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Keywords:

  • Water Framework Directive;
  • Cost-benefit analysis;
  • Scotland Heavily modified waters;
  • Environmental valuation

Abstract

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

The Water Framework Directive seeks to achieve the enhancement of aquatic ecosystem quality for all waters across Europe. However, it recognizes the need to accommodate social and economic considerations, so Heavily Modified Water Bodies may be designated where achievement of the Directive's objectives may result in disproportionate costs. This necessitates the use of cost and benefit data, which may be hard to acquire. Benefits and costs of implementing the Directive also can be tested at the national level. This paper considers the evaluation of costs and benefits at 2 scales in Scotland: microlevel analysis, for case studies on the rivers Tummel and Dee and the Forth Estuary, and macrolevel analysis. In the microlevel analysis, costs of hydroelectric power generation are compared with the marginal benefits of increased fisheries revenues and are argued to be disproportionate. In the estuary environment, the benefits of returning an area to ecologically productive salt marsh and mudflats are compared with the costs of lost agricultural revenue. In the macrolevel analysis, we compare the costs to impacted industries with the estimated national benefits from implementing the Directive. We find that for Scotland as a whole, implementation is predicted to result in positive net social benefits.


EDITOR'S NOTE:

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

This paper is among 6 peer-reviewed papers published as part of a special series, Ecology in a Cost-Benefit Society. Portions of this paper were presented by the author at an international conference on this topic held at Roskilde University, Denmark, in June 2004.

INTRODUCTION

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

The Water Framework Directive (WFD) (European Commission 2000) entered into European law on 22 December 2000 and has been hailed as the “most significant piece of water legislation for a generation” (ENDS 2004). It has a 15-y implementation timetable, applying to all countries in the European Union (EU), the European Economic Area, and accession countries intending to join the EU. For all water bodies, which comprise individual surface waters (rivers, lakes, reservoirs, and wetlands), groundwaters, estuaries (transitional waters), and coastal waters, the key objectives of the WFD are to prevent deterioration of environmental quality and to achieve “good status” in ecological and chemical terms by 2015.

Central to the implementation of the key objectives is the introduction of a system of river basin management planning, widely recognized as a means of introducing, for the 1st time across the whole of Europe, a mechanism for achieving integrated catchment management. The WFD requires the effective participation of all stakeholders in the management of each river basin, and it leaves each state to make arrangements for the definition of its river basins. Stakeholder participation in the planning process is seen as vital, because the plans for river basin management will impact stakeholders in areas such as water supply undertakings, hydroelectric power generation, and industrial water use, potentially leading to capital and operational costs. Plans are implemented over a 6-y cycle, which includes characterization, identification of pressures and impacts, setting of environmental objectives, and programs of measures to allow those objectives to be met. The WFD requires a system of licensing abstractions and impoundments, and whether or not states have existing legislation for such purposes, it will require existing uses of water to be subject to a process of review.

The WFD is sensitive to economic and social considerations, such as in identifying the “business as usual” scenarios, which need to be generated for each river basin to identify water bodies that are not likely to achieve good ecological status, and in requiring full environmental, financial, and resource cost recovery from water users. Good ecological status is the default target for all waters, but it is recognized that economic or social considerations must, in some cases, take precedence, such as the need to maintain the provision of public water supplies. The principal mechanism by which these considerations can be accommodated is the designation of Heavily Modified Water Bodies (HMWB) (Article 4). This designation is made only where

  • A physical modification has been made to the water body (e.g., construction of a dam or reprofiling of river banks);

  • Ecological status is assessed by the responsible authority as failing to meet the target level of “good” (Annex 5);

  • The measures necessary to achieve good ecological status would have a significant effect on the specified use of the water body (e.g., hydroelectric power generation or water supply) (Article 4(3)(a)); and

  • “[T]he beneficial objectives served by the… modified characteristics of the water body cannot, for reasons of technical feasibility or disproportionate costs, reasonably be achieved by other means, which are a significantly better environmental option” (Article 4(3)(b)).

If a water body is designated as an HMWB, the target of good ecological status is replaced by good ecological potential, which is equivalent in all respects except as necessary to allow the continuing effects of the socially and economically desirable activities justifying the HMWB designation.

Agencies and governmental departments responsible for implementing the WFD across Europe were required to make their provisional designations of HMWBs by December 2004, with full designations by December 2009 and again thereafter at 6-y intervals as part of the river basin management planning process. Ecological assessment is approached in an incremental manner, beginning with the risk assessments carried out as part of the characterization process due for completion in December 2004 and followed, as appropriate, by more detailed monitoring to allow definitive assessments of status. Biological monitoring data are used as available. Exhaustive monitoring to support status assessment is not possible, however, so estimation methods may be used, including hydromorphological characteristics if necessary. If alterations in hydromorphology are responsible for failure to meet good ecological status, this must be identified specifically for the designation process (Kampa and Hansen 2004). For the formal HMWB designation required in 2009, information regarding the costs and benefits of measures also must be collated and compared within the context provided by the WFD.

The present paper examines the practice of making and analyzing estimates of cost and benefits associated with the WFD by reference to 3 microscale case studies, which were part of a coordinated European process of trial designations (Kampa and Hansen 2004), and 1 macroscale study (Andrews et al. 2002). All 4 studies were undertaken in Scotland. Insights regarding the strengths and weaknesses of the methods applied are presented.

COST-BENEFIT ANALYSIS

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

To a considerable degree, the WFD requires environmental regulators to compare the costs and benefits of achieving improvements to good ecological status. For economists, this is consistent with testing the efficiency with which society's scarce resources are allocated to environmental improvement, because formal cost-benefit analysis (CBA) has, for many years, been seen as a practical way to test the efficiency of resource allocation, a criterion for judging public policy choices that is seen as important by economists. Indeed, CBA is a means of applying the main conceptual economic test for whether a project or policy improves social well-being, namely the Kaldor-Hicks Compensation Test (Hanley and Spash 1993). This states that a project/policy improves social well-being if the gainers could compensate the losers and still be better off. Because actual compensation for losses is not required, however, passing a CBA test can mean that an improvement in the efficiency of resource allocation is accompanied by a worsening of the equality of well-being.

The practice of CBA was established during the latter half of the 20th century, and it now follows widely accepted principles that have their grounding in economic theory. First, the relevant population must be identified. This is the group of gainers and losers over whom effects are to be evaluated. Typically, this is taken to be the population of a nation state. Second, all the potential impacts of the project/policy are set out. Because CBA usually is carried out ex ante, these will be predictions in terms of, for example, what improvements in water quality are expected, how these will be achieved, and who will benefit. Third, those impacts that are relevant from the viewpoint of economics are selected. Relevance here is defined with regard to social costs and benefits, so any impact that has a negative effect on the well-being of anyone in the relevant population, such as increased electricity prices caused by restrictions on hydroelectric power development, is counted, along with all impacts that are socially beneficial. Whether these impacts are valued by markets is unimportant in terms of whether they should count as economic benefits/costs. Economic costs will include the use of scarce resources in completing the project, such as construction costs for a fish pass, whereas economic benefits can include improvements in recreational opportunities.

The next phase of the CBA is to transform economically relevant impacts into monetary values. This allows impacts to be aggregated into the same unit of value. For those impacts that are valued by markets, such as changes in electricity outputs or h of labor during construction of new sewage treatment plants, market prices can be used to effect this conversion, because they show both the marginal social cost and the marginal social benefit of these changes. Where markets are distorted, however, such as by governmental intervention, shadow prices must be used that better estimate these marginal social costs or benefits (see again, Hanley and Spash 1993). Even so, many environmental impacts that are economically relevant (because they have effects on human well-being) are not valued by markets, because no markets exist for these resources. A good example is water status. The main benefit of actions taken under the WFD will be to improve water status, yet no market exists for “clean rivers,” or for the ecological status of rivers. Therefore, no market price exists with which to place monetary values on improvements in river ecology. In this case, economists make use of environmental valuation methods to estimate the marginal social benefit/cost of changes in environmental quality, based on the same principle of value that underlies all of CBA—that is, that the economic value of something is given by what people are willing to give up to acquire or hold onto it. This notion of willingness to pay (WTP) can be applied in a wide range of environmental valuation methods (see e.g., Bateman et al. 2002; Haab and McConnell 2003) to arrive at estimates in monetary terms of environmental costs and benefits.

Once all economically relevant impacts have been transformed into monetary values, the benefits and costs must be adjusted for when in time they occur. Because CBA normally is carried out on projects/policies with impacts stretching over multiple time periods, allowance is made for how long society waits for a benefit or cost to happen. Individuals have been found to exhibit positive time preference (i.e., we prefer benefits sooner rather than later), whereas the productivity of capital means that benefits acquired now can be invested at a positive rate of interest. For both of these reasons, future benefits and costs must be discounted into “present value” terms, using a formula typically given by

  • equation image

where Xt is a benefit or cost received/paid in year t, i is the discount rate, and PV represents present value. The effect of discounting is to reduce future values, because the term ([1 + i]−t) lies between 0 and 1. The further away in time the cost or benefit occurs, or the higher the discount rate, then the greater the reduction of the future value into present values. Clearly, the choice of discount rate is crucial here, but economic theory has failed to identify a single, theoretically correct value (see Sheraga and Sussman 1998; Weitzman 1998; Pearce et al. 2003). Typically, governments set discount rates nationally for use in public policy appraisal.

After discounting costs and benefits into present values, these cost and benefit streams can be aggregated together. A net present value test is then applied, whereby the aggregate discounted costs are subtracted from the aggregated discounted benefits. If the resultant value is positive, the policy/project is deemed to have passed the CBA test, in that it has been found to generate social benefits that are in excess of social costs. Aggregated, discounted benefits also can be divided by the cost equivalent to produce a benefit to cost ratio. These ratios can be used to rank policies/projects that are competing for scarce public funds, with the option of having the highest benefit to cost ratio being the most preferred. Finally, sensitivity analysis should be undertaken, whereby all value/flow/impact projections are subjected to scrutiny and the analysis redone using alternative estimates for these parameters. For example, the CBA could be repeated using alternative plausible estimates of changes in fish populations should a river improvement program go ahead.

It is important to note that the CBA process itself can be equally as valuable as the results that are generated. The process forces the analyst to think through, in a rigorous, consistent fashion, what the economic impacts of a policy/project will be, who will be affected and when, and the relative magnitude of any gains and losses. Also, CBA is an exercise in economic democracy, in that the benefits and costs to everyone in society are what count rather than just the effects on special-interest groups (although the strength of one's potential vote in a CBA exercise depends on one's income). The use of income-derived weights in a social welfare function is 1 partial solution to this problem of unequal money votes.

When applying CBA to environmental policy making, however, many problems arise (Hanley and Spash 1993), such as the complexity of ecosystem responses making estimation of benefits and cost difficult, the choice of the discount rate, the problem of dealing with irreversible environmental impacts, and the possibility of institutional capture. As will become clear in the discussion that follows, even attaining a full picture of benefits and costs can be difficult in the institutional setting, within which CBA is commonly applied, because of constraints on time and other resources.

We now proceed to explain how the CBA principles set out above were applied to analysis of the impacts of the WFD, 1st at the catchment level and then at the level of Scotland as a whole.

MICROLEVEL ANALYSIS: HMWB CASE STUDIES

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

As part of the European process for trial identification and designation of HMWBs, 3 case study areas were selected in Scotland: The river Tummel in the central Highlands (Black et al. 2002a), the Galloway river Dee in the southwest (Black et al. 2002b), and the Forth Estuary (Black et al. 2002c). The Tummel and Dee areas were both selected as studies concerned with the physical changes to rivers caused by hydroelectric development and the resulting environmental impacts, whereas the Forth Estuary was chosen as an example of an estuary subject to a mix of physical change types. An advisory group was established with representatives of stakeholders in the 3 areas to advise the research team undertaking the studies.

Tummel case study

On the basis of biological and surrogate assessments, ecological status was assessed as good for 1 of the reservoirs of the Tummel catchment hydroelectric scheme and moderate (the class below good, i.e., failing to achieve the WFD ecological target) for the remaining 7 reservoirs. As a direct result of physical modifications, almost all rivers immediately downstream of hydroelectric reservoirs, and also rivers and streams subject to physical diversions, were assessed as being of moderate status or worse. Pending completion of official assessments, ecological status assessments were undertaken using a system developed to account for such biological and physical monitoring data as were available and were guided by expert judgment from members of the advisory group (Black et al. 2002a). All water bodies failing to achieve good ecological status were therefore identified as candidate HMWBs. A policy-relevant question under the WFD is, then, would improving status by altering these modifications have costs disproportionately greater than benefits?

To achieve good ecological status in the water bodies of the catchment, the expert opinion of stakeholders indicated that the removal of all hydroelectric structures in the rivers and lakes of the area would be necessary, but that this would not guarantee the restoration of the previous, naturally high ecological status or that these measures would be technically feasible. Of necessity, these measures would cause the cessation of hydroelectric power production (i.e., the most profound possible impact on the designated use). This option was discounted, because it was judged that no better environmental option was available to serve the specified use—in this case, the generation of peak-load electricity—than the use of hydroelectric power. All viable alternatives to the generation of peak-load electricity would be thermal generation options, each involving the emission of greenhouse gases, and it was assumed that to replace hydroelectric power generation with such alternatives would not provide an environmentally preferable option.

This assessment gives rise, under Article 4(3) of the WFD, to consideration of alternative scenarios involving the continued generation of hydroelectric power in the catchment while adopting measures designed to raise the ecological status toward good. The principal mechanisms identified that would help to achieve this goal were, 1st, revision of compensation flows released to the rivers immediately downstream of hydrocontrol structures and, 2nd, upgrading of fish passes to enhance their effectiveness. These measures were considered to be practicable for the rivers Garry and Errochty, which form part of the east of the Tummel catchment, and were investigated accordingly.

Four probably viable options were considered. These are summarized, along with their annualized costs, in Table 1. Electricity generation was valued at £26 per megawatt-hour, which was understood to be a mean value for the Scottish wholesale electricity market in 2002, and capital costs were annualized at 6% over a 10-y period. Options 1 through 3 were considered as alternatives for the river Garry; option 4 was the only option considered for the Errochty system.

The benefits arising from the 4 options were identified as being of 2 types: Ecological benefits, including the restoration of missing species and changes to species abundance, and economic benefits accruing to individuals as a result of the ecological changes. The 1st type of benefit was excluded from the numerical assessment for the study, because information on WTP for improved ecological status was not available for the catchment area of the Tummel and no appropriate surrogate data were available.

Table Table 1.. Measures designed to restore good ecological status in the Garry/Errochty system: Annualized costs
OptionDescriptionCapital + running costs (£, thousands)Foregone electricity values (£, thousands)Total (£, thousands)
1Support restoration of fish population in river Garry10975131610
2As option 1, but lower Garry only49329378
3As option 2, but with enhanced measures designed to increase the chance of success49342391
4Restore good ecological status to river and Loch Errochty21762176

As regards the economic benefits, the only assessments made were those for increased fishing benefits. The anticipated return of migratory salmonids to the rivers Garry and Errochty would be expected to lead to increased letting values (and, hence, revenues) on the river Tummel downstream of the study area, as well as, perhaps, providing new fisheries letting opportunities on the rivers Garry and Errochty. Table 2 shows how the increases in annual salmon catch and the commensurately increased fisheries incomes were estimated, using confidential advice from appropriate local experts. Data in each of table rows A through E and G were based on local assumptions, which were then compounded to obtain the estimates of increased salmon catch numbers and fisheries incomes shown in table rows F and H. Therefore, the results must be subject to considerable uncertainty. Also, the values relate to 1 type of benefit only, and they cannot be taken as a guide to the total benefit. Nevertheless, these data provide benefit estimates of the only element of the total benefit for which quantification can be attempted.

Table 3 shows a comparison of these benefits with the costs associated with each of the options in annual terms. Given the uncertainty that exists concerning ecological outcomes, option 2 is assumed to deliver benefits with a 75% chance of success, whereas option 3 is expected to deliver the same benefits with a 100% chance of success. Option 4 is expected to lead to no economic benefit in terms of increased fisheries income. Restricting the assessment to fisheries benefits only, all the proposed options are considered to be disproportionately costly, and in the absence of any better environmental option, it is proposed that the affected water bodies should be designated as HMWB. Clearly, the benefits are lower-bound estimates, but the lack of any better environmental option for satisfying the current use of the water bodies (peak-load electricity production) is a persuasive factor in arriving at the proposed designation. The analysis was based only on eastern parts of the whole hydroscheme area, but fisheries interests selected this area as offering the maximum scope for fisheries benefit. It was suggested that fisheries benefits elsewhere in the catchment would be minor (Black et al. 2002a), so it was concluded that the designation should apply to all water bodies failing to reach good ecological status within the overall scheme area described. This seems to be acceptable for the Tummel catchment, if fisheries-related benefits are expected to be lower elsewhere in the catchment than in the water bodies studied (as maintained by local ecologists) and if we can assume that no matter where in the catchment hydroelectric power generation is foregone, the substitute source of power is likely to be more environmentally damaging. We suspect that exactly these kinds of judgment calls will need to be made during actual implementation of the HMWB provision throughout the EU.

Table Table 2.. Estimated increases in numbers of salmon caught and fisheries income following options 1 through 3 for Garry/Errochty system
 Option 1Options 2 and 3
(A) Length of river in which fishery restored (km)1711
(B) Area of spawning habitat created (m2 × 103)430240
(C) Smolt production (per m2)0.050.05
(D) Marine survival rate0.100.10
(E) Rod and line exploitation rate0.100.10
(F) Estimated increase in salmon caught (per annum) (= B × C × D × E)215120
(G) Increased fisheries income per fish caught (£)250250
(H) Increased fisheries income per annum (£) (= F × G)53,75030,000
Table Table 3.. Annualized costs and benefits of changes to hydroelectric power operation for the rivers Garry and Errochty
OptionCosts (£)Benefits (£)Net benefits (+ or -)
11,610,00053,750
2378,00022,500
3391,00030,000
42,176,0000

It should be noted that the results arrived at in this and the following 2 case studies are dependent on the ecological status assessments undertaken. If the estimated status of water bodies was revised from below the good/moderate threshold to above it, then the proposed HMWB designation would cease to be applicable. Under such circumstances, many of the hypothetical costs and benefits of status raising also would cease to apply, and the CBA would no longer to be valid. Some uncertainty does surround the proposed status assessments, but they have been substantiated by the risk assessments undertaken by the Scottish Environment Protection Agency (SEPA 2005).

Dee case study

The Dee case study followed a similar approach to that undertaken for the Tummel study, in that the designated water use was the same and the associated physical and ecological impacts were comparable. One feature of the Dee study not found in the Tummel study was the presence of acidification problems, caused partly by an acid geology and exacerbated by coniferous afforestation. This means that prospective achievement of good ecological status by restoration of prehydroelectric physical conditions may be subject to even less chance of success than in the Tummel catchment (Black et al. 2002b).

A greater range of possible improvements was identified as a means of achieving good ecological status for the Dee compared with the Tummel water bodies. These comprised increases in compensation flow, naturalization of heavily engineered river banks, physical changes to the design of fish passes, and changes to the flows of water passing through them; radically, the removal of 1 dam (Clatteringshaws) and a cessation of the use of its storage function also were considered. Costs and benefits of implementing these measures were calculated in a manner similar to that adopted for the Tummel study, again based on confidential consultation with local fisheries and hydroelectric power-generation personnel. The economic benefits of the measures, in terms of the economic value of enhanced recreational fisheries, were calculated to amount to £200/salmon/year, compared with £250/salmon/year in the Tummel study. The resultant estimated annual costs and benefits are shown in Table 4 for all the options considered. As with the Tummel study, however, removal of the morphological modifications may not guarantee achievement of good ecological status, and there may be further complications to the fisheries response detailed above (e.g., reductions in coarse fishing revenue). It is not possible to quantify such effects or to estimate their likelihood of occurrence; therefore, the proposed economic estimates must be regarded as approximate only and are best interpreted in the context of wider benefits.

The comparison of benefits and costs shows that for 4 of the 9 options, benefits (based on fisheries revenues alone) exceed costs. This might lead to an assessment that such options are not disproportionately costly and, therefore, are worthy of being pursued. For the sake of the HMWB designation, however, the pertinent question is whether the measures necessary to achieve good ecological status are disproportionately costly. For each water-body group shown in Table 4, the costs for the group as a whole must be compared with the overall benefits, because all would be necessary to achieve the ecological objective, and in the Drumjohn, Ken Cascade, and Glenlee water-body groups, overall costs were found to be disproportionate. Only the Tongland group, for which only a single measure was proposed (fish-pass upgrade), appears to have benefits that are not disproportionately costly. Accordingly, the designation of HMWB was not recommended for the Tongland group. For the other 3 water-body groups in Table 4, however, this designation was recommended on the basis of disproportionate costs, even though better environmental options were identifiable and technically feasible. In such cases, it will be left to the environmental regulator responsible for implementing the WFD to consider incorporating some of the proposed environmentally beneficial measures, particularly those that appear to be technically feasible and without excessive cost, as part of the program of measures required in the process of achieving good ecological potential (the alternative ecological objective that is provided under the WFD for HMWBs).

Forth Estuary case study

The Forth Estuary was chosen for study as an estuary with modifications resulting from dredging for navigational purposes, the development of port facilities, and extensive land reclamation for both these and agricultural purposes. For the present study, the estuary was divided into 3 water bodies between the outer estuary and the tidal limit to minimize problems that would arise in separating small areas from others nearby on which the ecology depends (e.g., mobile species such as fish and birds). Water body A, the farthest downstream, was classed as being of good ecological status despite the loss of 21% of the intertidal area in land-take and substantial alteration of the shoreline (e.g., by rock armor). Water bodies B and C, however, were classified as currently achieving only moderate ecological status, principally because of the extent of loss of intertidal area, at 42% and 68% of the original areas, respectively. These intertidal areas provide important elements of the habitat supporting many bird groups (e.g., 10,000–15,000 waterfowl and 30,000–40,000 waders) (Black et al. 2002c).

Table Table 4.. Annualized costs and benefits of measures proposed for the Dee catchment
OptionTypeaLength improved (km)Cost (£k)Benefits (£k)Net benefits (+ or -)
  1. a B = bank reprofiling; C = increased compensation flow; D = dam removal and site restoration; F = fish-pass upgrade; R = river flow regime alteration; S = cease storage function of reservoir.

  2. b These options may be considered as alternatives.

Drumjohn water-bodies group
Carsphairn LaneC, B11.4142.245.3
Water of Deugh and Bow BurnC7.219.928.6+
Ken Cascade water-bodies group
Water of KenR17 (river) + 6.5 (reservoir)2180.993.4
EarlstounF182.0371.7+
Glenlee water-bodies group
Black water of DeeC10145.539.8
Clatteringshaws DamF3172.0121.2+
Clatteringshaws ReservoirSb411359.2163
Clatteringshaws DamDb412038.7163
Tongland water-bodies group
Tongland DamF150 extra migrant salmon per annum2.0330+

The midestuary water body B was recommended for HMWB designation on the basis that no technically or economically viable or environmentally better option could be found to replace the industrial and navigation functions presently served by that water body. The reclaimed intertidal area supports major industrial facilities, such as an oil refinery and a power station, and the costs of relocating such large facilities were assessed as being prohibitive. In the case of the port, the impacts of switching cargo handling to the nearest alternative port (e.g., increased road or rail haulage to the industrial facilities adjacent to Grangemouth docks in water body B) were judged to be environmentally unacceptable.

Land-take to increase the cultivated agricultural area was the main physical alteration in the area of the upstream water body C. This change was achieved by the creation of earthen dykes, which have been faced with rock and could be altered readily to allow a return of intertidal conditions. A comparison of benefits and costs were therefore warranted to assess whether the measures needed to achieve good ecological status would be disproportionately costly. The principal cost element was judged to be the loss of agricultural production in the areas that would be affected, because the cost of physically restoring salt marsh would simply be the use of an excavator to make a small number of breaches in the dykes. The areas of cropping for each major crop/livestock activity in the affected areas were identified from census returns, then valued using standard farm accounting data (SAC 2000).

These figures represent private costs to farmers of lost output. However, the WFD implies that social costs are what should be considered (i.e., costs viewed from the perspective of the nation as a whole). These will diverge from private costs because of the presence of support payments (subsidies) to farming. An approximate adjustment from private to social costs was used, based on the Producer Subsidy Equivalent (Cahill and Legg 1990) calculated by the Organization for Economic Cooperation and Development (OECD) for the EU, which was 40% in the most-recent data before the study being undertaken, which was for 1999 (OECD 2004). Thus, the net social value of lost annual farm output was estimated as £76,447 per annum. It was not possible to quantify the monetary benefit of ecological restoration because of the absence of any suitable studies in the literature from which values could be transferred. Given the low cost estimate, however, it was judged that the benefit in restoring a large part of the intertidal area of the water body to a more natural condition, as a potentially sufficient step for the achievement of good ecological status, would not be disproportionately costly. Accordingly, it was recommended that water body C should not be designated as HMWB.

COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

In the year 2000, the Scottish Executive commissioned a study into the national costs and benefits of the WFD for Scotland. This was undertaken by the present authors, jointly with WRc (Andrews et al. 2002). The objectives of the study were as follows: 1) to quantify what benefits would be associated with implementing the WFD, and who would benefit; 2) to identify which sectors of the economy would face additional costs as a result of implementation, and what these costs would be; and 3) to compare the benefits and costs to Scotland during the period from 2002 to 2042.

A 1st step was to identify a business-as-usual scenario against which benefits and costs could be measured. This was done by examining current water-quality levels as well as improvements in these levels from planned infrastructure investment. A gap analysis was then undertaken to identify which water bodies would require investment to bring them up to good ecological status, which was taken as being equal to class A for rivers under the current SEPA classification system and class B for lochs, estuaries, and coastal waters (Andrews et al. 2002). The main reasons for the current existence of gaps in terms of the major determinants of poor or moderate status, such as nutrient pollution, were then obtained from SEPA. This allowed both the benefits and the costs assessments to be targeted at specific sectors and specific water environment problems. Gaps in current data, however, meant that it was difficult to identify/characterize gaps for groundwater resources and for hydromorphological impacts, which were thus excluded from the analysis.

Benefits assessment

The benefits assessment consisted of valuing changes in the lengths of water bodies where a gap had been identified between the baseline and the WFD requirements of good ecological status. Benefits transfer was used to generate the monetary values used in this exercise; that is, values (in terms of WTP) were adapted from existing studies in the literature rather than being obtained from original work (Barton 2002). We chose to restrict our benefits transfer exercise to using only studies done on water-quality issues in Scotland for these purposes. The use of benefits transfer resulted from the tight timescale within which the CBA had to be completed, but it also reflected what will be required as part of implementing the WFD because of the resource costs of original valuation studies. The benefits assessment was classified according to type of water body affected by the upgrading in status: Rivers, estuaries, and coastal waters. Changes in loch (lake) quality were not included, because to our knowledge, no estimates of economic benefit suitable to value improvements in loch water quality exist in the literature.

River quality changes

Estimates of river quality were based on values from a Choice Modeling study of the river Clyde (Hanley et al. 2001, 2004). These were converted into per kilometer values from the per household values originally obtained, then applied to all kilometers of the rivers that the gap analysis identified. The Clyde study gave separate values for improvements in the attributes of river ecology, aesthetics, and bankside vegetation, broadly ranging from fair to good status. These are shown in Table 5. For improvements from poor or seriously polluted to good status, we increased these values according to the law of diminishing marginal utility. This implies that increments from poorer quality to good status are more highly valued compared to increments from fair quality to good status. Values represent benefits to local residents from river quality improvements, whether or not those residents use the river for recreation. These estimates of economic benefit were then adjusted for variations in population density according to where in Scotland the gaps in ecological status had been identified. Note that these figures may well have underestimated total benefits, because we excluded both values to nonresident visitors and nonuse values to those who live outside the district in which any water body is located.

For angling, we used figures based on rental valuations by the Tay District Salmon Fisheries Board for the river Tay, obtained as part of the HMWD project described above. These can be viewed as estimates of producers' surplus (economic rent) per annum. This gives a central value of £2,250/km of river where salmon recovery takes place. It is important to stress, however, that these benefits will be realized only when salmon angling is restored where no such angling existed before, that this depends on actions being taken to ensure that habitat-water quality-physical modifications are changed as appropriate, and that salmon then actually return to the river in question.

Changes in estuaries

Very few Scottish studies exist regarding the benefits of improving estuarine water quality. A contingent valuation study by Dunkerley (1999) looked at the benefits of reduced sewage litter pollution at Broughty Ferry, on the Tay Estuary near Dundee. Respondents were asked the maximum increase in local authority sewerage charges that they were willing to agree to for a “bag it and bin it” program that removed sewage litter from the local beach. Some 317 responses were obtained, and a mean population WTP prediction made using an ordinary least squares bid curve. This gave a value of £16.31/household/y (equal to a 1.5% increase in council tax), which was equivalent to an aggregate value of £119,531/y. The area of estuary this relates to can be calculated as 10.5 km2, implying a value of £11,384/km2. Using the reported 95% confidence interval gives a value range of £9,492 to £15,962/km2/y.

The SEPA estuarine classification system includes aesthetic impacts as 1 determinant of status, but it does not include bacterial contamination. According to SEPA (1999), aesthetic impacts were the 2nd most significant problem for monitored estuaries, accounting for approximately 33% of pollution problems in Scottish estuaries. Calculations by WRc for this study showed 11.67 km2 of estuaries downgraded because of poor aesthetics, out of a total downgraded area of 25.38 km2. Using the value from the Dunkerley study, benefits from upgrading those estuarine reaches currently suffering from aesthetic quality problems ranged from £110,000 to £186,000/y, with a central estimate of £132,000/y.

Changes in coastal waters

For local residents, Day et al. (2001) used the contingent valuation method to estimate annual WTP to improve the local beach for residents in 2 towns in southwest Scotland, Irvine and Ayr. A payment-ladder format was used, with local council taxes as the payment vehicle. The data were analyzed using a nonparametric approach to avoid having to make assumptions about functional form. Results were used to estimate per-kilometer WTP values for coastal water improvements. Figures of £789 to £5,108/km/y, with a central value of £2,903/km/y, were obtained.

Table 6 gives a summary of the central estimates for benefits of implementing the WFD in Scotland, based on the values noted above. However, it is important to point out that because of data gaps, several potentially important benefits are excluded here. Besides the omission of benefits from improving loch water quality (already noted), no values existed for use in benefits transfer for the creation of wetlands as part of the implementation program, for changes in groundwater quality, for alleviations in low-flow episodes on rivers, or for improvements in acid mine drainage (SEPA 1999).

Table Table 5.. Values (£/km/y) to residents of water-quality improvements for the river Clyde
Attribute/reason for improvement being requiredImplied value per km of change from fair to goodImplied value per km of change from poor/seriously polluted to good
River ecology28,82672,065
Aesthetic appearance (e.g., litter)14,54836,370
Bankside habitat and erosion24,15760,392

Costs assessment

Estimates for the cost of compliance were made for sectors of the economy likely to be impacted by implementation of the WFD. These sectors were agriculture, the food and drink industry, pulp and paper production, mining, forestry, fisheries, power and water supply, wastewater treatment, households, and public bodies. Compliance cost assessments were made for each sector. For example, for agriculture, compliance activities include building improved storage facilities for farm waste and possible conversion to lower-input farming in areas where nonpoint nutrient pollution is expected to produce a gap between the baseline and good ecological status. For the hydroelectric power sector, measures needing to be taken include construction of additional fish passes, increased compensation flows, and bed restoration works. For mining, pumping and treatment of drainage flows are anticipated, while for forestry, liming of acidified lakes is included. For some sectors, measures could not be identified in sufficient detail or with adequate precision; examples include activities required by food and drink manufacturers as well as treatment of contaminated land by local authorities.

Costs were then aggregated across sectors and discounted over the period from 2002 to 2042. Agriculture was found to be the sector facing the highest compliance costs, at £270 million in present value terms. Relatively high compliance costs also were found for the mining, power, and water services sectors, each with costs exceeding £100 million in present value terms. Table 7 summarizes these cost estimates by sector, which are shown to total £836 million in present value terms.

Comparing costs and benefits

Taking our central estimates of relative costs and benefits, it may be seen that the present value of benefits exceeds the present value of costs by a ratio of 1,419:836, or 1.69:1. In terms of “best-guess” estimates of gains and losses, therefore, the benefits of introducing the WFD in Scotland outweigh the costs, and the policy thus adds to net social welfare. Taking the lower-bound estimates of benefits and comparing this with the lower-bound estimate of costs also shows a positive net present value; this same conclusion is reached if one compares the upper-bound estimates of benefits and costs as well. Only if either the central estimate or the upper estimate of costs is compared with the lowest estimate of benefits does implementation bring about net social losses.

These results, however, need to be viewed with caution. Much uncertainty and imprecision cloud the estimates of environmental benefits. Costs could exceed the upper-bound estimate if unforeseen difficulties arise in achieving good ecological status or if future changes in energy price increase the cost of pollution treatment significantly. Assuming a lower discount rate (e.g., using the recently adopted UK Treasury rate of 3.5%) increases benefits more than costs because of time-profiling and, therefore, increases the social value of implementation.

We also should note that this best-guess range of benefits clearly excludes many benefits that might be very important. Many of the expected benefits could not be converted into monetary values; thus, the actual total economic benefits likely will be larger than £325 million per year. This also will be the case in that even for impacts with monetary figures, these represent only part of the value of the impact. For instance, regarding improvements in coastal waters, we were able to give values only for impacts on bacteriological measures of water quality and were unable to include benefits to nonusers and nonresident beach visitors. For estuaries, we were unable to value reductions in persistent substances in biota and increases in dissolved oxygen. What this shows is how large an information gap exists between the ideal CBA laid out above and what is actually feasible in any kind of defensible manner in the context of the present study.

Table Table 6.. Annual Benefit Estimates for Improvements to Water Bodies in Scotland under the Water Framework Directive
 Length/area requiring improvementsAnnual Benefits Range (£, millions)Annual Benefits, central estimate (£, millions)
Rivers
 Residents4456 km120–262191
 Anglers5942 bankside km10.02–58.3234
Lochs0.6 km2Not monetizableNot known
Estuaries25.38 km20.11–0.190.13
Coastal waters71 km0.79–5.022.9
Total131–325228
Present value (at 6% discount rate)815–20231419
Table Table 7.. Costs of implementing the Water Framework Directive for Scotlanda
Broad sectorSubsectorPresent value of compliance costs, 2002–2042 (£, millions)
  1. a Andrews et al. 2002.

AgricultureIrrigated farming17
 Arable and livestock253
IndustryPulp and paper9
 Mining103
 Forestry16
 Fisheries15
 Power115
 Water services146
PublicFlood defense76
 Contaminated land17
 Urban drainage33
HouseholdsRural sewage treatment21
Total (at 6% discount rate)836
Lower estimate588
Upper estimate997

DISCUSSION AND CONCLUSIONS

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

One of the main points to arise in both the micro- and macroscale assessments outlined above is that data regarding the nonmarket value of environmental benefits are an important component of environmental CBA in general and of applications to the WFD in particular. Gathering the data to support a benefit assessment must be seen as a significant task, especially because data cannot be assumed to be transferable from the WTP values obtained for 1 population to another (Rozan 2004). This task also increases in complexity with the geographic scale of the CBA being undertaken, because wider-scale analysis implies a greater variability of both the ecological and socioeconomic factors that drive differences in environmental benefits. The microscale studies reported here were based on HMWB trial areas only (a catchment scale), whereas the macrolevel study was concerned with the likely costs of implementing the WFD across Scotland as a whole. Even at the microscale, a comprehensive assessment of benefits and costs could not be made, and this patchy information is likely to characterize use of cost-benefit methods in implementation of the WFD.

The present study makes clear the need for further development of both benefit transfer methods and a database of environmental benefit values for use in implementing the WFD. Indeed, this need exists more widely in environmental policy analysis. Uncertainty regarding the monetary value of benefits arising from the WFD is not just a matter of uncertainty regarding economic values, however, because it also relates to a lack of certainty in the achievement of the ecological objectives. For example, uncertainty exists over the restoration of salmon populations should hydroelectricity operations be altered.

Cost estimates are subject to considerable uncertainty as well. In the studies of hydroelectric power, a significant element is future changes in the prices for generation of renewable electricity. Government subsidies and targets for generation from renewable sources are likely to affect future price changes; therefore, this will affect the costs of regulatory measures. For all sectors, it remains to be seen exactly how Scotland's environment regulator puts the WFD into practice, and only then will the precise nature of programs of measures and, in turn, their costs be known. More generally, the economic costs of achieving good ecological status may well depend on flexibility in the implementation of measures, which would be enhanced by greater use of economic incentives. Considerable evidence now exists that by allowing more flexible responses across polluters, economic incentives can achieve environmental objectives at lower costs than purely regulatory approaches (see, e.g., Hanley et al. 1998; Stavins 1998). Such incentives in the present context could include tradeable abstraction permits in cases when good ecological status is compromised by low river flows or land-use subsidies, such as agri-environmental scheme payments, where ecological status is impaired by nonpoint nutrient pollution from farmland. Encouraging regulators to seek cost-effective solutions to river basin planning will be key to keeping down the macroeconomic costs of the WFD, which, indeed, calls for such cost-effective strategies to be identified and implemented as part of the process of drawing up river basin management plans. Estimating the social costs of these cost-effective strategies across all river basin management plans, however, will be an enormous task.

Significantly, the HMWB designations themselves are uncertain, because the ecological status assessments on which they depend are not yet certain. If water bodies prove to have good ecological status, then the results of the analyses will not be so much invalidated as simply inapplicable, because the target ecological objective will have been met and the measures to be adopted in the river basin management plan may be less demanding.

Finally, both the national-level and microlevel exercises illustrate how CBA can be used to quantify the distributional aspects of a policy change in terms of which economic actors gain and which actors lose, as well as the magnitude of these gains and losses. For example, the microlevel analyses of catchments affected by hydroelectric developments show that the main losers under a plan to raise ecological status would be hydroelectric power companies, whereas the main gainers would be anglers. In the national-level analysis, losses to certain sectors of the economy, such as agriculture and mining, turn out to be disproportionately high. This information concerning gains and losses from environmental policy initiatives can be useful in identifying areas of conflict between stakeholders and in developing suitable compensation mechanisms (although, as mentioned, the Kaldor-Hicks criterion that underpins CBA does not require compensation of those who lose out).

In conclusion, the macroscale assessment presented above shows that implementation of the WFD in Scotland is likely to generate benefits in excess of costs, with a central benefit to cost ratio of 1.69:1. In reality, benefits are likely to be considerably greater than those identified in the present study, because many gaps in the benefit assessment exist (e.g., regarding lochs and certain aspects of coastal water quality). For individual stretches of watercourse, however, imposing good ecological status may imply costs that are greater than the benefits: this was shown to be the case in several of the microscale studies. Thus, the option to be able to designate water bodies as HMWB and, thus, escape the requirement for good ecological status is likely to be an important aspect of maximizing the net economic benefits of the WFD for nation states throughout the EU.

Acknowledgements

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References

We gratefully acknowledge the support and input of the Scotland and Northern Ireland Forum for Environmental Research (SNIFFER) and the Scottish Executive as sponsors of the studies on which this paper draws. We also thank 2 anonymous referees for comments on an earlier draft of this paper.

References

  1. Top of page
  2. Abstract
  3. EDITOR'S NOTE:
  4. INTRODUCTION
  5. COST-BENEFIT ANALYSIS
  6. MICROLEVEL ANALYSIS: HMWB CASE STUDIES
  7. COST-BENEFIT ANALYSIS OF THE WFD AT A NATIONAL SCALE
  8. DISCUSSION AND CONCLUSIONS
  9. Acknowledgements
  10. References
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