfare impacts of these flux changes. Such assessments of the socio-economic costs and benefits involved will provide essential management information about possible resource and value trade-offs.
The assessment of the impact of changes in groundwater resources on human use of these resources (wealth creation) and habitation (quality of life aspects) requires the application of socio-economic research methods and techniques, which will be dealt with later on in this paper. Key issues related to the functioning of aquatic ecosystems and the assignment of values to ecosystem structures and functions at a catchment level, which must be
considered include:
- The spatial and temporal scales of ecological processes and socio-economic resource use.
- The structure, complexity and diversity which underlies ecosystem functions and their value to society.
- The dynamic (in space and time) nature of interactions between society and aquatic ecosystems.
- The uncertainties associated with these dynamics.
The essence of integrated assessment is to determine how society is affected by the functions the aquatic ecosystem performs, and changes in ecosystem functioning (possibly as a result of human use or exploitation of these functions). The key to valuing a change in an ecosystem function is establishing the link between that function and some service flow valued by people. If that link can be established, then the social value of a change in an ecosystem function can be derived from the change in the ecological value of the ecosystem service flow it supports (Figure 6). However, the multifunctional characteristic of aquatic ecosystems, including groundwater resources, at the catchment level makes comprehensive estimation of every function and linkages between them difficult. It will for example be
necessary to assess features of socio-economic activities and behaviour and how these respond to changes in ecosystem functioning.
Source: Turner et al. (2001).
Figure 6:Relationship between ecosystem functions and values (illustrated for wetlands)
In the context of water resources management, Mitchell (1990) has argued that efforts directed towards more integrated management have three related dimensions. In the case of groundwater management, integration can be interpreted as follows:
1) In systems ecology terms, i.e. to gain a better understanding of how each component of the aquatic ecosystem at catchment level influences other components, e.g.
groundwater dependent terrestrial ecosystems.
2) In wider biogeochemical and physical systems terms, i.e. where groundwater interacts with other biophysical elements, e.g. surface water.
3) In socio-economic and socio-cultural terms, i.e. where groundwater management is linked to relevant policy networks and economic and social systems with attendant
culture and history, so that the chances of a co-operative solution or mitigation strategy are maximised.
3.2.5 Evaluation of effects
Integrated assessment implies working with various values, reflecting different perspectives on water and wetland management problems. A core difficulty in integrated assessment is therefore to:
1) relate relevant single effects, values and criteria across fields of impact in a meaningful way;
2) make them comparable in order to be able to weight them and trade them off if necessary.
It is therefore important to identify and define all the relevant values or criteria that play a role in evaluating the effects of feasible (often politically negotiated) solutions. Given tripartite deliberation (e.g. regulatory agency, experts and stakeholders) to evaluate policy and policy measures, the extent to which effects are positively or negatively related to a chosen
reference situation needs to be examined. This is usually done on the basis of a number of pre-selected or emergent evaluation criteria. For instance, the extent to which impacts of various policy options contribute to the realisation of existing policy objectives can be examined. Various criteria may play a role when making a choice between different policy options and the corresponding policy measures. Although agencies seem to prefer to promulgate and enforce regulations based on quantitative criteria, qualitative descriptions of qualitative changes in for example community structure are often the best indicators of ecological disruption (Noss, 1990). In practice, qualitative descriptions of the intermediate changes or transitions between ecosystem states and ecosystem functions may sometimes prove to be the only way to assess the extent to which management restores, maintains or enhances the integrity of an ecosystem.
When pursuing integrated and sustainable water resources policy, this usually means that environmental, social and economic considerations play an important role in the decision- making process, and policy and enabling measures will be scrutinised for these types of concerns. For instance, in a recent communication (COM(2000)477) from the European Commission to the European Council and Parliament, water pricing is introduced as a measure to ensure sustainable and efficient use of water. Economic instruments should be an integral part of a set of measures in order ‘to guarantee achievement of the social, economic and environmental objectives in a cost-effective manner’.
Concern for public health and safety is usually one of the most important policy objectives and will generally be one of the main factors by which the effects of policy or policy measures are evaluated. Moreover, governments often believe that these safety and health concerns should apply to everybody in society. Hence, equity may be another important objective.
Although governments aim for safety and good health for everybody, they usually try to realise this policy at the lowest cost possible. In other words, policy options should be as cost-effective as possible. This example clearly shows that several criteria can apply simultaneously. Based on these values and criteria, the most appropriate method can be selected (Figure 7). In practice, an integrated assessment of effects based on prior
environmental, economic and social impact assessments may prove to be very difficult, due to the fact that (1) they are usually based upon fundamentally different starting points, values and norms and (2) the associated official (institutionalised) assessment procedures often follow their own set of rules or ‘procedural rationality’.
Figure 7: Values, criteria and assessment procedures and methods
Once the expected or observed impacts of policy and policy measures have been evaluated on the basis of one or more criteria, the importance of each criterion in the decision-making process has to be determined, in order to be able to make a choice between policy
alternatives. The degree to which various evaluation criteria will ultimately influence the decision depends on the political context (Susskind and Dunlap, 1981). Multi-criteria analysis techniques may be an important tool to support this decision-making (for an overview, see, for example, Janssen, 1994). The analytical hierarchy process (Saaty, 1980) is another helpful tool to translate qualitative judgements by decision-makers into quantitative weights.
Obviously, if all effects are expressed in monetary terms, this step is not necessary.
3.2.6 Risk and uncertainty analysis
Risk and uncertainty will be associated with both the physical outcomes associated with future environmental change and their social and economic consequences. Assessing the possible outcomes and the likelihood of perturbations to highly complex aquatic ecosystems will inevitably be fraught with difficulty. A range of possible impacts deriving from potential management actions needs to be identified, the relevant physical effects quantified, and
probabilities attached to each. A particularly important aspect relating to the uncertainty of physical effects, is the existence of thresholds beyond which disproportional and possibly irreversible effects occur. These will also be important in a social and economic sense due to the disproportional extent of the impact and the inability to reverse the consequences in the future.
Analysis should not assume that all current physical, social and economic conditions will hold in the future. For instance, land use changes might be predicted for the future, perhaps due to imminent regulation or long-term trends. This might affect, for example, the quantity of nitrogen in run-off and thereby the value of a wetland as a nitrogen sink. Behaviour of individuals could also adapt to change in water system functioning, for instance with farmers changing cropping patterns as a result of increased flooding, rather than foregoing land-use or yields altogether. These changes need to be incorporated into the analysis since they can influence projected effects (costs and benefits) and hence the outcome of the integrated assessment.
Uncertainty as to the correct value for environmental, social or economic variables employed and future trends can be addressed by employing sensitivity analysis or scenario analysis.
Sensitivity analysis gives more than one final answer using different figures for variables employed such as the rate of discount, the extent of a function being performed and shadow pricing ratios. This provides a range of estimates within which the true figure can be
expected to fall, which is less bound by particular assumptions but might result in ambiguous recommendations. Scenario analysis envisions a number of plausible future situations with a range of parameters within the valuation model, allowing a comparison of different future outcomes and policy response options.
A useful distinction is between risk, to which meaningful probabilities of likely outcomes can be assigned, and uncertainty, where probabilities are entirely unknown. It has been
suggested that the rate at which the future is discounted could be altered to incorporate a premium for risk, adjusted either upwards or downwards (Brown, 1983). However, risk is better dealt with by attributing probabilities to possible outcomes, thereby estimating directly the expected value of future costs and benefits (Boadway & Bruce, 1984) or their ‘certainty equivalents’ (Markandya and Pearce, 1988), rather than in some arbitrary and often
subjective addition to the discount rate which will attribute a strict (and unlikely) time profile to the treatment of risk.
As Costanza (1994, p.97) points out, ‘most important environmental problems suffer from true uncertainty, not merely risk’. Such uncertainty can be considered as ‘social uncertainty’ or
‘natural uncertainty’ (Bishop, 1978). Social uncertainty derives from factors such as future incomes and technology which will influence whether or not a resource is regarded as valuable in the future. Natural uncertainty is associated with our imperfect knowledge of the environment and whether there are unknown features of it that may yet prove to be of value. This might be particularly relevant to ecosystems where the multitude of functions that are being performed have historically been unappreciated. One practical means of dealing with such complete uncertainty is, for instance, to complement a cost-benefit criterion based purely upon monetary valuation, with a Safe Minimum Standards (SMS) decision rule (Ciriacy-Wantrup, 1952; Bishop, 1978; Crowards, 1996). It is important to recognise, however, that such rules are not panaceas for complex decisions with inevitable trade-offs, both ‘costs’and ‘benefits’ to society still have to be evaluated.
3.3 Integrated assessment methods
There are three main evaluation methods that combine information on the costs and benefits of alternative policy or investment options in order to support choices between these
alternatives. These methods are cost-effectiveness analysis (CEA), cost-benefit analysis (CBA), and multi-criteria analysis (MCA).
3.3.1. Cost-effectiveness analysis
Cost-effectiveness analysis (also known as least cost analysis) is used to identify the most cost-effective option for achieving a pre-set objective or criterion. The relevant objective is set, options for achieving it are identified, and the most cost-effective option is identified as that with the lowest present value of costs. It is implicitly assumed that the benefits of
meeting the goal outweigh the cost and action is therefore economically viable. The following steps can be distinguished in a cost-effectiveness analysis:
Step 1:Identify the environmental objective(s) involved
Step 2: Determine the extent to which the environmental objective(s) is (are) met
Step 3: Identify sources of pollution, pressures and impacts now and in the future over the appropriate time horizon
Step 4: Identify measures to bridge the gap between the reference (current/baseline) situation and target situation (environmental objective(s))
Step 5: Assess the effectiveness of these measures in reaching the environmental objective(s)
Step 6: Assess the costs of these measures
Step 7: Rank measures in terms of increasing unit costs
Step 8: Assess the least cost way to reach the environmental objective(s)
Cost-effectiveness analysis is suitable for use in situations where valid and reliable
estimation of the benefits of alternative options is not feasible. This is particularly relevant to actions that involve environmental change. Instead of attempting to identify and value the benefits, the most cost-effective means of achieving a desired objective is identified. Cost effectiveness analysis is suited, for example, to situations where clear and defensible conservation targets or other environmental goals exist which can be measured in terms of biophysical units such as minimum groundwater quality standards. It can also be used to identify the most effective option for a fixed amount of funding that has been allocated to achieve a policy objective. The drawback of cost-effectiveness analysis is that it does not identify the benefits of actions or the willingness of society to pay for improvements in environmental quality, which are important considerations in many decision contexts. For these reasons, cost benefit analysis is, if practicable, the preferred tool for decision support.
3.3.2. Cost-benefit analysis
Cost benefit analysis (CBA) is carried out in order to compare the economic efficiency implication of alternative actions. The benefits from an action are contrasted with the associated costs (including the opportunity costs) within a common analytical framework.
The benefits and costs are usually physically measured in widely differing units; comparison is enabled through use of the common numeraire of money. The benefits and costs of each
option are determined relative to the common scenario that would prevail if no action was taken. The net benefit of each option is given by the difference between the costs and benefits. The most economically efficient option is that with the highest present value of net benefit (i.e. net present value, NPV); economic efficiency requires selection of the option with maximum net present value. Options are economically viable only if the NPV that they
generate is positive. In general, the following steps can be identified in a CBA:
Step 1: Define problem and identify the main policy or project objective(s)
Step 2: Define the baseline scenario (the ‘without’ project situation) over the appropriate time horizon (including exogenous socio-economic and environmental developments) Step 3: Define alternative options or measures (the ‘with’ project situation) to achieve the
policy or project objective(s)
Step 4: Estimate the investment, operation and maintenance costs of each alternative option or measure compared to the baseline situation
Step 5: Assess the welfare effects of each alternative option or measure compared to the baseline situation
Step 6: Estimate and value these welfare effects
Step 7: Compare costs and benefits through time (using appropriate discount rate) through calculation of Net Present Value
Step 8: Sensitivity and uncertainty analysis
Cost benefit analysis provides a rational and systematic framework for assessing alternative management and policy options. It entails identification and economic valuation of all positive and negative effects of alternative options. This involves the translation of all benefits and costs into monetary terms, including where possible, non-marketed environmental, social and other impacts. It is based on the underlying assumption that the allocation of resources amongst competing uses in society should be determined by the preferences of individuals.
Selection of cost benefit analysis as a decision support tool is at the discretion of analysts and policy- and decision-makers. It is the responsibility of analysts to ensure that the underlying assumptions of a cost benefit analysis are appropriate to a specific situation and that the results are valid and reliable. Cost benefit analysis reflects a specific paradigm that is considered more or less appropriate in different domains of decision-making. The
appropriateness depends on culturally determined beliefs, norms and values, regarding, for example, the legitimacy of social-political organisation of decision-making, public consultation and the process of value elicitation.
3.3.3. Multi-criteria analysis
Multi-Criteria analysis (MCA) can be used to establish a ranking of alternative options by reference to an explicit set of measurable criteria that the decision making body has defined.
A MCA may accommodate a range of social, environmental, technical, economic and financial criteria. MCA is therefore applicable especially where significant environmental and social impacts are present, which cannot (easily) be expressed in monetary terms. MCA are often integrated with participatory approaches and tend to facilitate such input to a larger degree than the classical monetary assessment tools CBA and CEA. Unlike in a CBA, criteria do not need to be quantified in a common metric (i.e. money) and instead MCA provides a number of alternative ways of aggregating the data on individual criteria to provide indicators of the overall performance of options. The advantage of this approach is that costs
and benefits that cannot be expressed in monetary terms or can only be expressed on qualitative scales can still be included in the evaluation. The main disadvantages of using MCA are that information requirements are still high and that the evaluation process is vulnerable to manipulation. The steps in a MCA are more or less the same as in a CBA:
Step 1: Define problem and identify the main policy or project objective(s); in this case often multiple social, economic and environmental objectives are identified simultaneously Step 2: Define the baseline scenario (the ‘without’ project situation) over the appropriate time
horizon (including exogenous socio-economic and environmental developments) Step 3: Define alternative options or measures (the ‘with’ project situation) to achieve the
policy or project objective(s)
Step 4: Identify and the criteria against which the different options should be evaluated, along with their suggested indicators and measurement units
Step 5: Construction of an effects table in which the effect scores of each alternative option or measure (related to each individual criterion) are included
Step 6: Standardization of effect scores (different methods available)
Step 7: Identification and quantification of weights per criterion (by experts, stakeholders and/or policy or decision makers)
Step 8: Aggregation of weighted effect scores (different methods available)
Step 9: Ranking of alternative options or measures based on aggregate weighted scores Step 10: Sensitivity and uncertainty analysis
4. The economics of groundwater protection
14.1 Introduction
Groundwater resources are often used in a non-optimal way, either in terms of excessive extraction for domestic, agricultural or industrial uses, or in terms of being used for the disposal of pollutants. Private agents that use groundwater resources fail to establish optimal levels of use due to the occurrence of negative externalities and the fact that groundwater resources are often common property resources. Given the over exploitation of groundwater there is a strong argument for governments to intervene to control its use, and there are a number of alternative policy instruments available to achieve this, including taxes, subsidies, regulations, and tradable permits. In order to set efficient policies governments or water managers need information on the benefits and costs of groundwater protection.
4.2 Externalities and common property resources
Although the optimal level of groundwater use (both in terms of extraction and pollution) is unlikely to be zero, it is generally recognised that at many locations groundwater use exceeds optimal levels. In other words, it would be possible to make people better off by reducing the level of groundwater use. This begs the question of why people using
1 This chapter is based on Turner et al. (2003) and Turner et al. (2004).
groundwater do not reduce their use to optimal levels. There are two main reasons why markets fail to allocate groundwater resources efficiently, firstly groundwater use frequently generates external effects or ‘externalities’, and secondly groundwater is often a common property resource, which face well-known management problems.
4.2.1. Externalities
An externality (or external effect) is the direct effect of one party's action on another party without any payment or compensation. For example, in the case of groundwater, a farmer might contaminate a groundwater aquifer that is used by others as a source of drinking water. The contamination imposes a cost in terms of health effects or water purification that is not taken into account by the farmer when deciding how much fertiliser to use. If the overall costs to society of the farmer using the aquifer for waste disposal is greater than the cost to the farmer, this results in an inefficient or non-optimal use of the aquifer. If this external cost of the farmer’s activity were borne by the farmer he would choose to pollute less.
This partly explains why the use of groundwater is often excessive but does not explain why the parties involved do not reach a mutually beneficial agreement in which one party
compensates the other to the point that an optimal use of the resource is achieved. The reason why such private solutions do not succeed is related to the common property nature of groundwater.
4.2.1. Common property resources
Groundwater aquifers typically extend under large areas of land that are owned by many different parties. As such, they have the characteristics of a common property resource and share some of the management problems of other commons type resources such as
fisheries and communally held grazing land. Essentially the problem is that there is a lack of incentive for any one user to invest in the maintenance of the groundwater resource because any investment (in terms of reduced extraction or pollution) will be immediately utilised by other users. Similarly, in the case of a compensation scheme, there is an incentive for users to ‘free-ride’ on the compensation payments made by others as everyone can benefit from the reduced contamination.
In addition to the spatial nature of groundwater externalities and common property resources, there is also a temporal dimension. Groundwater pollution may persist for very long periods of time and will therefore affect future users of groundwater. These external effects are difficult to take into account, and it is not possible for future users to compensate current users for
reductions in their exploitation of the resource.
4.3 Costs and benefits of groundwater protection and remediation
As groundwater becomes increasingly scarce, both in terms of desired quantity and quality, the legitimacy of treating it as a free resource is being increasingly questioned. The absence of pricing, as well as the lack of cost recovery, has been a major determinant of inefficient and often inappropriate and excessive use of groundwater. In response to these problems, many countries and water management agencies are turning to pricing mechanisms to allocate groundwater. However, economists disagree on the appropriate means and methods by which groundwater should be priced and on the notion of an optimal
groundwater pricing policy. The methods used to price groundwater and the performance of these is dependent on the physical, social, institutional and political context. Nevertheless, since the pricing of groundwater affects the allocation decisions of those with competing wants, then through correct pricing, efficient allocation can be achieved. Pricing is used in this context to refer to the introduction of amended financial charges in situations where groundwater was previously free or under priced or to the consideration of the economic value of groundwater in decision-making through use of an appraisal and accounting procedure, such as cost-benefit analysis.
Because groundwater resources are often non-marketed, it is extremely important to ensure that wherever possible, the ‘true’ economic value of these resources is accounted for when making investments, or decisions concerning groundwater and environmental policy.
Shadow prices, determined through the economic valuation of groundwater resources, are employed in such decision-making in place of market prices. Unless groundwater resources are correctly priced, and those prices are internalised in decisions, distortions are created.
These bias investment and policy decisions against concerns about groundwater resource depletion and degradation, resulting in misallocation of resources and sub-optimal social welfare.
The opportunity costs of groundwater resource use and the economic value of the benefits can be compared in terms of whether the use is economically sustainable or socially optimal.
The opportunity costs and economic benefits of groundwater resource use are considered in more detail below.
4.3.1. Costs
Marginal Opportunity Cost (MOC) is an important and useful tool for conceptualising and measuring the physical effects of groundwater resource depletion and degradation in economic terms. MOC seeks to measure the full societal cost of an action or policy option that employs a natural resource such as groundwater. Economically efficient resource management requires that the price that users have to pay for resource use should equate with the MOC. If the price is less than the MOC then the resource is over-consumed or utilised. A price that is higher than the MOC results in the resource being under-consumed or utilised. Sustainable management can be achieved through sustainability pricing, which also includes a premium to cover the costs that accrue from any resource depletion.
The concept of opportunity cost is used to refer to the value of a resource in its best alternative use, i.e. other than the purpose being considered. This is the cost to society of use of the resource. It is considered in terms of a change at the margin i.e. the marginal opportunity cost, because management decisions usually entail relatively small changes in resource use. MOC comprises the following three components.
1) The first component is the direct economic costs of groundwater abstraction, such as the costs of labour, equipment and materials used for abstraction. Such costs require adjustment for any subsidies, taxation and market imperfections in order to reflect true opportunity costs (shadow pricing). These costs vary with the difficulty of extraction. In the case of groundwater quality deterioration, the main cost categories are:
- Directly reducing the level of activity that is damaging the groundwater resource. In this case, the cost is the decline in value derived from the harmful activity;
- Employing alternative production methods that are more expensive (e.g. in agriculture, controlling pests with natural or organic methods);
- Information gathering for better management. Collecting information and learning how to change processes can also be a substantial cost.
- Remediation of damage caused. Cleaning contaminated aquifers is not always technically feasible and is generally very expensive.
2) The second part of MOC is external costs that arise from groundwater use. This is the net value of any losses and gains in welfare that groundwater use imposes on individuals other than those engaged in the activity. External costs arise because changes in one component of the natural resource base affect other components and the efficiency with which other activities can be conducted. Costs that occur in the future require discounting in order to make them commensurate with present-day costs. Though information on marginal external costs is difficult to obtain and often imprecise or incomplete, useful approximations can be made. It is the external costs that arise from unsustainable resource use that are of particular interest.
3) The final component of MOC is relevant for non-renewable resources. If such resources (which are fixed in supply e.g. over-abstracted aquifers) experience a positive rate of exploitation, then use of a unit of the resource results in its non-availability for future use. A scarcity premium can be placed on the resource, the magnitude of which depends on the size of the resource stock relative to the rate of exploitation, the strength of future demand relative to present demand, the availability and cost of future substitutes, and the discount rate. This scarcity premium is known as the user cost (Conrad and Clark, 1987) and relates to the value of the opportunity foregone by exploiting and using the resource in the present period rather than at sometime in the future. It also incorporates increases in the costs of future resource use and exploitation that occur as a consequence of current use and exploitation (e.g. the increases in costs of future pumping of groundwater that occur due to the greater difficulty of extraction). Marginal user costs also apply to non-sustainable use of renewable resources.
The user cost of non-renewable groundwater resources is often ignored, especially where groundwater resources are treated as an open access resource and users behave in an individually competitive manner. This can happen in situations where property rights are ill defined or not enforced. Use of the groundwater is then governed by the law of capture, on a
‘first come first serve’ basis: each user tries to extract for instance as much as possible from the resource in fear that other users will exploit the resource first, and also in the belief that the amount they themselves use is only a small proportion of the overall stock. The
consequences of ignoring the user costs are that the costs of extraction are undervalued which results in exploitation rates that exceed the optimum. This is in contrast to the situation where a single user has rights to a resource: the user is forced to take user costs into
account because it is the user who faces the increased costs of extracting from a depleted resource in the future.
In summary, MOC= Marginal direct cost + Marginal external cost + marginal user cost.
Pricing based on MOC is a useful principle as it forces attention on to the externalities associated with natural resource degradation, and guides pricing policy in providing incentives for allocative efficiency. Groundwater is allocated to high value uses and high social costs provide a disincentive against excessive groundwater use (Dinar et al, 1997).
Failure to set groundwater charges for example for irrigation on the basis of either
opportunity costs or user benefits has been a classic cause of inefficiency in the agricultural sector (Repetto, 1986). Proper valuation of the socio-economic benefits derived from groundwater resources is therefore an important and often necessary condition for efficient and sustainable groundwater resource use.
4.3.2. Benefits
Protecting groundwater from excessive use or contamination has benefits in terms of the value of uses that were damaged. Groundwater provides a large number of services that are
of value to humans. Direct uses of groundwater include extraction for use as drinking water, irrigation, or in industrial production. Groundwater resources are also used indirectly for the disposal of waste chemicals, including excess fertiliser and pesticides from agriculture.
Groundwater may also have non-use values associated with it if people that are currently not using the resource place a value on its continued existence or pristine state. In addition people that currently do not use a groundwater aquifer might place a value on having the option to use it in the future, or for it to be available to future generations.
Evaluation of the trade-offs necessary to allocate resources between competing wants requires that we consider their economic value. In particular, one of the main principles of efficient and sustainable resource allocation requires knowledge of the marginal value or benefits of the resource in its alternative uses. Valuation is used in conventional economic terms to refer to the estimation of individuals’ preferences for the conservation or
improvement in quality of a resource, as well as individuals’ loss of welfare due to resource depletion or quality decline. An individual’s preferences are measured in terms of how much they are willing to pay, which is also referred to as the economic value or benefit. Economic value to society of a good or service is determined as the aggregate of all individuals’
willingness to pay. It is important to point out that the price of a good or service and its economic value are distinct concepts and can differ greatly: water can have a very high value, but a very low price or no price at all.
Though the economic value of a resource is most commonly determined by willingness to pay for gain or improvement in a resource, it is also theoretically valid to use willingness to accept compensation for loss or degradation of the resource. Theoretically, there should be no significant difference in the value of the two measures. However, empirical evidence suggests that in practice, willingness to accept compensation is often substantially greater than willingness to pay. WTP has become the most frequently applied measure of economic value and has been given peer review endorsement through a variety of studies (e.g. Arrow et al., 1993). The appropriate measure of economic value is determined by the specific circumstances and the property rights regime that is associated with the resource use in question.
Some values associated with groundwater may be directly observable in well-functioning markets (e.g. if water users pay a fee for the quantity of water they extract). Often, however, groundwater values are not readily observable and can only be estimated using appropriate non-market valuation methods. This topic is addressed in the next section.
4.4 Economic valuation methods
This section provides a general discussion of economic valuation techniques that can be used to value groundwater resources. Economic valuation is used here to provide a basis for evaluating the tradeoffs involved in the allocation of water resources between competing wants. The focus is particularly on valuation of changes in functions provided by groundwater resources under different allocation options. For the purposes of cost benefit analysis, the impacts of alternative options for groundwater resource use or management are specified in terms of the economic value using the common numeraire of money. Economic value is determined by the impact on social welfare, which is given by the aggregate impact on the utility of individuals in society. The utility to individuals is determined by their preferences, which individuals express in the amount that they are willing to pay for goods and services.
The key to valuation of groundwater resources is to establish the functions that they provide, i.e. the link between the structures and processes of groundwater resources and the goods and services they provide that are valued by society. If that link can be established, then the
