La hemodiálisis

One technique that is often promoted as being able to handle the complex trade-offs that are frequently required in any decision involving a mix o f financial and environ­ mental costs is multi-criteria analysis (MCA). This term is a generic one, encapsulating a range o f techniques with differing levels o f mathematical complexity, that originate in the field o f linear programming. Their comparative ease o f operation has greatly enhanced their popularity (Keeney & Raiffa, 1976).

From an operational point o f view, the major strength o f the techniques is the ability to handle conflicts both between intrinsic elements within the methodology and between various external interests. Thus, in a situation where there is a conflict between quantifiable and qualitative issues, these methods can provide systematic information on the nature o f the conflicts, highlighting the trade-offs for decision­ makers (Nijkamp et a /., 1990). This could involve the specification o f a number o f plan options, together with the identification o f the relevant decision criteria, which are generally broader than a simple desire to maximise the economic benefits. Those factors that can be quantified, and those that are presented in qualitative form are clearly distinguished. The latter are generally dealt with either directly or indirectly: In the former case, the information is used directly in a qualitative evaluation method, whilst in the latter it is transformed into a cardinal form, often utilising an appropriate weight. MCA can also involve setting specific targets for differing criteria.

However, there are a number o f areas o f uncertainty, from the identification o f the appropriate criteria and possible targets, to the eliciting o f appropriate weights, and to the choice o f appropriate decision rule, or rules. The latter issue, in particular, can raise significant difficulties, as often when utilising a number o f criteria, there will be no ‘dominant* alternative that outranks all the others. A more likely outcome is a number o f ‘non-inferior* alternatives. In these situations, the policy analyst is unable to offer any preferred option to policy makers and the judgement becomes essentially political (see Pearce et al. (1994) for a full exposition). The inherent difficulty o f undertaking such judgements, particularly when there are a

large number o f proposals, often leads to the adoption o f a specific criterion, the maximisation o f economic benefits, reflecting individual willingness to pay. This is social cost-benefit analysis.

3 .1 .4 Cost-Benefit Analysis

In theory, a social cost-benefit analysis (SCBA) allows a comparison, in commensu­ rate terms, o f all costs and benefits, both private and external, relating to a particular project. It involves the identification, measurement and quantification o f all the environment impacts. Such a procedure might also take due cognisance o f the social incidence o f the costs and benefits. Thus, EIAs identifying and measuring environmental impacts are inputs to SCBA, where the data are converted into commensurate terms through the use o f a monetary metric.

The advantage o f the SCBA approach is the resulting comparability o f both the financial and the environmental costs, measured in monetary terms, which means that both can be easily utilised when trying to determine the costs and benefits o f each waste treatment method. If the net benefits o f one method are greater than those o f the others (or the net costs o f one method is less than those o f the others), then that method would be favoured over the others and would be at the top o f the

‘waste hierarchy*.

The calculation o f the monetary values for environmental damage is based on willingness to pay (WTP) or willingness to accept (WTA). A measure o f environmental damage can be obtained through production or dose response functions which measure the change in crop yields, fish stocks, health effects, etc. By valuing these at market (shadow) prices, estimates o f WTP/WTA can be obtained. Alternatively, for goods for which no markets exist, the WTP measure can be obtained directly, or indirectly, from individuals preferences. One o f the direct methods is the hypothetical questioning technique contingent valuation, which asks individuals directly about their willingness to pay for a particular environmental good (or to avoid an environmental ‘bad’). An indirect method for estimating the value o f non-marketed goods is hedonic pricing, which establishes the WTP for

environmental goods, through the use o f surrogate markets, for example by analysing the effect on house prices caused by the proximity to particular environ­ mental goods or ‘bads*. However, one limitation o f indirect methods, in contrast to the direct methods, is that they are unable to ascertain the value o f passive use values.

The distinction between use and passive use values can be clearly illustrated through the use o f a simple example: Consider the case o f the recent purchase o f the Straithaird Estate, on the Isle o f Skye, by the conservation organisation, the John Muir Trust. An individual who has visited the area, and has thus derived benefit from it, might be willing to pay to assist the purchase o f the property. This payment could be held to reflect his direct use o f the estate (use value) or his intended use at a later date (option value). By contrast, a second individual, who has never visited the Isle o f Skye, and who never intends to, may also wish to contribute financially to the conservation o f an area o f outstanding natural beauty. This might reflect the individual’s desire to ensure that future generations also enjoy the area {bequest value)y a desire to preserve the environment for its own sake (existence value) (Pearce & Turner, 1990).

Thus, the essence o f the concept o f passive use values is that environmental services have value to many individuals who are not in a position to make direct use o f them. Thus, passive use values would be expected to be greater, in absolute terms, than use values, reflecting their greater potential hinterland, although some economists remain sceptical o f their potential significance. A 1989 U .S . court decision, in the case Ohio v. the Department o f the Interior, found that passive use were compensa­ ble under the Clean Water Act and the Comprehensive, Environmental Response, Compensation and Liability Act (Carson et al. y 1995). This decision, together with the grounding o f the Exxon Valdez in Prince William Sound on the 24th March 1989, has proved to be the catalyst for their inclusion into calculations o f compensa­ tion, at least in the United States, following an environmental disaster (see Arrow et û/.,1993).

There is a further problem in using a technique (SCBA) that reflects the WTP o f individuals; that is the incidence o f the costs. A simple example might illustrate the problem: Let us say that decision-makers have to make a decision between two sites for a waste facility, and the decision would have an impact upon the same number o f people, in absolute terms, in the two locations. The first prospective site is in a relatively affluent neighbourhood where the residents unanimously oppose the siting. The second prospective site is in a relatively poor neighbourhood, where the residents are less affluent, but still united in their opposition to the siting o f the facility in their neighbourhood. If no adjustments are made for social incidence, SCBA will favour the latter site, as the potential WTP to avoid the facility will be less. The favoured decision rule implies that if the individuals who benefit from the project, were able, hypothetically, to compensate those who suffer from the project, whilst still remaining better o ff themselves, then the project would proceed. In practice, actual compensation rarely happens, with the result that waste treatment facilities would often be situated in low-income neighbourhoods. Although society as a whole might benefit, unless actual compensation is paid to the affected households, they will be worse o ff in the ex p o st scenario^.

One way o f addressing these concerns is by weighting the ‘votes* o f the lower income groups. One way o f adapting the chosen methodology is illustrated below.

Without any such adjustments, the formal CBA model would be:

T

max J ((B , - C, - E ,) • e '" ) d t [3.1]

where s is the social discount rate; t is time periods {t = 1 ,..., 7);

are benefits;

See Been (1994) who suggests that the causal link might go the other way; it is the siting o f a LULU that attracts low income groups: "The presence o f a LULU in a neighborhood can lower the neighborhood's quality and thus its property values, making housing there more available to low-income families”.

Q are private costs; and Ef are environmental costs.

To account for social incidence, this can be modified^:

T ( n \ / / ( ( ^ - ■B,f - W; • C,_, - W, • - e - « ) d i \ 0 where 0 d t [3.2]

i is fth social class (f = 1, . . . , /i); and Wj is the weight attached to fth social class.

However, despite the theoretical ease o f addressing the issue, in practice, particularly in the developed world, such weights are not generally employed. Pearce et al. (1994) offer the following reasons for this neglect:

"This is in large part because there is no political consensus...on how such equity weights should be determined. More broadly, such equity weights rarely capture the ethical complexity o f policy choices..."

It also appears that there is little political consensus to actually employ explicit redistributive policies for projects that have impacts at a local level. Generally, it appears that such objectives are best dealt with by national macroeconomic policies. However, notwithstanding the resulting local inequity, the logic o f such a policy is further undermined, if one considers the ‘strategic’ impacts o f consistent use o f the methodology by a government agency. The utilisation o f the methodology on a series of, what appear to be discrete, local projects, might engender significant ‘strategic’, or aggregate, effects, that might be contrary to national macroeconomic policy.

Despite these caveats, cost-benefit analysis allows a systematic representation o f both the financial costs, and the environmental costs and benefits'^, and in the case o f the ‘waste hierarchy’, offers a framework for an assessment that is long overdue.

3.2

THE COST-BENEFIT APPROACH

The assessment o f the ‘waste hierarchy’ involves a comparison o f the net social costs (NSC) - or benefits (NSB) - o f each o f the tiers in the hierarchy^. Thus, to allow the assessment to take place, the costs and benefits o f each waste treatment method will need to be identified and estimated. The remainder o f this chapter sets out the methodology for the cost-benefit analysis and the economic valuation o f environmen­ tal damages necessary for undertaking a social cost-benefit analysis.

For each waste treatment method all social costs and benefits must be included in a full social cost-benefit analysis; this not only include the private (or financial) costs and benefits, but also any external costs and benefits, such as environmental externalities.

3.2.1 Landfill

For landfill, there are two main elements o f costs and benefits:

a) collection and transport to the landfill; b) operation o f the landfill.

The costs and benefits o f the collection and transport o f waste are then:

N S Cl, ^ = P C i , ^ . + E Cl. ^ [ 3 . 3 ]

"The main reason fo r doing social benefit-cost analysis in project choice is to subject project choice to a consistent set o f general objective o f national policy. The choice o f one project rather than another must be viewed in the context o f their total national impact, and this total impact has to be evaluated in terms o f a consistent and appropriate set o f objectives. "

Dasgupta et al. (1972) The waste treatment method with the lowest NSC (or highest NSB) should be at the top o f the waste hierarchy followed by the method incurring the next lowest NSC (next highest NSB) etc.

where PC denotes private costs and EC is external costs. Subscript ‘L* denotes landfill, and subscript ‘trans’ denotes collection and transport.

Similarly, the NSC o f landfill operation can be expressed as:

NSCl.op = PCl,op + ECl.op - EBl.op [3.4]

with the same notation before, and subscript ‘op* denoting operation.

Whereas it might be assumed that the external benefits associated with the collection and transport o f waste will be zero, this is not necessarily so in the case o f landfill operation. If methane gas is extracted from the landfill, and the energy recovered from it, then the latter will displace energy that would otherwise need to be produced on site or purchased from the national grid. Alternatively, it could be sold to the national electricity grid. In either case, the resulting saving or revenue would be deducted from the operation costs, PC^, op> such that this would be a net cost. There is also a positive externality associated with the recovery o f energy from landfill. If the energy thus recovered is marginal to the domestic energy supply, it can be assumed to displace a marginal source o f energy, and thereby save the environmental costs associated with the generation o f that energy. Thus:

[3.5]

where the same notation as before applies and subscript ‘dis.eng.* refers to displaced energy. Equation [3.4] then becomes:

NSCl.op = PCl.op + ECl.op - [3.6]

Note that the external costs o f both operation and collection and transport refer to not only the damage costs resulting from pollution but also the disamenity caused by transport and landfill operation.