Beyond the bin: The economics of waste management options

A Final Report to Friends of the Earth, UK Waste and Waste Watch

ECOTEC Research and Consulting Ltdi

BEYOND THE BIN:THE ECONOMICS OF WASTE MANAGEMENT OPTIONS


by ECOTEC Research and Consulting Limited


ECOTEC Research and Consulting Ltdii

1. INTRODUCTION.............................................................................................................4

2. THIS REPORT................................................................................................................6

2.1 A WORD ON LANGUAGE................................................................................................8

2.2 OUTLINE OF THE REPORT...............................................................................................8

3. PRIVATE FINANCIAL COSTS ........................................................................................9

3.1 INTRODUCTION............................................................................................................9

3.2 REVIEW OF EXISTING STUDIES / INFORMATION................................................................ 11

3.2.1 Capital Costs ..................................................................................................... 113.2.2 Costs of Recycling............................................................................................... 113.2.3 Costs of Composting ........................................................................................... 153.2.4 Home Composting............................................................................................... 183.2.5 Landfill and Incineration costs .............................................................................. 193.2.6 Revenues From Material Sales (Recycling, Composting and Incineration) ..................... 213.2.7 Revenues from Recovery of Energy.......................................................................... 223.2.8 Packaging Recovery Notes..................................................................................... 22

4. INFORMATION COLLECTION..................................................................................... 24

4.1 FINANCIAL INFORMATION AND SCHEME PERFORMANCE .................................................... 24

4.1.1 Scheme 1............................................................................................................ 254.1.2 Scheme 2............................................................................................................ 264.1.3 Scheme 3............................................................................................................ 264.1.4 Scheme 4............................................................................................................ 264.1.5 Scheme 5............................................................................................................ 274.1.6 Scheme 6............................................................................................................ 274.1.7 Scheme 7............................................................................................................ 284.1.8 Scheme 8............................................................................................................ 284.1.9 Scheme 9............................................................................................................ 294.1.10 Scheme 10.......................................................................................................... 30

4.2 COMMENT ON PERFORMANCE AND COSTS....................................................................... 31

4.3 COMMENT ON RECYCLING RATES ................................................................................. 33

5. EXTERNAL COSTS OF WASTE MANAGEMENT ......................................................... 35

5.1 LINEAR AND CIRCULAR FLOWS OF MATERIALS................................................................ 35

5.2 REVIEW OF EXISTING STUDIES ..................................................................................... 36

5.2.1 CSERGE et al (1993)........................................................................................... 375.2.2 Brisson and Powell 1995...................................................................................... 405.2.3 Coopers and Lybrand / CSERGE 1996 ................................................................... 415.2.4 Brisson 1997...................................................................................................... 425.2.5 Powell et al 1996................................................................................................ 44

5.3 STUDIES CONCERNING PAPER....................................................................................... 46

5.4 SUMMARY ................................................................................................................ 46

6. EXTERNAL COST ASSESSMENT................................................................................. 48

6.1 LIFE CYCLE APPROACH................................................................................................ 49

6.2 BACKGROUND TO OUR APPROACH................................................................................. 51

6.3 WASTE TRANSPORT..................................................................................................... 55

6.3.1 Residuals ........................................................................................................... 556.3.2 Recyclables and Compostables............................................................................... 556.3.3 External Cost Analysis ......................................................................................... 566.3.4 Results .............................................................................................................. 60

6.4 LANDFILL................................................................................................................. 62

6.4.1 A Note on Avoided Externalities Associated with Energy Recovery from Waste Treatment Facilities 646.4.2 A Note on Methane Emissions and Energy Generation from Landfill Gas ..................... 656.4.3 Results .............................................................................................................. 68

6.5 INCINERATION........................................................................................................... 71

6.5.1 Results .............................................................................................................. 766.6 RECYCLING............................................................................................................... 80

6.6.1 Results .............................................................................................................. 82


ECOTEC Research and Consulting Ltdiii

6.7 EXTRACTION PHASE ................................................................................................... 83

6.7.1 Hidden material flows and total material requirements.............................................. 856.7.2 Emissions during the extraction of raw materials ..................................................... 876.7.3 Emissions during the movement of raw materials...................................................... 88

6.8 FURTHER COMMENTS.................................................................................................. 90

6.9 SUMMARY COMMENTS............................................................................................ 91

7. PUTTING IT ALL TOGETHER - PRIVATE AND SOCIAL COSTS ................................ 91

7.1 SCENARIOS DISCUSSED................................................................................................ 91

7.2 COLLECTION AND DISPOSAL TO LANDFILL...................................................................... 92

7.2.1 Summary............................................................................................................ 927.3 COLLECTION AND DISPOSAL TO ENERGY FROM WASTE INCINERATION.................................. 93

7.3.1 Summary............................................................................................................ 937.4 COLLECTION AT KERBSIDE AND RECYCLING ................................................................... 94

7.4.1 Kerbside Materials Collection............................................................................... 947.4.2 Transport to Reprocessors .................................................................................... 957.4.3 Effects of Kerbside Recycling on Waste Management Costs......................................... 967.4.4 Summary.......................................................................................................... 100

8. MUNICIPAL WASTE IN THE NETHERLANDS........................................................... 105

8.1 NETHERLANDS......................................................................................................... 105

8.2 HOUSEHOLD WASTE ................................................................................................. 107

8.3 HISTORY OF TARGETS................................................................................................ 108

8.4 NATIONAL PROGRAMME ON HOUSEHOLD WASTE........................................................... 108

8.5 RESULTS AND COMMENTARY – RELEVANCE FOR THE UK?............................................... 110

8.6 COSTS.................................................................................................................... 111

9. CONCLUSIONS AND COMMENT................................................................................ 113

9.1 VALUATION OF COSTS AND BENEFITS OF WASTE TREATMENT OPTIONS .............................. 113

9.2 STUDY RESULTS - THE EFFECTS OF KERBSIDE RECYCLING................................................ 116

9.3 IMPLICATIONS FOR WASTE MANAGEMENT................................................................ 120

9.3.1 Policy Makers ................................................................................................... 1219.3.2 Waste Managers................................................................................................ 125

9.3 CONCLUDING REMARKS............................................................................................ 127

ANNEXES

ANNEX 1: QUESTIONNAIRE FOR RECYCLING SCHEMES USED IN THE STUDY.................. 130

ANNEX 2: EXTERNALITY ADDERS USED IN THE ANALYSIS .............................................. 141

ANNEX 3: ASSUMPTIONS CONCERNING GHG EMISSIONS FROM COMPONENTS OF

MUNICIPAL SOLID WASTE (USEPA 1998) ........................................... SEE TABLES.PDF

ANNEX 4: EMISSIONS FROM AVOIDED ENERGY SOURCES (ETSU 1997) ............................ 145

ANNEX 5: COMPOSITION OF MUNICIPAL SOLID WASTE USED IN THE STUDY .................. 146

ANNEX 6: RANGE OF VALUES FOR INCINERATOR EMISSIONS.......................................... 147

ANNEX 7: CALORIFIC VALUES OF COMPONENTS OF MUNICIPAL SOLID WA.................... 148

BIBLIOGRAPHY ............................................................................................................... 149

A Final Report to Friends of the Earth, UK Waste and Waste Watch_________________________________________________________________________________________

______________________________________________________________________________________ECOTEC Research and Consulting Ltd

BEYOND THE BIN -THE ECONOMICS OF WASTE MANAGEMENT OPTIONS


1. INTRODUCTIONThis report has been prepared by Dr Elisabeth Broome, Prashant Vaze and Dr Dominic Hogg of ECOTEC

Research and Consulting. It comes at an important time for UK waste management. The management of

municipal waste in the UK has to change, for a number of reasons:

1. the current pattern of resource consumption and disposal is unsustainable because it fails to account for the

needs of future generations; and

2. legislative drivers emanating principally from the European Commission are forcing it to change. The most

significant of these have been the Packaging Directive and the Landfill Directive, and with a Composting

Directive under consideration, the changes that may have to occur over the next 20 years or so will make

end-of-life materials management in the UK unrecognisable from its current form.

In the face of these challenges, the UK has issued a draft strategy for the Millennium ‘A Way With Waste’

(AWWW) (DETR 1999a). The draft strategy, the outcome of debates between Government Departments,

presents goals but no statutory requirements. The experience since the publication of ‘Making Waste Work’ does

not inspire confidence in the ability of those with the responsibility to do so to deliver on such voluntarist

targets. There may be a policy instrument developed to help move the UK towards Landfill Directive targets,

but this has yet to be announced. This may require authorities (explicitly or implicitly) to divert waste from

landfill, but on the basis of the document, Limiting Landfill, it seems unlikely to specify the mix of diversion

facilities in any way.

Rather than putting in place incentive structures (and the necessary finance) which might encourage a move

towards more sustainable waste management, the Draft Waste Strategy falls back on a technocratic approach

based on Best Practicable Environmental Option (BPEO) and lifecycle assessment (LCA) applied in parallel

with the Proximity Principle. There are two elements to BPEO as defined in AWWW. The first is that

decisions as to what is BPEO should involve consultation. The second requires an element of understanding of

the environmental costs and benefits of different approaches. Inevitably, in order for the latter element to have

any meaning at all, a mechanism for trading off these costs and benefits is required. This is important because

the different waste treatment technologies have very different environmental profiles.

On the one hand, there are many who would argue that making the trade-offs explicit by laying out each of the

physical consequences of the different options is desirable since nothing is hidden from the eyes of those

charged with scrutinising the impacts. There is much to recommend this approach given that specific facilities

may have local consequences for health and the environment, and since BPEO also requires consultation with

local communities, the potential consequences for them should be made transparent. This would be expected to

be particularly important where there were believed to be significant health impacts that were locally confined

(in which case, the phenomenon frequently referred to in somewhat disparaging tones as ‘NIMBY’ is an entirely

understandable response on the part of potentially affected parties).


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A drawback of this approach is that merely quantifying the emissions of pollutants to the environment is rather

different from understanding their impacts. Different pollutants will have more or less severe impacts for

different species at different concentrations or doses. It is clear, therefore, that some form of weighting of these

different impacts may be required. The problem that arises here, however, is that in many cases, the underlying

response relationships that move us from ‘emission’ to ‘impact’ are themselves subject to some uncertainty.

Again, in these circumstances, ‘NIMBY’ responses seem entirely understandable. Scientific experts do not have

a monopoly on truth. Citizens may ultimately prove to be wiser in their judgements than so-called experts,

whose title frequently saddles them with the burden of expectation of ‘certain’ knowledge. As such, LCA has

limitations as a tool upon which to base waste management decisions.

Ironically in the light of prevailing scientific uncertainties, and perhaps because the aforementioned approach

does not lead to a ‘clear decision’, there is considerable support for, and momentum behind, moves which seek

to bundle environmental and health impacts together into a common metric - that of money - so as to facilitate

the trading off of costs and benefits. This would, at least superficially, facilitate a reduction of the problem of

choice to a ranking of numbers. As implied above, however, this approach has its own drawbacks. Unless great

care is taken, the transparency of trade-offs implicit in the analysis may be reduced. Global impacts can be

lumped together with local ones, making light of the quite different political consequences which flow from the

impacts associated with different pollutants. Perhaps more importantly, individual components of the total

valuations can also be subject to uncertainty at least as great as the scientific uncertainties mentioned above.

Those engaged in valuation have tended to field criticisms concerning methodology with varying degrees of

success, but surprisingly few have appreciated the significance of more fundamental scientific uncertainties

which make it difficult to know exactly what effects one is seeking to place values upon. This means not only

that some impacts are difficult to quantify, but in some cases, the impacts, which may be real ones, are simply

ignored.

Valuation exercises can raise as many questions as they solve since the assumptions concerning values,

behaviour, and the underlying scientific relationships on which the exercises are based are rarely ‘beyond

dispute.’ These questions are at the heart of this piece of research, which seeks to shed light upon the utility of

valuation approaches in the context of waste management decisions. These decisions are being faced daily by

policy makers and local authority decision makers in the field of municipal waste management. There are clear

implications for the extent to which one can claim to know what might be the Best Practicable Environmental

Option under given circumstances.



2. THIS REPORTThis report aims to contribute to ongoing policy debates concerning the merits or otherwise of different waste

management options. The intention is also to raise questions concerning the basis for decision making in the

context of uncertainty. Such contexts are frequently encountered in addressing environmental issues.

The scope of the report is the municipal waste stream, as defined in the Draft Waste Strategy. We further

confine our attention to the following materials:

v Paper

v Aluminium

v Steel

v Glass

v Plastic

v Compostables

Furthermore, we limited our scope to the consideration of four options for waste treatment: landfill;

incineration; kerbside recycling and composting. The importance and significance of minimisation of the

municipal solid waste (MSW) stream should be mentioned at this juncture, but this is beyond the scope of our

analysis. Suffice to say that, in the context of Landfill Directive (Article 5) targets, a 1% increment in arisings

growth translates approximately (over 20 years) into a requirement for capacity to divert an extra 6 million

tonnes of biodegradable municipal waste within the UK (hardly a cost-free option). Lastly, it is worth stressing

that this is not an attempt to carry out a complete assessment of costs and benefits associated with waste

management options. We focus on environmental externalities only, and have not tried to capture a number of

variables which an extensive treatment would seek to capture (see RPA and Metroeconomica 1999 for a

discussion).

The Draft Waste Strategy talks of applying BPEO, but finding out what is ‘best’ is less than straightforward

where the scientific basis for such decisions is uncertain, and our understanding of how specific impacts should

be weighted is incomplete. More importantly, assuming that the Landfill Directive targets are applied in a way

such that they become a simple constraint upon the ‘choice set’ from which BPEO treatments are chosen, the

actual choices made would still be constrained by something else - finance.1

The actual choices that are followed through to implementation will be further constrained by another all-too-

often-ignored ‘factor’: people. People do not appear in LCAs, but they are an integral part of approaches aimed

at arriving at BPEO. After all, what is not acceptable to citizens is hardly ‘practicable’. Public resistance to

some waste treatment plants makes this an awkward strategy for meeting Landfill Directive targets where a

semblance of local consultation remains an integral component of the decision making process. Equally, it

could be argued, forcing citizens to engage in recycling activities suffers from the same shortcoming, though

there is at least some evidence suggesting that citizens are positively disposed towards recycling.

This heightens the importance of informing citizens about the options that currently face them. It is

undoubtedly true to say that the Landfill Directive will affect all citizens in the UK, directly or indirectly.

1 Note, in passing, that depending upon the degree to which minimisation of MSW arisings occurs over the next 20 years, the total tax

take from the Landfill Tax could (under optimistic projections) fall significantly. We estimate that the tax on MSW accounts for more

than half the total tax take. Supposing that zero percent growth occurred, the Treasury would see a decline in Landfill Tax revenue in

excess of £150 million (on the basis of £14 per tonne beyond 2004) relative to a business as usual scenario.



Equally all citizens can influence the nature of that effect by engaging in activities which help or hinder the

meeting of those targets. Paradoxically, very few citizens actually know this.

Cost-benefit analysis, of course, explicitly does include people. In particular, direct valuation methods do

attempt to elicit preferences from individuals that are deemed to represent the level of disamenity that might be

experienced owing to the presence nearby of, for example, a landfill or a reprocessing plant. On the other hand,

such methodologies are still the subject of controversies of both methodological, and more fundamental (i.e.,

related to the philosophical underpinnings of such an approach) natures. Some also argue for inclusion of direct

and indirect effects on employment arising from proposed changes in more comprehensive cost-benefit

analyses.2

The report starts out as a hypothesis. This is based upon work already carried out in the field and is based upon

what we already know, or think we know, about the financial costs of waste management. The hypothesis is

that:

although the financial costs of recycling are greater than that for other methods of dealing with waste, to the

extent that one is able to incorporate the environmental costs and benefits associated with all methods, the

overall economic analysis will show that when one accounts for all costs and benefits, the net result shows

recycling to be the best option in respect of materials recovered from the household waste stream.

There is effectively a second hypothesis applied to composting, the wording of which could remain the same,

though the financial costs of composting can be much less than those of recycling (depending upon technique

chosen). These hypotheses are important since, if they are born out by the analysis, the fact that these may be

the preferred options for managing waste (and surveys by organisations such as Waste Watch appear to confirm

that they are) would imply that public opinion and a basic environmental appraisal are ‘in sync.’ Note, however,

that particularly with regard to dry recyclables, the recycling of materials cannot occur ad infinitum so that the

simple ‘either / or’ characterisation of the argument is somewhat misleading. As noted in Ecologika (1998), a

more correct view may be that one is postponing the final disposal or recovery of the material.

It might be argued that to begin with such hypotheses is to start from a point of view which is in some way

‘biased.’ We have tried to be objective in this report, but we openly accept that, just as the pith of a mango

remains attached to the stone, so no such report can be completely cleansed of the perspectives of those carrying

out the research. For this reason, we welcome comments from those who read this document and feel either that

it misses something, or that there are inaccuracies in the analysis. It should be pointed out, however, that we

have tried to incorporate both sides of what are often quite polarised debates. Where people feel we are wrong to

do this, and where they feel they have evidence to support a particular view, we invite comment, but we suggest

that there is a lack of appreciation of the extent of the uncertainty which remains in our knowledge concerning a

range of environmental effects which are the subject of debate in the field of waste management. This is perhaps

the key underlying message of this work.

We note with interest the flurry of activity in so-called ‘science communication’ studies, the basis for which

seems to be that there exist ways of communicating ‘facts’ to citizens such that their perspectives and those of

so-called experts are aligned. Such communication faces inevitable criticism that it is guilty of ‘propaganda’

2 In the case of MSW treatment, any positive impacts on employment from using different treatment technologies has to be set against

any incremental expenditure. This will typically, in the case of MSW treatment, be taken from tax revenue. As such, any such increase

has to be set against the opportunity cost of the additional revenue requirement (e.g. alternative deployment of public funds, or direct and

indirect employment effects of reduced taxation at the margin).



(however even the balance of the information provided may appear to be). The fact remains, however, that the

views of the public will be shaped, for good or ill, by mediated experience. Indeed, the strategies of non-

government organisations and corporate entities may be regarded as closely aligned in this respect since they

attempt to shape media opinion (and hence, public perceptions) to suit their ultimate purposes.

There is also a debate going on about where the limits to recycling and composting might lie. Even if one were

to accept that BPEO for biodegradable dry recyclables and putrescibles were recycling and composting, and that

finance was no constraint upon the choice of BPEO, the fact that Landfill Directive Article 5 targets might still

not be met (because of ‘limits’ to recycling and composting and continued growth in arisings) would imply the

need for the ‘gap’ to be met by other forms of waste treatment such as incineration, gasification and anaerobic

digestion (see Section 7.6 below). In view of the public’s apparent lack of enthusiasm for incineration, certainly

relative to recycling, the question of where these limits might lie becomes significant. Furthermore, the size of

any ‘gap’ between recycling and composting, and the requirements of Article 5, will be influenced enormously

by the success or otherwise of the UK in constraining MSW arisings.

2.1 A Word on Language

The lexicon of waste management is a puzzling one. Something becomes waste on the basis of one’s intent to

discard the material. It is a moot question as to whether, when kerbside separation of waste is undertaken by

householders, they intend to discard the material given that their understanding of the process is that it should

lead to re-use of that material (and the disenchantment of citizens when hearing that their recyclables may be

being landfilled would seem to support the view that they do not feel that they are discarding them). The Draft

Waste Strategy for England and Wales makes the statement that waste should be perceived as a resource. It will

be very difficult to bring about the positive transformation that the statement envisages when even after

householders have become convinced that what was waste can indeed be a resource, that material remains legally

defined as ‘waste.’ This makes it very difficult to change the language used to describe what we do with

materials, yet this may be an important aspect of changing behaviour.

The definition of ‘waste’ as defined in European Law may therefore constitute an obstacle to changing

perceptions and attitudes towards waste. Given the centrality of the definition to a number of legislative

instruments, changing the definition is likely to prove difficult, since the wording of the legislation itself is

designed to prevent certain rather undesirable practices. Having said this, the perceptions of citizens in other

countries are clearly changing, usually, or so it would seem, for a variety of reasons encompassing historic,

geographical, cultural and legislative / policy influences.

2.2 Outline of the Report

In Chapters 3 and 4, we look at the financial costs of waste management in local authorities. We review some

of what has already been done in Chapter 3, and what is in the public domain. In Chapter 4, we look at

schemes for which financial data has been provided. Because some of the information is confidential, we have

restricted the nature of the information available. In Chapters 5 and 6, we discuss the valuation of the external

costs and benefits associated with the different schemes first through a literature review, then through our own

exploration of the issues. In Chapter 7, we bring together the financial cost and external cost data, and

summarise the results. Chapter 8 looks at the experience in the Netherlands with waste management, comparing

this with the UK experience (the Netherlands has a very high recycling rate by comparison). Chapter 9 draws

out some key issues from the work undertaken.



3. PRIVATE FINANCIAL COSTS

3.1 Introduction

Since 1995/6, the Department of the Environment, Transport and the Regions and the Welsh Office (and now

the National Assembly for Wales) have commissioned an annual survey of local authorities in England and

Wales to collect information on the collection, treatment and disposal of municipal and household waste (this

coincided with the period in which the Scottish Office ceased collecting similar data for Scotland). This survey

indicates that in 1997/8, two million tonnes of household waste were collected for recycling or composting,

comprising about 8% of total household waste (increased from 6.5% in 1995/6). Paper and card accounts for

37% of the household waste collected for recycling, and glass 18%. In addition, waste collected for centralised

composting has more than doubled in two years to 380,000 tonnes in 1997/8 (DETR 1999a). The estimates for

home composting are, for obvious reasons, a matter of some conjecture.

These recycling rates have been achieved through a combination of the introduction of "bring" sites for recycling

and, increasingly, the introduction of "kerbside" recycling schemes. There are currently around eight bring sites

available per 10,000 households. For kerbside collection, 38% of households are now serviced by a kerbside

recycling scheme of one type or another, up from 17% in 1995/6. The range of materials collected and the

quality of the service on offer varies considerably from one authority to another. There is a substantial difference

between a single material collection scheme involving somewhat tired vehicles and multi-material collections

using more modern vehicles and kerbside boxes, often using materials recovery facilities (MRFs) to separate

materials.

The Draft Waste Strategy for England and Wales (DETR 1999a) commits the Government to a "substantial"

increase in recycling rates. There are good reasons to believe that without increased efforts to collect at

kerbside, the costs for many households, and the effort required to become engaged in recycling will place

limits on the increases in recycling rates which can be achieved. This is supported by work undertaken in the

US by Jenkins et al (1999) which sought to understand (through econometric analysis) the determinants of

levels of household recycling:

As expected, the presence of a curbside recycling program has a positive and significant effect on intensity of

recycling activity for all five materials [newspaper, glass bottles, aluminium, plastic bottles and yard waste].

Regular curbside collection of recyclables lowers the out-of-pocket costs of recycling by eliminating the need to

transport recyclables to central (and possibly distant) collection points… Introducing a non-mandatory

curbside recycling scheme increases the probability that the average household recycles over 95 percent of

glass and plastic bottles by more than 36 percent; over 95 percent of yard waste by more than 28 percent; and

over 95 percent of aluminum by slightly less than 13 percent. (Jenkins et al 1999).

A dense network of bring schemes might provide an alternative approach, but unless the density is high, it will

most likely remain the case that such an approach will exclude some citizens who might otherwise participate in

recycling of those materials targeted by bring schemes. Kerbside schemes make recycling ‘easy’ for citizens by

reducing requirements for transport of materials.

On the other hand, many European schemes maintain a ‘bring’ component to their overall recycling strategy. In

the Netherlands, an attempt to standardise a model for MSW collection actually suggested that the basic model

might be:



q Glass collection by means of bottle banks at a density of one per 650 inhabitants;

q Paper and board collection at kerbside at least once a month

q Textiles collection at kerbside at least once every quarter, as well as by textile banks;

It was anticipated that this could raise collection rates for these materials to 90%, 85% and 50% respectively

(Ministry of Housing, Spatial Planning and the Environment u.d.).

Approaches which seek to separate materials for recycling after collection suffer from two shortcomings: the first

is that the quality of materials produced tends to be much lower than in the case where materials are separated at

source (so that at times where secondary materials prices are low the materials from a dirty MRF tend to be the

first to be rejected as merchants and reprocessors become more choosy about the inputs to their process); and the

second is that the potential to educate the public concerning waste is not so much lost as never really seized in

the first place.3 Schemes based on source separation effectively require those who participate to take some

responsibility for their wastes, and kerbside schemes in particular carry that message to the doorsteps of

households in the scheme (although the degree to which different participating households actually take this

responsibility varies under any scheme).

As part of its deliberations DETR is evidently considering the relative costs and benefits of recycling and of

increasing recycling rates, and the results of these deliberations will, in turn, be used in discussions with other

government departments. Annex C of the Draft Waste Strategy provided some insights into the possible costs

and benefits of different approaches to management of MSW in the context of the Landfill Directive, but the

costs and the benefits are subject to considerable uncertainty. The sources of this relate to:

q poor state of knowledge of the composition of UK waste (though this will always be subject to some

margin of error) and how this will change over time;

q poor state of knowledge concerning the financial costs of certain waste management options and how these

might change due to better or worse integration of system components, the influence of European

legislation, variation in location type (sparsity of population, housing stock, socio-economic status), and of

course, time (effects of innovation);4 and

q uncertainties in the measurement of the external costs of waste management options;

This makes any attempt to analyse costs and benefits over time somewhat difficult.

This Chapter reviews information in the public domain. Some of the figures are somewhat dated. We have

suggested what sorts of adjustment should be applied on the basis of exchange rate conversions and GDP

(Gross Domestic Product) deflators, but it should be recognised that not all costs will have followed inflation.

For example, in the collection and transport of waste, the role of the fuel duty escalator has been to move fuel

prices ahead of inflation. Note that where we have adjusted quoted estimates, we have done so simply by

accounting for inflation through straightforward application of GDP deflators. It goes without saying that there

may be reasons to expect these costs to vary in ways which are unrelated to inflation rates.5 Fuel duty, landfill

3 Following effective kerbside schemes, post-collection separation tends to generate only limited additional material collection.

4 Also deserving of consideration is the question of accounting for costs incurred and the financing mechanisms available. To the extent

that financing mechanisms such as PFI tend to favour (from the local authority perspective) schemes for which the capital costs of waste

treatment are more significant, there is an element of distortion in the financing mechanism towards capital intense schemes.5 In addition, it is sometimes the case that the year for which costs are estimated is not obvious. Generally, however, we are talking of

adjustments of the order 10%, so the effect of choosing the ‘base year’ wrongly might be expected to incur errors of the order 5% at

most. This is probably less significant than errors incurred in carrying through an assumption that all costs can be re-based to 1999 using

GDP deflators.



tax, and materials prices may affect costs of specific treatment options significantly but their movement will not

be closely related to inflation rates.

3.2 Review of Existing Studies / Information

3.2.1 Capital Costs

Published comparisons between the relative capital costs of different waste management or disposal options are

infrequent, and generally provide insufficient background information to be of value in any more detailed

assessment of waste recycling options.

DTI (1997) presented a table of the capital costs of various disposal options for household and commercial

wastes (Table 1 below), indicating the highest investment costs are required for incineration.

Table 1: Capital Costs Of Facilities for Household And Commercial Wastes - Processing 200,000 tonnes

per year (£ Million) (figures in parentheses are 1999 £)

Facility Costs (£ million / 200,000 tonnes)

Energy from waste plant 40 (43)

Anaerobic digestion plant 25 (27)

MRF and transfer station 10 (11)

Transfer station and Civic amenity site 5 (5)

Landfill 4 (4)

Civic amenity site and MRF 2 (2)

Civic amenity site 1 (1)

Source: J Holmes, as cited in DTI 1997

This partly explains the long-term nature of contracts usually required in order for the construction of an

incineration plant to be worth the risk of the capital commitment. It also helps illustrate the manner in which

local authority decisions are likely to be distorted by financing mechanisms such as PFI (Private Finance

Initiative) that effectively grant local authorities credits to cover the costs of borrowing for capital equipment.

More commonly, the available cost comparisons take account of both capital and operational costs to establish a

figure or range for the overall costs per tonne of dealing with waste through the various disposal options.

3.2.2 Costs of Recycling

Estimates of recycling costs are provided in Coopers and Lybrand's (1993) analysis, which provided the

marginal costs associated with different recycling options (Table 2). It is worth noting that the cost of collection

in bring schemes is given net of sales revenue (hence, the zero revenue figure for bring schemes). An important

point to note is that the separation cost is given as the same irrespective of the level of materials separation

prior to collection. Our work suggests this is highly questionable.

Table 2: Summary of Recycling Costs (figures in parentheses are 1999 £)



Collection system Collection cost

(£/t)

Separation cost

(£/t)

Sales income (£/t) Recycling cost (£/t)

Bring 16-36 (18-41) 0 (0) 0 (0) 16-36 (18-41)

Blue box 60-150 (69-173) 50 (58) 25 (29) 85-175 (98-201)

Green bin 25-40 (29-46) 50 (58) 20 (23) 55-70 (63-81)

Green bag 25-45 (29-52) 50 (58) 20 (23) 55-75 (63-86)

Source: Coopers & Lybrand, 1993

A study carried out by Atkinson et al (1996) estimated costs for bring and kerbside recycling schemes in five

cities in the UK. The costs of bring systems varied between £40-£49 per tonne (£43-£53 in 1999 terms) in

low-density bank systems (one site per 3,500 households) to £65-£133 per tonne (£71-£145 in 1999 terms) in

high-density areas (one site per 500 households). These are much higher than the Coopers and Lybrand

estimates. For kerbside collection schemes, the study differentiated between separate collections and integrated

collection schemes - the former costing between £140-£231 per tonne (£153-£252 in 1999 terms) and the latter

costing £79-£155 per tonne (£86-£169 in 1999 terms). Again, the higher end of this range is higher than one

sees estimated in other studies (see below). The higher costs were associated with the separate kerbside

collection system involving two collections per household, one for recyclables and the other for residual waste.

Yet even accounting for this, the range £140-231 seems high. Indeed, all the costs estimated incorporate (within

ranges) values that are higher than in the Coopers and Lybrand report.

Work by Brisson (1997), based on work in which she was involved (EC 1996) suggests financial costs for the

UK (given in Euro) as at Table 3 below. The presentation of a single figure is somewhat unrealistic, and other

studies are apt to make use of a range of values (see above). Even allowing for this, the point to be made is that

the figures are not easily reconciled with those from the Coopers and Lybrand report.

Table 3: Costs of Recycling, 1993 ECU (figures in parentheses are 1999 £)

Mixed collection and

bring system

Co-collection

(blue box system)

Separate collection

Recycling 13 (12) 62 (56) 160 (146)

Source: Brisson 1997

A number of councils that have established recycling schemes are starting to present more detailed economic

information about the operation of their scheme. One such example is given in Box 1.



Box 1: Adur District Council's Waste Recycling Plan

Adur has achieved a recycling rate of over 25%, through the following activities.

• Kerbside Blue Box collection of four types of recyclables - glass, plastics, cans, newspapers

• Materials processed at a MRF

• Bring bank system

• Community recycling schemes operated through community groups and charities

• Home composting promoted through free composting units

• Income through sale of recycled materials, recycling credits, grants and sponsorship

£/household/year

Costs:

Household waste collection cost 16.62

Household waste disposal cost 18.68

Net cost of Blue Box and mini recycling banks 4.64

Income:

Sale of materials and recycling credits 2.75

Grants and sponsorship (Blue Box) 8.33

Total paid by Adur council tax payers 4.47

Source: Adur District Council Recycling Plan Summary, 1996, as cited in Williams 1998

Ecologika (1998) report on a Coopers and Lybrand survey of four schemes, the best performing of which (in

terms of diversion) cost £108 in gross terms with revenues of £13 (so net costs of £95).

The Audit Commission's Recycling Cost Survey, carried out in 1996, provides a detailed analysis of the

relative costs of bring and kerbside recycling in 21 local authorities. The survey categorised the schemes into

different types:

• Householder waste presentation method (box / sack / wheeled bin / bundled / carrier bag, etc.);

• Frequency of collection (weekly / fortnightly);

• Co-collected or not with refuse for disposal; and

• Vehicle type (multi-compartment / split / standard etc.),

together with details of the population served, and the tonnes recycled.

The survey identified the range of materials collected, including paper, glass, metals, plastics, textiles, other,

and the costs per tonne for the various materials collected. These data, combined with information provided on

the various financial costs experienced by the authorities to establish and run the schemes, provided details on

the gross and net costs (i.e. including receipts from sales of materials collected) for each of the authorities.

Recycling credits were excluded.

The survey reported a wide variation in the costs of kerbside collection schemes. Average gross costs per

household were £8.99, but these costs varied between different collection schemes from as low as £2 per

household to as much as £18.



The report noted the wide variation in costs between kerbside schemes, and suggested that the principal reasons

for these variations may be attributable to a range of factors, most notably:

• frequency and method of collection

• materials and tonnages collected

• population density

• number of households served

• efficiency of the collection, transport and sorting system

• materials income and recycling credits (Audit Commission 1997)

However, with only 21 authorities investigated in the survey, the relative importance of each of these factors

could not be assessed in detail. The influence of ‘less tangible factors’ such as scheme age and the information

processes used to promote participation, are not mentioned as being of influence, though other country

experience suggests they may be important. There is no distinction made between schemes that do, and schemes

that do not operate a MRF.

Some analysis was carried out to demonstrate the effects of the tonnages collected from each household on

average costs per household. This is presented in Table 4 below. The Table appears to suggest that a) the costs

per household increase as the rate of materials capture increases, but b) that costs increase at a lower rate than

materials collection (so that the costs of waste collection per tonne fall even as the costs per household

increase). This is consistent with other countries’ experience, and would be in line with what is expected when

increases in participation rates occur, perhaps with the scheme’s maturity, though costs may also increase with

the inclusion of new materials (especially plastics). In the former case, the marginal costs of collection may be

close to zero, though costs will be incurred in materials transport to reprocessors.

Table 4: Kerbside Recycling Costs and Performance (figures in parentheses are 1999 £)

Kg recycled per household: 21-25 26-50 51-75 76-100 101-146

Sample size 2 3 12 1 3

Average costs:

Gross cost per household (£/household) 4.74

(5.02)

7.01

(7.43)

7.56

(8.01)

12.55

(13.30)

14.59

(15.47)

Net costs (excluding recycling credits) per

household (£/household)

4.66 6.20 6.20 8.37 10.63

Net costs (including recycling credits) per

household (£/household)

4.54 5.74 5.48 8.37 8.03

Net cost per tonne waste collected

(excluding recycling credits) (£/tonne)

200 (212) 160 (170) 100 (106) 90 (95) 90 (95)

(Excludes kerbside schemes covering less than 10,000 households)

Source: Adapted from Audit Commission Recycling Cost Survey, 1996

The same finding is reported in work conducted in the US on kerbside schemes. Stevens (1994) reports that

cities with relatively high recycling rates had costs about one third that of schemes with lower recycling rates.



The US Environmental Protection Agency reported similar findings for ‘multi-family recycling’ (high rise

buildings etc.) (USEPA 1999). Ecologika (1998) report that the costs of a UK blue box scheme fell over three

years from £95 per tonne to £50 per tonne through reducing collection and sorting costs and increasing income

through consortia selling. The same study sites evidence from North America that suggests the potential for

cost reduction over time. One suspects that reductions arise both through ‘learning by doing’, and to some

extent, through increases in participation, though net costs (as opposed to gross costs) are vulnerable to swings

in materials prices, which obviously affect revenue streams.

Though Table 4 may appear to show that there is a correlation between costs and materials capture, the range of

influential factors would appear to make the strong application of such a conclusion somewhat premature. In

many cases, one is probably comparing apples with pears. Glancing at the full dataset, one notes, for example,

that the scheme recycling most per household (146kg) has a net cost per household (£11.26) that is less than

that of one of the low performing schemes (65kg at £14.61). At the same time, one scheme reports negative net

costs per household with capture rates of 61kg per household. Though average gross costs of kerbside collection

are £143 per tonne, these range from £41 – £290, whilst net costs average £107 per tonne, but range from £-4 to

£277.

In general it is clear that variation in the costs of waste treatment exist not just between the different waste

management options but within those options too. The term ‘kerbside collection’ encompasses a whole range of

activities and approaches whose gross costs (expressed in terms of £/tonne of material collected) vary

considerably. For exactly this reason, it is very difficult to draw firm conclusions about ‘the costs of recycling’

without far more detailed investigation, and closer specification as to what it is whose costs one seeks to

understand (and in what specific context). There may also be variation related to the start date of the scheme.

This is suggested by studies in the US, one of which revealed a statistically significant effect of scheme age

upon the collection rates for different materials (Jenkins et al 1999).

Arguably, for recycling systems, the factors affecting the costs of the scheme are more numerous than for other

technologies. The logistics concerning collection method and frequency are subject to a number of variables, but

it seems fair to say that lessons are being learned in this area. Collection methods, decisions concerning

materials coverage and collection frequencies cannot be divorced from decisions concerning the nature of the

vehicles and containers being used in the collection. These then influence the labour requirements of the

scheme. Each of these inter-related decisions has implications for costs both for the scheme itself and

‘downstream’, as in the case of scheme-organisers’ decisions to deploy (or not) MRFs. There seems to be more

potential for ‘learning by doing’ in recycling than there might be in, say, the more traditional final treatment

options, if only because, certainly in the UK, we are still in the ‘steep segment’ of the learning curve.

3.2.3 Costs of Composting

One of the Government's key objectives is to increase the amount of organic material in the municipal solid

waste stream that is composted. This objective may yet be propelled by legislative force in the form of a

proposed Composting Directive. This might require local authorities to collect organic waste separately. Similar

issues concerning the cost of schemes are likely to arise here because of the nature of the collection process. At

the same time, there may be specific characteristics of organic wastes (for example, its density, the proportion

found in the MSW stream, its moisture content, its putrescible nature) which make it suitable for collection in

ways to which mixed dry recyclables are not so well suited.



The Composting Association (1997) estimates that each UK household sets out some 200-250kg of VFG

(vegetable, fruit and garden) waste each year. The term VFG is used (in their report) to maintain consistency

with the nomenclature of other European countries and is used to differentiate such waste from purely garden

waste. An estimated 100kg p.a. of green waste per household is also available for composting.

In 1997/8, just over a third of local authorities in England and Wales had established centralised composting

schemes. These receive organic waste from both bring schemes and from civic amenity sites. The Composting

Association (1997) and HDRA (1999) have investigated the costs of establishing and running a centralised

composting scheme in some detail.6

Estimates of composting costs were provided in Coopers and Lybrand's (1993) analysis, which provided the

marginal costs associated with different recycling options (Table 5). This study includes a value for separation

of materials for home composting. This is not a value for separation per se but an estimate of the total costs of

minimisation via this route. The Coopers and Lybrand report expressed the view that ‘kerbside collection of

organic waste is extremely expensive, averaging £95 per tonne.’ This is not clear from the work we have

undertaken, much depending on vehicle design and collection logistics.

Table 5: Summary of Composting Costs (figures in parentheses are 1999 £)

Collection system Collection cost

(£/t)

Separation cost

(£/t)

Sales income (£/t) Recycling cost (£/t)

Central composting (1) 0 (0) 14-24 (16-28) 9 (10) 5-15 (6-18)

Central composting (2) 85 (98) 14-24 (16-28) 9 (10) 90-100 (104-115)

Home composting 0 (0) 45-50 (52-58) 0 (0) 45-50 (52-58)

(1) Based on bring system; (2) Based on a split bin, kerbside collection

Source: Coopers & Lybrand, 1993 (in 1999 terms, the costs would be of the order 15% greater)

In respect of composting costs, the outcome of the Composting Association’s own survey of capital and

operating costs is shown in Table 6 below. Their comment concerning costs is worth quoting:

‘Unfortunately, a general answer to this question [how much does composting cost?] is impossible – costs are

heavily dependent upon site-specific factors such as land ownership, the nature of capital financing used, the

throughput, and the type of process required.’

Note that the costs do not include the collection process that is covered in the Coopers and Lybrand (1993)

study. Without collection, the operating costs range from £6 to £30. Note also that the figures above show gate

fees for composting sites. These are not so different from landfill gate fees (see below).

HDRA (1999) identified the principal composting systems currently available and presented example costs.

The open air turned windrow composting system is currently the most prevalent in the UK. In this system, the

compost feedstock is formed into long windrows, in the open air. The compost is then aerated by regular

turning of the material. HDRA (1999) suggests costs are typically in the range £8-£18 per tonne throughput,

depending on site requirements and equipment used. HDRA also carried out a more detailed cost analysis for

an example of a plant processing 18,000 tonnes per year, a summary of which is presented in Table 7 below.

6 This unpublished work was carried out under a sub-contract to ECOTEC.



Table 6: Operational and Capital Costs of Composting

Year Throughput

(tpa)

Gate fee (£/tonne) Operational cost

(£/tonne)

Capital Cost

(£)

Capital Cost

(£/tonne)

1992 50 4,000 80

300 21-30

1996 300 15,000 50

1995 500 5,000 10

1994 1,400 21-30 20,000 14

1996 1,500 11-15 90,000 60

1997 2,600 16-20 21-30

1993 3,000 6-10 100,000 33

1988 5,000 37,000 7

1994 5,000 16-20 65,000 13

1995 8,000 11-15 11-15 600,000 75

1993 9,000 11-15 11-15 250,000 28

1993 9,000 6-10 250,000 28

1991 10,000 11-15 16-20 170,000 17

1996 10,000 11-15 16-20 80,000 8

1993 13,000 11-15 6-10 250,000 19

1996 16,000 1,200,000 75

1996 18,000 11-15 16-20 500,000 67

1990 25,000 <10 250,000 10

1985 30,000 250,000 12

Source: Composting Association 1997

Table 7: Open Air Windrow Composting System

Total tonnes processed per annum 18,000

Total investment £596,000

Operational costs per annum (1) £263,491

Cost per tonne input £14.64

(1) Investment costs were annualised and included in operational costs

Source: HDRA 1999

The report also provided estimated costs of a smaller operation, processing 10,000 tonnes of green waste per

year. For this example, total investment costs were £281,733, with operating costs totalling £18.79 per tonne

to produce a bagged 10mm screened product. The relative costs of these two processing operations indicate the

potential for economies of scale with larger installations. An earlier assessment of costs of centralised

composting facilities at varying levels of scale was provided by Leeds University, and also suggests economies

of scale for larger scale operations - see Table 8 below. Such scale economies are likely to be realised within

specific technical options, but the costs across technical options will vary significantly (see above and below).



Table 8: Centralised Composting Costs

Plant capacity (tpa): 3,500 7,500 18,500

Gross costs (£/t) 24 18 14

Net cost (£/t) (1) 15 8 5

(1) The net cost is based on a sales value of £15 to £20 per tonne.

Source: Department of Civil Engineering, Leeds University, as cited in Coopers & Lybrand, 1993

HDRA also provided example costs for a variety of in-vessel composting systems. They identified five main

types of in-vessel system - containers, silos, agitated bays, tunnels and enclosed halls. Detailed specifications

for two types of in-vessel systems were provided, for a GICOM batch tunnel composting system and a Herhof

composting plant. These are summarised in Table 9 below. Evidently these facilities are more costly than open-

air windrows, but the in-vessel nature of the process has advantages over open air systems. These relate

(amongst others) to the speed at which material is processed (so less land is required), and the degree to which

odours can be controlled (allowing processing of malodorous wastes).

Table 9: In-vessel Composting Systems (costs per year)

System GICOM batch

tunnel

Herhof

batch

container

Total tonnes processed per annum 20,000 18,000

Total investment (1) £2,083,333 £2,140,847

Operational costs per annum £492,107 £527,069

Cost per tonne input £24.60 £29.28

(1) Financing of investment costs included in operational costs

Source: HDRA 1999

From this brief review, it would seem that for open-air windrows, the costs of composting technologies might

be of the order £10-£20 per tonne. To this one must add the costs of collection and separation, and it may be

true to say that lower cost collection incurs higher costs in terms of separation. Equally, better separation is

likely to lead to higher quality end-products. Hence, countries which have made most progress in composting,

such as Austria, have strict rules concerning what fractions of waste must be separated or home composted, and

a supporting system of standards designed to prevent land spreading of compost which has high levels of

contamination.7 Optimisation of collection systems has to occur in the context of broader considerations of the

integration of components of the waste management system.

3.2.4 Home Composting

Home composting reduces the quantities of organic waste for collection without adding to collection costs.

Costs associated with home composting are primarily associated with any costs for provision of home

composting (compost bins, wormeries, digesters), and any additional publicity or information material required.

The distribution of the financial costs between those undertaking the activity and the bin providers will be

determined by the level of subsidy provided by a given service provider.

7 See the Austrian submission in DHV et al (1997).



Few studies appear to have addressed the financial costs of home composting. An exception is the Coopers and

Lybrand (1993) study (see Table 5 above). We estimate that home composting can reduce disposals by 10-20%

(figures used by Local Authorities to estimate home composting rates tend to use values between 75kg and

150kg per annum per household that is believed to be engaged in the activity). A composting bin may cost

£35-50 unsubsidised. There is no obvious reason why basic bins would have limited lifetimes. The costs of

treatment per tonne will depend upon the way in which one accounts for the outlay.

Collection costs could not be expected to fall on the basis of marginal changes but disposal costs might. For

100kg of waste, the savings on disposal may be of the order £2-£3 (including Landfill Tax). Assuming straight-

line depreciation, and assuming the materials are used by the householder, the bin pays for itself if it lasts 10-

15 years even if the Local Authority pays for the bin (assuming bulk purchases result in low costs for the bin).

Issuing free bins may present its own problems. Under a 50% subsidy, a £30 bin would pay for itself if it lasts

5-8 years.

3.2.5 Landfill and Incineration costs

The work carried out by Coopers & Lybrand (1993) also evaluated the current (1993) and future (next 10 years)

costs of landfill and incineration, and their estimates are given in Table 10 below. Note that most of the figures

discussed below do not include landfill tax, which now stands at £10 per tonne for municipal waste, and will

increase by £1 per year to 2004. The figures for future costs represent the likely range of costs after existing

incinerators were to be fitted with emissions control equipment beyond 1996. Figures from Brisson (1997) for

the UK are given in Table 11. The two sets of figures are not comparable since Brisson’s effectively include

collection costs whereas the Coopers and Lybrand work figures do not. Note also that the two studies seek to

capture variation in landfill costs across different dimensions, the one looking at variation with gas collection /

energy recovery, the other looking at the rural/urban dimension. In practice, both will affect costs. It is worth

noting in passing that all landfills accepting biodegradable waste (in practice, probably all receiving municipal

waste) will be required to have gas collection equipment in place in the near future (as a consequence of the

Landfill Directive).

Table 10: Costs of Waste Disposal Options (figures in parentheses are 1999 £)

Option Cost per tonne (£)

Current (1993) Future (2003)

Landfill

Urban high cost 22.5 (26) 31-47 (36)

Rural high cost 15 (17) 22-35 (25-40)

Urban low cost 10 (12) 15-24 (17-28)

Rural low cost 7.5 (9) 11-18 (13-21)

Incineration

Mass incineration 15-20 (17-23) 20-25 (23-29)

Energy recovery 20-30 (23-34.5) 20-25 (23-29)

Source: Coopers & Lybrand 1993



Table 11: Financial Costs Of Waste Management, ECU/Tonne (figures in parentheses are 1999 £)

Mixed refuse collection,

recyclables through

bring scheme

Co-collection

(blue box)

Separate collection

of recyclables and

residual

Landfill – no energy 85 (77) 68 (62) 68 (62)

Landfill – gas recovered 85 (77)

Landfill – energy 85 (77)

Landfill – no transfer 68 (62)

Incineration – electricity 91 (83) 91 (83) 91 (83)

Ecologika (1998) report costs in terms of the breakdown between collection and treatment. For London, they

estimate these (per tonne) as:

Landfill - £25 for collection, £30 for disposal. Revenue from energy sales (where energy is recovered)

equivalent to less than £1 per tonne. This would give a total of the order £54-55 for delivery of waste for

disposal / recovery at landfill; and

Incineration - £25 for collection, £44 for incineration, £13 for the disposal of residual ash, and £15 from

the sale of heat and power. This would give a total of £67 per tonne for a new 300,000 tonne plant.

The Government's Draft Waste Strategy (DETR 1999a) gives details of a modelling exercise carried out into

municipal waste targets. The modelling used a set of assumptions of the costs of different waste management

options, as given in Table 12 below. These are the total resource costs, excluding the landfill tax. The costs

include collection, transfer and transportation to the disposal or recovery facility, as well as gate fees. They

include operational and capital costs (which are annualised for conversion to costs per tonne). The recycling and

composting costs assume a zero price for recyclables (no income is received from the supply of waste for

recycling or composting).

Table 12: Costs Of Different Waste Management Options

Treatment Cost Range (£/t)

Recycling (kerbside collection) 55-145

Composting (kerbside collection) 70-120

Incineration 45-100

Landfill (excluding tax) 45-65

Source: DETR 1999a

The wide range in costs demonstrates the uncertainty over present and future costs and variations between

authorities and regions, both in actual costs and in the approach to reporting costs (in particular, whether or not

collection costs are included, and whether or not what is being reported includes capital costs, and if so, how

this is accounted for).8 The costs of collection, transfer and transport, which are significant components of the

overall cost, are simply not uniform across different parts of the country, partly for geographic and demographic

reasons, put also because of the different systems used in these parts of the process. The location and scale of

8 It has been suggested that the ownership of facilities is a factor which can affect the way in which costs are accounted for.



the plant also affects costs, as does the sophistication in terms of monitoring and gas recovery equipment at

landfills, and the pollution control equipment at energy from waste incineration plants.

3.2.6 Revenues From Material Sales (Recycling, Composting and Incineration)

In some of the studies surveyed above, the costs incorporate the revenues from material sales. It is worth

considering these separately. Revenues from dry recyclables will depend on the weight contribution provided by

each recovered fraction and on the price (£/t) received for that fraction. Coopers & Lybrand (1993) reported the

relative weight contributions from each of the separated recycled materials (Table 13).

Table 13: Weight Contribution For Recyclable Materials (kg/tonne)

Paper and board 510

Plastics (bottles) 70

Plastics (film) 30

Glass (colour separated) 230

Steel 110

Aluminium cans 15

Textiles 35

Total 1,000

Coopers and Lybrand 1993

These compositional data, to the extent that they are indicative of the relative quantities of material collected for

recycling (and clearly, they will not be so for collections which do not collect all these materials – in any case,

see below for compositions of kerbside collections), show how important in weight based terms the collection

of paper and board is in recycling schemes.9 According to these figures, along with glass, these materials

account for three-quarters of materials collected (and in the kerbside schemes examined below, the proportion is

much greater). Clearly, fluctuations in the price for these materials are important in the context of mixed

collections of materials from the waste stream. Equally, however, the significance of the materials is effectively

weighted by their relative price. In this respect, because the prices paid for aluminium are typically much higher

than those for other materials, aluminium acquires some greater significance than its low proportion (by weight)

suggests.

More generally, the price which is secured under any given contract, along with the terms of any contract, will

be an important determinant of the net costs associated with any specific recycling scheme (and their stability).

These terms may be better or worse depending upon the timing of any contractual negotiation. It is for this

reason that many have argued for the need for longer term contracts and/or market development activity, to give

stability to materials revenues, as well as increased value for the material itself.

Where composts are sold, sales prices for waste-derived composts range from about £1.50 to £3 for 40-50 litres

(delivered), alternatively bulk delivery of one tonne or more of compost can also be provided at approximately

£10 per tonne.

9 Our figures suggest that these numbers grossly underestimate the significance of glass and paper in kerbside collection (and almost

certainly in collection at bring sites as well).



MSW incinerators will typically extract steel (and in some plants, aluminium) from bottom ash. This is now

typically done post-incineration since the incineration process effectively resolves problems associated with

mixed materials (e.g. disposable razors, etc.). There are quality issues associated with this approach to materials

extraction since the material can take up some slag during the incineration process. However these materials

have value and are cleaned and baled for sale. Increasingly, incinerator operators are also looking to make use of

bottom ash in construction applications. This also recovers value. However, whilst the environmental benefits

associated with metals recovery may be significant, they may be less so for bottom ash recovery (see below).

One of the rationales for making use of bottom ash is that this avoids disposal costs associated with landfilling

the material. As such, there may even be a net financial saving even where construction operators are paid to use

such materials, or given the material at zero cost. In this sense, the landfill tax affects bottom ash in a similar

way to traditional construction wastes, encouraging their recovery / utilisation rather than delivery to landfills.10

Recycling credits are paid to waste collection authorities (WCAs) and (outside Wales), at the discretion of the

waste disposal authority (WDA), to third parties. Guidance on the payment of recycling credits suggests that

these should reflect the avoided cost of disposal of the marginal unit of waste. This includes landfill tax.

However, although it can be argued that third party credits help avert expenditure on collection as well as

disposal, the cost of collection is never included in third party credits. One could argue that this ‘omission’ is

significant to the extent that third parties recycle significant quantities such that these reduce the cost of

residuals collection below that which would prevail in the absence of the scheme.

3.2.7 Revenues from Recovery of Energy

Both landfill and incineration generate revenues through the sale of energy. Both have benefited from the

existence of contracts and funding through the Non Fossil Fuel Obligation. Particularly for landfill (because of

the conversion of methane to carbon dioxide, reducing the global warming potential of the gaseous emissions

per atom of carbon), this makes the recovery of energy a beneficial process both in terms of reducing external

costs associated with the process (this would occur only indirectly for incineration through displacement of

alternative energy sources – see Chapter 4), and reducing private financial costs.

3.2.8 Packaging Recovery Notes (PRNs)

More recently, energy-from-waste (EfW) plants have benefited from their being able to issue PRNs against

waste delivered for recovery. This could change in the future depending upon the European Commission’s view

concerning the role of EfW in fulfilling the recovery obligation under the Packaging Directive. To our

knowledge, and on the basis of current and past work, few if any authorities are, or have been, seeking to

become involved directly in the market for PRNs. This statement requires some qualification.

Firstly, even if no authority were directly involved in this market, the price for the materials delivered could

increase. However, in this context, a number of points are worth considering:

• The products which, by weight, are most significant from the point of view of local authority collections

are newsprint / magazines and glass. The former are not classified as packaging, so that delivered material

would not be expected to benefit from any price premium related to PRNs (and Figure 1 provides an

illustration of the effect on packaging grades of waste paper).

10

Note that if one assumes that around 25-30% of treated material is left as bottom ash, then 3.3 million tonnes of capacity or more would

be required to generate 1 million tonnes of material. This is unlikely to destabilise aggregates markets greatly given that the aggregates

industry supplies a market of the order 200 million tonnes.



• In general, because the supply of PRNs has been in excess of demand, the value of PRNs required to make

up an obligated entity’s recovery obligation has been relatively low (at least, once it became clear that

supply would outstrip demand). This is partly due to the very fact that the large quantity of PRNs related

to reprocessing of paper and board packaging has ‘cross-subsidised’ other materials in respect of making up

the gap between materials specific recycling obligations and the overall recovery obligation. This may

change if the revised Packaging Directive increases materials-specific recycling targets.

• The material whose PRN price is most likely to rise in the near future (on the basis of material-specific

targets not being met – much depends here on the rate of growth in exports of plastic packaging to foreign

reprocessors) is plastic, a material which is collected in relatively small quantities through local authority

recycling schemes.

Overall, the very short-term prospects for enhanced materials value from the average tonne of material recycled

by local authorities is not great. Beyond this, however, as the current situation of PRNs surplus begins a

transition to possible deficit, the prospects for enhancement of materials prices begins to look more promising

(if, indeed, that is the right term to use in this context). Furthermore, the much-awaited revisions to the

Packaging Directive will have implications which, on the basis of the discussions which have been ongoing,

would seem to make this scenario more rather than less likely (and especially in respect of plastics).

Figure 1: Changes in Materials Prices for Waste Paper Grades

$0

$5

$10

$15

$20

$25

$30

$35

$40

Jun Ju l Aug Sept Oct Nov Dec Jan Feb Mar Apr May Jun Jul Aug Sept

1997-98

A2, mixed sorted P&B

A5, corrugated waste

A9, mixed news, pams

B1, once-read newspaper

Source: PPI 1998 (in Enviros RIS 1999)



4. INFORMATION COLLECTION

A detailed evaluation of the wider economic benefits and costs of recycling requires a more detailed

understanding of the financial costs and incomes associated with different recycling options. For most of the

studies identified above, however, the costs of the recycling schemes have generally been presented as an overall

cost per tonne of waste collected or disposed of. There is little or no supplementary information that breaks

these costs down further into their components.

In our analysis we have sought to obtain more detailed costs data about the key components of a waste

recycling scheme that determine not only financial costs but also link with environmental costs or benefits.

However, we are not in a position to make all this information public for reasons of commercial confidentiality.

The questionnaire developed to compile detailed information about the net financial costs of a selection of

recycling schemes in England, and data on the materials recycled, is reproduced in full at Annex 1. One

omission (rectified through follow-up) was the lack of a question concerning gate fees at composting facilities.

The recycling schemes selected from which information was requested were chosen to be representative of a

range of different types of recycling schemes in operation. Those factors considered were:

• location of scheme - urban, rural

• age of scheme - recently established, mature scheme operational for a few years

• type of organisation running the scheme – private company, local authority, community sector

organisation

4.1 Financial Information and Scheme Performance

The following sections give outline information of the financial costs and general performance of the schemes

chosen. Note that we have not been able to obtain responses from all authorities with which we spoke. All ten

organisations initially interviewed expressed willingness to participate in the survey, but some found the

questionnaire too onerous in the face of competing priorities. In all cases, we have followed up the

questionnaire with further questions to clarify issues arising from the responses.

We have tried to probe deeply into the analysis of costs to ensure that all schemes are being costed on an equal

footing. This has been a frequent criticism of past attempts to assess the ‘true’ costs of different options.

Indeed, one can argue that there are ‘pros’ and ‘cons’ in such an approach. There are different organisations

playing different roles in this sector. If one seeks to account for all outlays on an equal footing, whilst this may

place schemes on a level playing field in terms of the efficiency of the resources they deploy, the different

organisations will charge those to whom they provide a service in different ways. In other words, the efficiency

of resource use may not reflect the cost to clients since this will be affected by the cost of capital to the

organisation concerned, and the profit margins built in to any contract. The perceived opportunity cost of capital

investments would be expected to be rather different in the cases of not-for-profit entities, in-house operations

and commercial entities. To level the playing field in terms of accounting for such outlays risks downplaying

the potential strengths and weaknesses of different types of operator within different niches of the waste

management industry. Lastly, looking only at costs and revenues makes it difficult to know to what extent a

scheme is ‘successful’ in the sense either of its not making a loss, or its ability to return an acceptable profit

(dependent upon the organisation’s aims and objectives).



There are also questions of detail as to how one should treat expenditure on, for example, scheme promotion.

This may be an annual expenditure, yet it might be supposed that as with capital expenditure, money spent on

promotional activity might be front loaded (note that the Audit Commission asks respondents to its

questionnaire to treat the initial distribution of bins etc. in the same way as expenditure on buildings, so that

straight line depreciation is treated as occurring over 25 years). Arguably, once schemes are more mature, the

number of new participants would drop off over time. More empirical data would be useful here, but some

experience suggests that sustained information campaigns do increase participation rates and diversion. In more

mobile populations, presumably this is even more important. Better knowledge about why people do and do

not participate might facilitate targeting of efforts at the more ‘reluctant’ households.11

For the purposes of this analysis, we have followed Audit Commission guidelines in respect of capital

expenditure, effectively annualising this on the basis of straight-line depreciation over a period of time

dependent on the nature of the asset.

Note, lastly, that we are concentrating on the kerbside element. The authorities examined have bring schemes in

place as well, so that the rate of recycling is greater than that implied by the kerbside scheme alone.

4.1.1 Scheme 1

This scheme operates in a fairly small district authority in a rural area. The kerbside scheme operates fortnightly

collection and covers just under half of the households in the authority. The coverage is confined to more urban

areas. It is run by a not-for-profit organisation. Materials collected in boxes include paper and board,

aluminium, steel, glass and textiles using 7.5-tonne multi-caged vehicles. There is no MRF, but a fork-lift

truck and baler are also used by the organisation. On the basis of counters used on the vehicles, the estimated

participation rate is 40-45% with a put-out rate estimated at 34%. The rounds are typically 15 miles in distance

and the journey from the round to the depot may be 20 miles.

The amount of materials collected is 104.6kg per household covered, or 246kg per participating household

(taking this to be 42.5% of total covered). By weight, the majority is paper and board (63%) and glass (28%).

6% is steel, 2.5% is textiles and just over 0.5% is aluminium. The total costs of the scheme are £120.30 per

tonne of material collected, or £12.58 per household covered, or £29.60 per participating household. Home

composting is also promoted by the authority through a subsidised bin scheme operated through local garden

centres using a voucher system.

On the revenue side, the scheme receives positive revenue from all materials collected. In addition, it receives

recycling credits, and grant revenue from local businesses. When these are taken into account, the net cost

figures for the scheme are £72.33 per tonne collected. This translates to £7.56 per household in the scheme and

£17.80 per participating household. In some parts of the UK the concept of recycling credits has limited

meaning. In the absence of recycling credits, the costs would be £100.70 per tonne, £10.53 per household

covered and £24.77 per participating household. However, for unitary authorities, relative to the counterfactual

scenario in which materials are simply delivered to landfill, recycling saves on disposal as well as collection

costs.

11

Some theories of behavioural change suggest that decisions to act in a particular way may occur through staged processes which may

be more or less reversible at any given stage. Hence, for different individuals, sustaining educational messages will have greater and

lesser significance.



In terms of the authority’s overall performance, we estimate an 11.5% recycling rate (entirely down to dry

recyclables) with the kerbside scheme contributing roughly half this. The kerbside scheme’s capital costs are of

the order £120 per tonne and variable costs are of the order £86 per tonne.

4.1.2 Scheme 2

This kerbside scheme is a weekly collection of mixed recyclables taking place in an urban area and it is about

three years old. All households are covered in the authority and the scheme is run by a not-for-profit

organisation. Materials are put out in boxes and collected in purpose built trucks which typically carry 4-tonne

loads. The scheme also promotes home composting, on which compositional analysis has been carried out in

the area.

106kg per household covered are collected. Since there are no estimates of participation, using high

participation rate assumptions (70%),12

one can estimate a per participating household figure of 151kg. The

materials include paper and board, aluminium, steel, glass and other materials. By weight 69% is paper and

board, 25% is glass, 3% is steel, 3% is other and 0.4% is aluminium. No plastics are collected. Home

composting is promoted.

The distance travelled on rounds is between 1 and 3 miles, the mean distance to the round being 9 miles. The

revenue from materials sales is of the order £32 per tonne in total, but the costs of haulage are of the order £7

per tonne. The net revenue figures, therefore, are of the order £25 per tonne. The scheme benefits from neither

recycling credits, nor (in its view) the PRN system. The gross costs of the scheme are £85 per tonne, or £9 per

household covered. Net of revenue, these translate into £60 and £6.32 respectively.

4.1.3 Scheme 3

This is a similar scheme to that in the authority mentioned in Scheme 3 and is also in an urban area. It has been

operating for over two and a half years. Again, all households are covered. Materials collected, by percentage

contribution (again, no plastics are collected) are paper and board (66%), glass (27%), steel (3%), other (3%) and

aluminium (0.4%). The scheme collects 102kg per household covered, or making the same assumption as made

above (70% participation), 145kg per participant household. No plastics are collected, but again, home

composting is promoted. 4 tonne loads are collected on rounds of between 1 and 3 miles, with the mean

distance to the round being about 4 miles.

The revenue from materials sales net of haulage costs is of the order £27 per tonne. The scheme does not benefit

from recycling credits or (in its view) the PRN system. The scheme’s gross costs are £95 per tonne, or £9.65

per household covered. The costs net of revenue are £67.93 and £6.90. There are no figures for the overall

participation rate (hence the assumptions above).

4.1.4 Scheme 4

This kerbside scheme is a weekly collection of mixed recyclables and occurs in an urban area. It has been

operating for just over one year. All households are covered in the authority and the scheme is run by a not-for-

profit organisation. Materials are put out in boxes and collected in purpose-built trucks, some of which carry 4-

tonne loads, and others of which carry 2.5-tonne loads. The distance travelled on rounds is between 1 and 3

miles, the mean distance to the round being 9 miles.

12

We use this high (for the UK) participation rate assumption so that the ‘per participating household’ collection figures are not over-

estimated. On the other hand, calculating costs on this basis makes them look lower than they would be using what are probably more

realistic assumptions concerning participation.



75kg per household covered are collected, or using the 70% participation assumption, 107kg per participant

household. The materials include paper and board, aluminium, steel, glass and other materials. By weight 67%

is paper and board, 27% is glass, 4% is steel, 4% is other and 0.4% is aluminium. No plastics are collected.


from recycling credits or (in its view) the PRN system. Gross costs are £100 per tonne or £7.50 per household.

Net of revenue, costs are £87.83 and £6.59 respectively.

4.1.5 Scheme 5

This scheme involves a local authority in partnership with the not-for-profit sector and it collects across an

authority that is partly urban and partly rural. There is a weekly collection of mixed recyclables, in green boxes,

for the majority of households. This is supported by mini recycling centres at multi-resident locations, for

which wheeled bins with special frames are supplied that are emptied as necessary (typically, every two weeks).

About a quarter of the households in the authority are served by monthly collections in residents’ own bags.

The different approaches are of differing vintage. The first started in 1993, the second in 1994, the third (the

oldest) in 1985.

The bulk of collection (the weekly one) takes place in purpose-built 7.5-tonne multi-caged vehicles, each with

two operators. Participation in this scheme is estimated at 70% on the basis of random streetblock monitoring.

The scheme also promotes home composting (through, amongst other things, subsidising bins), and one tenth

of all households had purchased these as of Spring 1999.

The collected materials are taken to a MRF or to depots. The MRF includes sorting equipment for separating

out metals and plastic bottles only, so its capacity is well below the total collection tonnage due to the scheme.

In addition to the authority’s materials, the MRF accepts some materials from other organisations.

The scheme collects 110kg per household, or 157kg per participating household. Of this, 67% is paper and

board, 26% is glass, 4% is steel, 0.5% is aluminium, 0.4% is plastics and 2% is other materials. The scheme

benefits from both recycling credits and from revenue from material sales. This was the only scheme in the

survey that even suggested that it might be benefiting from the PRN scheme, and then only obliquely,

commenting that any such revenue was ‘not separately identified in payments.’

The scheme’s gross costs are £123.23 per tonne, or £13.51 per household covered, or £19.30 per participating

household. Net of sales and recycling credits, the costs are £51.66 per tonne, or £5.66 per household covered, or

£8.09 per household participating. The treatment of capital costs is not entirely clear in the scheme. We have

made an estimate as to the costs which would prevail if we account for these. The costs are the £68.82 per

tonne, £7.54 per household and £10.77 per participating household.

4.1.6 Scheme 6

This kerbside scheme is a weekly collection of mixed recyclables. It takes place in an area of fairly dispersed

population and it has been operating for just over one year. All households are covered in the authority and the

scheme is run by a not-for-profit organisation. Materials are put out in boxes and are collected in purpose-built

trucks, some of which carry 4-tonne loads, and others of which carry 2.5-tonne loads. The distance travelled on

rounds is between 3 and 8 miles, the mean distance to the round being 14 miles.



142kg per household covered are collected, or assuming 70% participation rates, at least 203 kg per participant

household. The materials include paper and board, aluminium, steel, glass and other materials. By weight 65%

is paper and board, 29% is glass, 3% is steel, 3% is plastics and 0.3% is aluminium. This is one of only two

schemes collecting plastics in our survey.

The gross costs of this scheme are of the order £140 per tonne. The revenue from materials sales net of haulage

costs is of the order £11 per tonne so that net costs are of the order £129 per tonne. These translate into per

household costs of £19.90 and £18.29 respectively. The scheme does not benefit from recycling credits or (in

its view) the PRN system.

4.1.7 Scheme 7

This kerbside scheme is a weekly collection of mixed recyclables. It takes place in an urban area and it has been

operating for just over one year. All households are covered in the authority and the scheme is run by a not-for-

profit organisation.

86kg per household covered are collected, or 123kg per participant household under an assumption of 70%

participation. The materials include paper and board, aluminium, steel, glass and other materials. By weight

63% is paper and board, 33% is glass, 2% is steel, 0.3% is aluminium and 1% is other materials.


from recycling credits or (in its view) the PRN system. The gross costs are £105 per tonne, or £9 per

household. The figures net of revenue are £91 per tonne and £7.78 per household.

4.1.8 Scheme 8

In this scheme run by a District Council, 25% of all households are covered by a green bin service collecting

organic materials on alternate weeks, with residuals being collected in the other weeks. The rationale for this is

that the Council incurs no extra collection costs. The same vehicles, crew, etc. are used. As such, the

incremental costs relative to the weekly residuals collection are low. Indeed, the questionnaire reported these as

zero, but the scheme has a cost in that each household covered by the scheme now has two 240-litre wheeled

bins where previously they only had one. In addition, households have been given 15-litre kitchen bins. There

may also have been costs in terms of scheme promotion (if only in the trial phase) and education (if only

through stickers on bins informing residents as to what should and should not be put in the bins - paper has

been excluded on the basis of trials).

Note this scheme was initially tested on 500 households. A survey of residents at the end of the year trial found

91% of the 72% responding were in favour of the scheme’s continuation. The scheme collects 500kg of waste

per household per annum. This is some 38% of the average household’s waste stream in the district (and this is

of significance in respect of the Landfill Directive Article 5 targets).

The District also runs a Blue Box scheme for mixed dry recyclables. This scheme covers 1,600 households. The

collection figures are difficult to estimate since the authority does not collect data separately for the kerbside

fraction and the recyclables coming from bring sites. The kerbside and bring scheme combined give figures, per

household, of 236kg per household across the District (or 18% of the average household’s arisings). The

estimate from the authority for the quantities put out in the Blue Box is 400kg per household per year.



This is an unusual scheme in that it appears to force householders, by the nature of the collection system, to

participate in the scheme. The materials are sent to a combined composting plant / MRF site.

4.1.9 Scheme 9

This scheme is a composting scheme run by a private company. Unfortunately, in respect of this scheme, whilst

we have good information on the waste treatment facility itself, we have not had a response from the authority

concerning the collection of waste for treatment. As such, we can only properly comment on the costs of this

stage of the process. However, we do know the following about the collection scheme.

Organic waste collections are made on a fortnightly basis on separate dedicated rounds. These rounds are larger

that the refuse rounds covering up to 6,000 houses, and require a driver and between 1 and 3 loaders depending

on the social demographics of the area. Householders put out the appropriate wastes in special paper sacks,

which can be shredded along with their contents, prior to composting. Information is printed on the bags

explaining what can and cannot be put into the bag. A calendar is issued to households with an order form for

new bags and/or compost on the reverse side. When the bags are delivered the householder is issued with a new

calendar. Leaflets and promotional information are also available from council offices and sign boards are

displayed at civic amenity sites. The composting takes place at a closed landfill. The collection stops from

December to January as there is insufficient material to collect.

The council purchases bags at 20p. These are sold to the public, by the contractor, at a fixed price of 40p each,

or £3.50 for ten. This price includes delivery of the sacks, which enables householders to make use of the

service without needing access to transport. When costs for printing, stocking, storage, delivery and labour, are

taken into account, there is little profit in bag sales at this price.

One advantage of this type of collection is that, as it is a ‘buy-in service’, the material collected tends to be of

high quality, reducing contamination problems. The council does however recognise that this limits

participation. There are also complications in accessing the service in that people must purchase the sacks and

be aware of collection dates. Even so, collection has grown at 20% per annum since the scheme started

operating (in 1993). There is some variation in the collection round costs as the number of loaders employed on

each vehicle changes according to the demand for the service. As the scheme has grown the cost per tonne

collected is believed to have come down from £130 per tonne to about £75 tonne.

The composting plant is run by a Local Authority waste disposal company (LAWDC). It is a 10,000-tonne

open-air windrow facility. Gross costs, including depreciation, are £15.30 per tonne. Net of recycling credits

and the sale of recovered materials, the costs are £10.64. The average sale value of the end-product is

approximately £3 per tonne. This, however, does not include revenue from gate fees. The plant accepts 620

tonnes from the council’s collections, 6,500 tonnes from Civic Amenity sites and an estimated 400 tonnes from

landscaping firms (7,520 tonnes in total). The gate fees it charges vary according to the customer and where the

waste has come from. However, these are rarely below £12, and for commercial / industrial wastes, may be as

high as £20.

The district sees the organics collection as a valuable way of reducing queues at these sites, which in turn helps

to prevent fly tipping.



4.1.10 Scheme 10

This scheme is operated by a private sector company in a unitary authority. All households are covered. There is

a separate collection of paper and board fortnightly that is run as a box scheme. The vehicles used are 17-tonne

refuse vehicles and 10-tonne caged lorries. In addition, there is a collection for putrescibles that takes place

concurrently with the collection of residuals. This weekly collection employs split-bodied refuse collection

vehicles (RCVs) and collects materials in 10-litre buckets and 55-litre boxes. Home composting is also

encouraged through a subsidised bin scheme, though this is believed to have a minimal impact on the waste

stream.

The composting scheme collects approximately 23kg per household covered, or about 108kg per household

participating. The collection part of the scheme costs some £36 per tonne, or £0.83 per household, or £3.88 per

participating household. Half of these collection costs are incurred in publicity for the scheme including

advertisements in the press, on the radio and in newsletters (this is part of the contract under which the scheme

is run).13

The low collection costs (per household) are based upon the fact that at the currently low levels of

collection, the accounting for labour, maintenance and fuel has assumed zero incremental cost. The capital costs

have been based upon straight-line depreciation over 5 years (which is the replacement period for the vehicles) of

the incremental capital costs of the split-bodied vehicles over and above the costs of the standard RCV. If one

accounts for labour, fuel and maintenance through pro-rating costs to the compostables collected, then even if

the costs of publicity are removed, the costs rise to £52 per tonne, or £1.60 per household, or £7.52 per

participating household.

The plant is an in-vessel composter of 15,000 tonne capacity. It receives 8,025 tonnes per annum, including

1,400 tonnes from kerbside collection. The capital outlays included the construction of buildings and a

substantial outlay on plant and equipment. This makes this scheme quite capital intense, and net of £4,000 in

sales revenue, the costs are some £86 per tonne. Depending upon how one accounts for the collection, therefore,

the total costs of collection and composting are between £122 and £138 per tonne.

The paper collection scheme collects 40 kg per household covered, and 144 kg per household participating. Per

tonne collection costs are £147, whilst the costs per household are £5.89, or £21.13 per participating

household. The nature of the contract obliges the company to maintain the paper collection even though under

current market conditions the paper is taken for use in the refuse-derived fuel (RDF) plant. Clearly, this is a

potential drain on resources, but we were told that if the market situation improves, the operators might

consider selling the paper to mills. There is a danger, of course, that this may have a negative impact on

residents’ views in respect of participation.

The RDF plant receives all residuals, as well as 4,950 tonnes of commercial waste. Separation of materials

leads to recovery of steel and aluminium, and these derive revenue. Approximately half the waste arriving at the

RDF plant is rejected and sent to landfill. Of the other half, some 25% is driven off as moisture. The remaining

pellets are used in power stations, and the users have the right to PRNs. There was no indication that revenue

was derived from the sale of the pellets, which would be strange given the value both of the PRNs and of the

material itself. The costs net of sales (we have attributed revenues from aluminium and steel sales) excluding

depreciation are of the order £45.90 per tonne of fuel produced. Allowing for depreciation, these are likely to be

of the order £60 (the capital costs were given to us in one figure, i.e. the components were not split out

according to their replacement lives).

13

There has been no specific attempt to understand the effectiveness of the information campaign in, for example, increasing

participation rates etc.



Whilst this approach may seem somewhat expensive, it has to be viewed in the context of the shortage of

landfill in the immediate area. There were three landfills operating in 1984, but two of these closed in 1985 and

1988 respectively. The council has been extending the remaining landfill, which is now believed to have void

that may last for 12 years. There is plenty of scope for increased composting given the plant capacity and it is

quite clear that as participation in the composting scheme increases, then the costs could fall quite dramatically

(since a high proportion of these are related to depreciation on an existing capital investment). We estimate that

the costs could fall by as much as £30 per tonne if the plant runs close to full capacity. At that point, the

recovery would be equivalent to close to around 20% of municipal waste, perhaps as much as 33% of all

biodegradable waste. Given that much of the paper will be used in the RDF plant, which also has plenty of

spare capacity, this particular authority is well placed to meet Landfill Directive targets 1 and 2 (i.e., depending

on the rate of growth of the biodegradable waste stream, the approach will potentially see the authority well

placed well into the next decade since the final Landfill Directive target will not bite until 2020).

For Table 14: Comparison of Kerbside Dry Recyclables Schemes see tables.pdf

4.2 Comment on Performance and Costs

Table 14 compares the dry recyclables schemes for which we have good data. Schemes 8 to 10 have been

excluded since they do not include dry recyclables collection. The following key observations can be made in

respect of performance:

• The schemes collect between 75kg and 142kg of material per household. Between 91% and 96% of this is

paper and board and glass, with the former alone accounting for two-thirds of all collected material (by

weight).

• Per participating household, the range is from a low of 107kg to a high of 246 kg. The former is probably

an underestimate since it is estimated on the basis of an assumed participation rate of 70% (which is about

as high as one sees participation rates go across even small areas in the UK at present).

In respect of costs, the key observations are:

• The variation in gross costs per tonne is from £85 to £140. This is broadly in line with those quoted for

schemes in the previous chapter, though at the lower end of the broad ranges quoted for separate kerbside

schemes. The highest costing scheme is in a rural area. Generally, the figures are higher than those reported

for the United States by De Rose (u.d.).

• Net costs vary more than gross costs since the effects of materials revenue and the payment or otherwise of

recycling credits has important effects on the schemes. The variation is from £52 per tonne to £129 per

tonne. It could be argued that even where recycling credits are not paid, some ‘cost-saving’ element could

be attributed to schemes because they reduce not only disposal costs, but also those for collection. Note

that the highest net cost is seen in the scheme that has highest gross costs. This scheme receives relatively

low materials prices under its contracts, so this compounds the effect of the higher gross costs in the less

densely populated area covered.

Other more general comments include:

• Where a MRF is used, this introduces an extra capital component. Some operators appear reluctant to use

MRFs suggesting that this introduces a new capital item into the budget. The view was also expressed that

unless well specified, the use of MRFs can involve significant down-time. Also, where separation occurs



on vehicle, the merit of any MRF might lie in separation of plastic polymers were they to be collected.

This then raises questions as to whether material throughput could justify the investment concerned. More

generally, the debate about the pros and cons concerning the use or not of a MRF might also include

considerations regarding the desirability of ‘cherry picking’ materials depending upon prevailing prices, and

the collection logistics. The latter might relate to how much time is spent separating materials on-vehicle,

and whether it is deemed to be undesirable to do this in congested areas / on narrow streets. Interestingly,

in Brussels, neighbourhoods are often asked to have their cars off the street in periods when waste

collection will take place.

• The costs of collection, as imputed by the organisations themselves, depend upon how they collect the

materials and whether they have some responsibility for collection of residuals as well. Where they are

operating alternate collections of compostables or where split-bodied vehicles are in operation, the

incremental costs of materials collection is marginal. More generally, this leads to an interesting point in

that all the recycling schemes, to the extent that they are collecting significant tonnages of material (and

they are), are reducing the costs of residuals collection. It is true, as MEL (1999) state in their report, that

recycling in the UK has failed, as far as we can see, to curb the growth in the residuals stream, but at the

level of individual authorities the same may not apply. It is also true, therefore, that recycling schemes are

more than ‘second’ collecting rounds. They are collecting significant amounts of material that would

otherwise be collected on the residuals round, and therefore, there are savings in resources devoted to

collection. Note that the capital costs of kerbside schemes that do not include a MRF are bound up in two

elements, the vehicles and the depot. Once a scheme has matured, the former element becomes a cost which

would be incurred anyway in collecting residuals if the scheme did not exist. Removing the recycling

scheme is then a far from cost-free process. One scheme manager we spoke to believed that removing the

scheme would imply a requirement for two extra RCVs costing something between £230-300,000. With no

extra costs considered (for labour, for example), then accounting for this through straight line depreciation

over five years, this amounts to something of the order of 15-20% of the net (after sales) operating costs of

the scheme.

• Most obviously apparent is the fact that schemes collecting almost identical materials (in terms of

percentage composition) actually receive quite different revenues (per unit collected) for the materials. The

most expensive scheme noted that they were using very high specification vehicles that they had more or

less been required to use in the context of the specific contract. For similar schemes, the revenue ranges

from £11 to £25 net of haulage costs. The issue of recycling credits also affects overall financial

performance as quoted by a scheme, although the overall costs of a waste management system should be

indifferent to the payment or otherwise of such credits.

• Dry recyclables collections are collecting almost identical materials in terms of composition (see Table 14).

• Dry recyclables collections are not always collecting only dry recyclables. They are certainly not collecting

only packaging materials and newspapers and magazines. They also collect clothing, and importantly, some

of them collect other materials purely out of an appreciation that they are hazardous when disposed of in the

wrong way. The latter is a very important function. Whilst the Packaging Directive and the Landfill

Directive have thrown the spotlight on packaging and biodegradable wastes, the potential environmental

impacts of disposing of hazardous materials incorrectly should perhaps attract rather more concern than is

the case at present.14

• Few of the schemes studied are collecting plastics. The relative paucity of plastics collection reflects:

14

The arguments for producer responsibility are particularly strong in the case of household hazardous wastes. A requirement to collect

these separately would certainly incur costs. It could also alter perceptions among households in respect of the desirability of source

separation of wastes in the general case.



1. The mass of plastic collected in a vehicle is low, giving rise to higher collection costs per unit of

saleable material.

2. The fact that since plastics tend to be separated by polymer for reprocessing, and since this is difficult

to do on the collection round, the collection of plastics incurs added costs, either in (quite skilled)

labour or in the operation of a MRF. The incremental costs associated with plastics collection at

kerbside, therefore, can be considerable. Partly for this reason, some recycling schemes (and this is a

general comment rather than one applicable to the schemes interviewed) have looked at the feasibility

of collecting PET bottles only.

4.3 Comment on Recycling Rates

A lot of discussion has taken place recently concerning recycling rates, and what can and cannot be achieved in

this regard. In particular, there have been questions raised as to where any limits might occur. M.E.L.’s (1999)

work for the Energy from Waste Association has claimed to establish a ‘maximum’ rate of recycling of dry

recyclables of 15% for local authorities in the UK. We would dispute any strong interpretation of that report’s

conclusion. There is a major difference between what one can regard as ‘current best practice’ and the point at

which any ‘maximum’ might lie. Also, one could easily lose sight of the fact that the report itself, though it

purports to have found ‘maximum performance’, does so with the qualifier that the figure of 15% relates to dry

recyclables only ‘at the level of the local authority as a whole given current operational, financial and

legislative conditions.’ In other words, were any of these conditions to change, the hypothesised 15% ceiling

might change also.

Changes concerning, for example, a requirement to design for recycling would have an impact on what any

theoretical maximum, such as one can be stated, might be. It is interesting in this regard to quote the important

Global Assessment by the European Commission. In the context of waste:

‘Priority in the future will need to be given to promoting an active product policy in order to make products

recyclable from their design phase as well as further preventing waste generation’ (CEC 1999),

and in the context of eco-efficiency,

‘An Integrated Product Policy should address the entire life-cycle of production and consumption, and be

based on a mix of instruments – such as labelling and eco-design, links to the Community’s Environmental

Management and Audit Scheme (EMAS), greening of public procurement and product standards, and product-

related taxes – thus addressing the whole product chain including the production, use, distribution,

consumption and waste phase of products’ (CEC 1999).

More generally, one can make the following points:

• Fundamentally, the amount of material which can be extracted at kerbside relates to waste quantities and

waste composition. We know that this varies considerably across authorities, and across different areas

within authorities. A glance at the results from Ecologika’s (1998) review of seventeen kerbside refuse

composition studies shows enormous variation in total wastes, as well as considerable variation in the split

by category of material (recyclables, putrescibles, and those suitable for reuse, reduction or refuse

collection). Immediately, one is faced with the notion that whether expressed in terms of absolute



quantities or percentages, the idea that there is a standard, uniformly applicable maximum rate is simply

nonsensical.

• Also, the amount actually extracted will depend upon which materials are collected. The range of these

tends to be a sub-set of what is possible. Note that according to the Ecologika (1998) study, the average

contribution of putrescibles and recyclables to the total was 80.7%. The figures in A Way With Waste

(DETR 1999a) suggest something of the order 60% as a maximum figure. Note that both these figures

appear to relate largely to bin waste only (and the average Ecologika figure for the total is 745 kg per

household per annum, in line with other studies such as MEL (1999)). This is only 60% of the total MSW

stream, and therefore an unsound basis upon which to extrapolate to the total MSW stream (as was implicit

in A Way With Waste). The more materials one tries to collect, the more one will extract from the waste

stream. To the extent that recycling is a dynamic industry (and experience elsewhere suggests that it is), the

idea of limits seems artificial because it implies stasis.

• At a time when UK waste management has to change anyway, it seems less than useful to be seeking to

understand where any maximum might lie on the basis of how the world looks today. The German

experience suggests that, in the context of the institutional structures in place, the recovery of materials can

provide a means of gaining control over input material supplies (without necessarily exposing the

companies concerned to price fluctuations in commodity markets).

Returning to discussions concerning maximum rates of recycling for dry recyclables, the figures for recycling

per participating household are of interest. MEL (1999) suggest that RCV waste averages

14.2kg/week/household, or 738kg/annum/household. If one were to add recycled materials, one might estimate

non-bulky and non-garden household wastes at about 800kg per annum on average. The highest rate of

collection per participating household is 246kg/annum/household, or almost one third of this waste. This

scheme does not operate authority wide. The best one that does, however, collects 140kg per household, or

conservatively (using the 70% participation estimate), 209kg/annum/participating household. This is equivalent

to slightly less than 25% of this waste.

Work we have done elsewhere suggests that in England and Wales, the average figure for all municipal waste is

of the order 1,200 kg/annum/household. If all participants were recycling at the rate of 246kg per household, the

kerbside scheme alone could attain a recycling rate of 20.5% of all municipal wastes. At

209kg/annum/household, the rate is closer to 17%. Note that there is a point which could be made concerning

the wisdom of calculating the rate of kerbside recycling relative to the total amount of MSW. It makes more

sense to consider how much of what could be extracted is being extracted. To that end, the figure of 25%

calculated above seems a more accurate reflection of what is actually happening.

We know that some of the municipal waste is waste that is not from households (it is from commercial

premises, and occasionally, from industrial ones), and not all of this will be sorted. Without knowing in more

detail the extent of the non-household municipal component, and its fate, one cannot know what would be the

recycling rate attainable were all households recycling at these rates in a situation where only household waste

was collected by local authorities. This is an interesting question since the Landfill Directive targets apply to

municipal waste (not just household waste) and many authorities may find that a cheap way of moving closer to

targets would be to stop collecting non-household municipal waste (though some authorities may balance this

view against the impacts of the loss of a revenue stream).

One hundred per cent participation may seem unrealistic in current contexts, but some other European countries

do not see it as such (composting in the Netherlands, Austria and Germany, for example - see Federal Ministry



for the Environment, Youth and Family Affairs 1999). It would clearly require changes in either formal (i.e.

legislative) or informal institutions (e.g. social norms of behaviour). Such changes are likely to take time, and

certainly at the informal level, better targeting of educational resources would be required. At present,

educational messages do not appear to be tailored to increasing rates of participation. A blanket approach is

more common. It may well be that different types of message appeal to different groupings.

5. EXTERNAL COSTS OF WASTE MANAGEMENTThis chapter reviews existing literature concerning the external costs of waste management. Chapter 6 takes

forward some of the lessons learned to generate what are still highly-flawed estimates of the external costs of

waste management options.

5.1 Linear and Circular Flows of Materials

Unlike options such as landfill and combustion, which are final treatments, recycling (and reuse) result in

material being returned to production processes, either in closed loop processes (where the material is made into

the same, or similar product from which the material arose) or in processes where the waste material is

fashioned into something completely different. This means that, for the economy as a whole, there is a reduced

need for primary extraction, and hence there is a reduction in the environmental effects from the production,

processing and transport of the raw material. To accurately assess the difference in impacts of the options

recognition needs to be given to the reduction in the overall level of primary production, but also to the increase

in environmental impacts associated with recycling itself.

Recovery of energy from waste can do the same thing indirectly by reducing the need for consumption of energy

sources, but it can do this only once. Although recycling cannot occur indefinitely (for example, owing to

shortening of fibres in newsprint recycling), the recycling can usually take place more than once. There is,

therefore, an element of circularity in the recycling process that distinguishes it from landfill and incineration.

For the sake of convenience only, landfill and incineration will be referred to henceforth as ‘linear waste

management options.’15

The change in environmental impacts as a result of recycling comprise not just those associated directly with

the alternative materials processed themselves but also those associated with primary extraction, transport and

production of the virgin material. Only the environmental impacts arising between the point at which the

primary and secondary materials are combined, and the wastes are separated, are common to the linear and

circular pathways, and need not be assessed for the purposes of a comparative analysis.

An analysis comparing two linear waste management options for instance incineration compared to landfill

would only need to assess the post-consumption costs of the two options since the pathways up to that point

are entirely common between the two options. Neither landfill nor incineration offers any means of reducing the

need for primary extraction of the material concerned though both may lead to recovery of energy. Figure 2

below makes this point.

15

Note however that some have proposed storing materials in landfills until suitable uses / technologies can be found to justify ‘mining’

them. Whilst this may seem fanciful, past consultations have shown that at least one power company has considered mining its own

landfills for previously landfilled ash which it is now able to sell commercially.



Figure 2: Contrast Between Linear and Circular Waste Management Options

Extraction Production Use Recycle

Disposal / Energy Recovery

If 30% of the material that is used is recycled then only 70% of the amount of raw material (or slightly more

allowing for wastage in materials recycling) needs to be extracted and processed into material for use in final

consumption. This reduces the rate at which primary resources are run down, and reduces the disruption of land

surface and water pollution caused during the extractive process for the economy as a whole.

Of course recycling also has environmental effects. Energy is consumed in separating, transporting, cleaning and

processing the recycled material to the point at which it is combined with the primary material stream.

Greenhouse gases and particulates, as well as dioxins are emitted in the process. However, if the process of

primary production is more energy intense than secondary production, recycling reduces the rate of energy

consumption.

It is, as we shall see, a characteristic of waste management options that (apart from minimisation) no option

performs better than another on all accounts. Because that is not the case, the decision as to which option is ‘the

best’ requires some way of making decisions regarding waste management. One such way is to trade off the

different pros and cons of the approaches through economic valuation of the different effects. This is, it should

be added, only one of a number of approaches to ‘valuation’, some of which have been summarised in Powell et

al (1995). It is important, at this point, to emphasise that waste minimisation is beyond the scope of this

analysis. We are, therefore, ignoring some very interesting debates about materials consumption in modern

lifestyles (we are implicitly assuming consumption continues as today).

5.2 Review of Existing Studies

A number of studies have sought to address the external costs and benefits of different waste management

options through the combined process of LCA and economic valuation.

In what follows, we make much of the inadequate treatment, or appreciation of, uncertainty in the scientific

domain where economic valuation is concerned. It is important in this respect to distinguish between

uncertainty, risk, and error. Uncertainty is frequently used in contexts of description of exposure to hazards, and



in the context of measurement. In the former case, it is worth juxtaposing the terms ‘risk’ and ‘uncertainty’. A

risk can be considered as a quantifiable probability of a particular consequence occurring. Throwing the number

‘one’ on an unweighted die can be ascribed a probability owing to the intrinsic characteristics of the die. The

probability of a road accident occurring can be ascribed a probability on the basis of actuarial approaches using

statistical data. When there is less in the way of past experience to guide us, or where a particular problem raises

new questions which are not amenable either to some fundamental law, or to actuarial analysis, then uncertainty

is likely to characterise our understanding of that problem. We are frequently not, in these cases, able to know

exactly what the outcomes of particular processes or changes might be, still less to know the probability of

certain specified outcomes. Uncertainty is the battlefield on which scientists do what scientists should do, that

is to say, dispute in objective fashion the available evidence and interpretations thereof. It can be argued that we

are slowly beginning to recognise that this is the normal state of affairs in science rather than an occasional

exception to a world of much-cherished certainties.

Where measurements are being made, errors can arise owing to imperfections in measurement instrumentation,

or to statistical variation where observations are distributed around a particular value. Sometimes, the term

‘uncertainty’ is used interchangeably with what appears to be error. On the other hand, in the context of

valuation, there are certainly disputes as to the epistemological basis for certain approaches to measurement of

the value of, say, a species under threat of extinction. For the purposes of this study, one could interpret this as

methodological uncertainty. Again, as we shall see below, the debate around how to value life might be said to

be dogged both by scientific uncertainty as well as those of an epistemological nature which translate into

methodological uncertainties. Application of these methodologies, even if one was to accept their

methodological credentials, are also likely to incur errors in measurement. Attempts to value, for example, the

effects of dioxin on human health (see AEA 1997) face the rather daunting task of having to deal with both

uncertainty and error where valuation is attempted.

As regards the conventions used in what follows, we are reporting externalities. Negative numbers represent

negative externalities (disbenefits) whilst positive numbers are used to represent positive externalities (benefits).

5.2.1 CSERGE et al (1993)

This work is particularly well known as it was used in the run-up to the introduction of the Landfill Tax and

was influential in the decisions ultimately made concerning the level at which the tax should be set. The aim

was to understand the externalities associated with landfill and incineration. The study made clear that the

potential benefits which might accrue if materials were recycled were outside the remit of the study rather than

considered unimportant per se. Also, although the study undertook a review of work carried out elsewhere in

assessing disamenity associated with landfills, these were omitted in the quantification of the external costs,

though the study noted these could be ‘significant’ and therefore their omission would affect the results.

Brisson and Pearce (1995) later suggested that employing benefits transfer techniques from US studies would

make a figure of £160 (about £173 in 1999 terms) per household per annum not unreasonable for houses within

a 4 mile radius of a landfill (which they also suggest is more or less equivalent to 3% depreciation on house

values).16

The results of the study are as indicated in Table 15. For Table 15: Externality Values for Landfill and

Incineration (£/tonne waste other than disamenity) see tables pdf. Note that the study sought to measure the

16

The lack of consideration of disamenity is more comprehensible when one sets the study in its policy context. In informing the level of

a landfill tax being levied on waste, the fixed rather than variable nature of the disamenity externality could be used to justify not

including the disamenity externality.



external costs associated with typical tonnes of waste being landfilled or incinerated in different types of plant,

and then averaged externalities across types of plant to arrive at a UK figure. It is clear from these results that

the relative rates of energy recovery from the two facilities are crucial determinants of their external costs as

reported here. This is due to the significance of the pollution displacement assumed to occur as a consequence

of that recovery. The study assumed recovery of 664 kWh/tonne MSW for incineration and only 79 kWh/tonne

MSW for landfills with energy recovery. We return to this issue in the next chapter.

The external costs that were covered were:

• For landfill: emissions of CO2 and CH4, casualties as well as CO2, NOx and TSP from transport, and

reductions in CO2, NOx, SO2, TSP and CH4 from displaced energy sources. A limited attempt was made

to estimate leachate externalities whilst acknowledging the limitations of the exercise.

• For incineration: emissions of CO2, NOx, SO2, TSP, casualties as well as CO2, NOx and TSP from

transport, and reductions in CO2, NOx, SO2, TSP and CH4 from displaced energy sources.

As such, omitted variables in the valuation work (apart from disamenity) were:

• For landfill: arguably, a full consideration of leachate (BOD, COD, heavy metals and SS), particulates

from landfill gas (burned and not burned), and CFCs. There is some discussion of the role of landfills in

causing birth defects but the evidence can be considered to be at the level of ‘not proven’ at present. It is

important to note however that some modern engineered landfill sites (e.g. Nant y Gwyddon and Trecatti)

are the focus of some of the keenest concern regarding health effects in the UK.17

• For incineration: CO, and air toxics such as dioxins, and heavy metals. In addition, to the extent that flue

gas is being ‘cleaned’, the fact that such cleansing simply tends to move the pollutant from one medium to

another (from air to water or land depending on the cleaning technology) means that the not quantifying

externalities associated with discharges to water or land implies that some pollutants whose impact could

be significant escape the analysis. More generally, the benefits of flue gas cleaning would be exaggerated

under such an approach. Lastly, external benefits associated with the recovery of metals were ignored.18

Note that the study did not ignore these since many were discussed in considering the emissions from landfill

and incineration. The problem arises in seeking to quantify impacts associated with these emissions.

The ‘omission’ of air toxics is an important one, as we shall see, as indeed might be the lack of accounting for

materials recovery. With regard to the former, the report noted that ‘This component of economic damage is

therefore left unvalued in the current exercise, with the balance of probability being that such a value would

be close to zero or zero’ (our emphasis). The statement is based upon the views of the 1993 Royal

Commission on Environmental Pollution report on incinerators that comply with (the then) Her Majesty’s

Inspectorate of Pollution standards. There is a respectable body of science which contradicts this view. The

wording of the statement is also somewhat misleading since the truth content of statements which refer to

matters of scientific uncertainty will not, in the general case, be amenable to any ‘balance of probability’ (see

above). There may be a majority view among experts, but experts are not always correct ‘on the balance of

probability’.

17

We have not mentioned the fact that the study omitted congestion costs because congestion is not a characteristic of the process per seand should probably be treated in a separate transport module.18

See previous footnote.



Transport externalities were incorporated into the process of incineration and landfill (through use of typical

distances). This makes it difficult to understand the implications of changing distances travelled or modes of

transport used. For example, although incinerators may be close to urban centres, larger incinerators may accept

waste from more distant conurbations as well. It is somewhat debatable whether consideration of externalities

related to waste management facilities should consider transport externalities as part of the ‘process related’

externality (implying that, in some way, a particular facility always has associated with it a more or less well

understood transport externality). To the extent that transport externalities can be altered through changed

transportation modes as well as distances, such an analysis appears to obscure the possibilities for

improvement, or indeed, for their being addressed by transport policies. Indeed, a number of instruments,

notably the fuel duty escalator, have been introduced since the CSERGE work was undertaken. These effectively

internalise transport related externalities to some extent.

This suggests an approach which separates out transport externalities from the more ‘process specific’ ones. This

is especially important in considering the potential changes that could occur over time in collection logistics

(including vehicle design) and transport modes that might reduce the transport-related externalities associated

with waste transport over time. Admittedly, this is more likely to be of importance in the case of recycling, but

it is not unimportant in the case of the two options considered by the study.

The study thought that a treatment that differentiated between biogenic and non-biogenic carbon would be

better, but that the likely impact upon the results was negligible. This is true if the carbon related externalities

are small and known with certainty. It is not true if the estimates are subject to uncertainty (in which case, by

definition, we may not be able to assume that they are small). It is clear from the review by Fankhauser and

Tol (1995) that attempts to value climate change costs vary with (amongst other things) the discount rate used

and the degree to which models incorporate the sorts of low probability high consequence event which are being

deployed as a mechanism for ‘dealing with’ uncertainty.

Another comment worth making is that the study only used a range for the externality adders associated with

carbon dioxide and methane, though a range was also used for the valuation of incidents of mortality. In other

words, for all other pollutants, only one valuation figure was used. This gives the impression of a level of

certainty that is not really warranted given methodological and scientific uncertainties involved, as well as

‘measurement’ error (in some of the emissions data, for example – see below). On the other hand, the study

being an early attempt to generate estimated externalities, the availability of other relevant work was rather less

than it is today.



5.2.2 Brisson and Powell 1995This work updated the CSERGE work discussed above. The results are shown in Table 16.

Table 16: Externality Values for Landfill and Incineration (£/tonne waste other than disamenity)

L1a L2

b L3c L4

d I1e I2

f

Global pollution CO2 -0.32 -0.46 -0.32 -0.46 -2.55 -2.55

CH4 -2.36 -1.36 -2.36 -1.36 N/A. N/A.

Air pollution Conventional

gN/A. N/A. N/A. N/A. -2.27 -2.27

Conventional h

N/A. N/A. N/A. N/A. -2.82 -2.82

Toxics N/A. N/A. N/A. N/A. Not est. Not est.

Transport impacts Pollution

g-0.09 -0.09 -0.51 -0.51 -0.25 -0.45

Pollution h

-0.09 -0.09 -0.56 -0.56 -0.27 -0.49

Accidents -0.31 -0.31 -0.67 -0.67 -0.27 -0.60

Leachate -0.45 0 -0.45 0 N/A. N/A.

Pollution displacement g0 +1.92 0 +1.92 +16.09 +16.09

Pollution displacement h0 +2.63 0 +2.63 +21.98 +21.98

Totalg -3.5 -0.3 -4.3 -1.1 +10.6 +10.2Totalh -3.5 +0.4 -4.4 -0.4 +15.9 +15.5a L1 is an existing urban landfill without energy recovery

b L2 is a new urban landfill with energy recovery

c L3 is an existing rural landfill without energy recovery

d L4 is a new rural landfill with energy recovery

e I1 is a new urban incinerator with energy recovery

f I2 is a new regional incinerator with energy recovery

g Conventional air pollution damage including UK based damage only

h Conventional air pollution damage including UK based as well as that in the rest of the ECE region

Source: Brisson and Powell 1995

Note especially the source of the major differences between landfill and incineration. These are:

• leachate;

• the negative externality for methane from landfills, though this is counterbalanced by a higher externality

figure for carbon dioxide emissions from incinerators; and

• crucially, as mentioned above, the pollution displacement effects associated with energy recovery.

We will return to the discussion concerning pollution displacement below. Here, we merely point out that the

underlying assumptions concerning both the amount of energy generated per tonne of waste, and the source of

energy which is being displaced, appear to be absolutely critical in determining the overall externality. As we

shall see, the fact that methane emissions from landfill and associated energy generation from those with energy

recovery are subject to some variation makes it difficult to be so certain that the relative benefits lie so strongly

in favour of incineration as these figures seem to suggest. The point is reinforced by the fact that the study does

not evaluate air toxics.

5.2.3 Coopers and Lybrand / CSERGE 1996

The study carried out by Coopers and Lybrand and CSERGE is of particular interest because it has almost the

same focus as our own. The report was conducted for the EU 12 and the base year is early 1990 (1993). Some

issues were not considered in the analysis of external costs, for example, the environmental costs associated



with different MSW management options, toxic air pollutants from incineration and landfill, and disamenity

impacts and leachate (similar points are made by RPA and Metroeconomica 1999).

Factors deemed to be important for determining external costs were:

• composition of waste stream

• size of the disposal site or facility

• physical characteristics of the disposal site

• age of the disposal site, or facility

• spatial location of the disposal site

• level of pollution abatement in a facility

An updated version of the study’s results (from DETR 1999a) is shown in Table 17. The ranking of waste

management options by total economic (financial and external) costs and by external costs alone is shown in

Table 18.

The important step that this study tried to make was the inclusion of recycling in the comparative assessment of

waste management options. It is notable that recycling of all materials with the exception of plastic film

generates positive externalities. Under the assumptions made by the study, these positive externalities are very

large compared to the magnitude of those (positive and negative externalities) associated with landfilling and

incineration.

Table 17: External Costs and Benefits of Different Waste Management Options

Waste Management Option External Cost Estimate, £ Per Tonne

Of Waste, 1999 Prices

Landfill -3

Incineration (displacing electricity from coal-fired power stations) +17

Incineration (displacing average-mix electricity generation) -10

Recycling

- Ferrous metal

- Non-ferrous metal

- Glass

- Paper

- Plastic film

- Rigid plastic

- Textiles

+161

+297

+929

+196

+69

-17

+48

+66

Source: Adapted from Coopers & Lybrand et al (1997), in DETR (1999a)



Table 18: Ranking of Waste Management Options by Specific Criteria

Rank Total net economic costs Environmental costs

1 Source reduction Source reduction

2 Recycling Recycling (excluding composting)

3 Landfill Landfill

4 Incineration Incineration

5 Municipal composting Municipal composting

Source: Coopers & Lybrand et al (1997)

5.2.4 Brisson 1997

Brisson (1997), who was involved in the Coopers and Lybrand study, also assessed the external costs of waste

management in the UK under certain conditions. These are shown in Table 19. These were added to the private

financial costs discussed in Chapter 3 so as to arrive at total financial and external cost figures for the UK. For

recycling, these are shown in Table 20 for the different materials studied.

Again, the significance of assumptions concerning energy recovery is revealed through the assessment of

externalities from incineration as displayed in Table 19. In particular, the question as to what energy source is

being displaced is shown to be an important one. As with the other studies discussed above, there appears to be

little or no attention given to the recovery of metals from incinerators. Given the high value of the positive

externalities estimated for each tonne of ferrous metal (see Table 20), these could be expected to have an

important effect, even though ferrous metal represents a small fraction of all MSW.



Table 19: External Costs Associated With MSW Management Practices In The UK

ECU/tonne MSW

Present _ mixed refuse collection, bring system for recyclable and organic materials

Landfill _ no gas recovery -4

Landfill _ gas flared -5

Landfill _ energy generation (displacing old coal) -4

Landfill _ energy generation (displacing average EU electricity) -4

Landfill _ no transfer -3

Incineration _ electricity generation (displacing old coal) +18

Incineration _ electricity generation (displacing average EU electricity) -11

Recycling +170

Composting ---

Present _ Co-collection of mixed refuse and recyclable & organic materials (blue box)

Landfill -3

Incineration _ electricity generation (displacing old coal) +18

Incineration _ electricity generation (displacing average EU electricity) -11

Recycling +176

Composting ---

Present _ separate collection of mixed refuse and recyclable & organic materials (wheelie bins)

Landfill -3

Incineration _ elec. gen. (displacing old coal) +18

Incineration _ elec. gen. (displacing average EU electricity) -11

Recycling +170

Composting ---

Source: Brisson (1997)

Table 20: Total External and Financial Costs of Recycling in the United Kingdom

Material ECU/tonne of recyclable material

Ferrous metal 167

Aluminium 1481

Glass 183

Paper & board 44

Plastic film -30

Rigid plastic 39

Source: Brisson (1997)

Note that at the material specific level, the external costs vary enormously. This is a characteristic of studies

that have been undertaken thus far and indeed, was suggested by earlier work undertaken for us by CSERGE in

the context of a study for DETR (ECOTEC 1999).



In addition, one should point out that some materials which are frequently included in household collections,

such as textiles, are rarely analysed in this way. As mentioned in Chapter 4 above, the rationale for schemes to

collect one or other material varies from scheme to scheme. To our knowledge, no scheme bases this decision

on an assessment of external costs (though one scheme we interviewed had carried out such an analysis on the

basis of numbers reported in ECOTEC 1999), since more pragmatic approaches and the application of logical

principles tends to suffice. Such principles might include a desire to prevent inappropriate disposal of, for

example, paints, or oils, or batteries.

5.2.5 Powell et al 1996

The study by Powell et al (1996) is exceptional in that it looks at specific recycling systems. A comparison was

made not only between recycling and landfill, but also between two different types of recycling scheme, a

kerbside operation in Milton Keynes and a bring system in South Norfolk. The former employed a MRF to

separate materials. It is notable that the MRF, whilst not exclusively used for the purpose, was used principally

for the separation of plastics from the mixed recyclables (see comments at the close of Chapter 4).

One of the assumptions – that the distance each tonne of each material is transported is the same – appears

questionable. Presumably, the overall distance travelled in the kerbside collection ought to be apportioned

across materials in relation to the densities of collection. In this case, the distance travelled per tonne of paper

collected tends to be much less than that for plastics. This is certainly what one witnesses in reality, and it is

this that influences decisions as to whether or not to collect specific materials (if not in terms of the external

costs of doing so, then certainly in the private costs).

Another assumption made was that the energy used in the MRF can be apportioned equally across materials

(with the exception of glass). Whilst less questionable in the context of the scheme, the discussion at the close

of Chapter 4 speculates that the inclusion of plastics may be an important (though by no means the only) factor

in arriving at decisions as to whether or not to operate a MRF. Again, it should be noted that the study chose

to use only one externality adder in the computation of external costs (i.e. no ranges were used).

The key results of the comparison between the two schemes are shown in Table 21 below. These suggest that

on the basis of this analysis, the kerbside scheme performs somewhat better than the bring scheme. However,

one should note that the figures in respect of private costs that the study uses for the two schemes are quite

different to those which are found elsewhere in the literature (see above, and also Audit Commission 1997;

Atkinson, Barton and New 1993). The study itself notes this and suggests that when different figures are used

for the private costs, the ranking of the schemes could change.



Table 21: Comparison Of Kerbside And Bring Schemes’ External, Private And Total Costs (£ Per

Tonne Material)

Material Emissions1 Casualties Congestio

n

Externalities

(Total)

Private

Costs

Total Private

and External

CostsKerbside

Glass -0.86 -0.71 -3.40 -4.97

Other Materials -0.89 -0.71 -3.40 -5.00

All Materials -0.88 -0.71 -3.40 -4.99 56.00 60.99

Bring

All Materials -5.62 -10.93 -6.40 -22.95 92.00 114.95

Differentials(Kerbside w.r.t.Bring)

+17.96 -53.96

Source: Powell et al 1996

In the comparison between landfill and recycling, the same study suggested that net benefits from recycling

arise in the case of aluminium, paper, steel, and glass but that small net disbenefits arise in the context of

recycling HDPE, PVC and PET (see Table 22).

Table 22: Economic Valuation of Environmental and Social Impacts Associated with the Use of

Primary Materials and Landfilling Waste, And the Use of Secondary Materials and Recycling (£/Tonne

Each Material)

Material Primary Secondary Net External

Benefit from

Secondary

% Change in

Secondary from

25% Reduction in

Distribution

Distance

% Change in

Secondary from 25%

Reduction in

Emissions from

Secondary Processing

Aluminium -1880.27 -111.41 +1768.86 -0.85 -24.63

Paper -299.85 -73.79 +226.07 0 -23.12

Steel -269.40 -31.64 +237.76 0 -20.04

Glass -254.78 -67.20 +187.58 -0.64 -24.03

HDPE -9.49 -12.07 -2.57 -2.07 -22.87

PVC -7.46 -11.55 -4.10 -4.94 -15.24

PET -13.98 -21.25 -7.28 -5.36 -19.15

Source: Powell et al 1996

It is interesting that the study does consider variation in private costs but not around the adders used to assess

external ones. There is sensitivity analysis conducted around the issues of transport distance and emissions

related to recycling and secondary processing, respectively, but the adders appear to have been given a special

status in the analysis, despite the fact that uncertainty here is significant.



Note also that the percentage reductions associated with secondary materials processes are closely correlated with

energy use. This reflects only partly the significance of energy. More precisely, it reflects both the significance

of energy use in materials processing relative to transport energy, and the fact that the externalities assessed have

been limited to gaseous emissions and transport.

5.3 Studies Concerning Paper

There have been a number of studies, not only focused on the UK, which have sought to answer the question of

what to do with paper. The life-cycle debate here is especially awkward since so any of the key effects are not at

all well suited to treatment through life-cycle approaches. The loss of biodiversity associated with the

replacement of natural forests with plantations cannot adequately be captured by a technique which at best

quantifies land disturbance, but more generally, focuses only on inputs and outputs to a process (as opposed to

losses contingent upon certain activities taking place). The valuation of biodiversity loss and / or types of

landscape are especially problematic, but extremely high values are frequently found in the literature so these are

important impacts. It is also worth stating that they do occur since there have been attempts to downplay the

impact of the paper industry on biodiversity. Some of these issues are discussed in Leach et al (1997),

Ecologika (1998), and Pearce (1997) amongst others (see also Carrere and Lohmann 1997).

5.4 Summary

In the main, the studies reviewed have indicated a favourable view of recycling on environmental grounds. In

the US, a study by Ruston and Denison (1996) has made similar claims for recycling in the sense of the

benefits in respect of resource conservation, pollution reduction, and energy conservation (as much as £110 per

tonne of waste recycled, though like the studies discussed here and own, which follows, this could not be

termed a complete analysis).

However, a lengthy list of caveats ought to be attached to any attempt to derive firm conclusions from the

analysis in these studies. None of the studies are ‘complete’ and most have flaws. Some important inter-related

points, which arise from the consideration of existing studies, are made below:

• There is likely to be some difficulty in measuring and specifying the emissions from different waste

treatment plants, as well as from the activities which are ‘avoided’ through the recovery of energy and

materials. This is because a) plants are not uniform (and different pollution control measures determine the

media to which emissions are ultimately sent); and b) the composition of waste entering the facilities will

vary and this will affect emissions (and different plants are more or less sensitive to fluctuating

composition of inputs). Hence emissions may fluctuate across plants, and within a given plant, over time.

• Equally importantly, to the extent that one might wish to use such analyses for policy making, the

‘typical’, or ‘average’ performance of a specific type of plant will change over time. Life cycle inventories

provide snapshots of what may be happening at a given moment in time, but technological change can alter

the picture quite quickly. LCA, with limits as mentioned above, provides a static picture the utility of

which for decision-making purposes falls with the longevity of the installations being considered (because

it is impossible to understand how alternatives will evolve over the period).

• Changing policies can affect the significance of analyses undertaken in the past. To the extent that past

studies incorporated transport externalities within the externalities associated with a given process, policies

designed to internalise these presumably affect the policy conclusions that might be drawn from those

earlier studies. Hence, to the extent that the work by CSERGE et al (1993) was used to inform the decision



as to the level of the landfill tax, and because this attributed transport- related externalities to waste

treatment processes, consistency might have suggested that the landfill tax should have fallen as fuel duty

increased. That government chose to do the opposite suggests, possibly, a welcome departure from the

somewhat rigid view that taxes must somehow be justified by economic valuation, especially since it is

known that such valuations are rarely complete or free from significant uncertainties.

• The state of knowledge concerning the effects of all the pollutants examined, atmospheric and otherwise, is

typically in a state of flux. The effects of particulates on health, for example, are now suspected to be

greater than had hitherto been assumed. Carbon monoxide is another pollutant whose impacts are believed

to be in need of re-appraisal in the context of air quality debates (not just inside the home). Recent studies

of stratospheric ozone depletion may also implicate carbon dioxide in the process. There is an ongoing

debate about the health effects of landfills. The role of NOx (through related ozone production) may be very

important in terms of impacts on human health.

• A general criticism, and an extremely important one that follows from the above point, is that the many

studies appear to take only one value of the externality adder for a number of the pollutants examined.

Given that a number of such estimates are available (many of which themselves use ranges for the

pollutants concerned) the studies are assuming that the level of accuracy of such estimations is beyond what

can realistically be accepted. The danger in such an approach is that where the possible range of externality

adders is high, or where the environmental effect is not understood with any certainty, one is implicitly

introducing an entirely subjective element into the studies. Indeed, the valuation of specific effects

approaches a sort of ‘lucky dip’ in which one’s choice of externality adder / dose response function

inevitably influences the study’s conclusions. In defence of studies undertaken, the earlier works had rather

less to draw on in the way of work undertaken in the field. This would have been another reason to stress

caution in interpretation of results, especially where it was intended to base policy upon them (as appears to

have happened in the case of the Landfill Tax).

• None of the studies comes remotely close to being ‘complete’ in the sense of valuing all impacts associated

with all emissions to all media from all options (arguably once again reflecting a lack of research in the

area). One reason one could venture for this, related to the previous point, is that they couldn’t hope to do

this with any degree of certainty. There is an ‘air emissions’ bias to those studies undertaken. Even here,

however, few studies look beyond SO2, CO2, NOx, and PM10. They tend to concentrate upon a relatively

narrow range of atmospheric pollutants for which externality adders, or dose-response relationships (in

respect of damage to health, buildings and crops) are readily available (if not always entirely ‘cast-iron’ in

their scientific validity). This means that assumptions concerning the displaced energy source are also

extremely influential (see next point). One can say that ‘valuation does what valuation can’.

• As regards landfill and incineration, the suggestion from past studies is that the data concerning the amount

energy recovered, as well as the assumptions concerning what environmental burdens might be being

avoided, are crucial in determining the externalities, such as they have been measured, from these two

technologies. Note that, with regard to the landfill tax, there would appear to have been a strong rationale

for differentiating between those landfill sites accepting biodegradable waste that had no gas collection

equipment in place, and those that did. This will become less relevant in the future since the Landfill

Directive will require installation of gas collection equipment at such sites. Even now, this approach could

speed up the introduction of this technology. It will seem odd, perhaps even bizarre, to many observers that

the externalities from landfill or incineration are so heavily contingent upon one’s assumptions about what

is (or is not) going on somewhere else. At least in the case of recycling, the decision is somewhat more

clear cut. However, even here, matters will become more complex once recovered materials of one type

replace primary materials of another (as would happen under some of the more positive market development

scenarios – see IWA 2000; Enviros – RIS 1999).



• The externalities associated with recycling of specific materials vary from one material to another. Because

of the way these studies have been carried out (specifically, the air emissions bias), the favourable view of

recycling (and the differences between materials) is closely related to the issue of energy saving.

• The impacts of reduced materials extraction – the most obvious of the benefits generated by recycling – is

almost routinely ignored as an environmental benefit in the valuation context. One reason for this is that

such quantification would be extremely difficult to do. Hence, the results of valuation studies concerning

waste management do not necessarily reflect the actual environmental burdens that are related to the

approaches under examination. Most studies address these externalities only through assessing the scarcity

of resources, and the assumption usually employed is that market prices accurately reflect scarcity.

• Because most studies account for transportation effects though a specific level of transport externality, the

potential for reducing environmental impacts through measures which have next to nothing to do with the

actual processes to which the waste is ultimately subjected is made more obscure. Equally, the significance

of policies designed to already address transport effects might be overlooked. Ecologika’s (1998) work is a

good example of a study that explicitly recognises the need for planning in relation to transport at the same

time as one plans for waste since each has implications for the other.

• No study has taken a look at the whole range of materials which kerbside collections collect. These can

include oil, paints, and almost always, textiles. It would be extremely difficult to look at textiles in the

LCA context given the diversity of materials (and their origins) used in manufacture. Secondary clothing

exports may also have ramifications for local textile manufacturers where they effectively compete in local

markets (giving rise to social effects through the price depressing effect this may have).

In response to some of these criticisms, commentators might claim that, for example, the most significant

pollutants are those related to health and that most of the more significant ones are covered in the studies.

However, this is certainly not the case where air toxics are concerned (a laudable exception in that it seeks to

tackle air toxics, is AEA’s (1997) study, and the same study applies more than the normal level of caveats.

ERM’s (1998) study, taking its cue very strongly from the AEA work, also included these impacts).

Furthermore, to the extent that the emissions of certain pollutants to water may lead to damage to aquatic

ecosystems, the externalities may not be small. Certainly, just because little is known about these externalities,

there is no obvious reason to assume this implies negligible magnitude, especially given the (likely) location

specific nature of the impacts. This is important since, for example, in the case of incineration, what would

otherwise become air pollutants are being extracted from flue gas and discharged to other media. Under the

current state of the art, this would lead to on overstated benefit (associated with cleaning flue gas emissions) in

economic terms since valuation work aimed at understanding what the impact of discharging these pollutants to

land or water might be has not progressed very far. This partly reflects the particular problems one encounters

concerning benefits transfer in this field.

6. EXTERNAL COST ASSESSMENTThis 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:

q steel

q aluminium

q paper

q glass

q plastic

q high density polyethylene (HDPE)

q low density polyethylene (LDPE)

and the processes we have reviewed are

q transport

q landfilling

q incineration; and

q 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

Resource Depletion Pollution Degradation of Ecosystems and

Landscape

Depletion of mineral resources

Depletion of fossil fuels

Depletion of biotic resources

Global warming

Ozone depletion

Human Toxicity

Ecotoxicity

Photochemical Oxidant Formation

Acidification

Nitrification

Radiation

Dispersion of Heat

Noise

Smell

Occupational Health

Dehydration

Physical degradation of

ecosystems

Landscape degradation

Direct human victims

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.’

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

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).



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

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

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.

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.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.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

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).

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

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.,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).



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

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.



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).

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

Low High Units

Valuation factor GW 0.05 0.81 ECU/vkm

Valuation factor noise / health 0.006 0.054 £/tkm

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



Table 25: Valuations of Life Considered in ExternE Project

Units: MECU, 1990 (1 ECU = $1.24) Europe US

Hedonic Salaries 2.8-3.5 3.5-5.5

CV 4.1-6.3 1.4-2.5

Expenditures 0.7-3.4 1.0-1.1

Average 2.5-4.4 2.0-3.0

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

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.

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.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).



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).

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.

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.’



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

Category of Impact Low Adders High Adders

Global Warming -0.40 -3.97

Noise / health -1.15 -10.37

Slight injury -0.01 -0.20

Serious injury -0.02 -0.51

Fatalities -0.01 -0.91

Congestion -0.01 -7.37

TOTALS (£) -1.44 -23.32

NB: Totals are subject to rounding

Table 27: Method 2, 10 Tonne Payload, 80km Roundtrip, Impact Per Tonne Of Waste Transported

(£/t)

Category of Impact Specific Impact Low Adders High Adders

Greenhouse gases: CO2-0.01 -0.19

CH40.00 0.00

N2O0.00 0.00

PM10 -0.02 -0.72

Acid gases: SO2-0.05 -0.26

NOx-0.09 -1.92

Noise Cars 0.00 0.00

Trucks -0.12 -1.03

Casualties: Slight -0.01 -0.20

Serious -0.03 -0.51

Fatalities -0.06 -0.91

Congestion: Trucks -0.01 -7.37

TOTALS -0.39 -13.11

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, th



e 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

.

• 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

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.



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

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

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.

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.



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.

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 WasteTreatment 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

36

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



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.

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/m

3 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.

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).



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 (%)

Readily Moderately Slowly Inert

Newspaper 19

Cellulose (pure) 73

Toilet Tissue 56

Brown Paper 48

Cardboard 31

Putrescibles 80 20 0 0

Textiles 0 0 100 0

Paper and Card 0 20 80 0

Unclassified 0 0 10 90

Fines <20mm 20 20 60

Glass, plastic,

metals and non-

combustibles

100

Combustibles 100

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

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

Degradability rate April Sample September Sample

No recycling With recycling No recycling With recycling

Readily 22.1 26.5 21.1 24.8

Moderately 11.8 12.2 13.1 13.3

Slowly 29.0 27.2 36.7 35.4

Inert 37.1 34.1 29.1 26.5

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

DA

Rate of Gas

Generation

Time

In the case of the faster decay curve, methane

gas equivalent to the area ABC (below the

curve) is uncollected. With a slower rate, the

uncollected gas increases to that equivalent to

area DEF (below the curve)

CB E F



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 visa versa).

Tables 30-33 show our results for the composition of MSW we have used. For tables 30-33 see tables.pdf.

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.

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

In other words,

it would be possible to have in place high assumptions for the

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).

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

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

• Emissions of gases other than CO2 and CH4 (ozone depleting chemicals, such as CFCs, are believed to

arise from landfills).41

• 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

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.39

For obvious reasons, there is no counterbalancing disamenity reduction at the ‘source’ of displaced energy (the plant is still there).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 NOx. These were from landfilling contaminated bottle fractions only.41

See previous footnote.



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 (see Table 34). 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).

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

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.



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.

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).

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).



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

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.

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.



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

Recovery rate for steel 50%Recovery rate for aluminium 33%Efficiency of energy conversion 20%

Option 1 Option 2

Energy Recovered GJ 2.08 2.06

Avoided externalities due to energy recovery Average fuel mix High 71.30 70.52

Low 7.32 7.24

Coal High 127.74 126.35

Low 13.48 13.33

CO2 emissions High -29.11 -29.11

Low -0.97 -0.97

N2O emissions High -7.25 -7.25

Low -0.80 -0.80

Other air emissions High -126.79 -126.79

Low -1.38 -1.38

Benefits from recovered steel tonnes 0.03 0.03

High 80.97 80.97

Low 1.23 1.23

Benefits from recovered aluminium tonnes 0.01 0.01

High 30.81 30.81

Low 1.65 1.65

Benefits from replaced aggregates tonnes 0.25 0.25

High 0.33 0.33

Low 0.33 0.33

Option 1 Option 2

Air Emissions and Energy Coal High 76.72 75.32

Low 13.53 13.39

Average Fuel Mix High 20.27 19.49

Low 7.37 7.29

None High -51.03 -51.03

Low 0.05 0.05

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)

Recovery rate for steel 50%Recovery rate for aluminium 33%Efficiency of energy conversion 20%

Option 1 Option 2

Energy Recovered GJ 2.08 2.06

Avoided externalities due to energy recovery Average fuel mix High/Low 11.69 11.56

Coal High/Low 62.62 61.94

CO2 emissions High -29.11 -29.11

N2O emissions High -7.25 -7.25

Other air emissions High -126.79 -126.79

Benefits from recovered steel tonnes 0.03 0.03

High 80.97 80.97

Benefits from recovered aluminium tonnes 0.01 0.01

High 30.81 30.81

Benefits from replaced aggregates tonnes 0.25 0.25

High 0.33 0.33

Option 1 Option 2

Air Emissions and Energy Coal High 11.59 10.91

Low 13.53 13.39

Average Fuel Mix High -39.33 -39.46

Low 7.37 7.29

None High -51.03 -51.03

Low 0.05 0.05

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

• 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

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.



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

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

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.



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

Low Adders Differential (£) High Adders Differential (£)

Aluminium 315.1 5256.9

Steel 49.2 3239.1

Glass 61.2 1947.1

Newspapers and

magazines

26.4 521.5

HDPE 41.1 460.2

LDPE 15.4 92.8

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:

q Ancillary material flow

q Excavated and/or disturbed material flow

q Hidden material flows

q Direct material input

q 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 essentiallyexpressed as ratios of the weight of ‘hidden flow’ material to the weight of finished good)

Ancillarymaterial

flow

Excavated /disturbedmaterial

Hiddenflow

Directmaterial

Finishedgood

Totalmaterial

requirement

Glass (1) 0.02 0.02 1.00 1.02

Steel (2) 0.72 3.10 3.83 1.00 4.83

Other material inputs to steel (3) 0.00 9.00 0.82Aluminium (4) 3.00 1.92 4.92 1.00 5.92

Paper (5) 2.30 2.30 1.13 1.00 4.43

Plastics NA NA NA NA 1 NA

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 Report1996(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)

Forestry Oil MineralsCO2 1006.7 99942.2 6017.9

CH4 0.1 725.3 0.4

N2O 0.0 0.1 1.6

SO2 1.1 12.2 9.6

NOX 12.5 684.2 66.4

NH3 0.0 0.0 0.0

Black smoke 4.5 1.9 22.4

NMVOC 10330.5 1098.8 9.4

Benzene 0.0 1.0 0.1

CO 3.5 353.1 22.5

Lead (mg/tonne) 0.0 29.6 2.5

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)

Forestry Low -10.38

High -30.41

Oil Low -2.29

High -28.53

Minerals Low -0.26

High -6.64

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

Iron Ore Pulp and NewsprintCountry %

ImportsDistance (miles)

and port of originCountry % Imports Distance (miles)

and port of originAustralia 36 10,871 (Western

port)

USA 19 4,825 (Houston)

Canada 17 2,028 (St Johns) Canada 16 2,028 (St Johns)

Brazil 13 3,960 (Belem) Finland 13 1,414 (Hamina)

South Africa 12 5,943 (Cape Town) Sweden 13 694 (Stenungsund)

Mauritania 6 2,178 (Novakshott) Brazil 9 3,960 (Belem)

Total 84 Total 70



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 fromMiddle East (kg/tonne freight)

C02 CO NOX VOC SO2 CH4

26.56 0.075 0.697 0.021 0.488 0

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.

Emissions and external costs due to pulp and paper freight from weighted average of exportingcountries (emissions in kg/tonne freight, costs in £/tonne freight)

C02 CO NOX VOC SO2 CH4 TOTALSEmissions 28.76 0.08 0.75 0.02 0.53 0.00

Low (£) 0.10 0.00 0.75 0.02 1.06 0.00 1.94

High (£) 2.62 0.00 16.61 0.06 5.28 0.00 24.57

Emissions and external costs due to iron ore freight from weighted average of exportingcountries (emissions in kg/tonne freight, costs in £/tonne freight)

C02 CO NOX VOC SO2 CH4 TOTALSEmissions 42.57 0.12 1.12 0.03 0.78 0.00

Low (£) 0.15 0.00 1.12 0.03 1.56 0.00 2.87

High (£) 3.87 0.00 24.58 0.09 7.82 0.00 36.37

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.

7. PUTTING IT ALL TOGETHER - PRIVATE AND SOCIAL COSTS

7.1 Scenarios Discussed

It seems clear from Chapter 3 that the actual costs of recycling schemes vary considerably between types of

scheme, and as far as kerbside collection is concerned, across different authorities (as Chapter 4 showed). In a

recent response to criticisms of recycling, the Environmental Defense Fund in the US, accepting that kerbside

collections might be relatively inefficient at present, made the point that:

Gaining efficiency in curbside recycling requires careful, objective testing of truck designs, vehicle routing,

public education, collecting new materials, setting fees for residential waste management services to reward

households that produce less waste, and other techniques. This is work that must be carried out at the local

level, city by city. (Ruston and Denison 1996).

One has to cast doubt, therefore, given the local factors that will influence collection costs, upon an approach

that seeks to compare schemes across different authorities. Different schemes collect different materials and have

different rationales for doing so. The incremental costs to a scheme of including or not including one or other

material are not uniform.

What we have sought to do in this Chapter is to understand the picture in terms of the options available to

specific authorities. Building on the work in Chapters 4 and 6, we have tried to develop scenarios in such a way

as we can model existing schemes against alternative schemes which rely more heavily on linear waste

management options. We have concentrated on the kerbside recycling of dry recyclables (for which we have



better data), and a comparison with the base case of simply landfilling or recovering through energy from waste

only.48

7.2 Collection and Disposal to Landfill

The collection of residuals is, in all schemes, undertaken in refuse collection vehicles, typically of 24 tonne

maximum weight. The load carried is typically of the order 10 tonnes, with a maximum of around 13 tonnes.

Using the ten tonne figure, it is possible to impute a per tonne external cost for this. In this case, we can

simply draw on the analysis carried out above. The externalities are between –£0.4 and -£21.8 for an 80

kilometre round trip.

For landfilling, under the assumptions of 12.5% oxidation at the cap, 50% gas collection efficiency and 35%

engine efficiency (see Table 34 above), the range of externalities is from £-14.6 to +£18.6.

It is not necessarily correct to be adding the results from landfilling and transport corresponding to particular

sets of adders. The ranges in the adders to some extent reflect the variation in the contexts in which emissions

occur. One might have good reason to believe that the impact of emissions from transport, and therefore the

likelihood of externality adders being high or low, will be different to those from coal fired power stations

which one may be assuming one is displacing. The significance of the assumptions embedded in the analysis

are enormous, the more so as one entertains the possibility of ‘high value’ externality adders.

7.2.1 Summary

The financial costs of this route typically lie in the order of £25 and £35 for collection and disposal,

respectively (including landfill tax). The externalities are summarised in Table 44. We have added in some

estimates of transport externalities for longer trips.

Table 44: Externalities Associated with RCV Collection and LandfillingRCV Collection

60 km roundtrip 80 km roundtrip 100km roundtripHigh -£17.49 -£23.32 -£29.15

Low -£0.30 -£0.39 -£0.49

LandfillWithoutLandfill GasCollection

With LFG andFlaring

With LFG andEnergy Recovery

Assumption regarding fuelbeing replaced under energyrecovery

No Factor for

Lifetime

Collection

Factor for

Lifetime

Collection

High and Average Mix High -14.63 -3.03 9.05 -0.42

High and Coal High -14.63 -3.03 18.61 5.31

High and None -14.63 -3.03 -3.03 -7.67

Low and Average Mix Low -0.93 -0.32 0.92 0.18

Low and Coal Low -0.93 -0.32 1.96 0.80

Low and None -0.93 -0.32 -0.32 -0.57 48

It is worth pointing out that the position of the Energy from Waste Association is that this should not happen, and that the appropriate

deployment of the technology is in the wake of front-end materials recovery.



7.3 Collection and Disposal to Energy from Waste Incineration

For the case of incineration, the option of simply disposing of all waste is one that has been publicly shunned

by the Energy from Waste Association in its response to AWWW (see EfWA 1999), though this is put into

context with statements concerning the need to consider what the BPEO might be. However, for the purposes of

this analysis, we have looked at the externalities associated with such plants as though they were being used as

the main form of waste treatment.

For the transport of the residuals, we have included smaller distances than were used for the landfilling scenario.

We have also included some of the longer distances since it may well be that planning for energy from waste

incineration in the future occurs at a more regional level (so that waste is received from a larger area). We know

relatively little about public perceptions of incinerators in respect of their scale. There may be some trade off

(from the perspective of the generation of externalities) between greater transport and disamenity externalities,

and lost economies of scale, depending upon the nature of the community a plant is being designed to serve.

7.3.1 Summary

The private costs of incineration are generally of the order £45 for the gate fee plus collection costs. We estimate

collection costs as being the same as for landfill (£25). The evolution of incineration gate fees may depend upon

the way in which the PRN market evolves which in turn will depend upon the form of any revisions to the

Packaging Directive. Also, efforts to recover bottom ash are likely to produce savings on avoided landfilling.

The results for the externality analysis are shown in Table 45.

Table 45: Externalities Associated with RCV Collection and Incineration (£ per tonne MSW)

RCV Collection

30 km roundtrip 50 km roundtrip 70km roundtrip 100km round trip

High -8.74 -14.57 -20.40 -£29.15

Low -0.15 -0.25 -0.34 -£0.49

Incineration Steel recovery

50%, aluminium

recovery 33%

Steel recovery

50%, aluminium

recovery 33%

Steel recovery

50%, aluminium

recovery 33%

Option A Option B

High and Average

Mix High

20.27 19.49

High and Coal High 76.72 75.32

High and None -51.03 -51.03

Low and Average

Mix Low

7.37 7.29

Low and Coal Low 13.53 13.39

Low and None 0.05 0.05



7.4 Collection at Kerbside and Recycling

7.4.1 Kerbside Materials Collection

The externalities associated with kerbside schemes can be measured in an absolute sense, or they can be

measured relative to the ‘baseline’ situation in the UK. It seems clear from our discussions that in the financial

context, some local authorities perceive the costs of collection of materials relative to what they would

otherwise have had to do anyway. Hence, where waste is collected on alternate collections, the approach has

been to assume that the costs of doing so are best accounted for through assessing the incremental change in

costs resulting from the scheme. In other cases, where the collection is entirely separate, the situation is less

clear. As discussed above, however, the more successful kerbside schemes do reduce the costs of collection of

residuals through extracting materials that would otherwise have to be collected. For similar reasons, it seems

appropriate to deduct from the absolute external cost of the kerbside collection the incremental reduction in the

external costs of residuals collection which would occur anyway where materials are collected as part of the

residuals collection.

It therefore becomes important to understand what occurs to the residuals collection as the recycling scheme’s

performance increases. Either the residuals round remains the same, but with less waste collected, in which case,

the external costs associated with collection of residuals as calculated here hardly changes (even though the

weight collected has fallen),49

or the residuals round adapts to undertake slightly longer collection rounds to fill

the vehicle. It seems likely that the former situation would apply only in the short term, yet it would be

difficult to imagine complete ‘one-for-one’ substitution (not least because the situation is changing) in terms of

the efficiency of collection of materials.50

As such, taking a middle case, the RCV round can adapt to undertake

journeys which are longer depending upon the success of the scheme, but not in a perfect manner. The total

journey distance will change, but not in direct relation to a scheme’s success, however, since the journey

undertaken will be lengthened only by the extension of the collection part of the total journey (the journey to

and from the collection round will not change at all). Thus, the change in distance travelled will be changed, in

relative terms, incrementally. The change in externality per tonne of residual waste collected is therefore likely

to become only incrementally more negative. For the purposes of our analysis, we will assume that it is

unchanged (the assumption will hold best in the case where waste is collected in densely populated urban areas

and transported considerable distances for final treatment.

Suppose we look at a scheme that is collecting 100kg of dry recyclables per household covered (and this is

typical of the dry recyclables schemes we have looked at). Suppose also that one assumes that the amount of

waste generated by the average household for refuse collection is of the order 750kg (from MEL 1999). In this

case, approximately 100/750 = 0.14 tonnes of each tonne of MSW which would otherwise be collected by the

RCVs are being collected by the kerbside vehicles.

The kerbside vehicles we examined typically have a maximum axle weight of 7.5 tonnes with a payload of the

order 2.5-3 tonnes (others are 11 tonne trucks with 4 tonne payloads). The actual collection part of the journey

may be from 1-3 miles in urban areas, and anything from 3-15 miles in less densely populated areas. The

journeys to the depot in the schemes we looked at averaged from 4-9 miles in urban areas to 8-14 miles in less

highly urbanised areas. The possible range in journey distance would therefore be 9-43 miles, or between 14.4-

49

In practice, there probably would be savings owing to marginal increased in fuel efficiency, and associated reductions in transport-

related emissions. These are not accounted for in our model.50

Also, from the perspective of private costs, any savings that actually occur may not be passed on to local authorities until contracts are

renegotiated. This highlights the need to integrate flexible terms into local authority contract specification to account for changing system

performance.



68.8 km. The externalities associated with the collection of each tonne of waste collected on such rounds would

be, for the 14.4 km round, between -£0.21 and -£11.66 and for the 68.8 km round, between -£0.39 and -

£40.88. Note we have used 0.27 l/km as opposed to 0.32 l/km for both the 2.5 tonne and 4 tonne recycling

payloads, reducing the per kilometre emissions from these vehicles relative to RCV emissions. Note also that

one respondent operating a kerbside scheme explicitly mentioned the fact that vehicles were fitted with

equipment to remove particulates from exhaust.

Tonne for tonne, it appears that the collection of a tonne of waste through kerbside schemes, in urban areas, is

likely to give rise to higher transport externalities than the collection of waste for delivery to landfill. This is

because of the higher density of collection in the case of collection of residuals. Since many externalities are (in

this analysis) related directly to vehicle kilometres travelled, the more waste that is transported per vehicle, the

less significant these externalities will become when expressed per tonne of material collected.

It has to be recognised that, as mentioned above, once the recycling scheme makes more than a marginal

contribution to waste collection, the appropriate way to deal with transport-related externalities is to assume that

some would otherwise have to be incurred through collection of waste in RCVs. Hence, one should properly

subtract the avoided (negative) externality associated with RCV movements from the negative externalities

associated with kerbside collection of waste (which has the effect of making the negative externalities associated

with kerbside collection ‘less negative’). Because of the lack of a complete ‘one-for-one’ substitution effect, it

would probably be more reasonable to subtract only half (as a reasonable estimate) of the avoided RCV

collection externality.

7.4.2 Transport to Reprocessors

More serious, in respect of transportation externalities, is the question of transport distances moved by

secondary materials. One of the schemes we spoke to sends paper 160 miles, glass to reprocessors 147 miles

and 205 miles away, steel to a plant 104 miles away, and aluminium cans 194 miles. If we weight these

distances by the typical composition of dry recyclables (as a proxy for journey numbers made), we find that the

average distance moved for these vehicles is some 162 miles. In this analysis, we use this as the typical

distance moved for the purpose of calculating the per tonne externality. Clearly, this is a simplification as the

transport arrangements for different materials (in terms of vehicle use, contracting out, etc.) will vary, as will

the distances. The astonishing thing is that this is an authority which is short of landfill void, and for which

disposal costs are high. The lack of outlets for materials in such areas reflects the (pre)dominance of large scale

reprocessing facilities.

For a 162 mile (=259km) journey in which a haulage vehicle is used carrying some 20 tonnes of material,

assuming the vehicle is not returning empty,51

then if we estimate fuel consumption at 0.5 l /km, we can make

an estimate of the transport related external costs.52

These are estimated using the approach described in the

previous chapter as between -£0.87 and -£45.44 per tonne of material.53

We are not in a position to know

whether these are typical distances. The policy implications of this type of movement – and we are speaking

principally of glass and paper (because of their significance in collection) are not necessarily that one throws

51

Where schemes are undertaking the transport of materials themselves, they are less likely to return full. The issue here may relate to

whether the haulage is contracted out or carried out in-house. In the latter case, a full vehicle on the return journey seems more likely.

On the other hand, the location of reprocessors may influence the decision concerning haulage.52

One of the schemes we spoke with had done some work on these logistics. The 20 tonne vehicle did 6.5 miles to the gallon, so our

approximation is a good one.53

We have not calculated this for a two-way journey since such haulage is, we understand, increasingly carried out by private

companies who usually plan to make a return journey with other materials.



one’s hands up in the air and despairs. Another way of looking at these figures is that they suggest a need for

local reprocessing capacity, especially for glass and newsprint, to reduce the distances moved by these materials

post kerbside collection. This has been discussed elsewhere (Ecologika 1998; Murray 1999; IWA 2000) in the

context of strategies for economic regeneration. One would imagine that under positive scenarios for market

development for recycled materials, the distance travelled to reprocessors would fall as regional capacities are

developed.

These figures should be compared with the emissions from extraction and from the importation of materials

estimated (albeit somewhat crudely) in the previous Chapter. It is clear that the air emissions alone from these

activities associated with primary materials extraction and transport can substantially offset these transport

movements. The major exception here is likely to be glass, many of the materials for which may be locally

sourced. The extraction of these materials however creates site-related disamenities that are difficult to estimate

(and in any case, it would be inappropriate to express these in ‘per tonne’ form).

7.4.3 Effects of Kerbside Recycling on Waste Management Costs

What has happened to the tonne of waste which was, in our ‘baseline’ scenarios, taken in its entirety either to a

landfill or an incinerator? It has been partially separated. The separately collected fractions will be used in the

manufacture of new products displacing primary materials. We saw in the previous Chapter that this can

generate external benefits. Suppose again we look at systems that extract 100kg, 130kg and 150kg per

household of a typical household waste stream of 1.2 tonnes per annum. This amounts to extracting, per tonne,

the amounts of waste shown in Table 46. The external benefits in using this material as opposed to primary

material are shown in Table 47. The potential benefits from recycling range from £3.25 to £91.68 at 100kg/hhld

to £4.87 to £137.53 at 150kg/hhld.

Table 46: Material Removed from a Tonne of MSW under Different Rates of Material Recovery

kg recyclables collected per household 100.00 130.00 150.00

Effective extraction per tonne MSW by Material

Paper and board 0.053 0.068 0.079

Aluminium 0.001 0.001 0.001

Steel 0.005 0.006 0.007

Glass 0.023 0.030 0.035

Textiles 0.002 0.003 0.003

Total 0.083 0.108 0.125

Residual Material to Landfill / Incinerator (tonnes) 0.917 0.892 0.875

NB: Assumes 1.2 tonnes MSW per household



Table 47: External Benefits Associated with Recycling at Different Rates of Material Recovery (£ /

Tonne Of MSW, Calculated On Basis Of Material Diverted At Different Diversion Rates)

Material Collected perHousehold (kg)

100 130 150

Low High Low High Low High

Aluminium 1.39 27.47 1.81 35.71 2.09 41.20

Steel 0.20 3.34 0.26 4.34 0.30 5.01

Glass 0.24 15.90 0.31 20.67 0.36 23.85

Newspapers and magazines 1.41 44.98 1.84 58.47 2.12 67.47

TOTALS 3.25 91.68 4.22 119.19 4.87 137.53

NB: No figures were generated for benefits from textiles recovery

When these materials are extracted from the typical MSW tonne, the amount of material landfilled falls, but its

character changes. The nature of this change is shown in Table 48. The effect on landfill externalities is given in

Table 49. For Table 49: Externalities of Landfill under Different Scenarios for the Extraction of Materials

from A Tonne of MSW see tables.pdf.

Table 48: Changes in Composition and Weight of Material under Different Rates of Kerbside

Collection

No collection 100kg/hhld 130kg/hhld 150kg/hhldNewspaper 13.0% 7.73% 6.15% 5.10%

Office paper 6.0% 6.00% 6.00% 6.00%

Corrugated Boxes 3.0% 3.00% 3.00% 3.00%

Coated Paper 5.0% 6.00% 6.00% 6.00%

Al Cans 2.0% 1.94% 1.92% 1.90%

Steel Cans 5.0% 4.51% 4.36% 4.26%

Glass 8.0% 5.69% 5.00% 4.53%

HDPE 4.0% 4.00% 4.00% 4.00%

LDPE 4.0% 4.10% 4.10% 4.10%

PET 2.0% 2.00% 2.00% 2.00%

Food Scraps 20.0% 20.00% 20.00% 20.00%

Grass 4.0% 4.00% 4.00% 4.00%

Leaves 2.0% 2.00% 2.00% 2.00%

Branches 2.0% 2.00% 2.00% 2.00%

Yard Trimmings 2.0% 2.00% 2.00% 2.00%

Screenings 8.0% 8.00% 8.00% 8.00%

Textiles 3.0% 2.80% 2.74% 2.70%

Miscellaneous combustibles 7.0% 7.00% 7.00% 7.00%

Total Residual MSW(tonnes)

1.00 0.93 0.90 0.89



Note that, for landfill, even in the no energy recovery cases, the net change in externality is negative under the

high externality adders scenario even though less material is being landfilled. This is a consequence of the fact

that paper is being taken out of the landfill, and the assumptions in the landfill model effectively treat paper as

a net sequester of carbon relative to the situation which occurs when the carbon cycle is allowed to take its

course. On the other hand, less methane is generated too (so there are competing effects). In the energy recovery

situations, in all cases, the change is negative. There is a trade off between the lost benefits attributed on the

basis of less energy generated (because less pollution is being avoided) and the loss of a net sequester of carbon

(contributing a net negative externality), and the positive effect of less methane generated. Note that these

changes are small relative to the changes in the externalities associated with recycling. Note also that there has

been a reduction in the energy generated by the landfill owing to the fall in methane generation. Between the

‘no recycling’ and 150kg per household scenarios, the decline has been from 98 kWh (59 kWh) to 90 kWh (54

kWh) (the figures in brackets are those which incorporate the ‘lifetime’ factor).

The net change in incineration externalities is given in Table 50. As with all numbers in this report, these

should be treated with caution. Even if one accepted the earlier estimates, we do not trace through (in our

model) changes in air emissions associated with the change in materials composition (and possibly, materials

volume).



Table 50: Externalities of Incineration under Different Scenarios for the Extraction of Materials from

a Tonne of Municipal Solid Waste

Collection per Household ExternalCosts

Change Relativeto Base Case

0 kg TOTALS Coal High 76.72

Low 13.53

Average fuel mix High 20.27

Low 7.37

None High -51.03

Low 0.05

100kg TOTALS Coal High 59.13 -17.59

Low 12.47 -1.07

Average fuel mix High 6.61 -13.66

Low 6.73 -0.64

None High -59.71 -8.69

Low -0.08 -0.13


Low 12.09 -1.44

Average fuel mix High 2.31 -17.96

Low 6.52 -0.85

None High -62.22 -11.20

Low -0.11 -0.16


Low 11.84 -1.69

Average fuel mix High -0.56 -20.83

Low 6.37 -1.00

None High -63.89 -12.87

Low -0.13 -0.18

The interesting points to note are the much larger net changes in externality (relative to what happened in the

landfill case). These are a consequence of the fact that there is less steel available to be recovered, and the

removal of paper (remember, there is not plastic being recycled in our scheme) reduces the energy generated

from 2.08GJ to 1.85 GJ (an 11% reduction). Hence, the net change in external costs such as we have been able

to measure them are ‘more negative’ for the case of incineration than for landfill.

The important question that arises is ‘is this good or bad?’ What has happened is that of two key sources of

benefits under the assumptions made, one – the recovery of metals – is being carried out elsewhere, and another

– the recovery of energy – is being affected by the loss of combustible paper. In practice, some of this might be

offset by a reduction in flue gas emissions (though equally, loss of materials volumes might affect combustion

efficiencies). If one were to pro-rate the air emissions externalities against the mass of waste combusted this

would have the effect of reducing the net ‘benefit loss’ by some £10-£16 in the high adders scenario.

Some might argue that this constitutes an argument for not engaging in front-end recycling where incinerators

are present. This, to some extent, is what happens in the Netherlands. On the other hand, the range of benefits

from the recycling still, apparently, exceeds the lost benefits from incineration when the same externality adders

are used.54

There would appear to be a role for materials extraction pre-combustion. This largely supports the

54

We can say this with some degree of confidence because the source of these benefits is the same.



Energy from Waste Association’s view expressed in its response to A Way With Waste (EfWA 1999):‘We agree

with the Government’s view that recycling and composting should be considered before energy recovery –

something that already happens in most cases.’

It is worth pointing out that although the decline in benefits due to recycling associated with energy from waste

incineration are less than the decline associated with energy from waste landfilling, the absolute levels of

benefits show different ranges. It would be dangerous, however, to draw the strong conclusion from this, even

on the basis of our partial analysis, that incineration is environmentally ‘better’ than landfill. The numbers are

suggestive, but the lack of any consideration of the many omitted pollutants (both incineration and landfill)

cautions against drawing such conclusions. In addition, it is clear that there are distributional issues associated

with the costs and benefits generated. It is important to emphasise that there is a direct trade-off here. The

benefits attributed to incineration from recovery of steel are directly proportional to the benefits from metals

recycling. The greater the benefit from recycling, the greater (within the significant limitations of our model) the

benefits attributable to incineration for energy recovery. This means that at these higher benefit levels, the net

improvement to metals recovery from recycling becomes much more significant.

7.4.4 Summary

The net effect of integrating kerbside recycling, relative to landfill only, is summed up below.

BOX 2: PRIVATE AND EXTERNAL COSTS OF KERBSIDE COLLECTION RELATIVE TO

‘LANDFILL ONLY’ SCENARIO

The following is a summary of the cost changes from the ‘landfill all’ base scenario. The changes are calculated

with the tonne of MSW as the functional unit under examination (i.e., we are looking at the net change in all

costs, or at least those covered in the study, when materials are collected at kerbside).

1) External Cost Savings Associated With Avoided Extraction of Primary Material

The rough estimate of air pollution externalities made in the previous chapter from the extraction of specific

materials was £10.38 - £30.41/tonne of output for forestry and £0.26 - £6.64 /tonne of output for minerals.

External cost savings from avoided forestry activity £0.55 to £1.62 at 100 kg/household and £0.82 to

£2.40 at 150 kg/household

External cost savings from avoided mineral extraction £0.00 to £0.04 at 100 kg/household and £0.00 to


External cost saving from avoided primary extraction £0.55 to £1.66 at 100 kg/household and £0.82 to




2) External Cost Savings Associated With Transport of Primary Material

At the 100kg per household collection rate, 0.053 tonnes of newspaper / magazines and 0.005 tonnes of steel

are being extracted from each tonne of MSW.

At the 150kg per household collection rate, 0.079 tonnes of newspaper / magazines and 0.007 tonnes of steel

are being extracted from each tonne of MSW.

In Chapter 6, we estimated crudely that the average transport externalities associated with imported newsprint

were –£1.9 to -£24.6 per tonne. For iron ore, these were estimated as -£2.9 to -£36.4.

External cost savings from avoided paper imports £0.10 to £1.30 at 100 kg/household and £0.15 to £1.96

at 150 kg/household

External cost savings from avoided iron ore imports £0.01 to £0.18 at 100 kg/household and £0.02 to £0.25

at 150 kg/household

External cost savings from avoided imports £0.11 to £1.48 at 100 kg/household and £0.17 to £2.21

at 150 kg/household

3) External Cost Savings (Avoided Residuals Collection)

This is calculated by assuming that one saves on collection of half the material that would otherwise be

collected at the RCV. One therefore saves in financial (see below) as well as external cost terms. The fraction to

be considered is one half of that fraction of waste being recycled which would otherwise have to be collected.

The external cost saving is therefore:

0.5 x (1/12 to 1/8) x (£0.30 to £17.49) where 60km trip avoided



Where 60km trip avoided Max £1.09 (150 kg scheme), Min £0.01 (100kg scheme) per tonne MSW

Where 80km trip avoided Max £1.46 (150 kg scheme), Min £0.02 (100 kg scheme) per tonne MSW


4) Benefits From Recycling

At 100 kg /household external benefits of between £3.25 and £91.68 per tonne of MSW

At 150 kg /household external benefits of between £4.87 and £137.53 per tonne of MSW

5) Transport-Related External Costs of Kerbside Collection Per Tonne of MSW

14.4 km round between -£0.21 and -£11.66 per tonne recyclables collected

68.8 km round between -£0.39 and -£40.88 per tonne recyclables collected

Within each tonne of MSW, where 100kg are extracted per household, approximately 1/12 of each tonne of

MSW is being collected. At 150kg, the fraction of each tonne generated is 1/8 being collected. The incremental

externality per tonne of MSW will therefore be:

14.4 km round -£0.02 to -£1.46 per tonne MSW




6) Transport To Reprocessors (Indicative Figure Only)

-£0.87 to -£45.44 per tonne delivered

For 1/12 of a tonne, this would be between -£0.16 and -£5.40 per tonne MSW


7) Change In Externality Owing To Change in Weight and Composition of Material Landfilled

These range from +£0.02 to -£3.09 per tonne MSW (see Table 49)

8) Savings in Private Costs of Residual Disposal / Recovery and Collection Costs

There is a collection (of residual) saving to be considered on the financial side that is rarely appreciated. This

may be estimated as half the collection cost (estimated at £25) of the 1/12 tonne or 1/8 tonne collected.

Savings in collection = £1 per tonne MSW for 100 kg scheme

Savings in collection = £1.50 per tonne MSW for 150 kg scheme

The disposal cost is saved in full. Hence, this is equal to the avoided disposal of 1/8 or 1/12 of a tonne. The

disposal cost has been estimated at £35. (Note that where recycling credits are paid, these should reflect the

marginal disposal cost, and will therefore be internalised in net cost figures for the recycling scheme. Note also

that only two schemes in our study were receiving such credits).

Savings in disposal = £2.92 per tonne MSW for 100 kg scheme


9) Costs of Scheme

The net cost of these changes (collection of 83kg – 125 kg recyclables per tonne) is, for all but one of the

schemes we have examined, between £6 and £8 per household. Per tonne of MSW, one must reduce this by a

factor of (1/1.2).

Costs of kerbside scheme -£5 to £6.67 per tonne of MSW

Balance Sheet

Though the kerbside scheme has a net cost of £5 to £6.67 per tonne, if one subtracts from this the savings in

disposal (where no recycling credits are in place) and the savings in collection, these net financial costs begin to

approach zero. At 100kg, these savings may be £3.92. The higher end net cost (of all costs and savings) of the

scheme would be £2.75 per tonne of MSW, the lower end, £1.08.

A variety of social costs and benefits are generated at the same time. The first thing to note is that the range of

positive benefits from recycling covers values far in excess of any of the negative contributions. Also note that

the negative externalities associated with the scheme relate to a) the change in externality at the landfill (due to

changed mass and composition of waste); and b) the collection of materials and their transport to reprocessors.

Whilst the former (a) is easily offset by the benefits from recycling, at least some of the latter (b) are already

internalised so that to the extent that recycling generates net disbenefits from transport, some of these are

already being ‘paid for’ through fuel duty. The non-UK based transport components (associated with air

emissions from sea transport of primary materials) are not internalised so remain attributable in full as a benefit

from recycling.



It is not ‘proven’ that kerbside recycling justifies the expenditure outlay made upon it. It should be restated,

however, that not all pollutants have been quantified and this analysis, like most others of its kind, remains

both incomplete, and plagued by uncertainty. However, the evidence is highly suggestive. Note that there are no

benefits assigned to recovery of textiles in the analysis.

The net effect of integrating kerbside recycling, relative to ‘incineration only,’ is summed up below.

BOX 3: PRIVATE AND EXTERNAL COSTS OF KERBSIDE COLLECTION RELATIVE TO

‘INCINERATION ONLY’ SCENARIO

The following is a summary of the cost changes from the ‘incinerate all’ base scenario. The changes are

calculated with the tonne of MSW as the functional unit under examination (i.e., we are looking at the net

change in all costs, or at least those covered in the study, when materials are collected at kerbside).

1) External Cost Savings Associated With Avoided Extraction of Primary Material

As Box 1.



2) External Cost Savings Associated With Transport of Primary Material

As Box 1

External cost savings from avoided imports £0.11 to £1.48 at 100 kg/household and £0.17 to £2.21 at

150 kg/household

3) External Cost Savings (Avoided Residuals Collection)

This is calculated by assuming that one saves on collection of half the material that would otherwise be

collected at the RCV. One therefore saves in financial (see below) as well as external cost terms. The fraction to

be considered is one half of that fraction of waste being recycled which would otherwise have to be collected.

The external cost saving is therefore:





Where 30km trip avoided Max £0.55 (150 kg scheme), Min £0.01 (100kg scheme) per tonne MSW






4) Benefits From Recycling

As Box 1

At 100 kg/household external benefits of between £3.25 and £91.68 per tonne of MSW.

At 150 kg/household external benefits of between £4.87 and £137.53 per tonne of MSW

5) Transport-Related External Costs of Kerbside Collection Per Tonne of MSW

As Box 1.



6) Transport To Reprocessors (Indicative Figure Only)

As Box 1



7) Change In Externality Owing To Change in Weight and Composition of Material Incinerated

These range from -£0.13 to -£27.14 per tonne MSW (see Table 50)

8) Savings in Private Costs of Residual Disposal / Recovery and Collection Costs

As in Box 1, there is a collection (of residual) saving to be considered on the financial side. This may be

estimated as half the collection cost (estimated at £25) of the 1/12 tonne or 1/8 tonne collected.

Savings in collection = £1 per tonne MSW for 100 kg scheme

Savings in collection = £1.50 per tonne MSW for 150 kg scheme

The disposal cost is saved in full. Hence, this is equal to the avoided disposal of 1/8 or 1/12 of a tonne. The

disposal cost has been estimated at £45. (Note that where recycling credits are paid, these should reflect the

marginal disposal cost, and will therefore be internalised in net cost figures for the recycling scheme. It will be

recalled that only two schemes in our study were receiving such credits).



9) Costs of Scheme

As Box 1.

Costs of kerbside scheme - £5 to £6.67 per tonne of MSW



Balance Sheet

As in Box 1, for a net financial cost of £5 to £6.67 one generates a variety of private and social costs and

benefits. Savings in this case may be as high as £4.75, almost equal to the lower end of the schemes’ net costs.

It is notable that under certain assumptions, the reduction in net benefits from incineration is much greater than

the reduction in net benefits from landfill (shown in Box 1). This raises the question as to whether, in terms of

the externalities included here, the inclusion of recycling at kerbside might not make a ‘landfill-based’ scheme

more attractive than an ‘incineration-based’ one. Such a conclusion would have to be treated with some caution.

Some of the comments made in Box 1 apply equally here. The range of positive benefits from recycling spans

values in excess of any of the negative contributions, though the reduction in benefits from incineration is

significant. Some of the transport externalities are already internalised so that to the extent that recycling

generates net disbenefits from transport, some of these are already being ‘paid for’ through fuel duty. The non-

UK based transport components (associated with air emissions from sea transport of primary materials) are not

internalised so remain attributable in full as a benefit from recycling

It is not proven that kerbside recycling justifies the expenditure outlay made upon it. We re-iterate the point that

not all pollutants have been quantified and this analysis, like most others of its kind, remains both incomplete,

and plagued by uncertainty. However, the evidence is certainly suggestive. Note that there are no benefits

assigned to recovery of textiles in the analysis.

8. MUNICIPAL WASTE IN THE NETHERLANDS

It is interesting to compare the UK approach with the Netherlands. Relative to the UK, this is one of the

countries furthest advanced in recycling and composting. There may be some lessons transferable to the UK in

seeking to meet Landfill Directive targets, not least in the extent to which the UK could achieve Article 5

targets through recycling and composting. Like some other European countries, the Netherlands focuses more

on a reproduction, almost by administrative fiat, of the hierarchy rather than a supposedly detailed analysis of

pollutants, and costs and benefits, for different streams which suffers from shortcomings that have been

discussed above. Such an approach will always be open to criticism that the level of recycling is above what is

‘optimal’, and that the financial and environmental costs are not justified by the supposed benefits. The

Swedish approach to waste management has recently been criticised for this reason (ENDS Daily 1999). But for

the same reasons as costs and benefits are difficult to estimate with certainty, so such criticisms are, in general,

quite difficult to sustain.

We are grateful for the co-operation of AOO (the Waste Management Council) of the Netherlands in helping

with this chapter. Much of the information comes from Ministry of Housing, Spatial Planning and the

Environment (u.d.) and Waste Management Council (1995), as well as from the AOO directly.

8.1 Netherlands

As environmental problems in the Netherlands came to be perceived as more serious, from the late 1960s,

sectoral environmental legislation was drafted. A reappraisal of the sectoral approach occurred in the 1980s. The

lack of what we might now call ‘joined up governance’ led to dysfunctional approaches between sectoral

policies as interdependencies were downplayed. More cohesive legislation was seen as desirable and this was



reflected in the 1980 General Environmental Provisions Act. The decision to convert this legislation into a

single integrated environment act led, in 1993, to the Environmental Management Act entering into force. This

effectively integrated the two major Acts concerning waste, the Chemical Waste Act of 1976 and the Wastes Act

of 1977. A ‘Wastes’ chapter to this act was added in 1994 (see below).

The amount of household waste generated in the Netherlands is believed to have tripled in the years between

1960 and 1990. This growth has kept pace with increases in national income. In 1998, household waste was of

the order 7 million tonnes. In the late 1980s, measurement of dioxin emissions resulted in closure of a number

of waste incineration plants and concern also grew for the lack of environmental protection undertaken at some

landfills. Because there was so little support for the development of new final waste management options, waste

actually began to be stored temporarily in inland shipping barges. The problem had reached crisis levels.

Reflecting the waste management hierarchy that has been more or less universally agreed (though in the

Netherlands, the order of preference is termed Lansink’s ladder – see Figure 4), the focus of policy became waste

prevention. Wastes that could not be prevented were to be re-used or recycled where possible. Note that this

appears to have reflected not just debate concerning the environmental impacts identified, but also the fact that

there was almost no social and political support for new landfills or incinerators. For the wastes that could not

be re-used or recycled, disposal has to occur in such a way as risks to the environment are ‘acceptable’ (the

concept of ‘leakproof disposal’ was developed).

Figure 4 : Lansink’s Ladder – Preferences for Waste Treatment in the Netherlands

As in the UK, the ladder is oriented by the desire to reduce landfill, and by the perceived desirability of waste

minimisation. Methane emissions from landfill are frequently cited as a reason for placing landfill at the base of

the ladder.

The publication, Waste Management in the Netherlands states that ‘Radical measures are required in a number

of cases to achieve the aims of waste policy.’ The Waste Chapter of the Environmental Management Act (1994)

makes it compulsory for local authorities to introduce separate collection for organic household waste. Further

reinforcing the ordering reflected in the ladder, the Waste (Landfill Ban) Decree came into force in 1995 and

prohibits landfilling of waste if there is a possibility for reusing, recycling or incinerating the waste. As of mid

1998, the Decree covered 32 types of waste including household waste, as well as tyres, car wrecks and

remediable contaminated soil. The Waste Incineration (Air Emissions) Decree regulates incineration plants

whilst the Landfill Decree includes rules for the establishment of landfills. The former sets some of the most

stringent standards for incineration plant anywhere in the world. It is believed that the effects of legislation on

the costs of final disposal have been important in promoting waste minimisation, re-use and recycling. In 1996,

the average costs of incineration were 220 guilders (£87.56 in 1999 terms) per tonne (including VAT) for non

hazardous waste. Hazardous and hospital wastes were charges at 1000 guilders (£398 in 1999 terms) and

10,000gld (£3980 in 1999 terms) respectively (again, inclusive of VAT). Landfilling of non-hazardous wastes

cost 145 guilders per tonne in 1996 (including VAT) (£57.71 in 1999 terms) in addition to a landfill levy of

Prevention

Product Reuse

Material Recycling

Recovery

Disposal with Energy Conversion

Disposal other than to Landfill

Landfill



29.2 guilders (£11.62 in 1999 terms) for non-combustible waste and 70 guilders (£27.86 in 1999 terms) for

combustible waste. Landfilling of hazardous wastes cost between 200-300 guilders per tonne (£79.60-£119.40)

at suitable ‘ordinary’ landfills and 800 guilders (£318.40 in 1999 terms) per tonne at the special storage

facilities known as C2 depositories.

An important aspect of Netherlands waste policy has been the discussions around the responsibility for specific

waste streams. The question of where responsibilities of specific actors begins and ends has been important in

shaping the approach to household waste management. The principal of co-responsibility is intended to be a

mechanism that gives some substance to the Polluter Pays Principle. Producer responsibility has been

introduced on a voluntary basis for car wrecks (where a similar system to the Swedish vehicle scrapping levy in

operation), PVC cladding, PVC pipes, photographic hazardous waste and paper/board. In general, producers

may be obliged to take back and reprocess products at their end of life under the Environment Act. Statutory

measures exist to enforce the principle of co-responsibility in the cases of car tyres, batteries and white and

brown goods. For plastic agricultural films and packaging, a combination of voluntary instrument and statutory

measure exists.

Note that the system of co-responsibility is tailored to the product under examination. Product groups seek a

balance between incentives for prevention, re-use and recycling, and the internalisation of environmental costs,

and the impact on the macro-economy, the operation of market forces, effect on free movement of goods, and

effects of possible fragmentation of disposal infrastructure (see VROM Fact Sheet 5 for more details). Note also

(in the light of UK attempts to meet Landfill Directive targets) that the Netherlands believes it has covered

some 80% of the bulky household waste items which could feasibly be covered by such schemes (i.e. where it

is possible to attribute the ownership to a specific producer).

8.2 Household Waste

In the Netherlands, the definition of household waste covers all wastes arising from private households with the

exception of wastewater and car wrecks. Bulky household waste items are usually analysed separately however.

Responsibility for collecting household waste lies with the local authority. The basic pattern is once a week

collection near each premises. As mentioned above, authorities are obliged to collect organic household waste

separately, door-to-door, though there may be deviations in specific circumstances. Local authority bylaws

mainly include rules on disposal of household waste, for example, which components have to be kept separate,

frequency of waste collection and the agencies carrying out collection.

Provincial environmental bylaws for the collection of household wastes by the local authority where the

disposal of household waste is a matter that extends beyond the interests of the local authority. As a result of

provincial environmental regulations, local authorities are obliged also to collect paper and board, glass, textiles

and small chemical waste separately. Under the Environmental Management Act, options such as compulsory

deposit return systems also exist.

The role of information is seen as critical: ‘the most appropriate instrument to get consumers to restrict the

amount of waste they produce and motivate them to separate their waste is the provision of information’. The

‘Less Waste – It’s In Your Hands’ campaign is believed to have been important in increasing recycling of one

way glass packaging from 40% to 74% between 1985 and 1996.



8.3 History of Targets

Several targets for materials re-use and recycling have been laid down in Netherlands environmental policy.55

Generally, waste policy has taken account of the European context and has sought to respond positively to this.

However, the 1988 Memorandum on the Prevention and Reuse of Wastes (MPRW) which includes targets for

household wastes, bulky household waste and other waste streams, coincided with the period of crisis in waste

management discussed briefly above. Obligatory targets for 50% materials collection (from household waste) by

2000 were set, with 60-70% being the target for the collection and useful application of bulky household waste.

Materials targets were also set, some of these (for non-returnable glass and ferrous metals) being set at 100% for

materials collection.

Packaging Covenant targets set in 1991 include a target (60%) for the reprocessing and ‘upgrading’ of materials

not reused, and a target for re-accepting waste that is not destined for re-use but is collected separately (90%) on

the part of industry. The former again contains specific materials targets whilst the latter is intended both to

encourage separate collection of wastes by local authorities, and to reinforce the notion of producer

responsibility.

8.4 National Programme on Household Waste

Between 1993 and 1995, work was undertaken to see whether a national, broadly uniform system could be

developed for separately collecting dry components from household waste in an attempt to meet targets

specified. The MPRW targets include targets for collection (the document specifies ‘re-use’) of paper / board of

70%, plastics 35%, ferrous waste 100%, non-returnable glass 100% and packaging waste 60%. The research

programme led to publication of a document (amongst others) entitled ‘Programme on Separate Collection of

Household Waste’ (Waste Management Council 1995). A basic model was developed in which:

• Glass would be collected in banks at a density of one per 650 inhabitants;

• Paper and board would be collected door-to-door at least once a month; and

• Textiles to be collected door-to-door at least once every quarter as well as by textile banks.

It was expected that for individual components of the waste stream, performance could be improved, though

equally, the targets of 100% collection for non-returnable glass and ferrous metals were unrealistic. For the non-

bulky wastes, the overall recycling target is 60%. Since the Waste (Landfill Ban) Decree prohibits the

landfilling of household waste except where there is temporary shortage of incineration capacity, almost all

household waste in the Netherlands is incinerated. A decision has been made to remove metals from ashes post-

incineration allowing an estimated minimum of 80% of metals to be recycled.

The discussions around this model are illuminating, and they echo some of the difficulties which UK systems

of recycling are already facing. In particular, one can note the following:

• In discussions concerning the responsibilities of the different actors (under the rubric of co-responsibility)

the decision was made to leave the collection function in the hands of the governmental authorities for the

time being since the municipal collection infrastructure offers the best opportunities for achieving recycling

targets by 2000. It was also thought that having separate industry-led collection structures might make

55

The document speaks of re-use targets. In fact, the re-use targets mentioned are actually targets for the collection of waste. The

language is rather confusing, so I have tried to interpret the targets unambiguously as collection and recycling targets, the ‘recycling’

target referring to the actual percentage of material converted as a fraction of the total in the waste stream.



matters more confusing for the public and reduce response rates. Hence, the costs of separately collecting

dry components are for the local authorities to deal with up to the point at which materials are transferred to

an agreed point (a depot or transfer station, or a local dealer in secondary materials). From this point

onwards, producers are responsible for the material. Authorities may offer the material at a positive market

price, or at zero cost, but the producers are obliged to accept the material at no less than zero cost (i.e. they

cannot charge the local authority a fee for accepting the material). Where producers impose conditions of

acceptance which go beyond what is regarded as reasonable on the part of the authorities, the incremental

costs incurred can be charged by the authorities to the producers (and the separation of glass by colour is

given as an example of this). The costs of transport, trans-shipment, treatment and reprocessing of waste

collected separately, or of post-collection separation from the residual component, final processing and sales

are to be met by producers. Arguably, to the extent that producer responsibility can be introduced, the costs

of waste management shift away from the waste levy and cleansing charges imposed by authorities and

towards the product price (so that there is a more direct relationship between the costs of waste management

for householders and the costs of dealing with the waste they themselves generate).

• Apart from the basic collection system for waste outlined above and the compulsory collection of organic

household waste, the discussion concerning waste separation (for metals, plastics, drinks cartons and other

small streams) tended towards consideration of post-collection separation from the residual. The report

states ‘post-separation is not yet a proven technique, however. In addition, there is still some uncertainty

about the yield and feasibility of recycling plastics. It is desirable to wait and see whether this waste

separation develops into a system that could be introduced nationally.’ The prospects for post-separation

were thought most promising for metals and for plastics but the report notes that in tonnage terms, the

yield is likely to be insignificant relative to the basic collection system.

• Other enhancements to the basic system considered were can banks, plastic bottle banks, can/bottle banks

and possibilities for collection only of drinks cartons (where producers, motivated by the Packaging

Directive, have guaranteed a payment of 50 guilders per tonne collected). It is interesting that, possibly

because of the fact that municipal waste goes principally to incinerators, there appears to have been little

discussion of the collection of metals at kerbside.

Here one has a model of household waste policy which appears to be the complete opposite to that of the UK.

In the Netherlands, the political opposition to new linear waste management facilities has driven policy towards

targets which one can safely say were unrealistic. In the UK, the continuation of landfilling, taking place

alongside the continuing extraction of minerals and ergo, the creation of holes in the ground, has led to an

apparent lack of concern regarding waste management matters.

In the Netherlands, the approach has been to set out ground rules, and to delineate basic approaches which local

authorities can follow. In the UK, the approach is more technocratic and is based upon the belief that technical

tools can guide decision making on a supposedly rational basis, hence, the development of WISARD. The

lesson from the Netherlands, and arguably elsewhere (including, increasingly, the UK) is that decisions

concerning waste management facilities are first and foremost political ones, and supposing that they are merely

technical in nature is likely to generate quite considerable resistance in the political sphere.



8.5 Results and Commentary – Relevance for the UK?

In 1996, waste arisings were 6.18 million tonnes from 15.5 million inhabitants (in 6.5 million households). Of

this, 44% was collected separately for recycling / re-use. Percentages for separate collection by material were:

Glass 74%

Organic household waste 58%

Paper and board 47%

Textile 21%

Plastic 1%

From waste incineration slag:

Ferrous metal 60%

Non-ferrous metal 33%

This is quite interesting for a number of reasons. Firstly, the Draft Strategy for England and Wales suggests

limits to the degree to which one can collect recyclables and compostables. With the exception of textiles and

plastics, all of the implied collection rates for materials, even in the high participation rate scenarios, are

roughly what the Netherlands is already achieving. If one then looks at what the programme on separate

collection regards as feasible, all of the targets apart from that for textiles are much lower than in the

Netherlands. Whilst there may be reasons for these differences, it is not entirely clear what those would be (the

composition of waste in the Netherlands is similar, though the arisings per household seem slightly lower). In

particular, the Netherlands appears to be more persuaded by the fact that if local authorities have to, and indeed,

do collect waste from all households, there is no reason why, in principle, they could not do this for specific

fractions of material.

Why is this important? In setting a framework for meeting Landfill Directive targets, it is widely believed that

there is likely to be resistance to the construction of large numbers of incinerators. To what extent therefore

could the targets be met not through incineration, but through recycling and composting. If one diverted

biodegradable fractions of municipal waste at rates already being achieved in the Netherlands (i.e. organic

household waste at 58%, paper and board at 47% and textiles at 21%) then under different scenarios for waste

arisings growth and composition, the following are true:

Assuming 30 million tonnes arising in 1998, and assuming that the total biodegradable fraction is 62%, the

level of recycling one achieves by applying the Netherlands rates is 45% of all biodegradable wastes. The

biodegradable fraction of household waste used here, which is that used in the Draft Strategy for England and

Wales, and which we used in work for DETR on the Landfill Directive, is openly disputed. There are those who

believe the putrescible fraction is closer to 40% of the total, and that probably, as a result, we are looking at a

higher proportion of MSW that is biodegradable. Indeed, the biodegradable fraction may be as high as 70%

with each of paper and putrescibles accounting for more than 30% of the total. Supposing one takes figures of

34% for putrescibles, and 31% for paper and card, with a total MSW fraction of 70%, then applying the

Netherlands rates, one obtains 50% of all biodegradables. Note that whilst there may be discards from

processing of these wastes, our understanding is that actually these would not then fall under the Directive (as

they would be industrial wastes). If this interpretation is correct, our own earlier work has been slightly

inaccurate. We will term the two scenarios discussed here the low, and high, biodegradable content scenarios

(LBC and HBC). We now address the issue of targets in the context of different growth rates of the municipal

waste stream.



0% growth rate

At 0% growth, under the LBC option, one needs to recycle or compost between 14 and 19% of all

biodegradable waste in 2010, 38-43% in 2013 and 53-58% in 2020 to ensure the landfill Directive targets are

met. The lower figures apply if one assumes all already planned capacity for EfW is on stream. Doing what the

Netherlands does today, we would meet all but the final target. We would be at most 13% short (in terms of

diversion of total biodegradable waste) of a target which does not fall until 2020. This is equivalent to a need

for capacity that can divert 2.3 million tonnes of BMW. Alternatively, with planned EfW capacity on stream,

the additional capacity requirement would be a mere 1.4 million tonnes of BMW diversion. The time horizon

allows for considerable thought and activity to occur in the meantime.

Under the HBC option, one finds a similar situation. Though the total biodegradable content is greater, one

recovers greater fractions of the putrescibles under the Netherlands figures, so higher proportions of putrescibles

improve the overall diversion of BMW. On the other hand, the allowable quantities to landfill fall as the

proportion of BMW in the waste stream increases (Landfill Directive targets are based upon 1995 levels of

landfilling of BMW).56

The additional capacity requirement actually remains the same.

1% growth rate

The effect of compounded growth rates is significant. In this case, one succeeds in meeting the 2010 target

under both LBC and HBC scenarios, but fails to meet the 2013 target under either scenario. The shortfall is of

the order 8-13% of total BMW at the time, which translates to some 1.9-3.2 million tonnes of BMW diversion

capacity in 2013.

Higher growth rates

Once compounded growth rates go beyond about 2.3%, one fails even to meet the 2010 target under either

scenario.

One of the most obvious messages of this exercise is that it confirms what was already known – that the most

important measures in meeting Landfill Directive targets are likely to be those that target the minimisation of

the (biodegradable) waste stream. In this respect, the focus is likely to fall upon paper and board, and possibly

nappies, especially since the sorts of producer responsibility approaches which one could adopt for these waste

streams are simply inappropriate for putrescibles. As regards putrescibles, the fact that collection of these from

households may become mandatory under the proposed Composting Directive makes it possible that collection

rates may increase to levels similar to those in the Netherlands. Education will need to play a crucial role, just

as it has done in the Netherlands. Note that the Waste Management Council estimated that by 1994 around 73

per cent of all households were separating their wet waste (kitchen and garden waste) at source, and

representatives now claim that this proportion has increased to almost 100 per cent.

8.6 Costs

None of this has happened free of charge. The costs of waste management in the Netherlands have increased

significantly since 1991. The per capita waste levy more than doubled between 1991 and 1997 and now stands

at 377 guilders (£122). It is expected to peak at around 435 guilders (£141). Table 51 gives more detail on

costs. Note that data on costs and quantities is collected in the Netherlands every two years, and that since

1995, more and more municipalities (two-thirds in 1997) transmit financial data electronically. The response

56

It would be awkward, to say the last, to argue to the Commission that the composition of household waste had changed dramatically in

the period 1995-1999.



rate for the surveys (by inhabitants covered) are 94% (waste quantities) and 79% (waste costs) (the survey is

split in two parts).

One of the most interesting observations appears to be that both the costs of collection and the costs of

treatment are greatest for large and small municipalities (in terms of numbers of inhabitants). In both cases, the

municipalities of 10,000-50,000 inhabitants perform best in terms of collection and treatment costs per tonne.

The economy of scale thesis appears not to hold necessarily. It is difficult to tell why this should be the case

without closer examination of specifics, but there may be diseconomies of scale in terms of management and

logistics.

Table 51: Costs And Income From Removal Of Waste By Main Activity And Region, 1997

Costs IncomeTotal

Collection Treatment Others As % ofCost

NETHERLANDS 1,023.62 498.07 426.97 98.58 810.95 25.53

Province

Groningen 42.02 19.07 18.75 4.53 38.14 29.41

Friesland 42.02 17.45 19.72 4.85 35.23 27.15

Drenthe 34.91 14.22 17.13 3.88 26.83 24.89

Overijssel 68.52 30.06 31.03 7.43 54.95 25.86

Flevoland 19.39 10.34 7.11 1.94 15.51 25.86

Gelderland 113.77 46.54 53.65 13.25 84.68 24.24

Utrecht 66.58 36.20 23.27 7.11 50.74 24.56

Noord-Holland 178.09 102.14 62.06 13.90 134.46 24.24

Zuid Holland 245.64 125.41 103.11 17.45 191.34 25.21

Zeeland 22.95 10.99 10.02 1.94 20.04 28.12

Noord-Brabant 125.73 55.27 55.27 15.19 108.92 28.12

Limburg 64.00 30.71 25.86 7.43 50.42 25.53

By Size of Municipality

Under 5,000 inhabitants 8.08 3.56 3.56 0.65 7.11 28.77

5-10,000 inhabitants 55.59 24.24 24.89 6.79 48.48 28.12

10-20,000 inhabitants 159.67 66.58 74.02 19.39 131.55 26.50

20-50,000 inhabitants 272.15 120.24 121.85 30.38 234.01 27.80

50-100,000 inhabitants 159.67 80.48 63.67 15.19 130.90 26.50

Over 100,000 inhabitants 368.14 202.98 138.98 26.18 258.90 22.63

By Degree of Urbanisation

None 148.03 59.79 71.11 17.13 127.35 27.80

Low level 185.53 77.25 85.65 22.63 158.70 27.80

Average level 203.63 96.96 84.68 21.98 164.84 26.18

High level 243.70 122.18 97.93 23.59 192.96 25.53

Very high level 242.41 141.89 87.59 12.93 167.10 22.30



9. CONCLUSIONS AND COMMENT

9.1 Valuation of Costs and Benefits of Waste Treatment Options

The quantification of external costs and benefits requires one to value them in some way. Valuation of external

costs and benefits has the attraction for policy makers of enabling a ranking of the outcomes of one policy or

action relative to others. The component pros and cons of one approach or another can, at least in principle, be

summed together to allow a judgement to be made as to which approach might be ‘the best’. Attractive though

this may be in principle the attraction is superficial. The significance accorded to the numerical outcomes of

valuation exercises is, on the evidence of recent studies, more significant than can be justified by the scientific

uncertainties and methodological shortcomings which cannot be assumed away in conducting such analyses.

That such emphasis on the need for valuation should be prevalent at the current time is somewhat surprising for

two reasons:

1. A considerable body of literature now exists which question the degree of certainty that can be attached to

scientific information. Note that this is not just a question of whether we can know things within

‘knowable’ error bounds or confidence intervals, or whether we can assign probabilistic risk to certain

outcomes, but also one of facing up to the not-so-simple fact of uncertainty in our knowledge even of the

impacts which we might seek to place values upon.

2. The methodologies themselves are subject to widespread criticism from both inside (albeit expressed

sometimes only privately) and outside the economics discipline. This is not the place to review the

available literature here, merely to make the point that critics can be sub-divided into those who levy

fundamental criticisms (which question the validity of the whole approach), and those whose criticisms are

internal to the debate about valuation (concentrating principally on methodological shortcomings).

On the other hand, there are enough opposing trends to make the prominence of valuation techniques

understandable.

The defence of various valuation methodologies usually takes the form that ‘they are the best we have

available.’ But this pre-supposes a need for valuation for a specific purpose, and frequently, it presumes the

necessity of using cost-benefit analysis as a decision making tool. It is not answer to the question, ‘Does

valuation have a role in the policy process?’ which is the question that critics frequently ask.

As Harvey (1997) points out:

‘the critic of money valuation, who is nevertheless deeply concerned about environmental degradation, is faced

with a dilemma: eschew the language of daily economic practice and political power and speak in the

wilderness, or articulate deeply-held nonmonetizable values in a language (i.e. that of money) believed to be

inappropriate or fundamentally alien’.

It has not been our intention to ‘speak in the wilderness’, hence the efforts to develop on the one hand a credible

analysis in the mould of earlier works, but on the other, one that maintains awareness of the ‘warts’ of the

analysis, rather than simply presenting the ‘’n’all’ aspect of the work.

It is noticeable that more honest attempts at valuation are always forced to grapple with a tension between the

desire to be influential in the policy domain (and as Bromley (1989) has argued, there are few professions that



seem so well-equipped to influence policy), and recognition that the numbers have limited value. RPA and

Metroeconomica (1999), having admitted that valuation methodologies have their faults, conclude a recent

report on cost benefit approaches by stating:

There is little justification, however, for omitting non-marketed goods from the cost-benefit equation on the

basis of accuracy alone. This may result in the decision-maker implicitly assigning them a value of zero,

which in turn will result in a mis-allocation from an opportunity costs perspective.

If we had followed this approach in this analysis, we would be doing little more than guessing at a large

number of the external costs which are involved. The obvious point to be made is that assigning non-marketed

goods a wrong value is as likely (perhaps more so – numbers have a habit of being quoted out of context) to

result in mis-allocations as attributing no value. The point to be made here is that the analysis should try to be

honest in its approach, and also, in the applicability of its results. What politicians do with numbers once they

are generated is (arguably) up to them, but they should be left in no doubt as to the severity of the limitations

that may be applicable.

When uncertainty is present, there is no way of "magicking" it away. Some attempt to grapple with the issue

was made by AEA in work for the European Commission (IIASA et al 1998). Alongside a form of Delphi

survey, which produced rankings of perceived confidence in estimates of damages related to specific types of

effect, the authors produced confidence intervals for the valuation of specific effects. In that analysis, the

smallest 95% confidence interval spans four orders of magnitude (the ratio of the high to the low end of the

range is over 22,000). It is worth quoting the authors’ views of the implications of such wide variations:

‘Overall this part of the analysis succeeds in providing quantitative data on uncertainty, but on its own fails

to clarify issues. [That is, beyond stating that the benefits analysis is subject to large uncertainty, which is

already widely appreciated]. Some commentators will no doubt say that the existence of large uncertainties

undermines the credibility of benefits analysis as a tool for policy makers. In fact we regard the converse as

true: the fact that possible errors are large makes it all the more essential that benefits analysis is carried out

so that policy makers develop an appreciation of the potential risks of their actions.’

The authors are probably right in their view that some commentators will disagree with their view. They justify

their own viewpoint by making the point that policy makers might be faced with decisions in different

contexts, with the same cost implications, which are more or less certain to generate low benefits, or far less

certain to generate potentially large benefits. Two points are worth making here:

• The authors do not suggest why this aids decision-making. The decision still becomes a political one based

on judgements as to the risks associated with inaction (or alternative action) in respect of either issue.

• More importantly, this is not always the type of decision facing policy makers. Typically, decision-makers,

including local authority waste managers, are trying to find the best course of action, or policy decision to

take in the context a specific problem. They are not choosing between a portfolio of policy ‘possibilities’

from completely different contexts.57

Typically, they will have to balance alternative approaches to dealing

with the same set of pollutants / wastes in circumstances where no specific course of action is ‘vector

superior’ to all others. This means that the uncertainties pervade the decision making process, and there is

limited scope for claiming that the uncertainties are less under specific courses of action.

57

At least, this will not be the case at the level of a decision-maker. There may be considerations regarding parliamentary time which

have to be taken into account by Government.



The difficulties in arriving at clear decisions remain.

In asserting that environmental valuation has such a role, valuation exercises have to face head-on a number of

criticisms relating to what it is they are valuing (the existence of scientific uncertainty, the location-specific

nature of some impacts), their accuracy (the existence of statistical, measurement and data errors), their

completeness (can attempts at valuation account for everything, and where it does not, how sure can we be that

what is omitted is not important?), their methodological flaws (of both fundamental nature, and those internal

to the approach), and their suitability for the purpose (even if valuation exercises were carried out perfectly, there

is no reason to believe that such an approach to decision-making should over-ride other legitimate approaches to

decision-making). Decision makers rightly have concerns beyond the outcome of valuation exercises. Policy

makers have to exert political judgements, all the more so in cases where more technocratic approaches (such as

LCA-based valuation) provide a relatively inconclusive guide (for all the reasons mentioned) to knowing what

should be done.

Waste management provides a fascinating arena for the application of valuation techniques. The presumption

has been that with life-cycle analysis as its hand-maiden, valuation can provide the route to clear decision

making in this regard. It is worth quoting the Draft Strategy on the matter:

‘Life cycle assessment is the systematic identification and evaluation of all the environmental benefits and

disbenefits that result from a product or function throughout its entire life. Applying LCA to waste

management identifies what the impacts on the environment from any waste management process are likely to

be and can take into account all the environmental effects of any proposed change . For example, this might

include changes in the amount of fossil fuels consumed in power stations and the change in pollution

associated with this. LCA thus provides a basis for making strategic decisions on the ways in which

particular wastes should be dealt with, in order to develop a truly sustainable strategy for the management of

our waste’ (our emphasis).

Yet even where we know a good deal about emissions to the environment, we do not always know the

associated environmental effects with much certainty. It is not straightforward to move from knowledge of

emissions first to knowledge of impacts, and then to some valuation of those impacts which enables a strategic

decision to be made in the absence of controversy. Hence, whilst tools such as WISARD can present an analysis

of emissions, the fact that different sources of the same pollutant, emitted in different locations, can have quite

different impacts is not appreciated.

Whilst valuation can claim to account for some of the shortcomings of LCA-based analysis (notably, the

absence of any consideration of disamenity in the siting of waste management facilities – see above), these

approaches undoubtedly open up new concerns relating to ethics and income distribution. Shall we, for

example, choose to site all waste facilities where willingness to accept the facility is lowest?58

If we are

assuming energy from waste plants displace coal at the margin, to what extent will local residents, faced with

potentially high ‘non-global’ air pollution externalities, be happy to accept the facility on grounds of the

societal benefits it generates (principally for those living in the vicinity of a coal-fired power station)? To the

extent that these are real, the so-called Coasean solution might be to ask those near such coal-fired stations to

pay those in the vicinity of a proposed energy from waste plant to accept them. It seems unlikely that they

would, though there are schemes in the US in which compensation is paid to residents in order to encourage

them to accept the presence of waste treatment facilities such as incinerators.

58

See Harvey (1997) for an excellent discussion of the whole issue.



Does it help or hinder engagement of citizens to tell them what value is being placed upon the deaths that will

‘be brought forward’ because of the siting a specific facility? Certainly, this reveals a limitation of ‘externality

adders’ approaches. These generate a proxy number rather than clear information concerning impacts. They do

not tell (and nor does a life-cycle inventory) how many deaths will or will not ‘ be brought forward’ under

specific scenarios.

The debate about valuing deaths brought forward, or statistical lives, risks losing sight of more fundamental

principles. If we are already accepting the scientific validity of studies that suggest direct links between the

emissions of specific pollutants and the increased likelihood of death, then there must come a time at which one

legislates heavily to phase out the pollutant concerned, all the more so since those people most likely to have

their death ‘brought forward’ are the most vulnerable. There are lines which we cross at our collective peril, and

if with hindsight and new information we recognise we have already crossed it, all the more reason to step back.

In discussing the issue of causal proof, Beck writes (from a German perspective):

‘In other countries, quite different norms apply to the validity of causal proof. Often, of course, they have only

been established through social conflicts… judges in Japan have decided they will no longer interpret the

impossibility of a rigorous proof of causality to the detriment of victims and thus ultimately against everyone.

They already recognize a causal connection if statistical correlations can be established between pollution

levels and certain diseases… In Japan, a number of firms were obliged to make enormous payments to injured

parties in a series of spectacular environmental trials. For the victims in Germany, the causal denial of the

injuries and illnesses they have experienced must seem like sheer scorn. As the arguments they collect and

advance are blocked, they experience the loss of reality in a scientific rationality and practice that have always

confronted their self-produced risks and dangers blindly and like a stranger’ (Beck 1992).

This ‘economic Cyclopia of techno-economic rationality’, in which ‘the really rather obvious demand for non-

poisoning is rejected as utopian’ (Beck 1992) increases the prevalence of the NIMBY phenomenon. There have

to be real concerns that the exercise of the same ‘rationality’, in the technical quest to determine the Holy Grail

of BPEO, will over-ride the quite legitimate concerns of local people if not for their health per se, then for their

‘otherwise uncurtailed lives.’

9.2 Study Results – The Effects of Kerbside Recycling

The limitations of the external cost analysis we have undertaken are significant. The key results of the analysis

are summarised in Boxes 3 and 4. These list together the positive and negative factors contributing to the

balance of costs and benefits, private and external, when kerbside schemes are introduced against a backdrop of

either ‘landfill only’ or ‘incineration only.’



BOX 4: SUMMARY OF PRIVATE AND EXTERNAL COSTS AND BENEFITS OF KERBSIDE

COLLECTION RELATIVE TO LANDFILL ONLY SCENARIO

POSITIVE CONTRIBUTIONS

A) External Cost Savings Associated With Avoided Extraction of Primary Material



B) External Cost Savings Associated With Transport Of Primary Material

Avoided imports £0.11 to £1.48 at 100 kg/household and £0.17 to £2.21 at 150 kg/household

C) External Cost Savings (Avoided Residuals Collection)

Where 60km trip avoided Min £0.01 (100kg scheme) Max £1.09 (150 kg scheme) per tonne MSW

Where 80km trip avoided Min £0.02 (100 kg scheme) Max £1.46 (150 kg scheme) per tonne MSW

Where 100km trip avoided Min £0.02 (100 kg scheme) Max £1.82 (150 kg scheme) per tonne MSW

D) Benefits From Recycling

At 100 kg/household £3.25 to £91.68 per tonne of MSW.

At 150 kg/household £4.87 to £137.53 per tonne of MSW

E) Savings In Private Costs of Residual Collection

Savings in collection £1 per tonne MSW for 100 kg scheme

Savings in collection £1.50 per tonne MSW for 150 kg scheme

F) Savings in Private Costs of Residual Disposal / Recovery

Savings in disposal £2.92 per tonne MSW for 100 kg scheme


NEGATIVE CONTRIBUTIONS

G) Transport-Related External Costs of Kerbside Collection Per Tonne of MSW



H) Transport To Reprocessors (Indicative Figure Only)



I) Change In Externality Owing To Change in Weight and Composition of Material Landfilled

Range from +£0.02 to -£3.09 per tonne MSW (see Table 49)

J) Costs Of Scheme

The net cost of scheme £5 to £6.67 per tonne MSW.



BOX 5: SUMMARY OF PRIVATE AND EXTERNAL COSTS OF KERBSIDE COLLECTION

RELATIVE TO INCINERATION ONLY SCENARIO

POSITIVE CONTRIBUTIONS

A) External Cost Savings Associated With Avoided Extraction of Primary Material



B) External Cost Savings Associated With Transport Of Primary Material

Avoided imports £0.11 - £1.48 at 100 kg/household and £0.17 - £2.21 at 150 kg/household

C) External Cost Savings (Avoided Residuals Collection)

Where 30km trip avoided Min £0.01 (100kg scheme) Max £0.55 (150 kg scheme), per tonne MSW

Where 50km trip avoided Min £0.01 (100 kg scheme) Max £0.92 (150 kg scheme), per tonne MSW



D) Benefits From Recycling

At 100 kg / household £3.25 and £91.68 per tonne of MSW.

At 150 kg / household £4.87 and £137.53 per tonne of MSW

E) Savings in Private Costs of Residual Collection Costs

Savings in collection £1 per tonne MSW for 100 kg scheme

Savings in collection £1.50 per tonne MSW for 150 kg scheme

F) Savings in Private Costs of Residual Disposal / Recovery Costs



NEGATIVE CONTRIBUTIONS

G) Transport-Related External Costs of Kerbside Collection Per Tonne of MSW



H) Transport To Reprocessors (Indicative Figure Only)



I) Change In Externality Owing To Change in Weight and Composition of Material Incinerated

Range from -£0.13 to -£27.14 per tonne MSW (see Table 50)

J) Costs of Scheme

Net costs £5 to £6.67 per tonne MSW.

Perhaps the most interesting point to make on the basis of these results is that the costs of recycling need not

be great. As discussed in Boxes 1 and 2 in Chapter 7, Those attaining 12.5% percent diversion are likely to

find that the avoided costs of collection and disposal begin to approach the net costs of running the scheme.

These positive contributions are not disputed (though as mentioned earlier, some schemes will be fortunate

enough to receive recycling credits in which case the avoided disposal costs are already incorporated in a

scheme’s net costs – to add the avoided disposal costs would therefore imply double-counting in these cases).

At higher disposal costs and / or higher participation rates, recycling schemes will begin to pay for themselves,

although the prices paid for materials will play an important role in determining whether that is indeed the case.



The experience from elsewhere suggests that market development schemes can lead to materials prices rising

rather than falling in an otherwise over-supplied market.

Table 52 presents a different representation for incineration and landfill with and without a kerbside scheme

with a hypothetical recycling rate for dry recyclables of 20.8% (equivalent to 250 kg per household, a high

performance scheme on the basis of those examined). For Table 52: External Costs Associated with Landfill

and Incineration, With and Without Recycling Schemes see tables.pdf. This is the highest rate at which dry

recyclables can be extracted in the proportions in which they were removed by the schemes we have examined

(since rates of extraction are limited by assumed waste composition). It is also the same as the highest rate of

extraction per participating household in the schemes examined by us.59

The scheme is assumed to have

collection rounds of 14.4 km whilst the incineration collection round is 30km and the landfill collection round

is 100km. The net costs of the kerbside scheme are assumed to be twice that of the schemes described above.60

Two points are worth noting:

• Whilst in the best case scenario, incineration on its own is preferable to landfill on its own (it generates

greater net benefits), when the recycling scheme is introduced, the situation reverses. The reason for this is

that more of the benefits associated with incineration are associated with the presence of materials (in

particular, steel, aluminium and paper) in the waste stream which kerbside collections remove. This applies

to a lesser extent in the case of landfill.

• The best case in terms of overall performance, in both low and high externality adder scenarios is the

landfill with recycling scheme. This is partly due to the point just mentioned, but also due to the assumed

lower costs of landfill. This situation could change over time owing to landfill tax and other factors.

Equally, some incinerators may have lower costs than we have supposed here. Much also depends, as will

be recalled from Chapter 6, on assumptions concerning efficiency of energy recovery (for both landfill and

incineration)), materials recovery (for incineration), oxidation rates at the cap of the landfill, efficiency of

landfill gas capture, flue gas emissions from incineration, etc.

59

Clearly, 100% removal of specific materials is a significant achievement. Rather than suggesting an ‘impossible rate of extraction’, the

fact that one kerbside scheme extracts 246kg per participating household suggests that the composition data we have used may be

inapplicable. It is straightforward to run the models with new waste compositions.60

We have not employed analysis of the likely change in costs associated with increased rates of materials extraction per household and

participation. From the schemes we looked at, however, one could make an estimate of this, and it is more than likely that the costs would

be less than double the current levels because of increased collection densities, falling fuel costs and labour time per unit of material, etc.

On the other hand, in the absence of growing markets for secondary materials, one could argue that were such schemes present more

widely, the materials revenue would fall relative to the counterfactual situation.



We have been careful once again to stress the incomplete nature of this work in the Table. The caveats we have

carried throughout the report remain.

Given the size of the uncertainties involved, it is debatable whether one should accord any significance to the

absolute ordering of these schemes. However, there are some pointers here as to how one might use LCA-based

valuations in the context of waste management. No model is a complete reflection of the reality it hopes to

capture. It will bear the hallmark of the judgements of the modeller as well as the flaws implied in seeking to

simplify the complex world we inhabit. Models can, however, help elicit the direction of effects, and prompt

discussion as to the reasons for these, when they are subject to certain ‘perturbations ’. Arguably, it is in

shedding some light upon these (as opposed to absolute values) that the sort of model we have sought to

develop may have some value. In this sense, whilst it is tempting to seize upon the headline numbers as in

some way representing estimates of the ‘absolute size’ of the externalities associated with waste management

schemes, the more interesting observations relate to the insights gained from considering the effects of the

kerbside scheme on the comparative performance of the ‘incineration’ and ‘landfill’ schemes.

Even this has to be treated with some caution. If looking at these results, a course of action suggests itself

because it would reduce the measured externalities, due regard would have to be given to the likely direction of

effects on the unquantified externalities. Just because externalities have not been measured here, that does not

mean that they are associated with insignificant effects. The danger here is that the numbers concentrate minds

only on the quantifiable effects, not on the broader picture concerning the effects of the plant concerned.

All the externalities missed for incineration and landfill are negative with the possible exception of any benefit

attributable to post-closure amenity benefits from landfill (likely to be small). For recycling, the picture is less

clear since one is effectively looking at the net impact of one process replacing another. Note however that to

the extent that benefits from aluminium and steel recycling fall, the benefits attributed here to incineration

arising from metals recovery would fall also.

Lastly, in relation to the comment about the absolute magnitude of the externalities involved, we stress again

that one should not be tempted to believe that minimisation of waste is a negative activity in that it prevents us

from acquiring net benefits through treating waste so generated. Minimisation is beyond the direct consideration

of the study, but it generates its own environmental (and, commonly, financial) benefits.

9.3 Implications for Waste Management

This type of report arguably has implications for two key levels of decision-maker. The first is the policy maker

with the remit to make alterations to the institutional framework for waste management. The second is the

planner / decision-maker, tasked with implementing strategies on the ground for dealing with municipal waste.

It goes without saying that neither can do exactly as they might wish. In particular, finance will be a constraint

and this is where both ‘levels’ of policy maker have a shared interest.

This Section is split into two and seeks to understand where this type of report leaves both types of actor.



9.3.1 Policy Makers

Government policy sets out, as a fundamental platform for launching policy initiatives, the requirement for

costs to outweigh benefits. Policy makers might hope that this type of study would suggest a clear course of

action on the basis of a cost benefit type approach. This study does not do that and it also suggests why other

studies that might appear to do so perhaps do not.

Ironically, and contrary to what we had expected, the costs of kerbside schemes in the context of integrated

management systems are relatively low. Implementing a kerbside scheme generating 12.5% diversion of

material on its own might cost as little as £1-3 (net) per tonne of MSW managed.61

This is the same as the

landfill tax increases over the next few years. As landfill tax increases, the economics will shift in favour of

recycling (certainly relative to landfill) as the avoided disposal costs increase.

As regards a ‘cost benefit analysis of recycling’ the study is not completely conclusive. Although the

environmental case (i.e. considering externalities only) for recycling a number of materials appears a compelling

one, the analysis remains incomplete in this regard. For each treatment option, there are unquantified

externalities, and the analysis remains somewhat crude. Even so, with the results as they stand, all one can say

is that the work is highly suggestive of external benefits that would close the gap between the net financial

costs of kerbside recycling and the avoided financial costs of such schemes. Strictly, the incomplete analysis

carried out here is inconclusive because of unquantified effects.

The study is a long way from being a full cost-benefit analysis. This is not to say that a complete analysis

would necessarily generate any more conclusive policy recommendations than those generated here. We have

sought to show that to the extent that one looks for very firm and conclusive pointers for policy on the basis of

an analysis of private and external costs, the unavoidable uncertainties and errors which plague attempts at

valuation of environmental externalities reduce the likelihood of this proving to be the case. Perhaps the

relatively unproblematic introduction of the landfill tax has boosted hopes in this regard. Riley 1996 notes the

role of externality analysis in setting the level of the landfill tax. But as our work has sought to show, one can

re-visit that analysis and formulate some quite different numbers, not to mention conclusions.

One of these is that, with hindsight, it would have made sense to levy different rates of landfill tax on sites

with and without energy recovery. A higher tax rate on the latter might have encouraged early installation of gas

collection equipment with significant environmental benefits. This might still be of some relevance today since

all sites accepting municipal waste will have to install this equipment anyway owing to the Landfill Directive.

Some of the most successful environmental taxes have been those which, through imposing differentials, create

incentives to switch from one course of action to another (e.g., differential taxes on leaded / unleaded petrol).

The case here would be similarly strong (and would generate benefits). As long as the differential tax lay below

the abatement cost, installation of gas equipment would be expected to occur rapidly.

In respect of the treatment options we have considered, the most glaring omissions would appear to have been:

• Disamenity (of all facilities);

• Emissions to water and land, including leachate; and

• The ‘non-environmental’ aspects of the cost benefit analysis. The latter include, for example, employment

effects which have been the subject of work undertaken by Waste Watch (1999b) and Murray (1999), as

61

This is the difference between the avoided disposal cost and the lowest net cost for the schemes we looked at (which are not in receipt

of recycling credits).



well as potential benefits from changes in balance of payments, competitiveness and so forth (some of these

being less than straightforward to quantify).

In respect of the first of these, Table 53 summarises some studies from elsewhere. For Table 53: Summary of

Waste Related Valuation Studies see tables.pdf. These can appear small relative to some specific examples. In a

affidavit for a Judicial Review of a recent Leicestershire MSW landfill site planning decision, for example,

Susan Reiblein provided evidence that her property value fell from £150,000 to between £20,000 and £40,000

as a result of interference from a landfill 340 m away. Even the operating company (Hepworth) accepted a

£40,000 reduction because of 'bad neighbour' development.



Benefits transfer is, as the studies has indicated elsewhere, highly problematic and should always be approached

with caution. The evidence here can be taken alongside the studies presented in Brisson and Pearce (1995)

(some of the studies are also featured in that work). The first thing to note is that much more work has been

done in the context of landfills than in the context of incineration. More importantly, and potentially more

relevant to the UK in the future, many of the studies look at hazardous facilities, and one would expect WTP

(willingness to pay) and WTA (willingness to accept) bids to be higher for these than for MSW facilities.

Currently in the UK, co-disposal occurs, but in the future it cannot because of the Landfill Directive.

One US study which looked at waste to energy plants is that by Kiel and McClain (cited in Brisson and Pearce

1995, but not referenced). They looked at the effects on house prices over time from the pre-rumour stage,

through to the rumour stage, the construction phase, the online phase, and later years of operation. House prices

were affected only once construction began and peaked in earlier years of operation before falling back slightly.

The effect was a reduction of the order 3% per mile in the vicinity of the incinerator. Gerrard (1994) notes that

in North America, ‘Compensation has been quite successful in siting MSW landfills and incinerators which

have much lower perceived risks than do HW/RW [hazardous waste/radioactive waste] facilities.’ A popular

form of compensation has been ‘value protection’ which compensates householders for depressed house values

(see above).

Where recycling is concerned, there may be site-related disamenities associated with industries using secondary

materials but one might argue that these, such as they occur, displace extractive activities elsewhere in the world

(see Chapter 6) as well as (in the longer-term) any disamenity related to primary processing facilities. Whereas

one finds studies, and thinking, couched in terms of willingness to pay to avoid, or willingness to accept most

waste treatment facilities, where recycling is concerned, surveys have been carried out to gauge willingness to

pay for the service. Jakus et al (1996) carried out studies to elicit willingness to pay for recycling, and

estimated this at £5.78 per household per month. Evidently, this is far in excess of the actual costs of kerbside

schemes in our mini-survey. Tiller et al (1997), on the other hand, report that in Tennessee, households would

pay $4 per month (on the basis of contingent valuation). Again, per household, this is far in excess of the costs

of kerbside recycling in our schemes.

Taken in the round, what these studies appear to do is reflect the public preference for certain approaches to

waste treatment. These are important. Sceptics concerning valuation techniques might at least recognise the

ordering of preferences that these appear to suggest. Quite how the disamenity effects of incineration, as

measured through contingent valuation approaches, would compare with landfill is unclear (and likely to vary

with location). To the extent that population densities are important, the fact that incinerators are usually in

urban or peri-urban locations would suggest that the disamenity would be larger (since more households would

be affected). On the other hand, we have only limited knowledge from studies seeking to elicit the disamenity

associated with municipal waste incineration so do not know the speed at which the disamenity experienced

falls with distance (and whether it varies with capacity, and if so how?). Brisson and Pearce (1995) suggest 4

miles as the domain of influence for landfills. It could be that this is less in the case of incinerators.

The public preference argument in favour of recycling is an important one. It is especially important given that:

• Local authority waste management is funded by local taxation, so it is appropriate that the views of the

public should be taken into account; and

• Technocratic approaches to waste management decision making are unlikely to lead to determinate

decisions as to what should be done. As far as we can discern, recycling is an environmentally beneficial

option, though again one must stress the incompleteness not only of this study, but of all others besides.



To add to arguments in support of recycling, one can point to the fact that balance of payment effects are likely

to be positive. Employment may also be generated. Lastly, there may be important softer effects from kerbside

recycling since it may imply an increase in awareness on the part of householders concerning the wastes they

produce.

If people are, in the general case – and this not universally true62

(one finds volunteer communities for waste

management facilities, frequently, it should be added, where there are gains to the community incorporated in

the package) – unfavourably disposed to the presence of large waste treatment plants, the question arises as to

how to go about minimising the need for these. This has to account for the costs of the approaches undertaken.

The key is perhaps not recycling, nor landfill, nor incineration, but minimisation. This is the area where

virtually all interests in waste management come to agreement, yet relatively little is spent on tackling the

problem, certainly as regards MSW.

A key issue in the UK at present is that householders have no incentive to minimise the waste they generate,

and nor is there much spent on educating them as to how best to do this. As long as this is the case, there may

be much to be said for moving the incentive upstream to retailers and the like. The Packaging Regulations will

have some effect in that respect through putting in place incentives for minimising packaging waste, but they

affect only packaging materials. Other components of the municipal waste stream could be addressed through

Producer Responsibility mechanisms, not necessarily of the same nature as were developed for Packaging

Waste. These could be used as a mechanism for tackling the recovery of some wastes arising in the household

stream such as direct mail, and as has been discussed elsewhere, newspapers.

The new emphasis on market development is also a welcome one. The initiatives in England, Scotland and

Wales, and the intention of some regions to incorporate plans for market development within their Structural

Funds programmes may generate an environment in which entrepreneurs can develop new industries in a more

conducive and receptive environment. But market development is specifically not a strategy for waste

minimisation

Part of the significance of minimisation, when set in the context of the Landfill Directive targets, is that, other

things being equal, a lack of it requires more facilities for biodegradable municipal waste diversion. These bring

with them the site-specific disamenity effects and the non-global air pollution effects that are directly traceable

as consequences of the behaviour of the community at large. It can be argued that it is an unjust distribution of

the effects of a ‘solution’ to what is currently a societal problem . Minimising waste will minimise the extent

of this unjust distribution of the burden of dealing with waste.

If, as Chapter 8 suggested, waste minimisation is an important strategy for meeting Landfill Directive targets,

there must be logic in extending the tax on landfill to cover incineration, particularly since energy from waste

has been exempted (somewhat strangely, given its contribution to global warming) from the climate change

levy. This would have the dual effect of tilting the financial equation in favour of recycling and promoting

minimisation. In the longer term, there would also be merit in extending the tax (albeit at a lower rate) to

recycling itself.

MEL (1999) suggested that the current maximum rate of recycling for dry recyclables might be 15%. But that

study carried with it the important caveat (overlooked by some) that this was contingent upon there being no

62

One finds volunteer communities for waste management facilities, frequently, it should be added, where there are gains to the

community incorporated in the package



change in the institutions, financial arrangements and legislation regarding waste management. It would be

difficult to see how these could remain the same for the foreseeable future given requirements for the UK to

change what is currently doing in order to comply with the Landfill Directive. The question then becomes not

‘what is best practice now?’ but, ‘if we have to change what we are doing, how can we ensure that the change

occurs smoothly, and in such a way as generates benefits such as we believe them to exist?’ Extending the tax

on landfill to incineration has the merit of giving incentives to both minimisation as well as recycling.

Revenues could be channelled into a diversion fund for which local authorities would bid (on the basis of

incremental costs of the scheme in excess of the costs of the counterfactual linear treatment option). This would

seem a more laudable use of funds and of administrative capacity than the current ENTRUST set-up. Note that

the French Landfill Tax on Household and Industrial Assimilated Waste (Déchets Industriels Banaux), now

included alongside other environmental taxes in the General Tax on Polluting Activities (TGAP), was

established in part to generate a modernisation fund for waste management.

Note that because recycling tends to save energy, there would also be logic in devoting some of the funds

generated by the Climate Change Levy to recycling and waste minimisation. This would help re-instate some

element of joined-up thinking (waste and climate change) in the design of the climate change levy that may be

lost as a consequence of the exemptions specified thus far.

9.3.2 Waste Managers

Decision-makers at the local level have been asked to pursue BPEO, and to do so frequently with very few

resources. It is something of an oddity that a key platform of waste management in the UK, BPEO, is not

particularly well understood. Privately, some waste managers cynically express the view that anything can be

the BPEO. In the absence of firm guidelines as to how this is to be applied, as well as some assurances that

their attempts to pursue BPEO will be rewarded with sufficient finance to implement it, local authority decision

makers are likely to lose some enthusiasm.

The suggestion of this work is that a technical application of costs and benefits to waste management decisions

does not lead one to conclusive results as to what the best option is. In order to even approach this, an

enormous amount of work would need to be done, much of location-specific in nature. Even then, the analysis

would, most likely, still incorporate uncertainties and wide ranges regarding various impacts. The Environment

Agency’s Life Cycle Assessment tool, WISARD, may to some extent aid decision-makers in assessing the

trade-offs which might exist across facilities. But the tool also obscures the effects of the pollution in the

environment in which it is emitted. It does not give an account of the severity of actual impacts of the pollution

concerned, nor does it consider other issues such as employment, and nor does it account for disamenity. Also,

the assessment of burdens associated with different waste management options are intimately related to what is

happening outside the treatment method itself. The question concerning displaced energy sources is such a

crucial one in assessing net burdens that it deserves to be revisited to understand how this form of system

expansion should be handled (the more so since decisions regarding energy from waste landfill and incineration

capacity are likely to have an impact a long way into the future).

In this context three points seem worth making:

• Waste management decisions (especially concerning siting of facilities) are not about to be cleansed of their

fundamentally political nature by application of technical approaches to determining BPEO. Like it or not,

waste management decisions are political ones. There is little point in either denying, or trying to



overcome this. For this reason, it is important to know not just quantities of pollutants, but which ones are

emitted where, and what will be their impact and upon whom?

• To the extent that finance is viewed as a constraint on pursuing options which are more politically popular,

the availability of finance to local authority waste managers will have significant political consequences (all

the more so if the questions advanced above are answered in transparent manner).

• The pressing need for minimisation becomes even clearer, and it is in this area that much effort should now

be expended.

The potential for minimisation remains under-explored. More pro-active authorities might take up the challenge

of minimisation rather than simply accept, in fatalistic manner, an unending growth in the waste stream,

especially following the passage of the Waste Minimisation Bill. Such authorities might seek, through

education and through awareness campaigns to influence the choice of products made by consumers, particularly

so as to maximise the recyclability / re-usability of packaging purchased.

The degree to which such education is effective might be increased by measures such as variable charging or

other imaginative incentive schemes, as well as allowing, as in the German Ordinance, consumers to leave

unwanted packaging in retail stores, a move which would bring home the message of Producer Responsibility

without bringing to the home unnecessary packaging.

In addition, authorities could encourage householders to:

• Sign up with the Mailing Preference service so as to reduce ‘junk mail’.

• Engage in home composting.

• Engage in softer ‘demonstration’ based approaches, such as using non-disposable nappies in local hospitals;

• Purchase goods with lower packaging content.

• Place labels on letterboxes stating that where this is the case, the residents have no wish to receive free

‘newspapers’ or other ‘junk mail.’ For many residents, this is simply unwanted material that goes directly

from the doormat to the dustbin, especially where there is no opportunity to return the mail to those

responsible for its being posted.

• Engage in ‘clear-out’ days on which streets become temporary waste exchanges.

Also, there might be merit in looking at the application of the Essential Requirements Regulations of the

Packaging Directive. When the Regulations came into force the Local Authority Committee on Trading

Standards (LACOTS) made a resources bid which DTI accepted and the money was added to the Revenue

Support Grant. However, because it was based on a 'soft touch' enforcement approach with low inspection

frequencies it was for not more than £120,000, divided between about 200 Trading Standards Authorities in the

UK, amounting to about £600 per authority. We understand that the lowest frequency for inspection is about 1

inspection every 14 years.

Guidance, developed by the DTI in conjunction with LACOTS, was published in August 1998. It did not give

a steer to LAs about inspection frequencies or enforcement approach. A "code of practice for optimising

packaging and minimising packaging waste" was produced by the Industry Council for Packaging and the

Environment (INCPEN) and endorsed by relevant industry associations, the Institute of Packaging and

LACOTS. A series of compliance statements for different types of packaging has been published by LACOTS -

available on the LACOTS open website at www.lacots.org.uk (under Information Services). Whilst the larger

companies have tended to ‘comply’, it is believed that there are still low levels of awareness among SMEs



about the Requirements. This is disappointing, since the only onerous requirement is record keeping, and there

may be distinct financial advantages to the companies themselves in achieving compliance.

Local Authorities have to be aware also of the possibility of a Composting Directive in the near future. This

could require that the organic fraction of waste is separately collected or home composted in the future. Local

Authorities need to be aware of these developments. Some local authority waste managers have not even heard

of the proposed Composting Directive, but were it to come in, it would affect every Authority in the land. This

argues for flexibility in strategies such that they appreciate this as a possibility. Taken together, the Landfill

Directive and the Composting Directive suggest that one logical approach for Local Authorities will be to work

together in minimising, composting and recycling materials in the foreseeable future. This would occur with a

view to reviewing progress made towards Landfill Directive targets at pre-specified dates, which would then

allow an assessment to be made of the need for other diversion capacity. The figures used from the Netherlands

in Chapter 8 suggest that much could be done to minimise requirements for other diversion facilities, but that

significant efforts in minimisation would have to made to eliminate such a need altogether.

Just as Local Authorities need to be informed about the possibilities for a Composting Directive, so it might

make sense for citizens in Local Authorities to be informed about the nature of the decisions having to be made

in their Authority. Some citizens are already being made aware of these, not always in the way they would

wish. The implications of the Landfill Directive remain ‘unperceived’ by most citizens even though every one

of them will be affected, directly of indirectly, by it. Part of a waste strategy at the local level will be to inform

all citizens of the nature of the decisions which need to be made, and the options available.

9.4 Concluding Remarks

It would have been relatively easy in this study to ‘prove’ that recycling is ‘unequivocally’ superior to other

options. Change a number here, massage the figures there, alter the ranges, and hey presto, recycling wins! But

the aim is a broader one. One lesson from this work is that however strongly one believes the quantification of

private and external costs and benefits should be an ultimate arbiter of whether or not something may or may

not be a good idea, one is likely to have to accept the fact that studies attempting to do this will be less than

conclusive. This raises questions as to how what has been stated as a ‘fundamental principle’ of Government

policy – that policies where the benefits do not justify the costs should not be introduced – can be put into

operation. Do we do nothing where the costs and benefits themselves are subject to considerable uncertainty?

Do we pretend that analyses which are less than complete, and which are trying to cope with uncertainty,

constitute the basis for policy formulation? Do we need cost benefit analysis at all to arrive at a decision? To

what extent do we fall back instead on other principles of policy making, such as the Precautionary Principle?

Or do we take actions, but ones which retain flexibility (and degrees of reversibility) whilst seeking to address

what one suspects (perhaps on the basis of a precautionary approach) is a problem in hand? These are, at heart,

epistemological questions. At what point do we know what it is we think we have to know in order to justify

action? What models of rationality should we be using in such cases (because the message is that old ones are

likely to fail us)?

Such a debate needs to be opened in the spirit of a non-dogmatic attempt to explore the alternatives to cost-

benefit based decision making. So much time and effort has been spent trying, apparently, to concoct a rationale

for one or other course of action even where the co-incidence of public desires, and what the cost benefit

analyses undertaken thus far suggest, are ‘in sync’. There will always be another dispute about the life cycle

inventories, about the value of life (not to mention the view as to whether this is an acceptable practice at all),



about the threshold value for this or that pollutant’s effect, and about the carcinogenic nature or otherwise of a

whole host of different pollutants whose impacts we don’t really understand (least of all when they are mixed in

a soup).

There will be those who are asking why, when economic analyses undertaken and the opinion of the public are

apparently aligned, are we still waiting? Why are mistrustful of other country experiences when we are ill-placed

as a nation to judge where the boundaries to recycling and composting might lie? Will the Treasury inject the

finance required to stimulate an activity which appears necessary and desirable for a variety of reasons? If it

waits for a strictly determinate outcome from cost-benefit analyses, we could be waiting for a very long time

indeed (or alternatively, as seems to be our destiny in the UK, until a new Brussels Directive emerges).

What institutional models should we design for the purposes of policy-making? It is less than clear that with 20

year objectives to meet (under the Landfill Directive), the current structures will serve us especially well.

Hukkinen’s (1999) work perhaps provides a starting point, exploring the extent to which cognitively dissonant

mental models of environmental problems are constructed by those in the policy making field.63

The suggestion

is that new institutions with less corporate perspectives are required. These have to include stakeholders:

‘the efficiency and the social good of a particular institutional design are defined by the stakeholders who

have to tolerate or enjoy the institutions… The only thing that planners can wish to do is stimulate the

creation of institutions that enable the stakeholders to search and find an acceptable concensus on

environmental issues, instead of continuing to redefine and disagree on them.’

Note that the comment on the social good resonates with the institutionally informed criticisms of Bromley

(1989). Instead of looking solely at costs and benefits, the distribution of which is dependent upon pre-existing

market structures, we need to look instead at how the structure of markets has to change in accordance with

conceptions of social welfare. Might it be pertinent to re-consider the structures of rights and obligations within

which citizens function? Should we, as citizens, have the right (as we currently do) to discard materials without

considering their potential value as a matter of course? More fundamentally, can we justify continued growth in

materials consumption in the context of debates concerning sustainability and the needs of future generations

(and projections of a 3% growth rate in the waste stream until 2020 are, to say the least, alarming)? Should, for

example, distributors of free newspapers and mailshots have the right to expect local authorities, and hence,

citizens to foot the bill for collecting and disposing of their materials? How can we justify the distributional

(il)logic of having large waste treatment plants in specific locations that deal with what is actually everyone’s

problem? Isn’t it right that everyone that everyone should be obliged to try to solve that problem? The oft-

quoted bottom line numbers of externality assessments are apt to downplay the significance of the distribution

of the costs and benefits which comprise the bottom line. (It should be added here that we can know something

about the nature of distribution of these even where their magnitude is unknown.) These are questions that seem

prior ones to consideration of costs and benefits. They imply the establishment of institutions, formal and

informal, which structure markets, and within which costs and benefits are assessed.

The focus in the UK thus far appears to be the costs of minimising the solution to the UK of meeting Landfill

Directive targets. This may not lead to particularly innovative approaches, and nor is it likely to lead to a

situation in which the potential to seize opportunities in respect of new jobs is grasped. Ironically, one obvious

way of alleviating the financial constraint, variable charging (in the United States, Pay as You Throw), in which

householders are charged directly for waste services (and which would allow them greater freedom to determine

63

The case study on Finnish waste management is particularly interesting in the context of this study.



what treatment approaches are used in their area), will not happen unless primary legislation comes onto the

statutes (despite the fact that there appears to be some degree of support for such an approach – Waste Watch

1999b). There has been some concern that this will have regressive impacts. Yet it is difficult to believe that

even if this were true, the remaining Council Tax element could not be adjusted to account for this supposedly

regressive effect.

We started this study with a hypothesis:

although the financial costs of recycling may be greater than that for other methods of dealing with waste, to

the extent that one is able to incorporate the environmental costs and benefits associated with all methods, the

economic analysis will show that when one accounts for all these, recycling will be shown to be the best

option for dealing with the relevant fractions of municipal waste .

The hypothesis is not proven conclusively. There are a number of persuasive arguments that one can present for

recycling. There are fewer for other treatment options (other than that they make the job of ‘dealing with waste’

disarmingly simple). There are more for waste minimisation (and energy efficiency).

Many other countries apparently see some merit, more than we do here in the UK, in the simple messages

conveyed in the waste management hierarchy. The simplicity and clarity of the message that the hierarchy

conveys has much to recommend it, even if it might not lead to perfect outcomes in environmental terms. It is

surely incumbent upon those who claim that it does not to state why, and to ensure that their answers are more

than simply theoretical. After all, if one is looking to give some guidance to decision makers, asking them to

follow BPEO is certainly not devoid of its own problems, and it opens up huge space for argumentation.

Countries like the Netherlands more or less reproduce the hierarchy through principles for local authorities and

householders. Denmark does the same with its waste tax. The Landfill Directive holds no great fears for these

countries, though to be sure, they have had to make awkward decisions in the past (not least, in the Netherlands

case, in terms of incineration capacity, where concerns about emissions from these led to imposition of

relatively stringent standards for pollution control).

Given this, from the point of view of Local Authority waste managers, one can do much worse than making the

hierarchy, and not BPEO, the guiding principle of decision making in waste management. Quite apart from

reproducing the ranking which most studies, albeit highly imperfect ones (as this one is), that have been carried

out suggest, this may have the twin merits of enhancing the flexibility of waste strategies, and aligning UK

waste management with European proposals for new legislation. A failure to do the latter risks painting the UK

into a corner in which it is continually dancing to a tune set in Brussels. It will be interesting to see how many

Authorities (and Regional Technical Advisory Bodies), for example, in setting out strategies to meet Landfill

Directive targets, take on board the possibility of a Composting Directive entering into force in the medium-

term.



ANNEX 1: QUESTIONNAIRE FOR RECYCLING SCHEMES USED IN THE STUDY

ECONOMICS OF RECYCLING QUESTIONNAIRE

Local authority:

Forms completed by (name):

Position:

Thank you very much for agreeing to complete this questionnaire. A brief description of the project’s aims and

objectives is given overleaf.

Please fill in all of the sections that are relevant to your areas of responsibility. We hope that most of the

information is readily available for you, however, we appreciate that some of the financial and cost data may be

more difficult for you to identify and provide. If you are unable to provide answers to any of the questions,

please be as specific as possible about why not - for example, because your accounting system cannot provide

data in this format, or you cannot release the data because of commercial confidentiality.

If you have any queries please contact either Lis Broome or Dominic Hogg at ECOTEC, while if you have

more general queries about the overall project then contact either Dominic Hogg at ECOTEC or Doreen Fedrigo

at Waste Watch.

Finally, if for any reason you are unable to complete the questionnaire, please inform Lis or Dominic at

ECOTEC as soon as possible, as we will need to identify an alternative recycling scheme instead.

Thank you again for your help with this study, your input is very valuable.

CONTACTS: ECOTEC: Waste Watch:

Lis Broome Doreen Fedrigo (project manager)

Tel: 0121 616 3656 Tel: 0171 253 6266

Email: [email protected] Email: [email protected]

Dominic HoggTel: 0117 924 9478

Email: [email protected]



SUMMARY OF CONTENTS OF FORMS

Form 1DESCRIPTION OF THE LOCAL AUTHORITY

WASTE MANAGEMENT SCHEME –

tonnes of waste disposed

Kerbside collection scheme(s) for

recyclables / compostables:• Form 2: Description

• Form 3: Costs

RESIDUALS COLLECTION:• Form 7: Description

• Form 8: Costs

TREATMENT FACILITY:• Form 4: Description

• Form 5: Costs

RESIDUALS FACILITY:• Form 9: Description

• Form 10: Costs

Form 6:

REVENUES - TREATMENT FACILITY

Form 11:

REVENUES - RESIDUALS FACILITY



FORM 1:DESCRIPTION OF THE LOCAL AUTHORITY MANAGEMENT SCHEME

TONNES OF WASTE DISPOSED

Total number of households:

Tonnes of waste recycled through kerbside:

Glass

Paper and cardboard

Aluminium

Steel

Garden and kitchen

Plastics

Other

Tonnes of waste recycled through bring schemes:

Glass

Paper and cardboard

Aluminium

Steel

Garden and kitchen

Plastics

Other

Tonnes of waste composted by Authority:

Kitchen waste

Garden waste

General putrescibles

Tonnes of waste delivered for disposal / recovery (please specify whether this includes waste arising from

kerbside / bring schemes but not recycled)

Of which, to landfill

Of which, to EFW

Please supply any results of waste analyses that have been carried out for the Authority:

Give details here, or attach any relevant documents.

Have you carried out any studies into the social or environmental costs / benefits of recycling schemes?

If so, please give details or attach any relevant documents.

FORM 2:



KERBSIDE COLLECTION SCHEMES FOR RECYCLABLES / COMPOSTABLESDESCRIPTION OF THE SCHEME

Please duplicate this Form if you have more than one recycling / composting scheme operating

Number of households serviced:

Date (year) the kerbside collection scheme started:

Types of households serviced (ACORN grouping /

density / income / other qualitative description):

Frequency of collection:

Separate collection round or at the same time as the

waste for disposal collection?

If separate collection, are the vehicles used for

recyclables the same as those for collection for

disposal?

Method of householder collection – box / bin / bags /

etc.:

Separated or co-mingled ?

Type of vehicle used – commingled / separated waste

streams? Capacity? Number of operators per vehicle?

Distance travelled by vehicle during rounds:

Immediate destination for waste (e.g. transfer station,

composting facility, etc.):

How far away is the treatment site (composting site

etc.)? (in miles)

Materials collected - please identify materials collected

and give estimates of recovery rate (%) and amounts

collected:

Collected?

Yes or No

Recovery rate

(%)

Amounts

collected

(tonnes)

Glass

Paper and cardboard

Aluminium

Metals

Garden and kitchen

Plastics

Other



FORM 2 (contd.) KERBSIDE COLLECTION SCHEMES FOR RECYCLABLES / COMPOSTABLES

DESCRIPTION OF THE SCHEME

Collected weight per household

Estimated householder participation (%)

(and how estimated)

Do you promote home composting?

If so, give details – type of system, numbers of

households involved, cost of system to household

Estimated waste reduction due to home compost

scheme (per participating household / total)

FORM 3: KERBSIDE COLLECTION SCHEMES FOR RECYCLABLES / COMPOSTABLES

COSTS OF COLLECTION (£)


10. CAPITAL COSTS

Vehicles:

Bins / boxes (cost per bin/box, or per household):

Other (including home composting bins where

relevant) - give details:

11. REVENUE COSTS

Labour:

Maintenance of equipment:

Fuel:

Publicity / information (leaflets, advertising etc.):

(please give description of what you have done to

publicise the scheme)

Capital servicing:



FORM 3 (contd.) KERBSIDE COLLECTION SCHEMES FOR RECYCLABLES / COMPOSTABLES



Other costs:

Total costs (£):

In-house calculation of cost per household / per tonne

/ per collection round (if available).

State whether this calculation is gross or net (i.e.excluding or including income receipts), and whether itis the total or marginal costs of recycling, i.e. excludingor assuming the current methods of collection fordisposal are maintained.

FORM 4:KERBSIDE COLLECTION SCHEMES FOR RECYCLABLES / COMPOSTABLES DESCRIPTION OF

THE TREATMENT FACILITY


Type of treatment facility - MRF / composting /

other:

Manufacturer:

Capacity (give units, e.g. t/year):

Operator (in-house / contractor / other):

Operator's contract start and end date:

Does the facility receive waste from kerbside

collections other than the one described above? If so,

from where?

Materials received each year: Received?

Yes or No

Tonnes received Tonnes sent to

waste

disposal/EfW

Glass

Paper and cardboard

Aluminium

Metals

Garden and kitchen

Plastics

Other



FORM 5:TREATMENT FACILITY

COSTS (£)

Please duplicate this Form if you have more than one treatment facility operating

Description of facility - MRF / composting plant /

other:

12. CAPITAL COSTS

Land acquisition:

Construction (give details, e.g. buildings / surface

improvements and hardstanding, landscaping, fuel

store, fencing, etc.)

Plant and machinery:

Other:

Total investment costs:

13. REVENUE COSTS

Maintenance (give details):

Labour:

Fuel:

Insurances:

Utilities:

Monitoring and analysis:

Other (please give details):

Total cost per year:



FORM 6:TREATMENT FACILITY – REVENUES (£)

Grants - give details:

Sponsorship - give details:

Recycling credits - £/tonne:

Sale of recovered materials:

List materials types: Revenue:

PRN revenue:

List material types: Revenue:

FORM 7:RESIDUALS COLLECTION AND TREATMENT DESCRIPTION

Number of households serviced (if all, please state

all):

Number of businesses from which waste is co-

collected (on the municipal round):

Frequency of collection:

Method of householder collection – bag / bin /

wheelie bin etc.:

Distance travelled by vehicle during rounds:

Immediate destination for waste (e.g. transfer station,

landfill, EfW plant etc.):

How far away is the immediate destination? (in

miles)

If the immediate destination is not the EfW plant or

landfill, how far is the destination from the landfill .

EfW plant to which the waste is ultimately taken?

Collected weight per household:



FORM 8:RESIDUALS COLLECTION AND TREATMENT


14. CAPITAL COSTS

Vehicles:

Bins / boxes (cost per bin/box, or per household):

Other (including home composting bins where

relevant):

15. REVENUE COSTS

Labour

Maintenance of equipment

Fuel

Publicity / information (leaflets, advertising etc.)

(please give description of what you have done topublicise the scheme)

Capital servicing

Other costs

Total costs (£)

In-house calculation of cost per household / per tonne

/ per collection round (if available).

(State whether this calculation is gross or net (i.e.excluding or including income receipts), and whether itis the total or marginal costs of recycling, i.e. excludingor assuming the current methods of collection fordisposal are maintained)



FORM 9:RESIDUALS FACILITY

DESCRIPTION

DESCRIPTION OF THE RESIDUALS FACILITY

Type of treatment facility – EfW / Landfill

Date the facility opened:

Manufacturer (EfW)/ Operator (landfill):

Capacity (if EfW):

Operator (in-house / contractor / other):

Operator's contract start and end date:

FORM 10:RESIDUALS FACILITY

COSTS (£)

Description of facility - MRF / composting plant /

other:

16. CAPITAL COSTS

Land acquisition:

Construction (give details, e.g. buildings / surface

improvements and hardstanding, landscaping, fuel

store, fencing, etc.):

Plant and machinery:

Other:

Total investment costs:



FORM 10 (contd.)RESIDUALS FACILITY

COSTS (£)

17. REVENUE COSTS

Maintenance (give details):

Labour:

Fuel:

Insurances:

Utilities:

Monitoring and analysis:

Other:

Total cost per year:

FORM 11:RESIDUALS FACILITY - REVENUES

Grants - give details:

Sponsorship - give details:

Is there a NFFO contract? Give date started, NFFO

tranche, scheme name:

Recycling credits - £/tonne:

Sale of recovered materials:

List materials types: Revenue:

PRN revenue:

List material types: Revenue:

ANNEX 2: EXTERNALITY ADDERS USED IN THE ANALYSIS



There are a number of studies that have looked at the external costs associated with different air pollutants in

particular. We have not made a completely thorough investigation of these. Below, we list studies we have

looked at and the values for the unit damage costs, or externality adders, that have been derived. Note that in

the studies reviewed in Chapter 5, there is some, though not complete, consistency in estimates. This is

perhaps unsurprising given the fact that some of the studies involved similar personnel.

For some key pollutants, we show some of the studies consulted. For others, we simply present the ranges used

and the sources consulted.

Particulate Matter (PM10)We have looked at UK and European studies (see Table 1). For comparison, RPA and Metroeconomica site US-

based studies by Rowe et al (1995) and Thayer et al (1994) which give values of 20534 ECU (approx £12,800)

per tonne and 46825 ECU (approx £29,000) per tonne), respectively.

Note that some estimates are for all particulates or Total Suspended Particulate matter (TSP), whilst some are

for PM10 specifically. We have used low and high values of £6,000 per tonne and £200,000 per tonne

respectively. We accept that not all these studies are strictly comparable – they tackle pollutants arising in

different contexts. The variation is therefore very significant. Clearly, one could seek to adjust values for rural

and urban areas, especially in respect of transport if the ECMT (1998) study is to be believed.

Sulphur DioxideSee Table 2 for studies reviewed. Note not all studies include all effects. We have used a low value of £2000

per tonne and high value of £10,000.

Oxides of Nitrogen (NOx)See Table 3 for studies reviewed. The values used in this study are £1,000 at the low end and £22,000 at the

high end.

Table 1. Estimates of Damages from Recent European Studies

Study Study Area Pollutant Damage

Low Central High

Krewitt et al (1997) (ECU/tonne) UK/Germany1 Particulates 22,046 60,439

CSERGE (1993) (ECU/tonne) UK Particulates 12,240

AEA (1997) (ECU/tonne) Birmingham, UK (50m

incinerator stack)

PM10


incinerator stack)

PM10


incinerator stack)

PM10

Pearce and Crowards (1995)

(£/tonne)

UK PM10 23,288 57,748

Beukering et al (1998) EU PM10 20,468

ECMT (1998) (ECU/tonne) UK (rural transport) PM10 0

ECMT (1998) (ECU/tonne) UK (urban transport) PM10 70,000

CIEMAT 1998 (ECU/tonne) UK PM10 8,000 22,917

Powell et al (1996) (£/tonne) UK PM10 8,980

Coopers and Lybrand et al (1997)

(ECU/tonne)

UK TSP (transport) 7,522


(ECU/tonne)

UK TSP (electricity

generation)

12,149





Low Central High


incinerator stack)

SO2 20,131a


incinerator stack)

SO2 18,715 a


incinerator stack)

SO2 18,243 a

CIEMAT 1998 (ECU/tonne) UK SO2 6,027 10,025

Powell et al (1996) (£/tonne) UK SO2 2,584


(ECU/tonne)

UK SO2 4,339 b

Davidson and Wit (1998) (£/tonne) SO2 2,000 4,000

a Includes acute health, chronic health and materials impacts.

b Includes impacts on health, buildings, crops and forests.



Low Central High

Krewitt et al (1997) (ECU/tonne) UK/Germany NOx 17,864 47,003

CSERGE (1993) (ECU/tonne) UK NOx 1,005 a


incinerator stack)

NOx 34,739 a


incinerator stack)

NOx 34,267 a


incinerator stack)

NOx 34,149 a

ECMT (1998) (ECU/tonne) UK (rural transport) NOx 4,000

ECMT (1998) (ECU/tonne) UK (urban transport) NOx 8,000

CIEMAT 1998 (ECU/tonne) UK NOx 5,736 9,612

Powell et al (1996) (£/tonne) UK NOx 1,270


(ECU/tonne)

UK NOx 3,076 b

a Includes acute health, chronic health and materials impacts.

b Includes impacts on health, buildings, crops and forests.

Tropospheric Ozone and Volatile Organic CarbonKey ozone precursors are NOx and organic carbon (TOC) (see below). It is not always completely obvious

whether ozone damages are included in assessments of the damages due to these products. The AEA (1997)

report gives a value of 2,530 ECU/tonne of ozone. The CIEMAT (1998) report, acknowledging the complexity

of the reactions involved, gives a value for the EU of 1,500 ECU / tonne NOx. Estimates for damage costs from

volatile organic carbons do not always obviously include estimates for creation of tropospheric ozone. For

Volatile Organic Carbon compounds, ECMT (1998) use a figure of 4,000 ECU/tonne in rural areas and 8,000

ECU per tonne in urban areas.



Because of the complexity of the chemistry involved, damage estimates are difficult to arrive at. Since, in the

contexts we are dealing with, we are looking to understand the ozone-related effects of emissions from NOx and

TOC, we have estimated low and high values of £500 and £2,000 per tonne of NOx and the same for VOCs.

For VOCs, we use high and low values of £500 and £4,000 (since these are suspected of having carcinogenic

effects beyond their impacts on crops and health via ozone formation). Evidently, these numbers are somewhat

arbitrary.

Greenhouse Gases (Carbon Dioxide, Methane and Nitrous Oxide)Evidently, placing values on greenhouse gas emissions presents particular problems. Theoretically, one needs to

know how climate will change because of anthropogenic emission of gases (relative to the counterfactual). The

uncertainty surrounding climatic projections and the dynamic path by which climate changes, specifically, the

frequency and severity of extreme events, makes easy quantification a rather distant prospect.

Carbon DioxideMarginal Social Costs for CO2 emissions from a number of studies are given in Fankhauser and Tol (1995).

Note that these vary over time so that typically, the shadow price of a tonne of CO2 rises over time. Where

ranges were given, they were given for 90% confidence intervals. Examples of these are:

• from Nordhaus (1991) $0.3 to $65.9;

• from Cline (1992) $5.8 to $124; and

• from Fankhauser (1994) $6.3 to $45.2.

All these are valued in $1,990 and are per tonne of carbon (so for values for CO2, one has to multiply by the

relative molecular weights, that is (12/44). Other studies include ECMT (1998) which, in the spirit of

precautionary approach, used 50,000 ECU / tonne CO2. Davidson and Wit (1997) (cited in ECOTEC 1999)

estimate damage costs at £30 / tonne CO2.

We have taken values ranging from £3 to £90 per tonne of Carbon. Ecobalance and Dames and Moore (1999)

used £3 to £109 in their recent DTI report.

MethaneThe two extreme values that we have made use of effectively come from Fankhauser (1995) and Davidson and

Wit (1997). Fankhauser’s range for a 90% confidence interval is £36.6 to £136.4 /tonne CH4 . This was the

range used in work done for us by CSERGE in ECOTEC (1999). The same study mentioned the work by

Davidson and Wit (1997). We have used the £700 per tonne value as an upper bound estimate. Hence, our range

is, at the low end, £36.6, and at the high end, £700.

Nitrous OxideThese come from the earlier work carried out for us by CSERGE (in ECOTEC 1999). The values are taken from

Fankhauser (1995) and cover a 90% confidence range as estimated there. The low value is £614.30, the high

value, £5,534.78 per tonne of N in N2O.

Carbon MonoxideThe damage costs for carbon monoxide come from Fankhauser (1994). The central estimate as given in Powell

et al (1996) is 0.6p/kg. We have used values of £2 to £10 per tonne. This is highly arbitrary. However, the

influence of carbon monoxide under these assumptions is minimal in our analysis.

Heavy Metals and DioxinsA fairly comprehensive treatment of benefits assessment associated with heavy metals from incineration plants

under different assumptions is given in AEA (1997). The reader is directed there for details of the derivations

and the discussions surrounding specific pollutants. The greatest variation is witnessed in the case of dioxins.

Here, the assumption concerning the absence or otherwise of thresholds has a massive influence on the results.

The values we have used reflect the variation with the assumptions employed by AEA (all values are £/tonne):

• For dioxins, a low of 0, a high of 6.8 billion;

• For cadmium, a low of 50,000 and a high of 700,000

• For arsenic, a low of 6,000 and a high of 4,000,000;

• For mercury, AEA use a value of 0;

• For chromium, a low of 600,000 and a high of 4,200,000;

• For nickel, a low of 11,000 and a high of 450,000.

These reflect not just variation in assumptions about effects, but also stack height (note these were derived for

an incinerator in Birmingham).

For lead, we have used a range from EFTEC (1996) of £3,000 to £9,000 per tonne.

CFCs, Water Pollutants incl. Leachate



We do not have data on emissions from any of the treatment routes for CFCs, and we do not feel that the

valuation work available allows for an easy quantification of impacts from water pollution. These are omitted

from the valuation work undertaken.

For Annex 3: Assumptions Concerning GHG Emissions From Components of Municipal Solid Waste (USEPA1998) see tables.pdf.



ANNEX 4: EMISSIONS FROM AVOIDED ENERGY SOURCES (ETSU 1997)

Average fuel mix

Emissions (g/GJ)

CO2 (kg) CO CH4 NMHC NOx N20 SO2 PMTotal 161.8 45.3 479.2 8.2 443 10.1 1356 32.5

Coal-fired

Emissions (g/GJ)

CO2 (kg) CO CH4 NMHC NOx N20 SO2 PMTotal 269.8 40.4 1006.1 7.2 783.6 14.3 2539.1 56.3



ANNEX 5: COMPOSITION OF MUNICIPAL SOLID WASTE USED IN THE STUDY

These are ‘guesstimates’, but are based on a number of studies which have carried out waste analysis at the

Local Authority level. Some of these studies are based on a ‘hands on’ approach, others are based on ACORN

data. There is a clearly subjective element in the final choice, but guidance from other studies undertaken in

Bristol, South Gloucestershire, Birmingham, London, and various authorities in Scotland and Wales has been

sought.

Material Percentage Composition

Newspaper 13.0%

Office paper 6.0%

Corrugated Boxes 3.0%

Coated Paper 5.0%

Al Cans 2.0%

Steel Cans 5.0%

Glass 8.0%

HDPE 4.0%

LDPE 4.0%

PET 2.0%

Food Scraps 20.0%

Grass 4.0%

Leaves 2.0%

Branches 2.0%

Yard Trimmings 2.0%

Screenings 8.0%

Textiles 3.0%

Miscellaneous Combustibles 7.0%

Mixed MSW 100%



ANNEX 6: RANGE OF VALUES FOR INCINERATOR EMISSIONS

Incinerator emissions in our modelling exercise are based on a small number of studies. This is a difficult

choice to make since incinerator emissions will vary over time. Even when one resorts to looking at targets in

the proposed Waste Incineration Directive, one finds targets specified over different averaging periods. Ideally,

one’s choice of emission values reflects knowledge of the nature of the effect. As we have discussed in the main

report, these themselves will be the subjects of some uncertainty so that we may not be able to be so sanguine

as to assume that short events of higher exposure are unimportant in causing specific effects. We have compared

a small number of quoted sources with the limits set in the Directive. At the low end, if quoted sources give

lower values, we have used them. At the high end, if quoted sources exceed the highest Directive target (which

will relate to the shortest averaging period) we have used. If all quoted sources give lower values, even at the

maximum range, we have used the highest quoted value.

For comparison, we show rounded figures for the ‘New Incinerator’ as specified in the Environment Agency

WISARD demonstration disk (and some appropriate caveats are specified therein). The High and Low values are

taken from the MAX and MIN values specified in the demonstration disk (which it suggests should be used for

sensitivity analysis only). There is considerable divergence for CO, VOCs, SO2 and cadmium. Given the actual

data used by Entec (1999b) in respect of CO, we believe our figures are defensible. Even the Maximum figures

for retrofitted incinerators in the WISARD demonstration are far in excess of the Entec (1999b) figure for

Tyseley. Almost identical comments can be made in respect of the figures that the demonstration disk for

WISARD quotes for SOx. From the point of view of the valuation work, it is the divergence in the SO2 figure

that is most critical. Figures for ‘dust’ in the WISARD demonstration lie within the same range as that we have

used for particulates.

For Annex 6 Table see tables.pdf.



ANNEX 7: CALORIFIC VALUES OF COMPONENTS OF MUNICIPAL SOLID WASTE

Atkinson et al(MJ/kg)

USEPA (MJ/kg)

Newspaper 13.24 16.78

Office paper 13.24 14.35

Corrugated Boxes 13.24 14.88

Coated Paper 13.24 11.08

Al Cans 0.04 -0.74

Steel Cans 0.05 -0.42

Glass 1.08 -0.53

HDPE 27.75 39.46

LDPE 27.75 39.46

PET 27.75 20.47

Food Scraps 6.12 5.91

Grass 6.12 5.91

Leaves 6.12 5.91

Branches 6.12 5.91

Yard Trimmings 6.12 4.96

Screenings 5.51 2.30

Textiles 16.29 3.50

Miscellaneous Combustibles 15.63 12.00

NB. Where estimates from one study are not available, we have used the calorific value of the closest matchingmaterial (and this is likely to incur errors)

Sources Atkinson et al 1996; USEPA 1998.



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