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6. EXTERNAL COST ASSESSMENT

This Chapter seeks to highlight the difficulties faced in seeking to estimate even a range of external costs which might be associated with different waste management options. It is by no means complete (so it is rather similar to studies considered in the previous Chapter in this respect). What we are seeking to do is to open up the discussion about how accurately we can really know what the external costs of different waste management options might be. Given that this is an oft-suggested approach to understanding how to aggregate the effects of different pollutants, we are implicitly raising questions as the suitability of such an approach for choosing between different waste management systems.

Our work suffers from many of the same shortcomings that have affected other studies. We have tried to show quite explicitly why extreme caution would have to be exercised by anyone seeking to make use of these (and not just our) estimates in considering policy options, or in trying to understand, through economic approaches, what the 'best' waste management option might be. Later in this study, we seek to understand the implications of this 'requirement for caution' for both policy makers and local authority decision makers.

As well as discussing environmental repercussions over the whole product cycle as far as possible, the depletion of non-renewable resources and 'ecological rucksack' are each briefly reviewed to capture environmental effects of the different options.

The materials found in MSW that are discussed in the chapter are:

·         steel

·         aluminium

·         paper

·         glass

·         plastic

·         high density polyethylene (HDPE)

·         low density polyethylene (LDPE)

and the processes we have reviewed are

·         transport

·         landfilling

·         incineration; and

·         recycling.

We had hoped to discuss compostables, but we have not done so as we have been unable to source data that we had hoped might be forthcoming in the course of this study (see below).

6.1 Life Cycle Approach

One can compare competing options by taking account of cradle to grave environmental and resource impacts. One such approach is lifecycle assessment (LCA), according to the ISO standard ISO 14040 (International Organisation for Standardisation (ISO), 1997). The Environment Agency (Environment Agency 1997) has issued research reports setting out what constitutes best practice in Life Cycle Analysis for Waste Management. The report suggests LCA should consist of the following four stages.

i) Goal definition and scoping, which defines the system to be studied and the functional unit on which the study is based

ii) Inventory Analysis, which compiles data on resources used and wastes and emissions generated in the form of an inventory table

iii) Impact assessment, which converts the inventory table into an understandable evaluation of the magnitude and significance of the potential environmental impacts; and,

iv) Interpretation, where the inventory and impact assessment results are assessed in line with the goal

and scope of the study. Carrying out a full LCA is time consuming and expensive because of the data collection requirements requiring all resource inputs and environmental discharges for all commodities and economic activities in the product

chain to be assessed. It is also extremely difficult to move from stage ii) through to subsequent stages. Some of the reasons for this will be highlighted in this Chapter.

The list of resources used and emissions generated that could be considered within the analysis is given in Table 23 below. This illustrates the magnitude of the challenge.

Table 23: Environmental Impacts Shown in An LCA

We are not attempting, in this study, to carry out a complete LCA analysis along lines proposed under the ISO standards.¹⁹ However, much of what we are doing applies the essence of the life cycle approach. We seek to use inventory analysis (and we have done no primary work here) to quantify the environmental burdens across the life cycle. We have no doubt that the inventory assumptions used will be questioned. This is the first of many reasons one can give as to why the LCA-based valuation approach will always be open to question. As Hukkinen (1999) puts it in an excellent study:

'The inherent systemic complexities of industrial ecology are compounded by the analytical complexities involved in conducting a life-cycle analysis (LCA), which aims to report the cumulative environmental impact of a product throughout its life-cycle…. The complexities of industrial ecology and the consequent analytical confusion can have a paralysing effect on decision making, when social groups with diverse political and economic agendas use conflicting mental models to understand the system. Scientific uncertainties and complexities frequently open up the platform of both public and corporate environmental politics for yet another LCA expert who can question the 'scientific' validity of all previous LCAs.'

[¹⁹ These include standards in respect of general approach (14040), inventories (14041), and two which are likely to be released soon on impact assessment (14042) and interpretation (14043).]

We are not questioning a specific LCA, but in the spirit of the comment above, we question the claim not so much of LCA, but of any LCA-based valuation, to represent some 'true' value (or even range of values) of the externalities associated with waste management options.

The goal of the LCA is to understand the environmental impacts associated with extracting materials from the waste stream through kerbside collection. The approach would in our view ideally take, as the functional unit, a tonne of municipal waste whose composition would be taken from local compositional studies. It would then use actual data from authorities (such as we have been able to gather) and compare the existing situation, in which some materials are extracted at kerbside, with the situation in which the whole waste stream is landfilled (or incinerated). In essence, we would treat, where possible, the tonne of municipal waste as a set of discrete components which may or not be separated out from the whole.

The boundaries of the analysis are such that we seek to compare, for each of the materials extracted, the environmental impacts associated with materials collection, reprocessing and 'remanufacture' with the alternative route of using one or other linear options (landfill or incineration) and then extracting, processing and manufacturing an amount of raw material required to generate an equivalent amount of the material concerned. Note that this means that we are not considering the environmental impacts associated with the material in use. Issues of functionality lie outside the scope of the study. Waste minimisation is also beyond the study's scope. ²⁰

[²⁰ These omissions are important. There are questions to be asked as to how LCA-based approaches, which seek to assist in some way in the making of waste management decisions, can do so in the spirit of 'sustainability' when to some extent, one takes as given the waste materials which enter the bin. Waste management starts well before the waste is generated. It might be argued that those making decisions concerning what to do with waste are not in a position to influence its generation. This is a matter for debate since some studies suggest that the actions of waste managers (e.g. the provision of wheelie bins - see DETR 1997b) do influence waste generation.]

The 'functional unit' against which impacts will be ultimately be quantified in Chapter 6 shall be a tonne of municipal waste, understood to include fractions of each of the materials listed above (steel, aluminium, paper, glass, HDPE, LDPE) as they arise in non-segregated municipal waste. We are necessarily dependent upon secondary sources of data, which are likely to be of variable quality. Needless to say, we are interested in updating our basic analysis on the basis of what might be claimed to be better information, though it is worth pointing out once more that inventory data is unlikely to be beyond dispute. The Environment Agency should soon be publishing reports that have underpinned its LCA tool, WISARD, containing inventories for different processes. The data in WISARD will also be subject to scrutiny and criticism, the more so since the Agency itself seems very keen to see the tool used by Local Authorities and at the regional level (in Strategic Waste Management Assessments).²¹

[²¹ It should be pointed out that full scrutiny of the tool will be made more difficult by the fact that it is being marketed at quite high cost by a private company. In addition, those who feel the tool can be improved might be reluctant to suggest improvements for the same reason.]

6.2 Background to Our Approach

In the course of conducting a (far from exhaustive) review of relevant work undertaken, we have sought to extract estimates for externality adders associated with different pollutants. These are figures that express the externalities associated with a pollutant or effect in a convenient 'per unit' form, such as £/tonne, or p/vehicle km. We have done this so as to simplify the analysis. All of the studies reviewed above employ this 'externality adder' approach. Purists will point out that this is an unsatisfactory approach and they are probably right. More sophisticated studies, recognising the problems associated with benefits transfer,²² will make use of techniques designed to capture as far as possible the impact at a location under study. For example, in the case of air emissions, modelling will be undertaken to establish the change in pollutant concentration due to those emissions, and to understand how this varies across space (so that changes in the level of exposure can be mapped across the receptors affected). Exposure response functions can then be used to estimate effects, and a final step involves valuing these effects (see e.g. IVM et al 1997; AEA 1997; IIASA et al 1998). It should come as no surprise (and indeed, one can argue that it is the corollary of the fact that benefits transfer is problematic) that these adders vary significantly. For this reason, we have used broad ranges of these adders.

[²² The benefits transfer problem refers essentially to the difficulties inherent in carrying location specific valuations across to different locations.]

There are two reasons for doing this. Both actually cast more fundamental questions about whether this simplified approach is really adequate in the contexts under consideration:

1) The first has to do with the presentation of damage costs in this simple way. Using externality adders implies one is transferring estimates from what are often (not always) location specific studies to other places (benefits transfer). It is well known that even if the underlying dose-response functions are known with certainty (and they are frequently not) and are readily transferable (and they might not be), the environmental effects may not be (and this will be implicit in the function) related linearly to emissions. It makes something of a nonsense of the effort involved in deriving location-specific estimates of the net benefits associated with changes in pollutant concentrations to then imply that a per unit emission factor, derived in local contexts on the basis of the effects of a specific source of emissions on local concentrations (and hence, exposure), can be transferred easily from place to place. This may be a tolerable approach where one is considering similar emissions sources in areas of similar population density and geographical characteristics, and where 'background' levels of the pollutants under investigation are similar. Even then, however, different authors make use of different estimates of the value of statistical life, and this will play a scaling role in quantification of the effects. Furthermore, where threshold effects are believed to be involved (and they may be for dioxins) the assumption breaks down more or less completely. By way of example, it probably makes little sense to make use of adders from studies which have modelled exposure to air emissions resulting from a 100m chimney stack, and then converted the external cost estimates to per tonne values, when the source of the emissions might be a car whose exhaust fumes are much closer to ground level.

2) The second raises more fundamental questions concerning the limitations implicit in the valuation approach. There are problems associated with uncertainties in the underlying science (affecting the reliability of dose-response / exposure-response relationships), the ability to model accurately changes in pollutant concentrations and their distribution across media (introducing errors), and methodological approaches to the valuation of life. Even if, therefore, one was dealing with similar emissions sources in areas of similar population density and geographical characteristics, it would be surprising to find agreement across studies upon the external costs associated with a specific pollutant other than in the statement that the approach is a problematic one (although in fact, this is frequently downplayed). Scientific uncertainty, properly understood, is not something that can be handled through probabilistic analysis. Within reason, one may speculate over the boundaries of that uncertainty, but this can be but speculation. The obvious examples here are the cases of dioxins, where the existence or absence of threshold levels in determining the effects of exposure are subject to debate, and climate change, where the significance of extreme events may yet be enormous - we simply do not know at present. Tinch (1995) refers to the latter as being of 'low probability', but this implies that what is not known - the probability of these events occurring -can be given some quantitative basis. Tinch goes on, however, to cast doubt upon the robustness of the damage estimates associated with global warming. This is in stark contrast to the view expressed in CSERGE et al (1993) where the authors express the view that such estimates are robust to variation on the basis of changes in random variables, even though these are again generated within a probabilistic realm. In a recent DTI-funded study, Dames and Moore adopted an approach used by the free University of Amsterdam where global warming externalities per tonne of CO₂ were assessed using a range £3-£109 per tonne (Ecobalance and Dames and Moore Group 1999).²³

[²³ A problem here is that whilst a study may be methodologically sound in the sense of covering all bases, and explaining key assumptions, at the end of the day, what one is seeking to place a value on is not a set of assumptions, but a real-world effect which may well differ in its manifestations to what was being assumed, and therefore valued.]

In essence, therefore, we are admitting that this approach (which has been adopted in all of the studies discussed in the previous Chapter) is a flawed one. To move beyond this, however, would require location specific modelling work, perhaps involving comprehensive work at 'exemplar sites' designed to facilitate benefits transfer to other sites suitably classified by type. Even this, however, would not overcome the second of the issues discussed above.

Unlike some other studies, we have separated out transport components from the specific options. One reason for doing this is that it allows some understanding of the significance of fuel duty in adding to the costs of different options. To the extent that one accepts this represents an internalisation of the external costs of transport, one can estimate the degree to which the total externality of one or other waste management option is already internalised through fuel duty.

This is a significant change in approach. The work done in 1993 by CSERGE et al which informed the setting of the level of the landfill tax did include transport costs within the different waste management options, but the work was undertaken in the year that the fuel duty escalator was announced. Using the landfill case addressed in that study, the fuel duty per tonne of landfilled waste is close to the mean value of the externality reported by the study for a landfill with energy recovery. In other words, some of the externality associated with landfill with energy recovery in that study is not associated with the landfill per se, but the transport to the landfill. To the extent that a) one believed the landfill tax should be set on the basis of externalities (and we have stated elsewhere reasons why it might not be - see ECOTEC 1997), and b) that the transport element has been internalised by fuel duty, one might suggest that the landfill tax should have been falling as the fuel duty increased. Evidently, similar comments could be applied in the case of incineration, though the transport component assumed in the CSERGE et al (1993) study is less significant than for landfill.

For recycling, to the extent that transport externalities may be significant (and for materials which are collected in lower density forms, as plastics are in kerbside collections, they will be especially so), the fact that transport externalities may be a significant component of the total is interesting. To the extent that the other externalities reported (e.g. those associated with greenhouse gases from materials processing) are not internalised, the current level of internalisation will act to skew choices between waste management options in such a way that the level of recycling is below that which would prevail in the absence of any internalisation at all. This is ironic since the results of most studies (see previous Chapter) seem to suggest that full internalisation would have the opposite effect.

Note here that the climate change levy could have had an effect which reinforced the waste management hierarchy. Yet the detail of its design, notably the outright exemptions (for primary aluminium processing as an electrolytic process) and the levels of exemption proposed for intensive energy users (steel etc.), will reduce the extent to which the price mechanism affects the balance between recycling and primary materials use. This will be further hindered by exemptions for renewable energy, including energy from waste.

We have tried to be reasonably accurate in converting and updating past externality estimates to ensure they are comparable, and are converted accurately into UK currency terms using appropriate deflators and exchange rates. However, the date to which the originals refer is not always absolutely clear. Any inaccuracies will be of limited concern given that:

• Most of them come from relatively recent work so that the impact of exchange rate movements and / or deflators will be relatively small; and

• We are using ranges of values, and the range is typically very large, so that any 'accuracy' lost in the conversion and updating is more or less spurious in the context of the ranges available (it seems rare to find externality adders which are all confined within a range of one order of magnitude).

With respect to the last point, mindful of the many caveats which need to be applied, we are aiming at illustrating ranges which are plausible on the basis of the existing literature, and with the understanding that the analysis is a long way from being a complete one. It is better, in our view, to indicate a broad range of possible estimates than either pretending that we can undertake valuations in possession of certain knowledge, or adopting the 'lucky dip' approach to the valuation of externalities (and indeed, this is the advice which is given - it seems rarely to have been followed - by the Better Regulation Unit).²⁴

[²⁴ See www.cabinet-office.gov.uk/regulation/1998/brg/brg_part2_section2.htm. Here, there is guidance on how to treat uncertainty in the context of Regulatory Impact Assessments. Whether the study goes far enough in appreciating the radical nature of uncertainty that can exist is debatable (one is still being asked to estimate the magnitude, or the extent of uncertainty, i.e. to say something about something one might know next to nothing about). There is still pressure for quantification.]

We have grouped 'high' and 'low' valuation factors together. There are two reasons for doing this. One is that several of the factors used relate to pollutants whose values are estimated using estimates of the valuation of life (see below). High values could be assumed to result from high values of life. The other is that the high-value externality adders, to the extent that these are related to air pollution, would perhaps be more relevant for places where population densities are higher (e.g. urban areas). In both cases, there would be reason to believe that where one adder affecting health is 'high', others might be high too (on the basis that the population exposed is significant, or that gas concentrations are low because of high chimney stacks, in all cases).²⁵ There are exceptions to this 'rule,' and in addition, where one is dealing with emissions from different locations, the assumption no longer holds true. We show below what can happen when one changes the adders as they are applied to specific plants. Generally, through creative manipulation and choice of externality estimates, one could make one's ranges broader by choosing high and low contributions where these work in the same direction to increase the net externality (i.e. one would choose high values for positive contributions and low values for negative ones).

[²⁵ There are exceptions to this 'rule' however. Ozone, as derived from VOCs, would be one as it tends to be generated in areas where specific carriers are not present. These carriers exist in the main in areas where NOx is present. Hence, tropospheric ozone may do more damage in areas which are less densely populated (even when its precursors are actually emitted in densely populated urban areas).]

We are not well-placed to know what might be the income elasticity of demand for avoiding the external costs being assessed. Coopers and Lybrand and CSERGE (1996) (and Brisson 1997) work on the basis of an income elasticity of demand of 0.3 (using a figure of 1 for sensitivity). A more elastic demand (as has been hypothesised in the context of some agri-environmental studies) would magnify the effects of increased real incomes over the time after the externality assessment was first made.

Lastly, whilst some studies seek to allocate environmental burdens associated with landfill across the whole life-cycle of the landfill, our principle focus is on marginal changes in the use of one or other type of facility. We have not made any attempt to attribute environmental burdens associated with, for example, landfill engineering, to materials landfilled. The assumption is that externalities associated with landfill engineering are fixed, and only those emissions directly associated with the waste landfilled are taken into account. This may limit the usefulness of this type of approach where the question being raised is one of whether or not to construct one or other facility, although the private cost analysis clearly takes into account the financial side of the equation. Whether the private costs internalise externalities or not, and how well they do this, will depend upon future decisions as to the adequacy of financial provisions for covering the potential for accidents, the requirements (if any) for compensating local residents in the context of decisions to site new facilities, and the ability to enforce operating standards for specific treatment plants.

The marginal / non-marginal distinction is an important point and one that is rarely addressed adequately in the literature. It raises questions as to how, when marginal changes are being considered, to account for disamenity effects where many of them may be poorly (if at all) related to the level of inputs to the facility. For example, where landfills exist, the disamenity associated with the existence of the facility is unlikely to be dealt with best through a pro-rating of that across landfill inputs since the disamenity will be relatively fixed irrespective of inputs (elements related to transport, for example, will not be - see ECOTEC 1998 and EFTEC 1999 for a discussion in the context of aggregates extraction - but most of the analysis undertaken thus far already accounts for some of the transport-related externalities. More local transport disamenity such as litter, dust and dirt, and noise associated with vehicles congregating at the site will not have been included).

6.3 Waste Transport

6.3.1 Residuals

The collection of residual waste for treatment in linear waste management options involves the use of vehicles on a journey which will take them from a depot on to the collection round to pick up waste from bins and bags. Waste will then go either to a transfer station or direct to a landfill site or incinerator. In addition, some bulky waste is collected either for a fee or at zero cost to its producer, from households. There is also the collection of waste delivered to Civic Amenity sites to be considered and the collection of litter from parks, public places, highways etc.

The majority of the waste is collected at the door (although the non-doorstep fraction is a significant fraction of the total). We focus on that element of the waste in this study. As such, we are assuming that the collection approach does not affect the volume put out by the householder. This will not be strictly true. The provision or otherwise of kerbside schemes for dry recyclables will have a bearing on the amount of material taken to bring sites where these are already available. Equally, the provision or otherwise of kerbside systems for organic waste collection will affect the amount of such material taken to civic amenity sites (less will be taken where provision of kerbside collection exists). Transportation of refuse typically takes place in vehicles which may weigh some 24 tonnes, but which have a capacity, typically, of 10-14 tonnes (since they have a dry weight of 10 tonnes or so).

6.3.2 Recyclables and Compostables

The collection of recyclables and compostables can involve a similar collection round to that for residuals, with materials typically being delivered to a depot for sorting (where this has not been done on the round itself). The materials might then be composted at site, or in the case of dry recyclables, transported to reprocessors for their use. Different vehicles may be used for dry recyclables but in our schemes, the vehicles tended to be between 7.5 and 11 tonne vehicles with payloads between 2 and 4 tonnes. Certainly for dry recyclables, vehicles are unlikely to be as fully loaded in weight terms. This may increase the private costs of transport, but it will reduce the associated external costs per load since some externalities associated with transport are related to vehicle weight (e.g. those for road damage and, to some extent, emissions through the relationship to fuel efficiency).

6.3.3 External Cost Analysis

There are a number of impacts associated with transport which should be accounted for in a complete analysis of environmental impacts. These include:

• health effects of vehicle emissions (local);

• effects on global warming through greenhouse gases (GHGs);

• transport related accidents, fatal and non-fatal;

• transport related noise; and

• damage caused to highways.

This is by no means an extensive review of all environmental impacts of transport (see Tinch 1995, Maddison et al 1996). Relatively few studies look at the issue of damage done to the road itself. The system of road tax is now arguably better structured to internalise this, but it is not clear that the nature of the vehicle (maximum load, number of axles etc.) is adequately accounted for in the analysis of external costs. In any case, we are principally interested at this stage in the external costs generated rather than their current level of internalisation (a point to which we return below).

Note that the activity in which those collecting waste are engaged, by virtue of its being carried out on public highways, may expose them to a higher probability of accident than those in other occupations. External costs associated with collection could account for accidents suffered by those engaged in the activity concerned. There is good reason to believe that not only the exposure to traffic, but also the handling of materials in waste, are likely to pose specific hazards. Powell (1992) noted that 'over-3 day injuries' (those which cause workers to be off work for more than three days) are much more common in waste collection than in comparable industries. She noted that both the physical handling of material and the nature of vehicles used could be an issue.

Whether this should properly be accounted for as an external cost depends, arguably, upon whether one believes those facing the hazards involved are well appraised of, and either protected from or compensated for, them. To the extent that one believes that they are, the externality is internalised (such a view is adopted in recent work by Ecobalance and Dames and Moore Group (1999) for the DTI). The belief that such a calculus is being made by employees underpins one approach to the valuation of a statistical life. The hedonic wage approach is founded upon the belief that those undertaking employment consider the remuneration in the context of their exposure to hazards. To the extent that alternative employment opportunities may be limited - and it seems fair to assume that they may be for those employed in waste collection - the remuneration might not reflect this increased exposure to hazards (labour market effects might act to counter the presumption in favour of increased remuneration). Therefore, the health related external costs of waste collection, to the extent that they are based upon average figures, could be understated since the nature of the job places workers at a particular risk, and these might not be reflected in wage rates.

The greater these health related risks are, the more potentially significant it becomes that kerbside collection often (though not always) involves greater time spent in collection activity (in the case of weekly collections of each) than would otherwise be the case. Although the same materials (more or less²⁶) may be being handled, the fact that workers are frequently working in the road itself adds to the dangers associated with the task. This is not explicitly included in the analysis. It is worth pointing out that we have, in speaking to those operating kerbside collection, discussed the issue of injuries. The general response has been that these are fairly rare, and that none were serious in nature. More empirical data on this would be useful (we have not actively sought this, so it may exist). One respondent mentioned the issue of morale in the job, but this was traced to more general concerns (possibly the time of year the research was being carried out - a month or so before Christmas).

[²⁶ The introduction of a kerbside collection for organic material may bring more material into the collected waste stream than in the absence of such a collection.]

We have evaluated transport effects in two ways.

Method 1

In the first approach, emission coefficients for health and noise which come from the Tinch Report (Tinch 1995) and the European Council of Ministers for Transport (ECMT 1998) were used. For the air pollution and noise costs, we have used HGV (heavy goods vehicle) per 000tkm estimates. The higher end of the range in Tinch (1995), taken from the 'urban driving' estimate, may not cover the ECMT figure (depending upon how one updates them). Tinch himself notes that his figures are ' "best estimates" drawn from a survey of the literature. They are intended to show the potential for valuation, and should not be interpreted as "the" value of those [noise and air pollution] effects.' The ECMT (1998) figure has been taken as 5.4p/tkm.

For global warming, the higher value used is from the ExternE estimates as cited in EFTEC (1999) (note the context is similar so the estimates are likely to be transferable). The lower value is the shadow cost estimated in the ECMT (1998) report. The estimates have been updated to account for exchange rates (where, as in the ECMT report, the estimates are in 1991 ECU) and for inflation. These values are shown in Table 24.

Table 24: Valuation Factors Used in Transport Analysis

It is worth pointing out that valuation of transport-related accidents are driven by the separate products of the number of accidents and fatalities, and a measure of the value of life. Whilst some knowledge of the former (at least in the road transport cases) may be gained through statistical analysis, the latter is subject to considerable uncertainty and debate.

Some of the estimates considered in the course of the ExternE study are given in Baranzini (1997) (see Table 25). Other reviews have found variation from ECU 360,000 to ECU 10 million (EFTEC 1996), or elsewhere, from 0.3 to 17.5 MECU (European Commission 1995). The UK Government uses a value of just below £1 million for the purpose of valuing transport deaths and casualties (£902,500 for June 1997 - DETR 1997a), whilst Pearce and Crowards (1995) suggest a value more than double this is more appropriate. The latter is consistent with Metroeconomica's (1996) estimate of ECU 2.8mn, and indeed, most Commission studies, including ExternE, have settled for figures between ECU 2.6-3.0mn - which is close to the Pearce and Crowards (1995) view. Few studies discuss the influence of the nature of the cause of death as a potential factor influencing the value used, which is increasingly seen as an issue in debates on the matter.²⁷

[²⁷ It would appear to be correct to vary one's valuation of life according to the nature of the risk to which people are exposed, at least to the extent that one is evaluating only people's preferences. Sociological studies of risk reveal that lay-people's rationality (and probably that of experts when they are 'acting as' 'lay-people') conflicts with expert assessment of 'risk' (which is not to argue the correctness of one or the other). Factors such as 'dread' and the control which people are able to exert over their exposure to the risk concerned appear to affect their perception of risk, but in ways which are poorly understood at present (for a discussion in the context of waste, see Kasperson et al 1992; Gerrard 1994; more generally, see, e.g., Starr 1976; Slovic 1981; Slovic et al 1994; Horowitz 1994). The approach in work undertaken by NERA (1997) appears to be to pluck multiplicative factors out of the sky to 'account' for these. These seem to have been chosen so as to arrive at results which are 'not too high' and 'not too low', once again downplaying the significance of the uncertainties (perhaps more correctly expressed as ignorance) involved in accounting for such poorly understood impacts upon risk perception. In any case, one suspects (from some detailed consideration of the matter) that were one to find some relationship between risk perception and the nature of hazard, that perception of risks varies in a non-linear manner in relation to the potential consequences.]

Table 25: Valuations of Life Considered in ExternE Project

Source: Baranzini (1997)

Although the closeness of the agreement in some studies and amongst some authors may appear to suggest some form of convergence, in our view, it would be wrong to suppose that such values are necessarily 'correct' by virtue of this agreement. It remains appropriate, in the face of continuing debate, to retain high and low estimates. There appears to be no means of validating these estimates, the only form of validation being that associated with how well a particular study, approached using a specific methodology (which one may or may not accept as valid for the purpose) has been performed. Methodological approaches are still the subject of disagreements, not to mention the question of whether this should be done at all.²⁸

[²⁸ The classic argument for doing so is that policy makers need to allocate resources and therefore make decisions across competing claims. One might ask why, if this is what policy makers do, do they need consultants and economists to do this for them? There is an interesting debate to be had about whether, once aspects of CBA start trying to account for the nature of the hazard to which an individual is exposed, the approach has not finally discovered its own limitations. CBA proponents always claim that individuals, in making decisions, make them on the basis of a cost-benefit analysis. This has always been a questionable assumption (see Sagoff 1989). The fact that a decision has been made cannot lead, ineluctably, to a deduction about the nature of the process by which the decision was arrived at. We live in an extremely complex world, and bombarded by information which carries competing messages, we have, in Scott Lash's words 'no choice but to choose.' In seeking to account for more psychological characteristics, valuation techniques are now trying to capture the spirit of a far more complex rationality through which it hopes to elicit people's preferences. In doing so, it is likely to encounter limits to the degree to which it can be assumed that individuals' preferences correspond to what is assumed to be 'rational' behaviour in the economic context. An exploration of these issues from a different perspective can be found in Sagoff (1994) (see also various chapters in Foster (1997)). Quite apart from the fact that cost-benefit analyses are likely to contain uncertainties and omissions which are not always made completely explicit, such approaches to decision making risk reducing the significance of what some might argue are more fundamental moral and political issues. Some of the responses to contingent valuation questionnaires given by those who are questioned provide a testament to the extent of unease felt by many in going down this route. Interesting examples of the more sceptical attitude to valuation and its use in the field of policy-making are Foster (ed.) (1997) and Vatn and Bromley (1994).]

With respect to data on accidents, we have used two estimates. Both are from the DETR. The first (high) set is from DETR (1997a) and relates to casualties from trucks. The second (low) set is from the DETR (1998) and is calculated from figures for HGV traffic volumes and accidents to HGV drivers and passengers. Evidently, this would be expected to be low since it does not include data on pedestrians and the like who may be involved in accidents involving HGVs.

The other aspect where the valuation of life plays an important role (albeit in a non-transparent way in our analysis) is in the health impacts of transport emissions (and air emissions more generally). Here, economic effects are elicited by establishing the effects of emissions on concentrations, usually through atmospheric modelling, and then using dose response functions to estimate the effects on health of these changed concentrations. The modelling of atmospheric concentrations is far from being a precise science partly because the dispersion of pollutants is likely to vary under specific topographical and other local conditions. Dose-response relationships are also the subject of varying degrees of debate (depending on the pollutant). The final step involves valuing the mortality and morbidity effects of the specific pollutants.

Current debates in the valuation literature take the view that mortality effects of air pollution should be treated differently from those associated with, for example, car accidents since in the former case, it is argued that the effect will be to bring forward the deaths of those who would have died soon after anyway.²⁹ Discussions have taken place, therefore, concerning whether, in the case of air pollution, the most appropriate measure for valuing life might be one based upon the value of life years lost (VLYL) or on the value of a statistical life (VOSL) (for a useful discussion, see RPA and Metroeconomica 1999). In the context of air pollution, a recent Department of Health publication decided to use a range of estimates for willingness to pay to reduce the risk of a death brought forward from £2,600 to £1.4 million (DH 1999). Department of Health Ministers subsequently decided that the currently available data 'do not allow the benefits of reducing air pollution to be converted into monetary terms with a sufficient degree of certainty to allow the results to be used in the cost benefit analysis of the NAQS [National Air Quality Strategy]' (DETR 1999b).

[²⁹ The strange thing about this is that, from the perspective of external costs, deaths caused by pollution are less of a concern than those that are treated as accidents on the road. Apparently, because the deaths from pollution are those of vulnerable people, they are attributed less value than if the person were 'less vulnerable.' Moral outrage at murder works in the opposite sense. The more vulnerable the victim the more repugnant the crime. It would be difficult to counter the view that what the pollution is doing is not actually killing the most vulnerable people (this is what the science tells us). It is paradoxical that this is seen as (in relative economic terms) less of a worry than killing younger people. The whole notion of 'bringing death forward' seems to be an attempt to sanitise what is actually a rather unpalatable situation in which we are seeing vulnerable people killed by pollution. Presumably, no self-respecting defence lawyer would state in Court 'sorry, m'lud, my client was merely bringing the deceased's death forward.']

Note that in this analysis, the valuation of mortality only enters into the analysis directly in the context of accidents and injuries related to transport. Indirectly, a valuation of mortality is implicit in externality adders used, however. To the extent that these have been based on studies that made use of mortality estimates based on VOSL (as opposed to VLYL), they will be higher than would be the case had VLYL estimates been used. In this work, we have taken, as high and low estimates for mortality, £6 million and £500,000 respectively.

We have valued congestion using the estimates used by CSERGE in our work for DETR (ECOTEC 1999), these coming from Newbery (1988; 1990).

Method 2

Method 2 is very similar but goes back to first principles in respect of emissions from transport. We have then applied externality adders (used elsewhere in the study - these are shown in Annex 2) to a sub-set of the range of pollutants emitted. Both congestion and casualties are treated in the same way as in Method 1, so that all that changes are the health and global warming estimates which are now derived through estimates of vehicle fuel efficiencies and emissions associated with the relevant fuel type.

6.3.4 Results

We illustrate below, in Tables 26 and 27, the externalities by category for a waste collection system carrying waste in RCVs with average payload 10 tonnes travelling 80km.

Table 26: Method 1, 10 Tonne Payload, 80km roundtrip

NB: Totals are subject to rounding

Table 27: Method 2, 10 Tonne Payload, 80km Roundtrip, Impact Per Tonne Of Waste Transported (£/t)

NB Totals are subject to rounding

The only real differences in these analyses are in the valuation of global warming externalities, and in health effects of pollutants. This illustrates the different results which can be obtained through adopting, on the one hand, the more bottom-up approach in Method 2, and the approach in which one chooses to lump together transport-related impacts on a per kilometre, or per tonne kilometre basis. When one looks at the full analysis, the following points can be made:

• A trivial, though nonetheless important observation is that if we take the low externality adders, the total externality is small. One might suggest that this is tantamount to saying that we are more or less indifferent to what distances waste is moved and in what vehicles when we are not bothered about effects on people's lives, or where we do not think that global warming etc. will have important ramifications. In a sense, if nothing is important, if pollutants do not really cause any harm, and nothing is really changing in ways that need bother us, the issue need not concern us. Whilst this case should not be ruled out as a possibility, policy based on the low externality adder case (given the prevailing uncertainties) is somewhat cynical and potentially leads to significant levels of regret in the future.

• Less trivial is the fact that when aggregating different types of externality, one has to be very careful to account for externalities correctly. The reason for this is that some of the external costs are not directly related to tonnages per se. Some are derived from the number of kilometres a vehicle has travelled. It is inappropriate, in such conditions, to believe that the external costs associated with the transportation of a vehicle carrying ten tonnes of waste will be ten times the externality associated with the movement of one tonne of waste where one has attributed to each tonne a 'distance travelled' equivalent to that moved by the whole vehicle. This poses no great analytical problems, but the problem has not been properly treated in other studies. By way of example, externalities associated with casualties tend to be related to distance travelled. In our earlier work with CSERGE (ECOTEC 1999), taking the tonne of waste as the functional unit, the view adopted was that since each tonne of waste was being transported the same distance (i.e. the whole journey), this distance could be used in calculating the externalities which are related in some way to distance. But clearly, if a ten tonne load is being transported, this approach erroneously attributes some external costs related to each load to each tonne being transported. In this way, externalities associated with casualties owing to movements of waste were over-estimated. Properly treated, the external costs relating to casualties are 'diluted' by the weight of the vehicle's load. As such, one finds that for a 100,000 tonne collection scheme using trucks carrying ten tonnes of waste (they may be 24t RCVs, but they do not carry 24t of waste) the significance of the value attributed to life is small in the calculation of per tonne externalities.³⁰ It increases as one shifts to lower 'payloads'. However, the lower the payload, the more likely fuel consumption is to improve, reducing externalities associated with health and global warming (see below). Even so, at the higher payloads (i.e. especially for residuals), the way in which 'valuation of life' affects our analysis is through the number of casualties associated with transport, and since these numbers are quite low, the effects are relatively small. Note that casualties per tonne of material collected will be far more significant in the case of bring schemes in which journeys are made specifically for the purpose of delivering materials (since the casualty and accident rates are not significantly different for cars, but the number of journeys made to collect one tonne of material may be quite high)³¹.

[³⁰ Obviously, the 'high' and 'low' range in respect of impacts on health (as opposed to casualties) are in some way related to differing valuations placed upon life so this does enter the analysis indirectly through other routes.

³¹ Obviously, at higher densities of bring sites, journey distance for those doing the 'bringing' will fall, and indeed, use of a vehicle may become completely unnecessary.]

• The greatest contributions to the total externality associated with moving waste from one place to another come, potentially, through congestion, global warming, and the effects on noise and health. Clearly, congestion effects will vary depending upon the route the vehicle takes/has to take, as well as the time at which the journey is made (some studies attempt to account for variation owing to the these by differentiating by level of urbanisation and on and off peak periods. This is not easy especially since these change, and they can be different in different areas). The same things can be said for effects on noise and health (since these will be related to population exposure). Noise and health effects will, however, also be amenable to influence through adequate vehicle maintenance and use of modern vehicles with suitable emissions abatement equipment. The great unknown is global warming, and the effects of transport upon this are invariant with respect to location of emissions. They do, of course, vary with distances travelled, and also with the vehicle load. In the case of kerbside schemes, these externalities will fall (per tonne of material collected) as participation rates increase. At the same time, depending upon the relative rates of growth of residuals and the kerbside scheme, and depending upon whether the collection takes place at the same time as the residuals, the external costs per tonne of residual collection will rise.

Note that in an earlier study (ECOTEC 1999), we did attempt to understand the extent to which transport related externalities were already internalised through fuel duty. This was not done in the study by CSERGE et al (1993) since the escalator was only introduced in 1993. Effectively, we can calculate an implied level of internalisation per tonne of waste transported in a given phase (on the basis of the fuel consumed per tonne of waste transported and the existing level of fuel duty). We estimate this to be some £1.2 per tonne of waste (assuming a round trip for a 10 tonne load of 80km).

Since transport externalities were used in the CSERGE et al (1993) assessment of the external costs of landfilling and incineration, then to the extent that the assessment of external costs was used in support of the tax level, there would be reason to believe that the internalisation of some of these 'landfill' externalities would suggest lower levels of landfill tax. The current level of fuel duty, applied to the rural landfill scenario in the CSERGE et al (1993) report (return journey of 80km, so total of 160km), would effectively internalise around £2.31 per tonne of waste assuming 16 tonne trucks with a 10 tonne load returning empty from the landfill (i.e. 16km per tonne landfilled). This assumes a fuel consumption of 0.32 km/l of fuel (from White et al 1995). This is interesting since mean values of the externalities from landfill as measured by CSERGE et al (1993) are lower than this for urban landfills with energy recovery and only marginally greater for all other types of landfill examined. Increased fuel efficiency and lower transport distances would, of course, lower the duty per tonne of waste.

6.4 Landfill

Emissions associated with landfill are a subject of some debate. Estimates of landfill gas generation have been given in Aumonier and Warren Spring Laboratory (both reviewed by CSERGE et al 1993), Powell (1992), USEPA (1998) and Entec (1999a) among others. Relatively little information exists concerning the external costs of landfill on the environment, a somewhat surprising statement since it lies at the bottom of the waste management hierarchy, and is therefore arguably deserving of attention. Indeed, if there is uncertainty about its impacts, one might reasonably question the logic behind its position at the base of the hierarchy. Reinforcing its position at the base, however, is the view that however well engineered they may be, landfill liners (natural or otherwise) will not contain waste indefinitely. Quite apart from the issues associated with (temporary) land-take, therefore, there is a perception that at a more fundamental level, the practice of landfilling simply passes on a problem created in one generation to another in (possibly very many) years to come. This has been debated more seriously in connection with hazardous and radioactive waste landfills and depositories (see Gerrard 1994 for an account of the US experience).³²

[³² In the case of radioactive wastes (where the 'future generations' issue is most pertinent, Gerrard (1994) reports that the US Department for the Environment 'spent several million dollars designing a "keep out" sign for WIPP [the Waste Isolation Pilot Plant] that would be effective for 10,000 years and recognisable by any future earthling.']

The work that was undertaken by CSERGE et al (1993) prior to the landfill tax concentrated primarily upon GHG emissions and upon transport to the landfill site (see above). That study did not distinguish between the CO₂ emissions that arise from biogenic sources and those that do not. The argument given was that this would not alter the analysis significantly. Yet this assumes that the estimates of damage associated with GHGs are fairly well understood (and implicitly, that they are believed to be small). Furthermore, it is a statement that has to be made relative in the context of an analysis which focuses only on a subset of the total external costs, and where sensitivities in respect of landfill gas collection and combustion are ignored. Collection and combustion of landfill gas has the net effect of converting CH₄ to CO₂, making the question of how one accounts for proportion of the GHG emissions which emerge as CO₂ rather more important (since more CO₂ is produced, but much of this may be from biogenic sources).

The CSERGE et al (1993) report offers the view that estimates of the valuation of damages associated with global warming as have been made are relatively robust, yet it alludes to studies which seek to deal with 'uncertainty' through use of random variables with a triangular probability distribution.³³ This view is at odds with that of ECMT (1998) (and in spirit, that of Tinch 1995 - see above) who, in taking what one might call a precautionary approach to the issue of uncertainty, used a value of 50ECU/kg rather than the $20 per tonne used in the CSERGE et al (1993) study. These are three orders of magnitude apart. It seems that if one believes that one must attribute values to phenomena whose outcome is uncertain, the use of wide variations is likely to be if not the appropriate way, then the only way to deal with that uncertainty (rather than to pretend that one has certain knowledge of something about which one has admitted one does not) if indeed one believes one can within this sort of analysis.³⁴

[³³ If the so-called uncertainty is being approached through probabilistic analysis, it loses the characteristic of being uncertain. Uncertainty as defined here is qualitatively different to inaccuracy, or error in measurement. One might be able to ascribe boundaries or probabilistic assessments to the latter, but not to the former.

³⁴ A study for the World Bank by Hagler Bailly et al (1997) made use of shadow price values of $5, $20 and $40, but even this choice was arbitrarily made.]

Elsewhere, it has been usual in valuation of the effects of biodegradation under landfill conditions to ignore the releases of CO₂ on grounds that these are emissions which would have occurred anyway and that they are part of the carbon cycle. The argument is that these sources of CO₂ are not the consequence of anthropogenic releases into the atmosphere per se, but are releases that would have occurred anyway (USEPA 1998). The methane component, on the other hand, can be considered anthropogenic in character. It would be consistent with this view not only to ignore the CO₂ emissions from landfill (on the basis that all are biogenic), but also to subtract from any valuation of the emissions of methane from landfill the value of the equivalent emissions of CO₂ which would have occurred had the material been biodegrading outside landfill. As far as we can see, this has not been done in any external cost study thus far.

Our analysis has tried to shed light on an important question to consider as the composition of waste being sent to different options changes over time. Indeed, since consideration of the matter might shed light upon the desirability of sending different wastes to different disposal options, we have sought to model the externalities from landfill in such a way as the model can incorporate changes in waste composition sent to landfill.³⁵ In doing this we have relied on estimates of methane emissions which come from only one source (Barlaz 1998), which is recognised as a problem by the USEPA in its work (from where these estimates are taken - see Annex 3). It is interesting to note that some materials are treated as net sequesters of carbon in this model since their carbon is deemed of biogenic origin and is assumed to degrade incompletely in landfills.

[³⁵ It is also worth questioning at a more fundamental level the presumption that the emissions of CO₂ which occur outside landfill conditions are truly not of anthropogenic origin. The activities which constitute the cycling of carbon are, it could be argued, being artificially speeded up. The rate at which photosynthetic product is extracted from the land (and the way in which its production may be speeded up, for example through the use of synthetic fertilisers, production of which uses natural gas to fix nitrogen) has consequences for the cycling of carbon and for the fluxes of GHGs. However, for the purposes of this study, we do not investigate the matter further.]

The model also allows for varying estimates of the rate at which methane is oxidised through the landfill cap, though estimates used in USEPA are 10%. We have also allowed for flexibility in terms of performance in respect of gas recovery from the landfill, and hence, in the case where the collected gas is used for energy recovery rather than flaring, efficiency of energy conversion. This allows us to model the situation for three types of landfill:

1. one where no gas collection occurs (so there are net emissions of CH₄, with any CO₂ emissions assumed to be biogenic).

2. one where gas collection occurs and all the collected gas is flared (converting CH₄ to CO₂, hence reducing the costs associated with collected gas since there is oxidation to CO₂ of biogenic origin).

3. one where collected gas is used for energy recovery (so the same oxidation effect occurs, with the added benefit of displacing energy).

Displaced energy is treated in a separate module, which uses high and low emissions factors for air pollutants to arrive at high and low estimates of avoided costs per MJ of energy generated. This is done for three cases - that where one assumes the marginal energy source displaced is coal-fired, that where one assumes the marginal source displaced is from the UK average mix (avoided emissions from these were based on ETSU 1997, see Annex 4), and that where no displacement is assumed. Hence, there are four non-zero values for the avoided external costs associated with energy generation (each of the two displacement cases with high and low adders, respectively, applied).

Note that in this system expansion, the avoided externalities associated with gas collection and flaring / gas collection and energy recovery depend upon a number of factors:

• the volume of gas generated;

• the composition of the gas (in particular, its calorific value, dependent principally on the proportion that is methane);

• the gas collection efficiency;

• in the case of energy recovery, the efficiency of that recovery process; and

• the assumption made about which source of energy, if any, is being 'displaced'; and

• the emissions data pertaining to that source.

We noted in the previous Chapter that these figures and assumptions are crucial in arriving at figures for the net externality attributable to specific waste treatment options. Methane generation is discussed in more detail below. The assumption concerning avoided external costs deserves further comment, however, because of its critical influence on the analysis.

6.4.1 A Note on Avoided Externalities Associated with Energy Recovery from Waste Treatment Facilities

It has been customary in analyses of this nature to treat energy recovery as having beneficial impacts through the displacement of other sources of energy. There are a number of issues that one would need to account for in dealing with the issue. The first involves whether one is really displacing anything, and whether it might not be the case that, because energy use is expanding, nothing is being displaced as such. One is simply sourcing energy from different (new) sources. It makes sense to recover energy from incineration plants since the process itself implies generation of energy (if not necessarily its recovery). This approach would hold that no energy source is being displaced per se.

If energy use is expanding, or more generally, if one looks at the longer-term, the question of which if any source is really being displaced, or replaced (even in a shrinking scenario, plant is replaced), might be reduced to one of 'which source is not being introduced that would otherwise have been introduced?' Two approaches might be relevant here. The first would be to make the observation that the principal new source of energy is gas. There might, therefore, be an argument that one should consider gas-fired power as the source being displaced. The second might look to the longer term. Which energy sources are we seeking to develop in the future? In this case, the answer might be 'renewables of one or other type', especially if the Government is keen to meet its target of 10% for the proportion of energy supplied by renewables in the future. The displaced source could, therefore, be an alternative source of renewable energy. Even here, however, it could be argued that government renewables targets are set on the understanding that energy from waste will be a contributor.

More commonly, in studies of this nature, it has been common to focus on marginal changes. The incremental increase in energy from waste capacity would displace the marginal source of electricity. It is this perspective that has led those carrying out this type of analysis to treat recovered energy as though it were displacing coal, or the average source of energy supplied.

We understand that this view was also adopted in the Environment Agency's WISARD mode though for different reasons. The argument that seems to have been employed is that energy from waste displaces coal since it replaces energy sources which are not base load. Some argue that this is a difficult argument to sustain since incinerators are generating more or less continuously.

We have used both the standard assumptions, as well as the assumption that no displacement effect is occurring. We do this since the effect of the 'replacing coal' and 'replacing average energy mix' assumptions are controversial. Using the no displacement scenario not only allows one to see the effects of these assumptions on the results but also reflects our belief that the standard assumptions are controversial and likely to generate disagreements. It may, in any case, not be appropriate to apply the usual assumptions where one is considering non-marginal changes in the supply of energy from waste treatment plants.

6.4.2 A Note on Methane Emissions and Energy Generation from Landfill Gas

Methane emissions from landfills are not incredibly well understood. A range of estimates could be generated from different studies in the public domain. CSERGE et al (1993) looked at estimates from Aumonier and from Warren Spring Laboratory (WSL), and found ranges for best estimates of methane generation of between 53-81 m³ per tonne of MSW. The full range, from the low estimate assuming 20% methane oxidation, to the high estimate from Aumonier, was from 25-117 m³ per tonne. Powell's (1992) mini-survey estimated recoverable quantities of the order 100 m³ per tonne (in which case, the actual quantities would presumably be much higher). Entec (1999a) on the other hand, used much higher figures of the order 400-500 m³ landfill gas per tonne of MSW of which 50% was assumed to be methane (i.e. 200-250 m³methane per tonne MSW).³⁶ Using the composition figures we have taken, the USEPA (1998) methane generation figures give 50 m³ at 5% oxidation rates, and only 42 m³ at 20% oxidation rates. It should be noted, therefore, that these are relatively low estimates of methane generation. Methane emissions in our analysis come from USEPA figures, not because we feel these are 'correct', but principally because they allow us to link methane generation to specific components of the waste stream.³⁷ No methane emissions factors are given for emissions from screenings, textiles and miscellaneous combustibles, which together comprise 18% of MSW in our compositional data. Hence, methane emissions per tonne of MSW are sensitive in our model (and obviously in practice too) to the waste composition. In particular, looking at the USEPA data by material type (see Annex 3), methane generation is sensitive to the distribution of 'paper' across paper and board types, as well as to the distribution across putrescible components, especially the relative proportion of food scraps.

[³⁶ EIRU (1992) report similarly large ranges in a review of theoretical studies.

³⁷ It has been suggested that the USEPA (1998) figures are low partly for political reasons (since this reduces the US contribution to global warming from landfill gas).]

Discrepancies are magnified when one looks at the assumed energy delivery from MSW landfilled. Calorific values for 1 m³ of landfill gas of the order 19MJ/m³ landfill gas have been quoted (Entec 1999a). Almost equivalently (if one assumes 50% of the gas is methane) a calorific value for methane of 39.75MJ/m³ was used by Manley 1990 (in Powell 1992). Entec (1999a) then estimate energy content of landfill gas per tonne of MSW by:

1. estimating gas collection efficiencies; and

2. the percentage of landfill gas utilisation over the lifetime of the landfill,

and then multiplying these factors together, along with the calorific value mentioned above, to arrive at a calorific value of the gas collected. This is then further reduced by a factor representing the efficiency of the engine used to generate electrical output (Entec (1999a) use a figure of 40% for the engine efficiency).

Using the approach taken by the US EPA, for every Metric Tonne Carbon Equivalent (MTCE) of CH₄ collected one generates 646 kWh of electrical energy, which translates to approximately 115-150 kWh per tonne MSW (depending upon assumptions concerning oxidation rates). This can be compared with an estimate of only 79 kWh in CSERGE et al (1993) and 298-475 kWh (worst and best case scenarios) in Entec (1999a). In our study, we have used US EPA (1998) methane data and then followed an approach similar to that used by Entec (1999a) to derive energy recovery figures. This has meant converting from MTCE CH₄ from the US EPA report to m³ of CH₄ for the purposes of understanding energy generation (and we have assumed a conversion factor of 238 m³ per MTCE CH₄). This means that at 35% engine efficiency rates and 30% landfill gas collection efficiency, the energy generated is 59 kWh, whilst at 60% collection efficiency, the energy generated is 118kWh (in our model, this is independent of oxidation rates).

Note that this approach does still introduce the question about whether one is interested in marginal changes or those over the lifetime of the landfill. Arguably, if one is landfilling in the period after gas collection has begun, but well before closure, the distinction is irrelevant. Early in the lifetime, and late in the lifetime, the issues are more important since then, the composition of landfilled waste affects the efficiency of gas capture. Different materials decay at different rates, and degrade more or less completely under landfill conditions. EIRU (1992) note that factors affecting methane production include particle size, size of lysimeters, refuse composition (including nutrient content, i.e., carbon, nitrogen, phosphorus; and pH), density, temperature, moisture, size and depth of landfill, site geology, nature of intermediate cover, nature of lining or capping, local climate, and the presence or otherwise of sludge or other methanogenic inoculum, or biomethanation inhibitors.

Materials such as paper are known to degrade less well than putrescibles under landfill conditions (see Table 28). What our model does not account for in any way is the rate of landfill gas generation. This is important in considering the economic feasibility of landfill gas collection, and consequently, the net environmental impact of landfilling. Removal of relatively inert materials from the waste stream can increase the rate of methane production, bringing forward the onset of economically feasible levels of gas generation and reducing the time period of landfill stabilisation. This means that over the life of the landfill, a greater proportion of the gas generated would be collected, and more would be available for electricity generation and conversion of CH₄ to CO₂. Work by EIRU (1992) suggested that degradability of waste changed significantly after the introduction of a recycling scheme in Stockbridge. This is shown in Table 29. The key changes are that the readily degradable fraction of residual waste landfilled increases whilst the inert fraction falls.

Table 28: Biodegradability of Waste Components

Material Biodegradability (%) ¹ Degradability Category (%)

Sources: ERL (1990), EIRU (1992) and Mosey and Mistry (1991)

Table 29: Changes In the Degradability of Household Waste Collected in Stockbridge

Source: EIRU (1992) calculated using data from WSL and Poll (1991)

In the ideal world, one would model gas generation with more dynamic profiles. The nature of waste landfilled influences the completeness of gas collection, though this is also influenced (for a specific waste fraction) by the period at which one landfills the material relative to closure. This is illustrated graphically in Figure 3 below, in which it is assumed that landfill gas collection becomes 'cost ineffective' below a certain rate (note the curves are drawn for illustrative purposes only and are not intended to be perfect representations of the post-closure situation). The volume of emitted gas is equivalent to the integral under the decay curve once the rate of generation has fallen below the cost-effectiveness cut-off.

Figure 3: Effect of Rate of Gas Generation Post-closure on Uncollected Gas Volumes

In environmental terms, the smaller is the area ABC, then other things being equal, the better will be the performance of the landfill. On the other hand, removing paper would, under the USEPA assumptions, remove a net sequester of carbon. Arguably, it then becomes important to know what alternative use is being made of the paper, but as we shall see, a clear-cut decision as to what is likely to be the 'best' option is likely to be elusive.

6.4.3 Results

The DETR estimate regarding waste composition are the subject of considerable disagreement amongst those who believe that the significance of the putrescible fraction in particular has been understated in that compositional analysis. This appears to be a common view among those who have conducted analysis of actual waste streams as opposed to conducting 'waste analyses' on the basis of what may be outdated, or simply incorrect, linkages between Acorn social groupings - themselves outdated - and waste generation.

Some information on composition can be found for London in Ecologika (1998). Like that study, work by Network Recycling in South Gloucestershire also suggests a putrescible fraction of the order 40%. We have used a composition as shown in Annex 5 which we suspect is a reasonable approximation to the composition of municipal waste. It should be pointed out. However, that there is no obvious set of statistics to use in this area. The actual typical tonne will vary across authorities.

Questions can be asked as to which values for the key variables should be used. The USEPA (1998) reports sources and commentators as suggesting that oxidation at the cap could range from 5-40%, whilst gas collection efficiency might range from 60-95%. We have accepted the former range, but the figures for the latter are relatively high. Willumsen (1997) suggests that only about 25 to 50% of the gas produced in landfills is recoverable. ETSU (1996) also suggests that collection efficiency was unlikely to be greater than 50%. We show

cases with 40% and 70% gas collection efficiency. On the efficiency of the engine, Entec (1999a) use 40% as the 'best case'. We have used a range from 25% to 40%. The total externalities are shown for differing combinations of the energy recovery and landfill gas modules (we have combined the high externality gas estimates with the high externality estimates for displaced energy, under different assumptions concerning the energy source displaced, and vice versa).

Tables 30-33 show our results for the composition of MSW we have used. What we have done is to start with the low oxidation, low gas collection efficiency and low engine efficiency scenarios and we have changed each of these, in turn, to the higher figure (generating the 4 tables). The effect of using wide ranges of external cost estimates produces results that are, unsurprisingly, rather different to those in which relatively no ranges in per unit externalities were attributed to the emissions of specific pollutants. In particular, the high externality adders highlight the significance of undertaking measures to collect gas, and to recover energy from it since this places a higher premium on the replacement of other energy sources.

Table 30: Estimates of Some of the External Costs of Landfilling a Tonne of MSW (£ per tonne MSW)

Table 31: Estimates of Some of the External Costs of Landfilling a Tonne of MSW (£ per tonne MSW)

Table 32: Estimates of Some of the External Costs of Landfilling a Tonne of MSW (£ per tonne MSW)

Table 33: Estimates of Some of the External Costs of Landfilling a Tonne of MSW (£ per tonne MSW)

The following comments seem relevant:

• It will be seen that whilst the externalities are all negative in the assumptions made in Table 30, the situation is rather different in Table 33. The landfill with no gas collection still generates negative externalities. Others are, however, under this analysis, generating positive ones. This is true in this analysis for both the high and low externality adders. The only exception is under the scenarios where no energy source is assumed to be displaced.

• Certain materials have the effect (under the USEPA assumptions) of being net sequesters of carbon. Consequently, removal of these materials, can, ironically, incur negative externalities (the implication is that they can contribute more to global warming out of the landfill than within it). Hence, waste composition will have an influence on greenhouse gas emissions as measured in this model. Note specifically that in the cases where gas collection is more efficient, the negative global warming externality associated with methane is actually less than the positive externality from sequestered carbon dioxide (so there is a net positive value associated with global warming). This is an extremely controversial result and stems from assumptions concerning how to treat biogenic and non-biogenic sources of carbon dioxide, as well as the emission factors for methane which are in the USEPA (1998) model. It should be recalled as well that we have no figures (for either carbon sequestration or methane emissions under landfill conditions) for 17% of the waste composition.

• The rates of gas collection are important for obvious reasons. The effect of gas collection efficiencies, and flaring or energy recovery is to convert the more potent GHG, methane, into a less potent one, CO₂. Investments to capture landfill gas generate net benefits which may be quite significant (irrespective of the amount of energy generation). In the high externality adder cases, the net benefits may be as much as £10 per tonne of MSW. As such, requirements in the Landfill Directive for the installation of gas collection equipment would appear to be justified.

• The externalities associated with avoided energy generation can be the largest contributing factor. Hence, the data concerning gas collection and energy recovery, not to mention the assumptions about which (if any) energy source is actually being displaced, are crucial in determining total externalities as reported here. One can see how the avoided externality increases as gas collection and engine efficiencies increase due to the higher rates of energy recovery.

• The uncollected gas that escapes may be oxidised at the cap. This will be affected by a number of factors, and the results are influenced, though not greatly, by the assumptions regarding oxidation for similar reasons to that concerning gas collection and flaring/energy recovery (methane is being converted to carbon dioxide).

• Composition, as well as factors internal to the landfill itself, will affect rates of degradation which may in turn affect the collection of gas over the landfill's lifetime. Methane gas generation is determined by composition in this model. The combination of oxidation rate, gas collection efficiency and engine efficiency, as well as the externality adders, determines the net effect of the additional methane generation (there will be a negative effect associated with methane emitted to air, but a positive benefit from all methane used to generate energy where one assumes that the recovered energy displaces a specific source).

• As discussed above, the assumption concerning the avoided energy source is crucial. A significant percentage of the 'benefits' associated with landfills recovering energy are traceable to the assumption concerning the avoided externalities of energy production. In practice, the external costs of landfills recovering energy will be quite location specific, not (in the case of the limited range of types of externality we have examined here) that of the landfill itself, but that of the displaced energy source. If the displaced (non-renewable) energy source is located in remote locations, the benefits gained from recovering energy could be assumed to be lower than if it was located in a more densely populated location.³⁸

[³⁸ Note that again, this would not necessarily be true when considering non-marginal changes, i.e. where energy from waste capacity potentially offsets construction of new plant. Here, disamenity effects are likely to be important, and are may be as significant for e.g., large-scale wind power as for gas-fired stations.]

Equally important, notwithstanding the fact that the ranges in our estimates are already large, these estimates omit a number of external costs that may well be significant. Hence, these results should not be interpreted as an accurate measurement of the external costs of landfill. Many factors have been omitted. These are:

• All the relatively fixed externalities, such as the disamenity effects (including reduction in asset prices),³⁹ the impacts associated with landfill construction and engineering, any changes in non-use values of specific sites, and possibly, any non-market benefits from recreational uses post-closure (though these might have to be considered against counterfactual land-uses).

[³⁹ For obvious reasons, there is no counterbalancing disamenity reduction at the 'source' of displaced energy (the plant is still there).]

• All impacts associated with the use of on-site vehicles.

• Leachate impacts - leachate may have significant effects owing to high biochemical oxygen demand.⁴⁰

[⁴⁰ Bez et al (1998) list 23 different emissions to water as well as emissions to air of VOCs, hydrogen sulphide, sulphur dioxide, dust, carbon monoxide and NO_x. These were from landfilling contaminated bottle fractions only.]

• Emissions of gases other than CO₂ and CH₄ (ozone depleting chemicals, such as CFCs, are believed to arise from landfills).⁴¹

[⁴¹See previous footnote.]

• A number of other impacts whose status is 'unproven' as yet, for example, the possible problems in respect of birth defects that been mentioned in the context of landfilling (mentioned in the previous chapter).

The possibility remains for heavy metals (from, for example, fluorescent tubes) to enter water courses through breaching landfill liners in the future. This is possibly one example of the 'low probability, high consequence risks' which social theorists have recently sought to come to terms with. All of these (apart from the possible benefits from non-market recreation and amenity post-closure - likely to be heavily discounted) are negative externalities. As such, the net externality is a more positive reflection of the true situation than is warranted.

It is important to recognise that these figures, where they are positive, should not be taken to imply that there are environmental benefits associated with 'the landfilling of waste used to derive energy'. This is because the benefits are entirely contingent upon what is happening elsewhere, and reflects the setting of the boundaries around the LCA (and system expansion within it). The net effects of what the landfill itself does are negative. Net benefits have been attributed on the understanding that more damaging impacts, which are occurring elsewhere, may be being avoided. Therefore, the figures should not be interpreted as, for example, a disincentive to avoid generating waste. Nor should the numbers be used to imply that those living near a landfill should bear the consequences because the landfill 'provides society with benefits.' This would be a complete misinterpretation of the results, and of the caveats which come with them. For local residents, the most important impacts are those such as disamenity that we have not even tried to measure (see Chapter 9). Residents may also have reason to be concerned about as yet unproven health effects. Benefits, such as they are attributed in this analysis, derive from the fact that worse things are happening elsewhere. This is as much a comment on the current system of energy generation as it is one about landfill energy generation per se.

In the general case, we do not have very good information on a number of key parameters in seeking to model what is going on. To re-emphasise the difficulties in arriving at a 'true' value of the external costs of landfill, we suggest that there will be disagreement about all of the following, each of which determines the external costs of landfilling as we have calculated them:

• Waste composition (varies considerably, each component +/- 50% around the mean, also seasonal).

• Methane generation by components of landfilled waste (relatively few studies done - difficult to replicate landfill conditions - we have nothing here for 17% of the waste).

• Net carbon sequestration associated with components landfilled (the comments in the previous bullet apply).

• Oxidation rate of methane at the cap (varies with a number of factors - see above).

• Efficiency of landfill gas collection (and ideally, to the extent that one is looking at effects at the margin, one might wish to understand the effects of the waste over the lifetime in the landfill, this being affected by the composition and the time at which the waste is landfilled in the context of its life) (significant disagreement - varies over the lifetime of the landfill).

• Efficiency of engine operation (likely to be better known in specific case, but still exhibiting variation);

• Emissions from displaced energy source such as one believes the assumptions to be correct. Depending upon one's assumptions, these may be changing, though for a given assumption, the data ought to be reasonably accurate at a given time. However, it is worth pointing out that individual coal plants, for example, differ hugely in their emissions of sulphur dioxide and particulates. Using these values in the coal case would have significantly altered any benefits attributable to displacing coal as an energy source.

These difficulties are merely those that exist in carrying out the calculations as we have made them. As regards finding a true value, or even a true range, these difficulties are compounded (and one's efforts are confounded) by the various omissions listed above, as well the uncertainties in placing values upon the emissions such as have been quantified. Quantifying the external costs of landfilling is no 'stroll in the park.'

For the purposes of the rest of the report, we will use external cost figures for an 'average' landfill under the assumptions we have made. This has 12.5% oxidation at the cap, 50% gas collection efficiency and 35% engine efficiency (details can be found in Table 34; if you need a copy, please contact Waste Watch). This would appear to be representative of a landfill performing well. These are indicative figures only. The range is from £-14.6 to +£18.6.

6.5 Incineration

Similar sources of uncertainty apply regarding the emissions from incinerators as apply to the emissions from other sources of pollution. These points are well made in the Spanish study undertaken in the context of the

ExternE programme. Commenting on the uncertainties involved in deriving external cost estimates, the authors state:

Several aspects should be improved, mainly the estimation of global warming damages. Atmospheric dispersion models, which, at least for the Spanish case, should account for the complex topographical conditions are also a controversial aspect. An important issue which should also be studied is the relationship between atmospheric pollution and chronic mortality. Regarding global warming damages, its range of estimated results is so broad that it dominates the results for fossil fuel cycles… Considering that chronic mortality is, by far, the major externality besides global warming damages for fossil fuel cycles, the fact that there is only one exposure-response function for its estimation, and that this function comes from the US, without being checked in Europe, adds a lot of uncertainty to the final results. … Controversy still exists around [the issue of valuation of life], and in spite of the modifications introduced in the valuation of life by the Core project, the values assigned are still contested outside the project. (Linares et al 1998)

A particular issue for the valuation of externalities associated with incineration is population density. Hence, in the context of the ExternE project, the following observation was made:

The influence of large cities is shown mainly for the waste incineration plants which are usually placed near or in large cities. This location produces very large damages, as shown especially in the French case, where particulates produce damages around 57,000 ECU/t in the Paris area. These large damages per tonne of pollutant emitted require then that emission factors are kept to the lowest so that the external costs of electricity generated by these plants are not excessive. (CIEMAT 1998).

In Paris, the external costs of MSW incineration were estimated as 52 ECU (£34) per tonne of waste excluding CO₂ emissions, and between ECU 67-92 (£44-60) when the CO₂ emissions are included (Spadaro and Rabi 1998). Most of the damage costs were attributable to nitrate and sulphate aerosols. However, these results are raw externality estimates and do not account for displaced externalities associated with the generation of electricity and other energy (which the authors suggest can roughly halve the estimates). In Italy, the ExternE Implementation study suggests that contrary to the waste management hierarchy, in the (location-specific) case studied, landfill has lower external costs associated with it than incineration. It is not clear that in either case, the avoided externalities associated with potential energy recovery were accounted for, although it is clear that landfill disamenity was accounted for through a specific hedonic pricing study. The net externality (i.e., the amount by which external costs of incineration exceed those of incineration) is given as 7.5 ECU per tonne of waste (£5 per tonne) (Crapanzano et al 1998), an almost complete reversal of the situation found in CSERGE et al (1993) in the UK.

The significance of population densities is reflected in the COMEAP recommendations (DH 1998; see also DETR 1999b; IVM et al 1998) regarding exposure-response relationships for specific pollutants. These are expressed in percentage increases in deaths and respiratory hospital admissions per incremental increase in concentration.

We have tried to model the external costs of incineration of waste in a similar way to the approach taken in the landfill module. In other words, transport is excluded, and we have attempted to model the processes such that the composition of the waste stream is incorporated. Again, we have discounted CO₂ deemed to be from biogenic sources. However, we have included those which are non-biogenic in origin (e.g. from plastics). The accuracy of this assumption might have to be questioned if plastics derived from, for example, genetically engineered crops becomes common, in which case, the analysis of benefits associated with recycling would have to be measured with reference to the avoided external costs associated with such a production process (as opposed to more familiar ones). Hence, although the performance of incineration of plastics vis a vis landfill might improve, it is not clear how this would affect the situation vis a vis recycling (see below).

The question of what emissions arise from incineration is obviously central to this part of the analysis. A number of points need to be made here:

• In any plant, the emissions are likely to vary over time. Hence, limit values on plant tend to specify the period over which the measurements must be taken. This makes it somewhat difficult to understand the value that one ought to use in any attempt to evaluate the external costs of such plant. This is especially true to the extent that certain effects might be triggered by threshold values, the exceeding of which might be obscured when average values are taken. This limitation to the analysis applies even with more complex approaches to modelling than the 'externality adder' one taken here.

• There is, in any case, some variation in reported emissions from incineration plant. This will be partly due to the fact that the plants themselves are different, because they use differing technologies to address emissions from the flue gas, and because the wastes they receive may be different too. These may, in turn, lead to different amounts of specific pollutants in the emissions to different media (e.g. wet scrubbers are likely to lead to more emissions of chlorine in the form of effluent than in the form of solid waste, the latter being more likely where dry lime injection is used). Not just the level of emissions, but also the media to which they are discharged, varies with the technology used.

• As with landfill gas emissions, the emissions from incineration are dependent upon the material combusted. We have less good information here in respect of links between materials and micro-pollutants. However, some work has been done on the effects of removing dry recyclables and compostables from the waste stream (Entec 1999a; Atkinson et al 1996). This shows that the calorific value of the remaining waste can be increased when such schemes are in operation, increasing the efficiency of the energy recovery process. Clearly, the removal of organics (because of moisture content), metals and glass (because both effectively absorb heat) increase the calorific value of the remaining material.

• Recent work by Entec (1999b) suggests that existing MSW incinerators do not meet all the standards likely to become law under the Incineration Directive. Arguably, once the Directive becomes law, to the extent that enforcement is effective, emissions will fall in line with what is required by the Directive.

We have taken an approach, which may be controversial, in which we have used high and low values for all emissions. Annex 6 explains the choice of emissions levels for the different pollutants. For CO₂ we have used the USEPA (1999) figures. Values for other pollutants come from comparing various sources including CSERGE et al (1993), Carroll (1995), Environment Agency guidance (1996) and the values for Tyseley in Entec (1999b).

As regards calorific values for the various fractions of MSW, we initially used two sets of data which, though they are slightly different, are broadly consistent. These were from Atkinson et al (1996) and from USEPA (1998) (yet another slightly different set can be found in the US EPA's MSW factbook). With typical compositions of waste, the energy content of a tonne of MSW are within the ranges typically quoted (usually between 9 and 10.5MJ per tonne of waste). Hence, plugging in the efficiencies of energy recovery used by Entec (1999a), the output energy is suitably close to the values derived in Entec (1999a). Because the externality analysis as carried out here is broadly unaffected by the different calorific values used, we have chosen the Atkinson et al (1996) values (see Annex 7).

The externalities we have valued are only those related to air pollution. For these, we have again used high and low values. Included amongst these are some heavy metals and dioxins, but we have no information on HCl and HF. It is well known that emissions of the former are associated with the presence of PVC (amongst other things) in the waste stream, and that for these reasons, there is some merit in pre-sorting waste to extract this fraction. The high and low values for some of the micro-pollutants vary enormously. One reason for this is that there is no unanimous agreement on the existence of thresholds, let alone where any threshold effect might lie. Furthermore, the pathways through which receptors, particularly humans, are exposed to these micro-pollutants are not so 'straightforward' as with the direct inhalation of gaseous emissions. This is a more honest (if not correct)⁴² way (using high and low values) to deal with uncertainty and it is, in our view, surprising that such large ranges do not appear more often in the valuation literature (especially, for example, with respect to global warming, where the extent of warming, let alone the impacts of this, is uncertain).

[⁴² Arguably, the most 'correct' way of dealing with the uncertainty is to accept that it will not go away, and to acknowledge therefore that presentation of any figures is simply disingenuous.]

Given their omission in the externality analysis, it is worth pointing out that in addition to the gaseous emissions from incinerators, about 30% by weight of the original waste arises as 'bottom' or grate ash (i.e. ash and unburned residues from glass, masonry, ceramics, metals etc.). This is typically quenched in a water bath and may subsequently be used as a material in construction applications (see below). Fly ash, on the other hand, arises through the process of controlling stack emissions of air and dust and contains materials which are far from inert. The ash is mainly silica or alumina enriched in heavy metals and organic products such as dioxins. The particles also act as condensation nuclei for volatile matter. AEA (1997) cite two studies looking at the problems associated with leachability of chlorine and heavy metals from the two types of ash.

In the AEA (1997) study, six possible approaches to pollution control were considered, each giving different emissions of trace pollutants to air, water and land. The study elaborated upon the differences in emissions to the different media across the options but made no attempt to value the external costs involved (which the same study did carry out for air pollutants). With respect to leachate, the study fell back on the work undertaken by CSERGE et al (1993) discussed above, in which clean-up costs were used as proxies, and in which an assumption was made that leachate was unlikely to occur in the near future so pollution arising could be heavily discounted. As the AEA (1997) report elaborates, there are a number of problems with this approach.

Possibly partly as a reflection of the fact that more is known about the effects of air emissions on health, and more valuation work has been done in this area, the focus of cleaner incineration technologies has been the flue gas. Cleaner technologies may, in part, involve changes in the emissions of pollutants themselves. However, a number of approaches simply result in the removal of pollutants from the flue gas for disposal to other media (land or water depending upon the mechanism). Hence, as long as the emissions relating to discharge to water and land are essentially ignored, the net effect on the 'bottom line' figure for the total externality of shifting pollutants from air to land is equivalent to the pollutant having disappeared, even though the net effect has been to shift it from one medium to another (and where it is disposed to landfill, then also from one generation to another).

Note that neither the CSERGE et al (1993) work nor (apparently) that of Coopers and Lybrand et al (1996) considered materials recovered in the incineration process. Both steel and aluminium are extracted, now usually post-incineration. This means that there may be an additional environmental benefit from the recovery of materials (depending upon the net externalities associated with recovering metals through this route). The quantities recovered in the Netherlands for 1996 were 33% of all non-ferrous metals and 60% of all ferrous metals (taken as steel). This turns out to be broadly consistent with the figures supplied by the Energy from Waste Association in their response to the Draft Strategy for England and Wales:

'EfW plants recover both ferrous metals (3-5% of total by weight) and non-ferrous metal (0.5 to 1% by weight -mainly aluminium.) During 1998, the EfW sector is understood to have represented the largest single contributor to UK ferrous metal recovery from MSW -in the order of 75,000 tonnes were sent to British Steel for reprocessing.'(EfWA 1999).

We use 50% of steel (which would generate 3% of total from our composition figures) and the Netherlands figure for aluminium (33%, which is effectively within the 0.5-1% range using our composition figures). Note that the financial benefits from this recovery are less than that associated with materials recovered pre-incineration. This is because the quality of the materials recovered is much lower (owing to contamination from the incineration process), so that whilst materials may be recovered in significant quantities, the quality imposes constraints upon its use.⁴³ At the time of writing, loose steel scrap sells for £8-13 whilst steel from incinerator ash fetches £0 per tonne (from Materials Recycling Weekly). Part of the incentive to make use of this material arises from the fact that use of the material enables the issuing of PRNs.

[⁴³ This will be especially true for aluminium where the desirability of closed loop processes stems from the fact that specific alloys are used for specific purposes. Lack of knowledge concerning the alloy content is likely to reduce the value of the metal considerably (ECOTEC 1999).]

We have added an environmental benefit which is attributed on the basis of an assumption that this material is recycled and that it displaces primary material (see analysis below). This is a controversial assumption since we do not know about the external costs associated with the processes of extracting the materials (magnetic extraction for steel, eddy currents for other metals), and then cleaning and them (we effectively assume the materials recovered are used in the same way as metals recovered from kerbside collection). Including these would reduce the estimated benefit associated with the materials recovery (and though one suspects there may still be a net gain here, further analysis would be required to confirm this, especially given the lower quality of the material extracted).

Energy from waste incineration plants are increasingly seeking to make use of their bottom ash, often displacing primary aggregate consumption. This is not happening at all plants at present, but there are construction projects making use of bottom ash, and supposing that this practice becomes more widespread in the future, one might expect an additional external benefit associated with displaced aggregates extraction.

It would be wrong, in our view, to simply multiply the mass of bottom ash by the estimated external costs of aggregates extraction which were quoted in the work on the aggregates tax carried out by London Economics. As discussed both in ECOTEC (1998) and EFTEC (1999), this estimate is composed of both variable and fixed elements. EFTEC (1999) estimates the variable component of the total as approximately 55%, or 18p per tonne for hard rock outside national parks, £5.79 for hard rock inside national parks, or £1.08 for sand and gravel. Note that on the basis of London Economics (1999) work, less than 5% of UK aggregates come from quarries located in National Parks (and this may fall over time owing to agreements in which operators are engaged).

This would imply that for each tonne of municipal waste, then assuming 0.25t of bottom ash used, a net external cost saving of the order 33p might be made.⁴⁴

[⁴⁴ This is calculated as a quarter of the weighted average (by production) externality from the three possible sources assuming no preference for any specific source / type of material. Also, it assumes that bottom ash replaces aggregate on a 'tonne for tonne' basis.]

Note this does no account for any differential transport externalities in transport costs which may arise when one switches from aggregates to bottom ash. Note also that other materials now competing in this market are recycled construction materials. To the extent that one might, at the margin, be replacing secondary aggregates, any additional benefit could (and this is arguable) be reduced to the equivalent of the avoided variable externality associated with secondary aggregates production. Lastly, note that we have not accounted for externalities arising from the removal of contaminants (some of which effectively involves the removal of metals discussed above). Also, heavy metals can be leachable so that in the absence of utilising chemical stabilising agents (at a cost), there may be longer-term effects from the use of bottom ash as substitute for aggregates. These considerations suggest that whilst our analysis suggests a small net benefit, the reality (i.e., if one were able to account for all these impacts) may be rather different.

6.5.1 Results

The results of the analysis are shown in Table 35. Options 1 and 2 refer to the use of different calorific values for materials concerned (see Annex 7). These do not affect the analysis tremendously, but their effect is noticeable (as one would expect) where the higher externality adders are used. The following points are worthy of note:

• First of all, under these levels of materials recovery and efficiency of energy generation, the externalities are positive where one assumes displaced energy source as either coal or the average fuel mix. The benefits associated with recovery of steel and energy offset the major negative externalities of air pollution. This yet again highlights the significance of the 'displaced energy source' assumption, as well as the recovery of materials. We re-iterate our caveat here that we may be 'over-attributing' the positive externality associated with metals recovery from incineration.

• Again, when one uses the low externality adders, the balance of costs and benefits inevitably leads to relatively low totals for the externality as measured here. However, the externality is still positive (there is a net benefit), even (in this case) under the 'no energy displacement' assumption.

• When the high externality adders are used, the figures are completely beyond what we are used to seeing in this form of analysis. This reflects the fact that we are entertaining, effectively, the possibility that the effect of pollutants may be more severe than has typically been implied. Under the high adders scenario, the benefits are much larger than for landfill because of the greater energy generation.

• Some air pollutants, notably particulates, arsenic, sulphur dioxide and NOx, generate significant disbenefits under the high externality adder assumptions. It is these pollutants to which the results are most sensitive to variation (and the same applies to the avoided energy source). This is very important since the first three of these produce localised disbenefits, and under the high externality adders assumptions, they are very large indeed. NOx effects are felt more widely. It is for reasons associated with these types of pollutant that citizens are reluctant to accept incinerators in their vicinity. Our analysis indicates the possible magnitude of these (though the possible range of these is enormous).

• The effects of aggregates recycling are not greatly significant.

Table 35: Externalities Associated with Incineration (£ Per Tonne MSW)

Note that we have only looked at cases where all the externalities are treated with the same set of externality adders. In the cases of the pollutants that cause disbenefits which are not global ones, there would be reason to expect unit damage costs to vary with the location of the incineration facility and that of the displaced energy source (principally, but not only, because of differing population densities). In this case, the unit damage costs for global pollutants would be the same in each case, but the figures for the pollutants with more localised effects would be different between the facilities (depending upon localised population densities).

It is interesting to see what happens when one makes assumptions concerning the variation in externalities with the hypothesised location of the facilities. At the extreme, one could posit an incinerator in a densely populated urban area, and a displaced source in a sparsely populated area. We do this to show the extent to which the outcome can vary with location / externality adders used. The way this is done is through using:

• for global pollutants, the high externality adders in both cases; and

• for local pollutants, the high externality adders in the incinerator case, and the low externality adders for the avoided energy source. The results are as shown in Table 36.

Table 36: Externalities Associated with Incineration with Low Adders for Local Pollution Associated with Displaced Source (£ Per Tonne MSW)

As one would expect, the situation changes such that in the average fuel mix case, the total externality begins to approach the 'no energy displacement' scenario (because the benefits from displaced energy are less). The net benefit in the 'replacing coal' case has fallen significantly. Relatively minor changes in the rate of recovery rates of steel and aluminium would make the figure negative (for steel, a fall from 50% to 42.6%, and for aluminium, a fall from 33% to 22.6%). For similar reasons, so would changes in the composition of waste entering the incinerator or changes in the benefit attributable to metals recovered from the incinerator. This illustrates how changing specific assumptions can give rise to quite different interpretations of the situation. It suggests, furthermore, the location specific nature of many of the impacts we are seeking to elicit (and the problems which are inherent in any attempt to make use of this type of analysis for the design of policies designed to be applied nation-wide). Seeking to base policy on externality analysis alone lets the cat of 'location specificity' out of a bag into which it is reluctant to return (and not just in the case of waste management).

It should be re-stated that this is a far from complete analysis. The following impacts have not been covered:

• all emissions to land (including disposal of fly ash, and bottom ash when not used as replacement for aggregates) or water;⁴⁵

[⁴⁵ Kremer et al (1998) list several waste materials, emissions to air not covered by us, and other residues arising from incineration of municipal waste.]

• some air emissions for which no externality adders were available;

• fuel use associated with on-site vehicles;

• impacts associated with extracting and cleaning recovered materials, and transporting them to reprocessors;

• any impacts associated with replacing materials traditionally used with materials recovered from the plant;

• extraction of primary resources (such as lime used in cleaning flue gas, and water); and

• disamenity effects associated with the siting of facilities (including reduced house prices);

As with the landfill case, we do not have very good information on a number of key parameters in seeking to model what is going on. Also as in the landfill case, all the unquantified externalities are negative ones. Hence, the net figure is not an accurate reflection of the true situation, which would ideally incorporate the negative externalities mentioned. To re-emphasise the difficulties in arriving at a 'true' value of the external costs of incineration, we suggest that there will be disagreement about all of the following, each of which determines the external costs of incineration as we have calculated them:

• Waste composition (varies considerably, each component +/- 50% around the mean, also seasonal).

• An exact computation of the links between waste components and emissions to different media. USEPA (1998) data were used for CO2 and N2O emissions. However, in the general case, a number of factors will affect emissions from incinerators (inputs by composition, but also by quantity, depending on how the incinerator has been specified).

• The relevance or otherwise of less frequent exposures to higher emissions of specific pollutants in determining ultimate effects upon which externality calculations are based.

• Efficiency of energy recovery (likely to be known for certain conditions in a specific case, but still exhibiting variation across plants and according to, e.g., completeness of combustion).

• Emissions from displaced energy source such as one believes the assumptions to be correct (depending upon one's assumptions, these may be changing, though for a given assumption, the data ought to be reasonably accurate).

Again as with our landfill, these difficulties are merely those that exist in carrying out the calculations as we have made them. Finding a true value, or even a true range, is made very difficult indeed by the various omissions listed above, as well the uncertainties in placing values upon the emissions such as have been quantified.

Therefore, as in the landfill case, extreme caution is urged in using not just these, but other results that have been generated in this field. Variation can be enormous. The same comments as made in the landfill case regarding any net benefits can be made here. In considering new facilities, to the extent that one accepts that the potential magnitude of the externalities are plausible (and scientific uncertainty, along with the fact that incinerators tend to be located in relatively densely populated areas suggests that these are difficult to downplay), the fact that the 'other air pollutants' may generate external costs varying between £1.50 to £127 per tonne of MSW suggest that a 250,000 tonnes per annum plant may generate 'non-global' external costs between £345,000 and £32 million per annum. Faced with such a scenario, local opposition to incineration plants is perhaps more easy to understand. Entec (1999a), in work for DETR, calculates figures for 'deaths not brought forward' as a consequence of implementing new operating standards for MSW incinerators under the proposed Incineration Directive. The figures for deaths not brought forward owing to ozone-related effects (via NOx) and for SO₂ were 3.8 and 1.1 respectively.

Note that this shows the importance of disaggregating the contributing components of the overall external costs of a given waste treatment facility. For local residents, these effects are likely to be of greater concern than the attributed benefits, which effectively arise (as mentioned above) because one is assuming that one is preventing something even worse from happening elsewhere. This illustrates how hypothesised benefits arise through the re-distribution of external costs associated with energy production (and the same is true of landfill with energy recovery, whilst similar issues apply in respect of recycling).

6.6 Recycling

In this part of the work, we take the same approach as we adopted for the work for DETR (ECOTEC 1999). The work carried out then was conducted by CSERGE and we have leant on that work in considering the implications of extracting materials from the waste stream. In essence, that work concentrated on the net external costs and benefits of activities involved in recycling relative to the external costs and we adopt the same approach here. The assumptions behind the inventory analysis are laid out in that report. The key differences in the approach are that:

• Since we are only considering kerbside recycling, we have adapted the approach so that there is no contribution from bring schemes; and

• We have left in the model the assumptions in respect of material separation, although an obvious point to be made is that many kerbside schemes have made MRFs, clean or dirty, redundant -vehicles with separate compartments can take their place, particularly where plastics are not part of the scheme. We conducted some analysis where we tested the effects of removing the separation energy from the analysis. We found that this had a negligible effect on externalities under both high and low externality adder scenarios. The principle reasons why MRFs are omitted in some schemes would appear to be financial and logistical.

The reader is referred to the original work for the key inventory assumptions. We had hoped to make use of new work on inventories. We have only been able to do this for steel (see below).

The Environment Agency has done work on composting within its considerable programme of LCA research and the USEPA is in the process of completing a major LCA study which also looks at compost. This work, as well as other inventory data used in the compilation of WISARD, should be published in the near future. We understand from Corus (the new name for the merged British Steel and Hoogovens companies, which now has interest both in aluminium and steel) that there are moves afoot to carry out similarly detailed work on aluminium (the most recent work being that conducted by the European Aluminium Association on primary production but we are not aware of work on secondary production). The European Commission, in the context of revisions to the Packaging Directive, is reviewing the environmental impacts of different means of reprocessing and recovery of plastics, and this will almost certainly include LCA-type approaches. The Fraunhofer Institute has already done work of relevance in this regard (see Heyde and Kremer 1999). Regarding paper, the BNMA is funding an LCA-type study on newsprint manufacture. As regards glass, this seems to be the material where relatively little work of an LCA nature has been conducted. Suffice to say that the inventory data is open to question. We welcome any inputs and comments which would serve to update the analysis. We ourselves are now contracted to the Commission (DG Environment) to carry out work on composting.

Note that all these approaches assume one is substituting like materials for like. As we pointed out in our earlier work (ECOTEC 1999), this assumption breaks down once market development for secondary materials outlets leads secondary materials of one kind to substitute primary materials of another. However, if one were to extend the analysis to account for this, the analysis becomes increasingly unmanageable.

On steel, the International Iron and Steel Institute (IISI) has carried out a major study of steel making processes and we have been given access to some of this information (see below).⁴⁶

[⁴⁶ We are extremely grateful to those at the Swinden Technology Centre of Corus, especially Louis Brimacombe, who enabled us to make use of the IISI work.]

The way we have treated this information is as follows:

• We assume that under the 'no recycling' scenario, the steel is landfilled or recovered/landfilled at an energy from waste plant. New steel is produced in a basic oxygen furnace (BOF).

• Under the recycling scenario, the steel is assumed to go to an electric arc furnace (EAF).

This is slightly questionable since in practice, the recovered metal could go to the BOF plant. BOF plants effectively involve two stages in the production of steel, the first involving the melting of (primary) iron from the ore, the second, in which the scrap is added to the furnace to make steel. It is this first stage that accounts for most of the energy used in making steel through the BOF route. In this case, one could compare:

• A secondary route, to which one allocates a fraction x/y of the emissions from the 'iron and scrap' smelting process, where x is the amount of secondary material and y is the total amount of iron and scrap (primary and secondary) material, leading to the production of z tonnes of steel.

• A primary material route, to which one allocates a fraction x/y of the emissions from the iron melting process, and a fraction x of the emissions from the 'iron and scrap' smelting process used to create one unit of molten iron.

In this case, the emissions from the production of steel through primary material, and an equivalent amount of steel through secondary processing are made comparable.

Arguably, it is the former result that may be more likely in the current UK situation where steel scrap is exported, probably to mainly EAF plant (in, e.g., Spain). However, one ought to incorporate associated transport emissions in this case (and we have not done this). In any case, the second approach is made rather more difficult to handle since the information from IISI comes in the form of averages across a number of plants (with maximum and minimum values from these) of emissions from different processes. Because consistent data across the same the plants are not available, the effect of subtracting the liquid iron emissions from the Gross BOF emissions sometimes generates results which suggest negative emissions from a process where this cannot be the case. Hence, we have resorted to the more straightforward approach through method one.

6.6.1 Results

The results showing the differential external costs of using secondary materials as opposed to primary are shown in Table 37. These appear different to earlier results, including our own (ECOTEC 1999). In particular, the equivocation concerning paper and plastics appears to have disappeared. The reasons for this are that a) we have taken the transport analysis out of the analysis and treated it separately (see above), and b) the range of external costs used has changed. The other major change is that the steel external costs are calculated using the IISI data (these are not so different from those calculated using the other inventory). As with landfill and incineration, there are no valuations of emissions to land and water.

Table 37: Differential External Costs Associated with Use of Secondary Materials as Opposed to Primary Ones

Note that we have not included any costings for the time that householders might spend separating wastes and cleaning them. Depending upon the assumptions made, these can be significant factors in determining the viability of source separation schemes. For example, a recent Swedish study has been critical of Swedish policy in respect of recycling (ENDS Daily 1999). A key reason for this is that the study accounted for the time spent by householders in separating materials on the basis that these should be costed at prevailing wage rates.

There are two reasons why one might question the assumption. The first would be that on basic economic grounds, the suggestion that individuals place equal values on their leisure and work time would appear to imply an assumption that they are able to choose freely the times at which they work, and that their wage rates are determined on an hourly basis. It is not clear that this is always the case. It may be that leisure time is valued in excess of wage rates, but equally, it may be that certain activities are 'discounted' from such a calculus on the basis that they are things that the person engaging in the activity 'should do' anyway. RPA and Metroeconomica (1999) cite a report by Markandya (1998) which valued non-working time at 15% of the gross wage rate (though the basis for the figure is not made clear in the context).

This leads neatly onto the second point which is basically one which follows from a more institutionally informed perspective. One might reasonably ask, where possibilities exist to make use of certain materials, why prevailing rights structures should allow citizens the freedom to dispose of materials without giving any thought to source separation? Indeed, some countries have, through legislation, introduced sanctions (or at least, the threat of them) to ensure that source separation routinely occurs. This is tantamount to altering the rights structure facing citizens so that it becomes a duty of citizens to source separate waste materials. This is an entirely defensible position, irrespective of whether materials are used or not, since a) it makes options available which otherwise would not be, and b) through changing the rights structure, what is defined as the acceptable norm is transformed. Elsewhere, such formal sanctions may not be necessary as norms of behaviour change, in which case, the same effect can occur through the medium of informal institutional changes. Under either circumstance, the fact that separating wastes can become a duty (dependent upon the rights structure) makes it awkward to impute a labour cost element for the activity. At the same time, those designing recycling schemes (or for that matter any scheme which seeks to elicit public participation, for example, responsible handling of litter) must make the process easy for the public to participate in.

Other omissions and limitations in this analysis are:

• The reliance on one set of estimate for the differential impacts of secondary materials reprocessing relative to primary materials processing. The choice of primary and secondary materials plants is obviously important in this respect. Ideally, one chooses (for the analysis we are involved in) the secondary materials reprocessing plants to which materials are sent, and the plants whose output is being, at the margin, displaced. This would involve significant work.

• Whilst in systems in which the secondary material replaces primary material of the same type, the issue of what is displacing what appears to some extent more straightforward than in the energy case, as soon as markets for recycled materials become more diverse, the problem becomes much more complex (since the materials are not being substituted in a like for like process).

• Also, in the same way as we looked at possible variation in the externality adders of incinerators vis a vis displaced energy sources above, similar variation with location could be expected between primary materials plants and those dealing with secondary materials.

• The omission, again, of all externalities associated with emissions to land and water.

• The omission, again, of disamenity impacts associated with either primary or secondary materials processing and reprocessing infrastructure.

• The lack of attempts to capture the external costs of primary materials extraction and transport.

An attempt to understand transport impacts of primary materials is made below. We also make an attempt to illustrate (if not to quantify - this would be extraordinarily difficult and would require location specific work) the potential impacts of materials extraction.

6.7 Extraction Phase

Relatively few analyses have accounted satisfactorily for the extraction of raw materials within LCAs carried out thus far. By extraction, we mean a broader concept of the stages leading to the input of primary material into the manufacturing process. There are three key phases that one ought to be considering in this type of analysis:

• The setting up of the infrastructure for the extraction (which may include roads to facilitate access), and in which we might include aspects of site-related disamenity (on the basis that such externalities are not, in general, variable in the sense of their varying in direct proportion to output). This would also, in a complete analysis, consider the issue of counterfactual land uses. Such questions would be especially important where one was considering the use of timber for paper production. Whilst this may be sustainable once in operation, the use to which the land was put before might have had considerably greater value as a biodiverse habitat. It is known that some paper and pulp is manufactured (not necessarily in the UK) from fibre coming from 'sustainably managed plantations' which were once biodiverse tropical, or temperate forests. Such 'losses' might also have to consider the differential impacts on fluxes of emissions and nutrients, and more controversially, the direct effects on climatic change itself through altered rates of evapotranspiration where plantations exist over large areas.

• The extraction processes themselves, most of which will be related to output. Some of these are related to the so called hidden material burdens discussed below.

• The transport of the materials (presumably, but not necessarily, overland) in the country / continent concerned from the point of extraction to the point of evacuation (the port). Depending upon the country concerned, the externalities associated with this transport phase will be more or less well internalised by duties and levies on fuel, road use etc. The degree to which this occurs may increasingly relate to targets agreed in multilateral negotiations (such as those laid down in the Kyoto Protocol) in respect of emissions.

• The transport of the materials from the port to the UK port. Since these occur in international waters, they do not contribute to any nation's specific targets in respect of emissions. As such, the level of internalisation tends to be low.

• The transport, within the UK, from the port of extraction to the point of processing. This transport will also incur external costs. These will depend upon the distance travelled and mode of transport. They may be more or less well internalised by existing duties and levies.

As regards a) and b), we could account for each of these through:

1) Average environmental effects of UK extraction / harvesting where this occurs

The rationale for this approach might be that one has no reason to believe that impacts are more or less great outside the UK than within it, or that, for one or other reason, the quality of UK based production is lower (or its price is higher) than that of competing countries. However, there are good reasons to believe that UK-based recycling would be more likely to substitute for imports than to substitute for domestic raw materials production. Furthermore, there are good reasons to believe that the environmental costs associated with extractive activity do vary across (and within) countries and that any UK based assessment would misrepresent the net effect of the recycling. Evidently, in cases where all, or most of consumption is from domestically sourced raw material, and where UK-based impacts are more or less uniform, the approach increases in validity;

2) Weighted average (on the basis of consumption) environmental effects of extraction / harvesting in countries exporting to the UK

This might be deemed appropriate if one felt that the effect of recycling would be to reduce imports, but with no obvious location for that reduction in use to occur. Evidently, a difficulty of this approach is that, to the extent that extraction-related externalities are location specific, the analysis has to be repeated a number of times in a number (arguably all) locations from which imported material originates;

3) Weighted average (on the basis of consumption) environmental effects of extraction / harvesting in countries exporting to the UK as well as the UK itself

This is similar to the above, but applicable if one believed that there was no reason to believe that domestic production, as opposed to exports, would not be substituted for, at the margin, as recycling increased. This is, in effect, how external benefits associated with the replacement of energy from current fuel mixes with that derived from energy from waste are calculated at present;

4) Weighted average (on the basis of global consumption) of environmental effects of extraction / harvesting in all producing countries

From a more macro perspective, to the extent that the markets for the raw materials under consideration are globally integrated - and in general they are - the issue is not so much that one country from which UK consumption is derived will see its market reduced, but that the global market is reduced, leading to a reduction in extraction somewhere in the world.

5) Environmental Effects of extraction in 'marginal producing country'

If it were possible to highlight marginal producers for specific materials, and it would require a study of some detail to do this if it could be done at all (other than in an ex post fashion on the basis of empirical information), one could assess avoided externalities on the basis of the marginal extraction costs.

Note that a limitation of any of these approaches is the assumption that displacement occurs in a closed loop context. This is likely to be more significant for those recyclables competing with low cost raw materials than for those, such as aluminium and steel, which have relatively secure outlets as a result of the value of the secondary material. In respect of the above approaches, there will be associated transport externalities (c, d and e) to be considered. These will vary with one's preference regarding methodological approach.

A full treatment (not to mention exploration of the issues) is beyond the scope of this study. What follows below is a very limited attempt to shed some light on the nature and/or ranges of externalities that might be associated with a), b) and d) above. We believe this is important since arguably, it is these externalities which appeal immediately to the minds of those who undertake recycling. What are we saving in terms of the extraction processes by recycling materials as opposed to treating them in linear treatment modes?

6.7.1 Hidden material flows and total material requirements

The extraction of virgin materials requires the movement and mobilisation of matter that is incidental to the recovery of the economically valuable product. Often these incidental flows of matter can be of tremendous environmental significance. They can disturb natural habitats, result in the death of non-target species, mobilise heavy metals into the water system and in the case of mining activities release greenhouse gases. Such impacts are frequently excluded from conventional life cycle analysis and environmental assessment work, because they are difficult to quantify and do not always vary linearly with the amount of material extracted. They are of significance to this project because the substitution of recycled material for virgin material causes a reduction in the amount of virgin material extracted for every tonne of product that is used by the household. As a result a rise in the proportion of municipal solid waste recycled will tend to decrease the amount of hidden material flows caused by household consumption.

The types of perturbations that make up 'hidden material flows' include disruption to the land surface from the excavation during mining or forestry, soil erosion due to the reduction in vegetation cover, lifting of soil / stone during the extraction of ores. Using the terminology used by the Wuppertal Institute (WRI, 1997) these impacts can be broken down into the following sorts:

·         Ancillary material flow

·         Excavated and/or disturbed material flow

·         Hidden material flows

·         Direct material input

·         Total material requirement

Ancillary material flow is the matter bound to the material of economic value that is extracted alongside the material and removed from the environment. It is released from the material during the first stage processing of the material. Often it is chemically and physically altered during the separation process. Examples of ancillary material include the components of a metal ore is that is discarded after the pure element has been refined, or the bark and brash from trees that are removed from the environment.

Excavated material flows are the matter that is physically displaced from the extraction process but is not transported away from the site of extraction. For instance, in an open cast mine, topsoil and earth are lifted from the excavation site to reveal the ore-bearing seam. Excavated material flows also include soil erosion arising from the loosening of soil structure caused by digging and clearance of vegetation.

Hidden material flows comprise the summation of ancillary and excavated material flows. They are all the non-economic flows of material that arise from the extraction of valuable products.

Direct material inputs are the materials that are economically important materials recovered from the extraction, forestry, fisheries and agricultural activities (the last two not being relevant to this study). These include the matter that is produced domestically and also the matter that is imported (less exports).

Total material requirement is the summation of the hidden and direct material inputs and therefore comprises the total materials that are mobilised by an economy.

By convention total material requirement analysis measures all material flows in terms of their total mass. The Table shows an estimate of the total material requirements including the hidden flows of ancillary material and excavated material that are mobilised during the extraction process. Note that the term direct material is taken to mean the material that is actually traded within the economy prior to its processing into a finished good. In the case of paper this would be the timber that is sold to the pulp mills. The finished good is the printed material itself.

Table 38 below excludes the hidden flows associated with the fossil fuels and minor materials that go into the finished goods. These may not be insignificant. For example, as a reminder of the scale of the use of these other inputs, the Table shows us that a ton or steel requires inputs of a 0.92 tons of coal, fuel oil and limestone. The production of aluminium requires even greater tons of oil equivalent of electricity.

Table 38: Total Material Requirement for Materials Found in Household Waste (these are essentially expressed as ratios of the weight of 'hidden flow' material to the weight of finished good)

Notes: total material requirements quoted from World Resources Institute (1997)

(1) USA total material requirement

(2) German total material requirement

(3) Coal, limestone and fuel oil consumption for UK steel production reported in British Steel Environment Report 1996

(4) USA total material requirement (5) Wood Raw Material Equivalent data from UK Forestry Commission, unpublished

6.7.2 Emissions during the extraction of raw materials

This section calculates the amount of atmospheric emissions caused during the extraction and harvesting of raw materials from the environment in the UK. Data are used from the UK environmental accounts. Data are only given for oil, non-metals minerals (sand) and forest products since bauxite and iron ore are not mined in the UK.

The data shown below were calculated using the UK environmental accounts (Vaze and Balchin, 1996) which give data on air emissions from all industries in the UK. This was divided by the total domestic production of oil, non-metal minerals and timber in the UK. The environmental accounts unlike life cycle analysis do not allow one to see the emissions caused by the production of individual products but rather of industrial sectors. For instance the oil sector conflates emissions from oil and gas production. Similarly, emissions from the minerals sector include not just sand but emissions from all other minerals (gravel, hard stone etc). For the sake of this analysis it has been assumed that emissions from the oil industry can be split between oil and gas production on the energy content of the two fuels. It is also assumed that emissions from all non-metal minerals can be split on a mass basis. These are clearly fairly gross simplifications but they provide some insight into the impacts associated with raw materials extraction.

Table 39 shows the emissions from the extraction of the virgin materials from the environment within the UK per tonne of economic product. The product of these emissions and our externality estimates give a range of estimates for the external cots associated with virgin materials extraction.

Table 39: Air Emissions from Selected Industries (g/tonne of output)

Applying valuation factors to these leads to the following results (Table 40), again using high and low values for the externality adders (ammonia was not valued as no suitable factors could be found).⁴⁷ These show that potentially, the avoided emissions from the extractive phase are significant. Perhaps as expected, they are not so great in the case of minerals as they are for forestry and oil. To the extent that these are omitted in some LCAs, these extraction-related externalities are seen to be potentially important.

[⁴⁷ We have followed Pearce and Crowards (1995) in assuming equivalence between black smoke and PM₁₀. This is a slightly controversial assumption (see the same paper for a discussion).]

Table 40: Externalities Associated with Air Emissions from Extraction of Materials (£/Tonne Output)

6.7.3 Emissions during the movement of raw materials

The raw materials that go to make up steel, aluminium, plastics and paper usually originate from outside the UK. The distances travelled by secondary materials are usually much smaller and typically remain within the EU if not within the UK. These are accounted for (albeit in some cases through approximation) in our study. There are two points worth bringing up here. Firstly, the greater distances travelled by virgin materials mean that the environmental impacts of their movement are typically much greater, although the mode of transport (and the bulk of material moved) is an important consideration. Secondly, secondary materials are largely transported within the UK by road. The price of road diesel is high compared to the underlying price of the oil. About 75% of the price is road fuel taxes and VAT. One could argue that these effectively internalise the environmental costs of road transport. The same cannot be said for international transport fuels. Marine bunker fuels are not subject to fuel taxes, and because emissions in international waters are not attributed to any nations there has been little pressure for international shipping to reduce the amount of emissions into the atmosphere. Marine bunker fuels often have a high sulphur content and are responsible for significant emissions of acid rain precursors.

Table 41 below gives the country of the five biggest sources of iron ore and newsprint for the UK in 1993 (Vaze, Schweisguth and Barron 1998). The numbers are the percentage of UK imports from each country, and these are given with figures for transport distances by sea (to Southampton). Similar data could be drawn up for bauxite (aluminium). With regard to plastics, though the UK is self sufficient in crude oil, large volumes of oil are still imported and exported in order that refineries have the appropriate grade of oil to produce the products required by the local markets. The raw materials for glass production are not widely traded on a significant scale.

Table 41: Iron Ore Imports by Country of Origin, And Associated Distance Moved

 As the Table shows, iron ore and newsprint are transported thousands of miles from their site of extraction to the UK. The most usual means of transport is by container ship. Sea freight, especially in the larger ships, requires relatively little energy for movement. ETSU (1996) calculate that 8.3 kg of marine bunker fuel are required to transport a tonne of oil from the Middle East. However as mentioned before emissions standards in ships are crude compared to road transport and the emissions factors much higher.

ETSU (1997) calculate the emissions from shipping a tonne of oil from the Middle East to the UK using a 250,000 gross registered tonne oil tanker. It is assumed that shipping of other (dry) freight will be of a similar scale. These are shown in Table 42.

Table 42: Emissions of Selected Atmospheric Pollutants from Transporting 1 Tonne of Freight from Middle East (kg/tonne freight)

The ETSU report does not specify the distance (or the port) from which the journey was made. We have assumed this to be Bahrain, the distance to which (by sea) is 6,175 miles. If we assume that dry freight incurs approximately the same emissions per tonne as oil, one can, on the basis of the distances travelled by imports, make a rough estimation as to the transport externalities incurred in sea transport from major exporting countries. In reality, the fuel used per tonne of cargo will depend upon:

• the shape of the freighter (dry cargo ships may be more streamlined);

• its size (dry cargo ships may be smaller);

• its age;

• the speed at which it travels; and

• the weight of the cargo.

In addition, the externality per tonne of final product will depend upon the degree to which the material as already been processed (the more highly processed materials have lower weights per unit of end product).

We have normalised the above import concentration ratios to 100% (effectively assuming all materials come from the countries mentioned) and derived a weighted average for the distance travelled by the average imported tonne. We have then calculated emissions by simply multiplying the per tonne distance by the emissions per mile under the assumption that the Middle East distance is 6,175 miles (as mentioned above). In the case of iron ore, we have multiplied by 1.48, since the amount of iron ore used, on average, in a BOF plant in making one tonne of steel is 1.48 tonnes. We have no such information for pulp and newsprint, but we assume that the import is of finished material (which is probably not such a poor assumption in the case of UK imports of newsprint).

The results are shown in Table 43 below. Although this shows that the movement of materials around the globe is not a cost-free process in environmental terms, one might argue quite reasonably that since many of the emissions will be occurring in areas where few people are found, then the higher estimates should be ignored (other than in the case of global pollutants like CO₂). The upper bounds may be closer to £3 in the case of pulp and paper, and £4 in the case of iron ore.

Table 43: Emissions and External Costs Associated with Freight of Pulp and Paper and Iron Ore.

6.8 Further Comments

The complexity of this process highlights some of the issues with which this type of analysis should be concerned. To the extent that one can construe the lack of internalisation of externalities as an implicit subsidy for their production (and therefore, a guiding rationale for understanding what they might be), one might also be interested to explore the explicit subsidies implied in tax breaks, government grants / subsidies, the process of awarding concessions (for mineral or timber harvesting), and others which are rarely applied to recycling per se, but are frequently applied to natural resource extraction.

Increased levels of recycling might save revenue (or reduce implicit losses) associated with such subsidies. On the other hand, this could displace jobs in primary production. However, these are likely to be offset (and most studies suggest they will be more than offset) by increased employment in materials collection, and secondary materials processing (see Waste Watch 1999b). Furthermore, the UK is a net importer of primary materials so the job displacement will occur in foreign countries. One reason why recycling attracts rather less government subsidy and support than the primary industries is that jobs in, for example, collection and separation of waste materials will be more dispersed across the country.

Note that similar comments apply as with the other treatments in that benefits arise not from the fact of the process itself, but because recycling is assumed to be replacing something worse (primary production), at least insofar as this analysis is indicative of net environmental effects of each. As such, just as questions concerning the growth or otherwise of energy demand may have implications for the calculation of net benefits, here the issue appears to be whether or not one assumes the materials economy is continuously expanding. Certainly, the these of eco-efficiency, and the concepts of Factor Four and Ten suggest materials recycling, but more fundamentally, materials reduction, as a way forward to reduce materials use in the economy over the longer term.

6.9 Summary Comments

This chapter has shown that externality estimates associated with each of the options are:

• very difficult to carry out;

• subject to a great deal of uncertainty;

• incomplete; and

• likely to vary with the specific choice of certain key parameters and assumptions which may or may not be used as a means of projecting the specific interests of certain interest groups (the point is that they can be).

Perhaps what can be said with some degree of certainty is that minimisation of materials use, of waste, and of energy use are strategies which are deserving of greater attention than they currently receive. Irrespective of 'bottom line' figures for externality estimates, which (for energy from waste options) incorporate the results of assumptions concerning displaced energy sources, all treatment options (including secondary materials reprocessing plants) have negative impacts. Waste minimisation activities that reduce the amount of municipal waste generated are urgently required.

It is tempting to suggest that, on the basis of the potentially huge externality savings associated with recycling that recycling is the best option. This might be the case, but so many caveats have to be kept in mind in reaching such a conclusion on the basis of this analysis (which, in the context of the problem in hand, is extremely limited). Some further analysis on the issue follows in the next Chapter.

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