Policy Coverage of Environmental Impacts of Materials

A research report completed for the Department for Environment, Food and

Rural Affairs by AEA Technology. July 2006

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AEAT In Confidence

Title Policy Coverage of Environmental Impacts of Materials

Customer The Department for Environment, Food and Rural Affairs (Defra)

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Suggested citation Bates, J., and Watkiss, P. (2006). Policy Coverage of Environmental Impacts of Materials: A report to the Department for Environment, Food and Rural Affairs. AEA Technology. Defra, London.

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25/07/06

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Approved by Dan Forster 25/07/06

Executive SummaryThis report presents the findings from the SCP study on the Policy Coverage of Environmental Impacts of Materials. The project aims were to: Investigate the environmental burdens associated with 3 materials at each stage

of their life cycle, including their use in products and their final disposal; Assess the potential externalities associated with these burdens; Review the policies in place to address these burdens (the policy coverage),

considering both economic instruments and other forms of legislation.

The first phase of the study identified a large number of materials for consideration. These included primary materials (e.g. metals, aggregates, wood) and secondary materials and intermediate products (e.g. plastics, paper). After review, and consultation with the study steering group, the study selected the following materials: Plastics, specifically PET (PolyEthylene Terephthalate). This material was

chosen to capture the growing use of plastic materials, and to investigate the emerging plastic recycling industry.

Iron and steel. This material was chosen as it is an energy intensive material, used in construction and products, and an existing recycling market exists.

Wood. This material was chosen because it is a renewable material and is a major construction material.

Existing life cycle data was found for the extraction and processing of all three materials (although not always for the UK situation), though less useful data were found for each material product. However, the study found that existing LCA studies are not useful for subsequent policy externality analysis. This is because: These life-cycles are set up to assess a particular problem or question, and do

not easily lend themselves to an examination of burdens for resource flows. Data are often aggregated across the life-cycle. Moreover, life cycle impact

assessment methodologies are often used to aggregate individual burdens into a few categories. This makes it difficult to assess individual stages or burdens in detail, which is needed to examine externalities and policy coverage.

LCA studies are aimed at answering a different question to externality analysis. LCAs are primarily undertaken as a comparative (relative) way of looking at different options, using burdens which are usually calculated on average effects. In contrast, detailed externality assessment focuses on (absolute) impacts (not burdens), calculated based on marginal (not average) effects.

Most LCA studies do not contain the detail on the location of burdens. This information is important for externality studies, both in relation to UK versus non UK effects, and because externalities vary significantly between locations.

This finding is one of the main policy conclusions of the study. It suggests that any subsequent analysis of resource flows would need dis-aggregated life cycle inventory data i.e. emissions of individual pollutants, split by individual life cycle stage.

The study initially planned to use existing LCAs to identify burdens for the three material flows. Due to the reasons above, this was not possible. Instead, the study used an LCA tool (SimaPro) to model simplified life-cycles for each material. This included raw material extraction and processing, one main product use for each material (wood in construction, steel in a car and PET in a drinks bottle) and end of life management of the product. This allowed the output of disaggregated life cycle inventory (LCI) data, split by life cycle stage. Two recognised impact assessment methodologies were then applied to the inventory to identify the most important environmental impacts (as a screening tool).

Using this inventory burden data, the study scoped the externalities for the three material life cycles. Two potential approaches were considered. Firstly, to use life cycle impact assessment (LCIA) output and try and monetise this directly. Secondly, to use the LCI (inventory) output, and apply existing externality estimates using damage costs (simplified £ per tonne values). Regarding the two approaches, the study has found: It is difficult to apply valuation estimates directly to LCIA output, as most units do

not relate directly to the economic valuation estimates in the literature. There are also problems of consistency, because LCIA approaches do not follow standard UK guidance for impact assessment, e.g. for health impacts of air pollutants. However, the use of LCIA data has a major advantage, as it has complete coverage across all life cycle steps and burdens.

The use of the LCI data and an impact pathway approach (or simplified impact approach using damage costs) is much more satisfactory in terms of policy consistency, and is methodologically more robust. It also allows a direct match between impacts and valuation endpoints. The main disadvantage is that there are not agreed approaches for analysis and valuation of all burdens from a LCI (though this problem also applies to the use of LCA output). In addition, applying this approach to all the burdens from an LCI is time consuming.

The two points reflect the trade-off between trying to value everything (with inconsistent methods and high uncertainty) versus only valuing a few areas robustly, but leaving potential gaps.

This study adopted the second approach, using LCI data on burdens. This provided the most robust and evidence based approach for subsequent externality analysis. Our approach has been: To produce dis-aggregated LCI data output; To use the LCIA as a screening tool to identify important steps and burdens; To select the priority burdens and assess these in detail using an impact pathway

approach to estimate externalities (in this case a simplified approach using damage costs), using agreed impact and valuation approaches for UK policy.

To screen the LCI data to see if there are any potential major burdens that have been overlooked. We also highlight that it would be possible to use the priority burden estimates, in combination with the LC impact assessment data, to calculate the relative values for all burdens. This approach should be investigated in any subsequent studies.

The analysis revealed significant externalities for each of the three material flows. The externalities are dominated by greenhouse gas emissions and air quality pollutants arising from energy and electricity use, though this reflects the choice of materials (high embodied energy or energy intensive materials). For iron and steel, some trace pollutants (heavy metals) may be potentially significant.

During this analysis, a number of specific methodological issues were identified for externality assessment using LCI data. These were: LCI data does not (generally) specify the location of burdens. It does not specify when burdens occur (there is no time dimension). Note this is

important for economic analysis and discounting. A number of potential externalities are not included in LCA frameworks (as LCA is

focussed on quantification of resource and emission flows). These include impacts on ecosystems and biodiversity, social impacts, accidents, and amenity.

A large proportion of burdens are associated with transport and electricity, and the burdens from these sectors change over time (e.g. LCI may be using current information to address a future policy question). The assumptions on the benefits of energy recovery from waste, or recycling are also dependant on energy mixes assumed, and the same issue applies.

The study assessed the policy coverage of the three life-cycles against the burdens identified. It was found that economic instruments in the UK do focus on a number of the key impact areas for the materials – including energy/electricity use and waste. The study also assessed whether these polices internalised the impacts.

For landfill as a waste option, the existing economic instruments seem to cover the potential externalities (i.e. the degree of internalisation suggests prices reflect the appropriate signals). For energy/electricity, and for incineration as a waste option, the analysis suggested that important externalities remain (i.e. prices do not fully reflect environmental burdens), particularly for CO2 emissions, and the conventional air pollutants (SO2, NOX and PM10). Following from this, and the analysis of wider policies, we suggest there might be a need for: Full policy coverage to address the total externalities of CO2 emissions, and the

conventional air pollutants (SO2, NOX and PM10). Market signals to recognise the benefits of recycling (i.e. to reflect avoided

externalities). The analysis suggests that recycling avoids significant externalities across the life cycle (i.e. from avoided extraction, processing and waste disposal).

Policies to tackle burdens associated with resource extraction, e.g. from mining of raw materials, or the potential for marine eco-toxicity from oil/gas extraction.

Policies to ensure optimal natural resource depletion (implying reliance on market forces is insufficient).

In addition, the results of this study suggest that not all the externalities of incinerating plastic in an energy from waste plant would be directly addressed by current policies. However, plastic would not normally be burned on its own, but as part of a mixed household/commercial waste stream.

It would be useful to examine whether for the incineration of mixed wastes in energy from waste plant, there are externalities which are not addressed by polices (such as the Waste Incineration Directive and local planning regulation), which impact upon waste incineration facilities at the moment.

The study has reviewed the transferability of this type of approach, and concluded: The use of a detailed ‘step by step’ life cycle model allows the level of detail

needed for a proper resource flow analysis. However, even with relatively simple lifecycles, the volume of data generated by such an approach will be large. For this reason, there is a need to identify and focus on priority burdens (as here).

The overall approach developed here is, we believe, transferable to other studies. For example, the externalities associated with material production (e.g. PET, iron and steel, wood) would be transferable to other products provided a proper in-depth analysis was undertaken (rather than the scoping analysis here).

This approach might require some additional work for each product. This would be needed to set up the life cycle stages from material production through to disposal for the specific product. It would also require work to set up for the actual disposal route (e.g. for the material to landfill, incineration, or recycling).

There is one final issue on transferability. This is over the year of the study and the assumed energy / electricity mix in place, and whether this should reflect current or future baselines. Different policy questions or analysis might want to work with different policy baselines, which would require the adjustment of the life cycle set up. Similarly, it is clear that the assumptions about waste disposal, e.g. energy recovery assumed, type of energy used in recycling, etc. have a big impact on determining the externalities, and again these might change for different studies. This is one area where further consideration is needed for transferability, and further work is needed to improve the analysis.

In relation to research priorities, the study recommendations are: Further testing with other LC impact assessment methods would be useful,

particularly some of the more advanced tools that include a more impact driven approach.

The externalities and life cycle impact assessment are generally closely correlated in identifying the most important impacts, but do differ in some areas (notably over the importance of trace pollutants). It would be useful to reconcile these differences in future material flow analysis.

More work is needed to attribute damage cost values (impact pathway approach) to all LCI burdens, or to develop approaches to screen all burdens more robustly. Similarly, work to progress the valuation of LCA output directly would be useful. We stress that these gaps are not a barrier to the successful implementation of a follow on phase of this study.

Sensitivity analysis would be useful to examine the potential effect on the results of uncertainty (e.g. on impacts and values) and over the choice of assumptions (e.g. on which energy source is displaced by waste to energy schemes, on future rather than current technologies/policies, etc).

Contents

1. INTRODUCTION 3

2. IDENTIFICATION OF MATERIALS 3

3. LCA/ENVIRONMENTAL IMPACT ANALYSIS 3

3.1. Introduction 3

3.2. Plastic 3

3.3. Iron and Steel 3

3.4. Wood 3

4. EXTERNALITIES 3

4.1. Methodological approach for assessment of externalities 3

4.2. Calculation of Externalities from Life Cycle data 3

4.3. Scoping of Externalities for the Three Material Life Cycles 3

4.4. Specific issues on using the LCI data 3

4.5. PET 3

4.6. Iron and Steel 3

4.7. Wood 3

4.8. Findings 3

5. POLICY COVERAGE 3

5.1. Introduction 3

5.2. Input to policy development 3

5.3. Policy Development Outside the UK 3

5.4. Other Policies 3

5.5. Internalisation 3

5.6. Findings 3

6. RECOMMENDATIONS AND RESEARCH GAPS 3

6.1. Key Findings 3

6.2. Transferability and Further Work 3

7. REFERENCES 3

8. Glossary 3

APPENDICES

APPENDIX 1 Screening of Materials

APPENDIX 2 Life Cycle Studies for Candidate Materials

1. Introduction

Historically, environmental legislation has focused on production (e.g. electricity generation, industrial manufacturing), and been driven by action towards individual environmental burdens (e.g. air pollution emissions). It has also focused on the use of command and control legislation to address these environmental problems.

However, there is a growing recognition that policy might need to adopt a more holistic approach. This includes the consideration of resource flows in the economy – and the need to consider activity across the whole life cycle, from extraction to final disposal. It also reflects the need to address multiple-pollutants and multiple-effects across these flows, i.e. to consider overall environmental issues rather than focus on narrow individual burdens.

At the same time, there has been recognition that activities, such as electricity generation, transport, material extraction, etc. have environmental and social burdens. These include air pollution, greenhouse gas emissions, waste, water emissions, and noise. These burdens lead to environmental impacts that are typically not paid for by providers or users, known as external costs or externalities. Alongside this, there has been a regulatory shift towards the use of economic instruments (e.g. green taxes, trading schemes) to tackle externalities.

While these changes are starting to be translated into policy, there remain some important perspectives that need to be addressed, to ensure optimal policy. The current practice of using specific instruments for selected processes or individual burdens may miss some important environmental externalities up- or down-stream. It may also lead to overlapping policy instruments that actually duplicate the correct pricing regime. Finally, it may lead to perverse effects, where action to address one isolated problem actually increases overall environmental burdens (and externalities) elsewhere in the system.

This scoping study (Policy Coverage of Environmental Impacts of Materials) aimed to investigate these issues as part of Defra’s Sustainable Consumption and Production Programme (SCP) – Development of an Evidence Base. It has investigated the environmental burdens associated with 3 materials1 at each stage of their life cycle and looked at the policy coverage (economic instruments and other forms) in place to address these burdens.

This report summarises the work undertaken in the study. The first task in the study was to identify three materials for analysis. The selection process is described in Section 2 (and appendix 1). A large number of potential materials were initially screened to produce a short list of five materials, which were considered in more detail. From these, in discussion with the steering group, three (steel, wood and plastic) were chosen for analysis.

1 The study focuses on materials, rather than on products: this is an important and innovative distinction from previous work, and ensures the project will build the evidence based for the SCP programme.

Section 3 of the report presents information on the life cycles of these materials and products in which they are incorporated. It also includes information on the environmental burdens impacts arising over the lifecycles of these materials.

Section 4 of the report discusses the monetary valuation of these impacts, the externalities. It summarises the externalities, state of quantification methods, and methodological issues with quantification and valuation. It scopes out the potential externalities for each of the three materials, using the life cycle data presented in Section 3, using a number of tools developed for policy applications for Defra and the European Commission.

The policy coverage of the environmental impacts from the three materials’ lifecycles is reviewed in Section 5. The lifecycle stages at which policy is addressed are assessed and gaps in coverage are identified.

Finally, Section 6 of the report summarises the findings of the study. It also makes recommendations on the transferability and consistency of the approach, and includes an analysis of critical issues and uncertainties or gaps in the evidence. It provides recommendations for further research to improve the evidence base.

2. Identification of Materials

A large number of potential ‘materials’ were identified for initial screening in the study:

Aluminium; Iron and Steel; Heavy metals (e.g. cadmium, mercury); Aggregates; Quarry materials (e.g. limestone, slate); Cement; Bricks; Glass; Wood/timber; Paper; Natural textiles (e.g. cotton and wool); Man-made textiles; Rubber; Plastics; Fertilisers; ‘Toxic chemicals’ (e.g. pesticides).

Materials excluded at the kick-off meeting were: Primary fuels, Transport fuels, Electricity and Agricultural products/food.

The environmental impacts of primary fuels, transport fuels and electricity over their lifecycles have been studied in considerable detail previously, and the externalities assessed (e.g. ExternE, 1998 and 2005: Friedrich and Bickel, 2001; Sansom et al, 2001; Watkiss et al, 2005). It was therefore felt that their inclusion would provide little added value to this study. Agricultural products and food were excluded as they are the focus of other parts of the SCP research programme

The materials were screened by assessing them against the selection criteria identified in the Terms of Reference for the study: Volume; Hazardous nature; % of total volume of material disposed of by landfill; Evidence of environmental concern (key environmental impacts as identified by

other studies); Sufficient data, especially LCA data; Transferability of learning.

Assessment of these criteria was important to enable the study: To obtain a good coverage against major areas of environmental concern, and try

to select materials that have major environmental burdens in each area, e.g. in

each of climate change, air quality, water quality, resource depletion, biodiversity, volume of waste produced, hazardousness;

To review the quality of LCA information. This is an extremely important factor, because it will dictate the potential for later analytical steps in the project.

Other aspects which were evaluated were: Trends in use (growth/decline): It would be less useful to study a material

whose use was in decline; Level of recycling; % of material imported; Number of product uses; this is an important factor in determining the

complexity of analysis needed.

The results of the initial screening (see Appendix 1) were combined with the main conclusions of other studies that have prioritised products/sectors/materials by examining their environmental impacts on a life cycle basis. This resulted in the initial long list of potential materials for further study shown in Table 2.1.

Table 2.1 Initial ‘long list’ of materials for further study

Material characteristics Suggested Candidates

Pros/Cons

Renewable WoodPaper

Wood: broader category than paper (would include paper), Flows could be complex, but allows examination of labelling schemes

Local amenity impactsupstream policy coveragerecyclable

Aggregates, Other quarried products

Aggregates allow an examination of the aggregates levy, specifically designed to address externalities.

Energy intensive (so broad range of impacts) recyclable

Iron and steel,AluminiumGlass

Iron and steel: good data sources availableGlass: mostly UK production

Growth in usage Plastic Plastic: interesting from a recycling perspective. Complex sector; necessary to pick specific a plastic with limited product uses.

Construction/building products High volume products

Cement, brick Cement: several recycling routes

Impacts during useImpacts at disposal phase Impacts on water bodies

Heavy metals – probably Cd,Pesticides,Fertilisers

Heavy metals and pesticides are toxic/hazardous.Pesticides/fertilisers are interesting from a policy coverage point of view (e.g. taxes in other countries). Best data set for fertilisers

Following further discussion and guidance from the steering group, it was agreed that construction/building products would be covered if iron and steel were chosen for

further analysis, particularly as other studies have generally identified steel as a construction material with higher environmental impacts than cement or bricks. Heavy metals were excluded as their use is likely to decline with introduction of RoHS legislation. One material was then chosen from each category. The materials, and reasons for the choices made are: Wood/paper (chosen as a renewable material) Aggregates (chosen as it is the target of a specific levy) Iron and steel (chosen as an initial literature survey suggested some suitable data

sources) Fertilisers (chosen on basis of likely availability of data, examples of policy

coverage) Plastic (PET identified as meeting criteria of limited number of uses);

The suitability of the five candidate materials for detailed study was examined by looking at: The flows of the materials through the economy in more detail, to establish the

diversity of ‘routes’ through the economy and the potential to construct representative lifecycles, which capture the majority of the material use.

The potential availability of life cycle data and studies, both for the raw materials, intermediate products and final products. This was assessed by looking at the life cycle literature, and also at publicly available life cycle inventory data repeated words (for materials not products).

These aspects are discussed in full in Appendix 1. It was concluded that there were good reasons to choose any of the five materials: The terms of reference specified that a renewable material should be included.

This will bring in ecological issues (e.g. impacts on biodiversity) which may not be captured if only renewable resources are chosen. This means that either wood or paper must be included.

Aggregates have well established policy coverage (aggregates tax for extraction and the landfill tax for disposal). As such they are an obvious choice to demonstrate how environmental taxes can be applied to internalise externalities.

PET has lower environmental burdens than iron and steel, or fertilisers. It is therefore less interesting from the perspective of the externalities, but its use is growing and it could be interesting to look at waste management.

The energy intensive nature of iron and steel production means that it has relatively high greenhouse gas and air quality burdens. There is established recycling of this material. While good LCA data exists for iron and steel production, the data on product use (for the five main product areas) is not so robust. The larger number of product uses means a more complex assessment.

Fertilisers have a single product use, and are the only material that has significant burdens during the use phase, and the only one with significant impacts to water. They are therefore interesting in extending the scope of the study to look at policy coverage in other areas.

The choice of materials was discussed at the interim project meeting with the steering group and it was decided to proceed with:

Plastics (PET) Iron and Steel Wood

Aggregates were not chosen for further study, as it was felt that the externalities associated with amenity impacts during extraction had been studied in some detail at the time the levy was introduced (and again more recently). It was unlikely therefore that this study would produce much new material of interest.

In the case of fertilisers it was felt that their environmental impacts were well known, and that the externalities associated with their use had been investigated previously. The other three materials touched on more policy areas (e.g. waste management, climate change, and pollution control) and were therefore of more interest to the steering group in relation to policy.

3. LCA/Environmental Impact Analysis

3.1. Introduction

This section provides information on the flows of the three materials: timber, plastic (PET), and iron and steel in the UK, and the main product uses for each of these materials. It also provides data on environmental impacts from each stage of the life cycle for one of the main product uses i.e. from raw material extraction, through material use, production and processing, incorporation into a product and end of life management of the product (Figure 3.1).

Figure 3.1 Lifecycle stages

The intention at the outset of the study was to use results from existing LCA studies of products to compile information on the emissions to air, water, and soil, at each stage of the lifecycles for key product uses. For the purposes of this study, it is important the information is:

In the form of emissions of individual pollutants/substances, rather than aggregated (using weightings) into impact categories. The latter is typically ‘end result’ of life cycle studies, as they are used to help reduce the large number of individual substance flows in the life cycle inventory to a number of environmental impacts, which can be interpreted.

By lifecycle stage. This is important because of the need to:(a) consider the location of the lifecycle stage (inside or outside the UK);(b) consider the source of the emission as in the estimation of externalites, this

can be an important factor in valuing it2 ;2 For example, emissions of NOx from high stacks in the electricity industry have different impacts from ground level emissions from transport

(c) distinguish impacts from end of life stages: depending on the lifetime of the product these can occur many years in the future, and this factor should be taken into account when valuing the externalities of these impacts.

For each of the materials, the published results of existing LCA studies reviewed did not provide information in the above format. Reasons for this varied; often the studies were product focused, and were comparing different products, or the influence of different end of life scenarios for products (e.g. recycling versus disposal). They therefore concentrated on results for the life cycle as a whole, rather than considering results for each lifecycle stage (and if a life cycle inventory was presented then it was for the whole lifecycle).

Furthermore, in order to help interpret the results of the lifecycle inventory (which typically gives the quantities of at least 200 substances emitted to air, water and soil, or consumed as raw materials), impact assessment methodology are often used to understand and evaluate the magnitude of the potential environmental impacts. Impact assessment methodologies typically aggregate, using some form of weighting factor, the pollutants which contribute to a particular environmental impact3, enabling them to give ‘scores’ for various environmental impacts. Many of the studies only presented the results of the impact assessment, rather than the inventory. All the studies identified for the products chosen were from outside the UK.

For this scoping study therefore, simplified life cycles for each of the materials for one main product use were created using the LCA tool, SimaPro. The system boundary for each lifecycle was taken as raw material extraction to final disposal of waste products (‘cradle to grave’). The infrastructure necessary for processing of materials (e.g. the blast furnace in the iron and steel lifecycle) was included. Information from existing lifecycle studies on material and product production and disposal was used to help define the lifecycles, and to identify processes which could be excluded from the modelling as they do not contribute significantly to environmental impacts. As this is a scoping study (and compilation of LCAs for the materials was not originally foreseen), use was made of existing inventory data sets4 for extraction and processing of raw materials and in some cases for production. These data sets (e.g. for the manufacture of PET) are often based on a number of European production plant, and use of an average European electricity mix is assumed in the data set. It is known that where there is significant energy use in the production of materials/products, then burdens form that energy use are often one of the most significant sources of environmental impacts. As the impacts associated with different forms of electricity generation vary significantly, it is important to ensure that the appropriate electricity mix is used. The data sets were therefore modified to use a UK rather than European electricity mix where appropriate. Any transport steps were also changed to be representative of distances in the UK. No modification of other parameters were made to reflect UK rather than European conditions, and this is acknowledged as a limitation of the results. If further study

3 (For example a total score for climate change, can be derived by multiplying the emissions of greenhouse gases by their global warming potential). 4 These included publically available data sets and those that are included in the lifecycle software tool SimaPro.

were to be undertaken then a proper life cycle could be undertaken, and UK specific data gathered (see Section 6).

Results have been calculated for 1 tonne of the material embodied in the product. Where alternative end of life options were available (e.g. landfill, incineration and recycling) these have been analysed separately, in order to allow the influence of disposal route on overall environmental impacts to be assessed.

Two life cycle impact assessment (IA) methodologies5 (CML2 baseline 2000 v2.1 and Ecoindicator 99) were used to analyse the life cycle inventory data and establish which were the most significant environmental impacts. The problem orientated CML IA methodology assesses the contribution of emissions to environmental problems by weighting them to allow their aggregation. So for example, the contribution of emissions to global warming is assessed by multiplying the mass of emissions by their global warming potential. In order to provide an indication of the significance of the different impacts assessed, the results for each impact are normalised against a score for that impact. For this study, impacts were normalised against scores for Western Europe in 1995. Eco-indicator 99 is a damage orientated approach. In a characterisation stage, a series of damage models are used to convert emissions of pollutants, land use and land-use conversion and extraction of minerals and fossil fuels into three indicators:

Damage to human health – expressed as the number of year life lost and the number of years lived disabled. These are combined as Disability Adjusted Life Years (DALYs) an index also used by the Worldbank and WHO. Damage to human health is assessed as damage from carcinogens, respiratory organics, respiratory inorganics, climate change, radiation and ozone layer depletion.

Damage to ecosystem quality, expressed as the loss of species over a certain area, during a certain time. This is assessed through impacts of acidification/eutrophication, and land-use/conversion.

Damage to resources – expressed as the surplus energy needed for future extractions of minerals and fossil fuels.

Each of these categories are normalised against a score for Europe (for 1993).

5 A number of impact assessment methodologies are available. The CML, problem orientated approach is fairly widely accepted, and is used here as it covers a wide range of environmental issues.

3.2. PLASTIC

3.2.1. Material flows through the UK Economy

The material flows of plastic through the UK economy were mapped as part of a report on plastics in the UK economy (Wastewatch, 2003). The report clearly shows the complexity of the plastics sector –in terms of the number of types of plastics andend uses, and in relation to imports and exports of plastics components and plastics in products. This means that it is necessary to consider a single plastic type separately in order to be able to map the resource flows. The example chosen was PET (Polyethylene Terephthalate) as it is manufactured in significant quantities in the UK, has a single predominant use (drinks bottles), and is a plastic that is recycled (as bottles are an easily segregated waste), with relatively low levels of contaminants.

Figure 3.2 Material Flow of PET through the UK economy

3.2.2. Material/product life cycle

The simplified lifecycle for PET used in drinks bottles is shown in Figure 3.3. The key assumptions and data sources for the lifecycle are given in Table 3.2. Production of PET, and recycling of PET (into granules which were assumed to offset the use of amorphous PET granules for applications such as the production of fleeces) were based on a dataset included in the SimaPro tool for average European production. The main source of this data in the tool is an eco-profile for PET produced by PlasticsEurope, the association of plastic manufacturers in Europe.

Crude oil and Natural Gas

Ethylene glycol (EG) and TerephthalicAcid (PTA)****

235kt of PET plastic products manufactured in the UK in 2000**

Total supply of PET plastic products in 2000: 261 kt**

Raw Materials

Intermediate products

Production Consumption

*British Geological Survey, 2005.

** Plastics in the UK Economy

*** Digest of UK Energy Statistics 2005

**** Lehmann2005.

Sources of crude oil in 2004:

Produced: 95,374 kt Imported: 62,516 kt Exported: 64,504 kt

Sources of natural gas in 2004 (in GWh):

Produced: 1,115,744 Imported: 133,035 Exported: 114,111***

Synthesised PET

A number of minerals needed to produce synthesised PET, including:

Limestone (6799 kt produced in UK in 2003 for industry)

NaCl(6000 kt produced in UK in 2003; 217 kt imported, 537 kt exported)

Bauxite (mainly imported; UK imported 271 kt in 2001) *

Net import of 26 kt**

The eco-profiles produced by PlasticsEurope are a widely accepted and widely used source of life cycle data for plastics production. The eco-profile data has been modified within the tool to include the infrastructure associated with plastics manufacture and transport stages. A dataset for mechanical recycling of PET from the SimPro tool was used to model recycling of PET. The data in the tool is based on a 2001 report by TNO-MEP in the Netherlands report (Ecoefficiency of recovery scenarios of plastic packaging'). The quality of this report and dataset is not known, and this is therefore a potential source of uncertainty in the analysis. Collection of PET waste from the consumer and any sorting of waste necessary for recycling of bottles was not modelled. This is a limitation of the analysis; inclusion of thee stages would reduce any benefits associated with recycling.

Data was available to support derivation of most of the parameters required to construct the lifecycle. Estimates of the distance PET was transported for incineration and landfill are based on experience of modelling waste management scenarios for local authorities, which suggest that this is a typical distance. The transport distances assumed for recycling are estimates only and have a higher degree of uncertainty.

Figure 3.3 Simplified life cycle for use of PET in drinks bottles

Bottle grade PET

Production of PET bottle

Retailer

Household

Crude oil & natural gas

Transport assumption

)

Waste

Loss of product

Recycle LandfillIncineration

Avoided product: Amorphous PET (production of fibres)

Recycle LandfillIncineration

Avoided product: Amorphous PET (production of fibres)

Amorphous PET

Ethylene glycol

Terephthalic acid (PTA)

Table 3.2 Key assumptions and data sources for the simplified PET LCA

Assumptions

Manufacture of bottle grade PET

Based on data from several European sites as included in SimPro. Data originally based on Eco-profile from Plastics Europe, the European Plastics Association) the Association of Plastics Manufacturers in Europe (Boustead, 2005).

PET plastic bottle blowing

SimaPro data set used which is based on data for bottle blowing at a Swiss factory; the data set was modified to use a Uk electricity generation mix rather than a European electricity generation mix. Transport distance for PET plastic granules to be blown into plastic bottle (assumed to be at location bottle is filled) was assumed to be 100km.

Transport to Retailer

The transport distance of the filled PET plastic bottle to retailer was assumed to be 112km based on freight transport statistics (Department for Transport, 2005a).

Transport to Household

The average transport distance from retailer to household is 6.8km (Department of Transport, 2005b). This is assumed to be by car. The average shopping trip involves 11kg of shopping (Smith et al, 20045. A plastic PET bottle weighs around 43g (Lehman et al, 2005), and assuming 4 plastic bottle purchases per trip, 1.56% (or 0.106km) of the transport impacts are allocated to the PET bottle.

Recycling

55% of bottles collected go for reprocessing in the UK, 45% overseas (assumed to be the Netherlands who are large reprocessors of plastic bottles) (Wilson, 2002). It was assumed the bottle travels 100km by lorry to a port and then onwards by ship for reprocessing in the Netherlands.

Landfill Transport distance of 50km was estimated for plastic bottle to landfill site.

IncinerationTransport distance of 50km was estimated for plastic bottle to incinerator with energy recovery. Estimates of electricity production from incineration of plastic on data in Smith et al, 2001.

3.2.3. Key environmental impacts

The environmental impacts over the lifecycle were examined using the two impact assessment methodologies( the CML problem orientated approach and the Eco-indicators damage orientated approach) discussed in Section 3.4.3.

Figure 3.4 shows that up to the stage of delivery to the consumer, the CML methodology assesses the main impacts as arising from production of the PET. Abiotic depletion (i.e. depletion of non-renewable resources) scores more highly than other impacts due to the use of crude oil as both a feedstock and (together with other fossil fuels) as an energy source in production and in electricity generation. Global warming and acidification are also high, again reflecting energy used in the PET

manufacture and bottle blowing process. PET production is also the main contributor to freshwater and terrestrial ecotoxicity. In the case of freshwater toxicity this is due to emissions to air of nickel (from the combustion of heavy fuel oil to provide heat for the chemical processes and for electricity production) and of vanadium to water (associated with steel used in the construction of the chemical plant, and from the disposal of terephalic acid residues). In the case of terrestrial ecotoxicity, the main emission of concern is airborne emissions of mercury mainly from xylene production (one of the precursors for PET), chromium VI to water from the disposal of wastes in PET production and airborne vanadium emissions from heavy fuel oil combustion to provide heat in the chemical production processes.

Fossil fuel use in the manufacture of PET is also the most significant impact using the Eco-Indicator99 IA methodology (Figure 3.5). The next two most significant impacts are the impacts on human health which result from respiratory inorganics and climate change. The main contributors to these impacts are emissions of nitrogen oxides (NOx), sulphur dioxide (SO2) and particulates (for respiratory organics) and CO2 (climate change), from fossil fuel combustion, firstly to provide heat in the production of PET and its constituent chemicals (e.g. xylene) and secondly to generate electricity for PET blowing.

The additional impact of the waste management of the PET bottle is shown in Figure 3.6 (CML) and Figure 3.7 (Eco-indicator 99). In this simplified lifecycle, recycling reduces all environmental impacts, usually significantly. However the beneficial impact of recycling may be overstated as the collection (and any sorting) required has not been modelled. The results indicate that landfilling and incineration would increase the scores for aquatic toxicity and human toxicity (in the CML IA methodology). This is due to emissions of vanadium to water (aquatic toxicity) and of antimony and vanadium for human toxicity. With the Eco-indicator IA methodology, impacts from carcinogens increase with landfill and incineration due to emissions of cadmium and arsenic. All of these elements are contained in trace amounts in the PET. The global warming/climate change impact for incineration increases as CO2 released during incineration of the plastic is not entirely offset by the savings from avoided electricity production.

Overall the two impact assessment methodologies suggest that the main environmental impacts in PET production and use, arise from the combustion of the fossil fuels used in production of the constituent chemicals for PET. While recycling offers a reduction in all environmental impacts, landfilling and incineration of the PET increases burdens associated with human and aquatic toxicity. Generally transport was not a significant contributor to any of the key impacts, suggesting that even if transport distances were higher than assumed they would not alter the above conclusions.

Figure 3.4 Environmental Impacts for 1 tonne of PET bottles at consumer (CML Impact Assessment Methodology)

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Figure 3.5 Environmental Impacts for 1 tonne of PET bottles at consumer (Eco-indicator 99 Impact Assessment Methodology)

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Figure 3.6 Additional Impact of Waste Management Phase (for 1 tonne of PET bottles) (CML IA methodology)

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Figure 3.7 Additional Impact of Waste Management Phase (for 1 tonne of PET bottles) (Eco-indicator 99 methodology)

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3.3. IRON AND STEEL

3.3.1. Material flows through the UK Economy

While almost all the raw materials for iron and steel production are imported into the UK, relatively little (less than 5%) of primary or secondary steel is imported. Coke used in the process is either imported directly or made in the UK from imported coal. While four sectors (engineering, automotive industry, construction and metal goods) account for almost 80% of steel use; there are a wide range of uses within these sectors: Engineering: mechanical machinery and equipment (9%), directly earthed

appliances (2%), other electrical products (3%), and other mechanical engineering (10%)

Metal goods: packaging (5%) and furniture and other goods (9%). Construction category: structural steelwork (13%) and building and civil

engineering (9%). Other industries: wire drawing (6%), forging and stamping (5%), cold forming

(3%), oil and gas (3%).

Scrap steel from semi-finished and finished products is re-used to make secondary steel. In 2003 the UK exported 7000 kt of scrap steel. (British Geological Survey, 2003).

Figure 3.8 Material Flows for Iron and Steel

Iron orePrimary Steel: 10700 kt produced in 2004**

Semi-finished steel products

12,114 kt of finished steel products were produced in the UK in 2003 *

Raw Materials Intermediate Products Final Products

*British Geological Survey, 2005.

** UK Iron and Steel Statistics Bureau

24% Engineering

14% Metal goods

17% Automotive

22% Construction

23% Other industries **

A number of minerals are needed to produce crude steel, including:

Limestone: 2018 kt used in the steel industry, primarily produced in the UK (in 2003 a total of 79000 kt of limestone was produced in the UK)

Chromium: 55.4 kt consumed in steel industry, primarily imported (107 kt imported in 2003)

Ferro-manganese: 99.3 kt consumed in steel industry, primarily imported (79 kt imported in 2003)

Nickel: 15.8 kt consumed in steel industry, (26.8 kt produced in UK in 2003; 146.4 kt imported)*

Scrap steel

5,200 kt consumed 2004**, 139 kt of

which was imported*

Crude Steel

Secondary Steel: 3500 kt produced in 2004** 1306 kt

exported 2003*

540 kt imported 2003*

3.3.2. Material/product life cycle

The material/product lifecycle chosen for analysis was the use of steel in a car (Figure 3.9). Whilst not the largest product sector for steel, the automotive sector is less diverse than e.g. the engineering or metal goods sectors, where steel is used in a wide variety of products. A fairly detailed lifecycle study (Schweimer and Levin) was also available for a car, enabling the product manufacture stage to be modelled.

The production of steel was modelled using data from the Ecoinvent database, which is based on the European situation. As discussed in Section 3.1, the environmental impacts from energy use can be significant, and it is important that they reflect country specific information. The dataset in the SimaPro tool was modified to use the generation mix for the UK (as specified in the tool) for electricity consumed in the production process rather than the average European generation mix contained in the original data set. It was not considered necessary to modify data for other forms of energy used in the production, as characteristics for coal and coke production and combustion show far less variation between countries than for electriticy generation.

Figure 3.9 Simplified Lifecycle for Use of Steel in a Car

A wide variety of steels are produced using a number of different alloys, and varying quantities of alloys to give a range of functional characteristics. Several types of steel are used in vehicle construction. For this scoping study a low alloy steel was picked as an example; the influence of this choice on the results is discussed below.

Pellet plantSinter plant

Coke Oven

Coal

Blast furnace

Pig iron

Oxygen converter

Pellet plantSinter plant

Coke Oven

Coal

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Car manufacturer 10% ‘prompt scrap’

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Iron ore

Coke

Ferroalloys

In a more detailed study, it would be desirable to try and identify the quantities of different types of steel used in a car, and to perform the analysis on this basis. Car production data is based on LCA data for car production in Germany; only a limited data set (for some air pollutants and three water pollutants) was available. Emissions from car production were allocated on the basis of the percentage by mass of steel in the car (~ 60%). This is a gross approximation. Ideally allocation should be done on the basis of some physical, causal relationship, but where this is not possible, allocation by mass of the outputs of a process, or by economic value is often used. A rate for ‘prompt’ scrap arising from the manufacturing process for the car was taken from Moll et al, 2005.

At the end of its life, the car is assumed to be shredded, with 95% of the steel in the car being recovered. This scrap and the steel scrap from the shredding of the car at its end of life are assumed to go into an electric arc furnace and produce secondary steel. This is assumed to offset primary steel.

The only data on the shredding process which could be found within the project timescale was on the energy consumption of the process. Other emissions are not estimated, but are likely to be small compared to those from other parts of the lifecycle. Other dismantling processes are not modelled.

Assumptions made in modelling the lifecycle are summarised below.

Table 3.3 Assumptions and Data Sources for Simplified Life Cycle for Use of Steel in a Car

Source of iron ore and pellets

Iron ore is imported into the UK mainly from Australia, Canada, Brazil and South Africa (ISSB, 2005). A weighted average of transport distances by sea from these locations was used. Additional road transport of 100km once in the UK was assumed.

Coal and cokeCoal and coke are assumed to be transported from Australia by sea, which is the main source of imported coke for the UK iron and steel industry (ISSB, 2005).

End of life vehicle waste

disposal

95% of the shredded ferrous metals are recycled the remaining 5% are sent to landfill. Energy consumption is 40 kWh per tonne waste (Smith et al, 2001)

Landfill Distance to landfill 50km

Electric Arc Furnace

Distance to electric arc furnace 100km

Transport distances for raw materials and waste products within the UK are estimated based on geography and experience of the waste management sector. While these are likely to be the right order of magnitude, actual distances could vary by a factor of two to three from those estimated below. The uncertainty this introduces into the results is discussed below.

The main limitation of the analysis is the limited data set available for car production, and the allocation methodology used for the car production data set. The results indicate that the production stage is potentially a significant contributor to the overall lifecycle impacts, and in a more detailed study this is an area that would need to be examined further. As no existing data appears to be available this could require a significant amount of effort.

3.3.3. Key environmental impacts

As discussed above, a full set of emissions were not available for the car production phases, and furthermore, only crude allocation of the car production emissions on the basis of weight was possible. These approximations must be borne in mind when interpreting the following results.

Figure 3.10 and 3.11 show the environmental impacts (as assessed using the CML and EI99 impact assessment methodologies) by life cycle stage for 1 tonne of steel used in a car. The contribution of the car production stage could only be shown for global warming, photochemical oxidation and acidification for the CML methodology and damage to human health from climate change and respiratory organics and inorganics and damage to ecosystems from acidification/eutrophication in the EI99 methodology. as data on emissions and resource use which are relevant to other categories was not available. Even for the categories for which a contribution is shown data for all relevant emissions was not available and the contribution to eutrophication, photochemical oxidation and ecosystem damage is almost certainly underestimated, and there may be a small underestimation of the contribution to global warming and acidification. The contribution of the car production stage to the overall lifecycle for the limited range of pollutants for which data is available for that stage is shown in Figure 3.12.

Impacts in a number of categories are relatively significant. Steel production is known to be an energy intensive process, and so using the CML methodology global warming and acidification impacts (which are primarily associated with pollutants from fossil fuel combustion) are high for steel production. Within steel production, the blast furnace process to produce pig iron, and the sintering of iron are significant contributors to these impacts. It is interesting to note however, that car production makes a significant contribution to the overall lifecycle for these impacts. Figure 3.12 shows that CO2 emissions from the car production stage are as great as those from steel production. For SO2 and NOx, they are just under half of the emissions from steel production. This is on the basis of allocating emissions from car production on the basis of the percentage mass of materials in the car. As discussed above, this is a gross approximation, and ideally allocation should be based on causal relationships. However for the energy related pollutants (CO2, SO2 and NOx), this approximation is considered unlikely to invalidate the conclusion, as the fabrication of metal components is known to be energy intensive compared to, e.g. the fabrication of synthetic materials (the next most significant component by mass).

The ‘high’ score for fresh water aquatic toxicity is dominated by nickel and cobalt emissions to water from the disposal to landfill of smelter slags from ferroalloy production (particularly ferronickel) and the disposal of slag from the electric arc furnace (which reduces the ‘credit’ from recycling of scrap steel). Disposal to landfill of slags from ferroalloys (particularly ferrochromium) is also the main contributor to human toxicity (through emissions of chromium VI). In the case of terrestrial ecotoxicity, emissions to air of mercury from the electric arc furnace process is the main contributor.

Using the Eco-Indicators methodology, the most significant net impact is damage to human health from inhalation of inorganic substance respiratory inorganics, mainly particulates from iron ore extraction at the mine (so impacts are substantially offset by recycling at tend of life. Damage to human health via climate change is also significant, and as described above for global warming, key lifecycle stages are the blast furnace process and sintering of iron. Damage to human health from carcinogens and damage to ecosystems from ecotoxic chemicals are also relatively significant. In the case of the former this is due to emissions of arsenic and cadmium to water from iron ore extraction and pelletisation, and in the case of ecotoxicity to emissions of chromium and metal associated with ferro chromium and ferronickel production and the disposal of smelter slag.

The impact assessment results show that because of the energy intensive nature of the steel production process, transport does not contribute significantly to environmental impacts and the uncertainty in the overall results introduced by the assumptions made on transport distances is therefore likely to be low. The choice of the type of steel has more of an impact on the results as emissions from ferroalloy production and processing contribute significantly to a number of impacts.

The iron and steel materials system in the EU15 was studied in some detail in a report by the European Topic Centre on Waste and Material Flows (Moll et al, 2005). This found that the main environmental hotspots were:

Generation of overburden and other mining waste from the extraction of iron ore, coal and ferroalloys. The majority of these occurred outside the EU15.

CO2 emissions related to the production of steel, which is very energy intensive. These arise from the combustion of fossil fuels in the process, fossil fuels used to generate electricity used in the process, and process related emissions.

Emissions of heavy metals and harmful organic compounds from the electric arc furnace process used to produce secondary steel. These arise from impurities in the steel.

Emissions of pollutants associated with energy consumption (CO2, SO2 and NOx) in the ‘product manufacturing’ sectors.

The last three of these points were also apparent from the results of the simplified lifecycle. However, it is worth noting that the generation of overburden (the rock and soil above the mineral of interest which must be removed to mine the mineral) and other mining waste are not always included in the lifecycle inventory data.

The environmental impacts from disposal of this waste, and also the amenity impacts associated with mining could, however, be significant, and illustrate the necessity of looking at a wider range of environmental impacts than those typically calculated in the life cycle inventory and typically assessed in impact assessment methodologies, to ensure that the full range of environmental impacts are assessed.

Similarly, the lifecycle inventory tool does not calculate the total amount of waste disposed of to landfill, which can be an important consideration in terms of waste policy. The impacts of disposal in terms of landfill gas emissions and leachate production are accounted for however, and as discussed above were for this example a significant contribution to some impact categories.

Figure 3.10 Environmental Impacts for 1 tonne of steel used in a car (CML IA methodology)

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Figure 3.11 Environmental Impacts for 1 tonne of steel used in a car (EI99 IA methodology)

Figure 3.12 Contribution of emissions of air and water pollutants from life cycle stages for steel use in a car

Note: Scale shows values of more than 100% as there are both positive and negative offsetting) contributions to each impact. The net value of for each impact category is always 100%.

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3.4. WOOD

3.4.1. Material flows through the UK Economy

The material flow for wood in the UK economy was recently examined by Trada Technology (2005a) as part of the Biffaward Programme on Sustainable Resource Use. The study brought together the findings of individual mass balance studies undertaken by the five key sectors in the timber industry. It is a comprehensive and well researched report, which made use of Products of the Community (PRODCOM) data (available from the Office of National Statistics) on timber and timber products, supplemented by market research to complete areas where information was unavailable or insufficient. There was also considerable consultation with industry bodies.

The study found that the material flow for wood in the UK economy is complex (TRADA, 2005a) . As well as domestic production of timber and timber products, there are imports and exports of wood, sawn timbers, intermediate products such as plywood, veneer sheets, particle boards and fibre boards, and of finished products such as furniture and of packaging (mainly as pallets). A simplified view of the material flows is given in Box 1, together with details of the main uses of wood in the UK: construction, accounts for just over a third, and packaging and pallets, almost one-fifth each. About one-sixth is transformed into paper, and the remainder is used for joinery.

About 4 million tonnes of wood are estimated to enter the waste stream from construction and demolition sites6, together with over 1 million tonnes from packaging. Estimated wood wastes from the furniture industry are lower (about 0.3Mt). About 10% of waste wood in the UK is estimated to be recycled via a number of routes including use as animal/poultry bedding, mulching/composting, use in panel manufacture and combustion to produce heat and/or power.

3.4.2. Material/product life cycle

Construction is the main use of wood in the UK, and within the construction sector, softwood (which becomes embodied in construction) is the main product used (TRADA Technology, 2005b). This was therefore the lifecycle chosen for analysis. Of softwood used in construction, 55% comes from the UK and 45% from overseas (TRADA Technology, 2005b), principally Scandinavia. The key assumptions and data sources used to model the simplified lifecycle are listed in Table 3.4. A major simplification is that no treatment of the wood (e.g. with preservatives) has been included. Transport distances for timber and waste products were estimated based on geography and experience of the waste management sector. While of the right order of magnitude, actual distances could vary by a factor of two to three from those estimated below. The uncertainty this introduces into the results is discussed below.

6 http://www.globaltrees.org/proj.asp?id=4

Box 1 Material Flows in the UK for the Production and Use of Timber

The finished products from the wood panel mills were primarily particleboard or oriented strand board (OSB). In 2004, 2.6 million m3 of the finished panels were particleboard or OSB and 0.9 million m3 were fibreboard (UK Timber Statistics 2004).

Consumption of Wood and Wood Based Products in the UK

Source: TRADA Technology Ltd, 2005a

Figure 3.13 Simplified Life Cycle of Softwood used in Construction

Home-grown Softwood

Scandinavian Softwood

Drying and planing

Drying and planing

UK Construction site

Fell tree Fell tree

UK Sawmill Scandinavian Sawmill

45% residues

45% residues

WasteWaste

39% waste

Landfill 37%

Incineration 17%

Re-use 15%

Recycle 31%

Embodied in house

Landfill 61%

Incineration 6%

Re-use 32%

Recycle 1%

Demolition

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Transport assumption

Loss of product

Table 3.4 Key assumptions and data sources for the simplified timber LCA

Assumptions and data sources

Softwood production

With 55% of wood used in construction is sourced from the UK and 45% from overseas (TRADA Technology, 2005b). The majority of overseas softwood comes from Scandinavia, and Sweden (as a major Scandinavian timber producer) was chosen as the country of origin for this example. Data for production of softwood in Northern Europe was used for both the Scandinavian and UK situations.

Forest to Sawmills

Transport from forest to sawmill is estimated at 50km for both the UK and Scandinavia as sawmills are typically located close to the forest.

Drying and planing

Low-grade construction timber is typically kiln dried, rather than air-dried. Kiln drying occurs at the sawmill using wood chips produced at the sawmill.

Sawmills

Only 55% of the softwood, which enters the sawmill, leaves as the primary product to the construction site. The residues left are used in other applications (TRADA Technology, 2003). The environmental impacts from the sawmill are allocated between the timber produced and other residues.

Sawmill to construction site

The Swedish timber is assumed to be transported 300 km by road to the port of Gothenberg and then onwards by ship to London, and then again by road (100km) to the construction site.

Construction site

39% of the wood that enters the UK construction site leaves as waste wood (TRADA Technology, 2005b). The other 61% is embodied in the house. It is estimated that at the construction site it will take 30 minutes with a power saw to cut a 1m3

piece of timber to size.Construction

wasteThe waste scenario breakdown is as reported in TRADA Technology (2003).

Demolition waste

Destination of timber reclaimed from buildings that have been demolished is taken from Thompson (2005). The energy and equipment required to demolish the building and extract the timber were not estimated, as information on this was not available within the time frame of the study. This is a limitation of the study.

Transport of construction and demolition waste

Is estimated at 50km to landfill and incinerator sites and 100km to recycling markets. This is an estimate for typical distances to waste disposal sites based on experience of the waste management sector.

3.4.3. Key environmental impacts

The environmental impacts each lifecycle stage up to embodiment in the building (as assessed using the CML IA methodology) from are shown in the figure. These show a significant benefit in terms of global warming due to carbon dioxide which is ‘stored’ in the wood. Production of wood is non-energy intensive (compared to other construction materials such as steel), and therefore the contribution of transport to impacts such as acidification and eutrophication is relatively high. As discussed above, estimates of typical transport distances were made, and while of the right order of magnitude, they could be up to perhaps three times higher in some cases. If road transport distances were three times greater than assumed, then transport would be responsible for about 60% of acidification and eutrophication impacts and the a very significant contributor to most other impacts. This suggest that more accurate modelling of the lifecycle, and improved data on average transport distance should be undertaken in any further analysis which is undertaken, to clarify the importance of transport distances in the environmental impacts of timber.

Figure 3.15 shows the additional impacts from the waste management stage – a considerable amount of scrap wood arises at both the construction phase (as about 0.6 tonnes of wood waste is created for every tonne of wood incorporated in the building) and (decades later) during demolition of the building. To show the differences between the waste management options clearly, results have been calculated assuming all of the waste wood arising at both construction and demolition is either recycled (into wood chips for use in particle board manufacture), landfilled or incinerated. Energy recovery is assumed for the last two options, and the conventional fossil fuel generation which is avoided by this energy recovery in the disposal phase reduces impacts in most categories, apart from eutrophication and freshwater ecotoxicity which are both increased by landfilling. The way that CO2 storage in the wood is treated within the IA methodology means that wood recycling shows a reduction in offset CO2 – this is because the recycled wood replaces additional wood which would have been growing in the forest that would have offset some CO2 in itself. In the real world, CO2 continues to be stored in the wood when it is recycled, so that it could be argued that the benefit remains.

The impacts under different waste management scenarios using the Ecoindactors IA methodology are shown in Figure 3.16. This shows that the most significant impact is land use (for forestry), followed by fossil fuel use. The latter is reduced when scrap wood is landfilled or incinerated due to the energy recovered during these processes. The most significant impacts in terms of damage to human health are respiratory inorganics and carcinogens. The main contributors to respiratory inorganics are particulates (from the combustion of woodchips at the sawmill to kiln dry the wood, and from cultivation and harvesting operations) and NOx emissions (from transport, from the combustion of woodchips at the sawmill to kiln dry the wood, and from cultivation and harvesting operations). Carcinogens are cadmium and arsenic, from road transport and combustion of woodchips. There are additional emission of these two carcinogens if wood is incinerated or combusted.

Figure 3.14 Environmental impacts from life cycle stages for 1 tonne of wood embodied in a building (CML IA methodology)

Figure 3.15 Additional environmental impacts from disposal of waste wood at construction and demolition for 1 tonne of wood embodied in building (CML IA methodology)

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Figure 3.16 Additional environmental impacts from disposal of waste wood at construction and demolition for 1 tonne of wood embodied in building (Ecoindicator IA methodology)

Neither of the impact assessment methodologies capture a number of environmental impacts, which can be associated with forestry. These include reduction of biological diversity and its associated values, water resources, soils, and unique and fragile ecosystems and landscapes. Impacts on biological diversity, ecosystems and landscapes are particularly important for harvesting from primary forests (as compared to commercial plantations).

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4. Externalities

Most activities, such as electricity generation, transport, agriculture, and mineral extraction have environmental and social burdens. These include air pollution, greenhouse gas emissions, waste arisings, water emissions and noise.

These environmental burdens – via the environmental impacts they cause - lead to economic costs that are typically not paid for by providers or users (of the goods or services in question). These are known as external costs or externalities7. In order for economic activity to be sustainable, it is essential that these environmental and social externalities are taken into account.

Historically, environmental legislation has focused on the use of command and control legislation to address environmental problems, i.e. through traditional regulation. However, there is a growing recognition that policy needs to shift towards the use of economic instruments 8(e.g. green taxes, trading schemes) to tackle these externalities. This has been recently recognised at a European level9, though it is the UK that has advanced the use of economic instruments over the last decade.

This chapter details the methodological approaches for externality assessment, and how these can be applied to life cycle analysis. It also undertakes a scoping analysis to assess the externalities from the material LCAs identified in the previous chapter.

4.1. Methodological approach for assessment of externalities

There is a long history of external costs studies. Several studies were published in the late 1980s and early 1990s which estimated some of the externalities associated with electric power production and fuel cycles10. Many of these earlier studies adopted a top-down approach, working with national estimates of emissions and national economic costs.

However, a growing recognition during the 1990s identified the importance of technology, location, and life-cycle thinking, particularly in accurately estimating marginal external costs. As a result a bottom-up approach was developed and advanced, notably through DG Research’s ExternE Project (1995:1999).

This has led to the development of a standard framework for externality assessment, known as the ‘impact pathway approach’ (or alternatively, and less accurately, the ‘dose-response’ approach). It is stressed that while the framework is now generally accepted, the specific methodologies used for impact relationships and valuation are still evolving. The impact pathway approach adopts a logical progression assessing: Environmental burdens; 7 Also often known as ‘social’ costs, i.e. relating to the full costs to society.8 We define economic instruments as including environmental taxes, charges and subsidies, as well as market-based instruments. Economic instruments in Government are discussed in the document: Tax and the environment: using economic instruments. HM Treasury. 2002.9 ‘Getting the prices right’ is one of the key indicators of EC sustainability strategy10 The most prominent of these were by Hohmeyer (1988), Ottinger et al at Pace University (1990), Bernow et al (1990) of the Tellus Institute, ECO Northwest (1987) and Pearce et al (1992).

Environmental impacts, by assessing the change in the state of the environment and the impact on receptors;

Monetary valuation.

The best example is in relation to air pollution, where the steps relate to the: Analysis of emissions (burdens); Analysis of impacts, calculated by combining concentration (following dispersion

modelling), exposure (using the stock and risk), and relationships linking changes in the state of the environment to impacts;

Valuation of these impacts.

The approach is shown in the figure below.

Figure 4.1 Illustration of the impact pathway approach taking the example of direct effects of ozone on crops

Source: Holland et al, 2005.

Following from the figure, impacts and economic damages under any scenario are calculated using the following general relationships:

In the case of air pollution, the impact may be expressed in terms of concentration or deposition. The term ‘stock at risk’ relates to the amount of sensitive material

(people, ecosystems, materials, etc.) present in the modelled domain that are exposed to the impact (in this case increased air pollution).

Although the impact pathway approach can be widely applied, the approach does vary from impact to impact. For example, the impact pathway that describes materials damage from acidic deposition requires consideration of climatic variables (such as relative humidity), atmospheric chemistry, and the need to account for several pollutants interacting simultaneously. In contrast, simpler approaches can be used for direct cause and effect pathways, e.g. as with traffic accidents.

For any type of receptor it is necessary to implement a number of these impact pathways to generate overall benefits. So, for example, in the case of impacts of ozone on crop yield, it is necessary to consider, separately, impacts on a series of different crops, each of which differs in sensitivity. For health assessment it is necessary to quantify across a series of different effects to understand the overall impact of air pollution on the population.

The final stage, valuation, generally uses the perspective of ‘willingness to pay’ (WTP) – this is the willingness to pay of the affected individual to avoid a negative impact. The rationale is that values should be based on individual preferences, which are translated into money terms through individual WTP (EC, 1998)11.

For some effects, such as damage to crops, or to buildings of little or no cultural merit, this valuation can be done using appropriate market data. However, for some impacts, such as human health, there are no direct market prices. In this case, alternative approaches are needed. Some elements of the valuation of health impacts can be quantified from ‘market’ data (e.g. the cost of medicines and care). However, these still do not consider other elements, such as willingness to pay to avoid being ill in the first place. In such cases, alternative methods are necessary for quantification, such as the use of contingent valuation, which elicits the WTP through direct questionnaire. For discussion of this and other valuation techniques, see ExternE, 1998; EAHEAP, 1999.

When monetary damages occur in the future, it is necessary to consider discounting. Discounting is a technique used to compare costs and benefits that occur in different time periods. It is a separate concept from inflation, and is based on the principle that, ‘generally, people prefer to receive goods and services now rather than later’ (HMT, 2004). Discounting (using a discount rate) is used to convert all costs and benefits to ‘present values’, so that they can be compared in equivalent units.

All of the information on exposure, stock at risk, concentration-response and valuation needs to be brought together for the calculations. These are usually made within a grid system, to avoid averaging of data.

11 There is an alternative, which is the willingness to accept payment as compensation if a negative impact takes place. There is evidence that suggests that the choice between WTP and WTA approaches can lead to rather different results – with WTA estimates being up to two orders of magnitude larger than WTP.

There has been a recent shift towards the use of automated analysis within Geographical Information Systems (GIS), using grid systems with calculations undertaken in a matrix of grid cells at a high resolution. This allows the analysis of, e.g. the pollution concentration, number of exposed individuals, number of calculated impacts, and monetary valuation of these impacts, at a highly dis-aggregated level. The total effects across the area can then be aggregated (summed) for the overall results.

This approach has been adopted in a recent European impact assessment, as part of the Clean Air for Europe (CAFE) programme by DG Environment, and supporting the adoption of the Thematic Strategy on Air Pollution.

It has also been the method adopted in the UK for the regulatory impact assessments on the air quality strategy and updates, as undertaken by the Inter-Departmental Group on Costs and Benefits (IGCB 1998:2001), and in the recent Air Quality Strategy Review (IGCB, 2006).

As set out above, the approach has been most commonly used in the analysis of air pollution. Variations are used for other types of impacts, sometimes representing simpler or more complex impact pathways. For example: The detailed impact pathway approach is extremely complicated for assessing the

full social costs of climate change (known in the UK as the social cost of carbon (SCC))12. The values used are based on Integrated Assessment Models, which do follow an impact pathway approach, but have to use a simplified form.

For some effects, such as amenity, a simpler impact pathway can be used, by using the results of detailed hedonic price studies linking, for example, landfills and local property prices or rents (as in the Defra funded study by Cambridge Econometrics et al, 2003).

The site specific nature of water pollution has proved extremely challenging for the impact pathway approach, particularly in using the results of detailed studies for aggregation at a policy relevant, or national level. While the approach is possible, there are issues over transferability, i.e. how accurate it is to take impact analysis or monetary valuation from one location and apply to other cases.

However, the impact pathway approach is very detailed and resource intensive.

Ideally, for every type of case study (e.g. each resource flow in this study), the analysis should undertake a rigorous analysis of the benefits or dis-benefits, using a detailed ‘impact-pathway’ approach. This would involve assessing burdens, modelling the change in the state of the environment (e.g. using dispersion models), quantifying impacts using GIS, and valuing those impacts.

Unfortunately, it is not practical or cost-effective to apply this methodology to all policy appraisals or studies. This is also the case when considering resource flows across the economy using life cycle analysis, as with this study.

12 The SCC is usually estimated as the net present value of climate change impacts over the next 100 years (or longer) of one additional tonne of carbon emitted to the atmosphere today. It is the marginal global damage costs of carbon emissions, i.e. the costs with climate change actually occuring.

To address this, recommendations from policy reviews, such as the Air Quality Evaluation (Watkiss et al, 2005), have led to sets of simplified benefit values based on detailed impact pathway analyses (often known as ‘damage costs’) for appraisal and other uses13. These are usually expressed as damage costs in £ per tonne (e.g. £ of health damages per tonne of particulate matter (PM) pollution emitted).

It is stressed that while these provide a consistent set of numbers for benefits analysis, their use is only recommended in certain circumstances e.g.:• As part of a filtering mechanism to narrow down a wide range of policy options

into a smaller number that are then taken forward for more comprehensive assessment;

• Where the policy is primarily concerned with other objectives, e.g. for air quality impacts that are expected to be ancillary to the primary objectives of say a transport policy.

They are ideal for the sort of assessment considered here. The specific damage costs we have used, by impact category, are described in a later section.

4.2. Calculation of Externalities from Life Cycle data

Historically, LCA and externality studies have been rather separate disciplines (environmental science and economics, respectively). Moreover, they have been used to answer different policy questions. LCA has tended to be used to look at product choice and options, e.g. for end of

life management, or comparing product choices. It is focused on producing relative values for direct comparisons between options.

Externality studies (based on the impact pathway approach) have been used for cost-benefit analysis or as an input into the design of economic instruments. They are focused on generating absolute values (rather than relative values as for LCA above). These have been targeted at policy questions of whether the benefits of policies justify the costs, or the level of taxes that might be appropriate to try to correct market distortions.

There are a number of other historical differences between the approaches.

Average vs. Marginal. While life cycle analysis is based on careful and holistic accounting of all flows associated with a system or process, it has tended to work with average data and analysis (i.e. average burdens)14. In contrast, externality studies are usually focused on marginal analysis, i.e. looking at the additional environmental impact at the margin, e.g. for one additional power plant, or one additional tonne of emissions. This focus on marginal analysis is needed to capture the fact that additional (or marginal) damages often increase with pollution. This argument is taken from a parallel with other goods that affect individual welfare, and is also essential for assessing the economic optimal point for policy (when comparing marginal abatement costs and marginal damages).

13 The EC’s Clean Air for Europe (CAFE) programme has produced similar unit pollution costs (see Holland et al, 2005b), following from an initial set produced (the BeTa values (Holland and Watkiss, 2002) for the EC. 14 Though we note that the latest LCA guidance is to use marginal analysis when looking at offsets. E.g. in electricity production.

Burdens vs. Impacts. Life cycle inventory / analysis has tended to work with environmental burdens (i.e. pressures). For example, the emissions of acidifying pollutants, or the emissions of toxic pollutants generated. This is necessary to capture the extremely large number of potential burdens across the life cycle. Externality studies (using the impact pathway approach) have tended to focus on environmental impacts. For example, the health impacts of air quality concentrations. This incorporates an analysis of the change in the state of the environment, and consideration of the stock at risk (e.g. population exposed to pollution). This focus on impacts is more detailed and accurate, but necessitates a narrower set of impacts to allow quantification. As a result, externality studies have tended to focus on specific activities (e.g. waste, transport), ignoring the complete wider effects across life-cycles, and the fully holistic and comprehensive approach in LCA.

Aggregation vs. Location specific. Life cycle analysis typically aggregates burdens across the life cycle, e.g. adding up the emissions across the life cycle to give total acidifying load. This is an extremely effective way of capturing the relative burdens from different choices over the full life cycle. Externality studies (at least those using the impact pathway approach) tend to assess impacts separately, and only aggregate monetary values. Related to this, life cycle analysis has tended to ignore location and so all burdens are treated equally, irrespective of where they are emitted15. In contrast, the impact pathway analysis (externality analysis) specifically addresses location, in recognition that the impacts of the same burden will differ by area due to site-specific effects. As an illustration, the health impacts (and externalities) of a tonne of air pollution emitted vary according to whether these occur in rural or urban areas, because the population density, and so exposure, is vastly different in these locations.

The above points are not a criticism of LCA. In practical terms it would be impossible to adopt the detail needed to accurately undertake impact pathway analysis for a full LCA (i.e. to assess marginal, location-specific impacts). Moreover, LCA has not been targeted at providing marginal social costs (or impacts) and so has not needed to adopt a similar line of analysis to the externality studies. However, it is important to recognise these differences, in order to combine the approaches here for the study objectives (assessing externalities, policy coverage and optimality).

Notwithstanding the differences above, there has been greater interaction between LCA and the impact-pathway approach over recent years, towards some harmonisation, for example: The better externality studies have used a life cycle approach to identify burdens

(as in the case of the ExternE project (1995: 1998) and its use of fuel cycle analysis). This approach is used to screen all the potential burdens across the life cycle to select priority impacts. These priority impacts are then assessed in detail using an impact pathway approach. This approach effectively uses LCA as a filter to focus on the most important externalities for quantification and valuation.

15 This has been recognised as a limitation in earlier life cycle analysis, and some life cycle inventory data is now using classifications (e.g. urban or low population density) to recognise the additional characteristics of the burden, to allow more sophisticated impact assessment methodologies to be developed.

Some of the more recent impact assessment methods have taken on board the findings of the externality studies – for example in adopting PM as a major health risk in human toxicity (whereas earlier LCAs ignored this pollutant).

Much of the LCA literature is trying to focus life cycle impact assessment on ‘impacts’ rather than ‘burdens’, using the exposure - response relationships that have emerged from the detailed impact-pathway studies (e.g. through the use of damage category IA methods, or through specific site-dependent characterisation factors for emissions from processes that are geographically specified in the inventories)16. There is also a recognition that population weighted exposures from air emissions are useful in deriving appropriate damage weightings.

For the analysis here, we have explored two broad approaches to use the information in the previous chapter, and estimate the externalities for the three material life-cycles. These are:

1. Take the LCA output (in terms of the traditional output of global warming potential, human toxicity, acidification, fresh water aquatic eco-toxicity, etc.) and convert these values to monetary estimates directly. This approach has been used in recent waste studies (e.g. IPTS, 2006), but does have a number of methodological problems and limitations.

2. Take the LCI output (i.e. the inventory data on burdens), and apply damage cost values to these numbers. This is a more rigorous approach, and is more consistent towards answering the study objectives, but raises issues over coverage of impact categories.

These approaches are discussed below.

4.2.1. Using LCA output and monetising

We have considered the LCA output from one of the more widely used LCA impact assessment methodologies used in Europe (the CML problem orientated method – see previous chapter). This uses weighting factors to allow emissions of pollutants17, which contribute to an environmental impact or problem, to be aggregated together. The environmental impact/problem categories in this methodology are: Abiotic depletion Global warming Ozone layer depletion Human toxicity Fresh water aquatic ecotoxicity Marine aquatic ecotoxicity Terrestrial ecotoxicity Photochemical oxidation Acidification Eutrophication

The analysis has also considered the Ecoindicator 99 impact assessment method (See previous chapter).

16 For example, IMPACT2002+ v. 2.1 (Jolliet et al. 2003) or the EDIP2003 methods (Hauschild & Potting 2005).17 Or in the case of raw materials, resource use

The major advantage of using the LCA output directly is the comprehensive and holistic coverage. The life cycle inventory generates hundreds of potential burdens, covering emissions to air, water and soil (and also resource use). From this: The approach allows the consideration of emissions from across the life-cycle, so

for example, it will capture all the different greenhouse gas emissions (CO2, CH4, etc.) from all activities associated with the material flow. These burdens can then be aggregated (e.g. all CO2 added together), and each pollutant multiplied by a relative indicator – in this case the global warming potential (GWP) of each greenhouse gas relative to CO2 emissions – to give the overall global warming burden. A similar approach is used for ozone layer depletion, or photochemical oxidation.

It can also compare the burdens of different pollutants from different media. For example, the LCA can estimate the relative human toxicity of several hundred potential emissions as emitted to air, water or soil. The individual burdens within each group are then weighted using indicator factors so that they can be compared against each other, and aggregated e.g. to give the total human toxicity. Similar approaches are used for ecotoxicity, acidification and eutrophication.

The burdens from each impact category can be normalised18 to give an indication of the importance of the effect. This can be by normalising against the total European levels of that burden (as in the CML method), or through alternative approaches such as the distance from political targets, or through expert panel values. Note, LCA does not usually compare between impact categories, i.e. to give the direct relative importance of global warming versus human toxicity, though some impact assessment methods seek do provide alternative weighting factors that can be applied to the normalised indicator results. Of course, one approach for normalising between impact categories is to use monetary valuation (i.e. to aggregate indicator results across impact categories), as in the impact pathway approach, though this is not commonly adopted in LCA.

However, there are some problems with LCA impact assessment as well. A major downside is that none of these relative toxicity values are based on agreed UK specific methodologies. For example, within the CMAL methodology, the approach for human health is based on the generic health impacts literature, rather than the specific Department of Health (DoH) guidance. This is particularly problematic where there is firm guidance in place for quantification from DoH (as with the combustion pollutants such as PM), or where the DoH guidance differs to the generic literature (e.g. over the potential risk of certain carcinogenic [cancer risk] pollutants).

If these specific methodological issues are put to one side, it is possible to undertake valuation using LCA output directly – at least for some categories – at least in some impact assessment methodologies.

18 Normalisation expresses the results relative to a reference value. Usually normalisation works by taking a result and either dividing it by the selected reference value, or multiplying by a normalisation factor.

A good example is with the use of human toxicity, as expressed in QALY (Quality Adjusted Life Years)19 or DALYs (Disability Affected Life Years)20. Here, one can directly apply valuation endpoints to the numbers generated from the LCA21.

This approach does generate monetary estimates of total health effects, and has the advantage of being able to cover a much wider range of potential pollutants. However, it is less robust than the direct estimation of specific health impacts and direct valuation of these impacts (the latter derived from primary CV studies to elicit WTP for specific health impacts), as is usually undertaken in the impact pathway approach. Essentially there is a trade off between a high coverage of all potential pollutants with a lower level of confidence, versus a focus on a smaller number of more robust estimates but with gaps for other pollutants.

However, other LCA impact categories are not so easy to value. It is more difficult to put a value on photochemical oxidation potential per se, or generic acidification potential. These areas have been (partially) quantified and valued in the impact pathway literature, through the quantification and valuation of specific impact categories, e.g. ozone formation and its impacts on crops (see Holland et al, 2005), or acidification and the effects on buildings (Watkiss et al, 2001), and have been used to provide the analysis of non-health benefits (though with a focus on air pollution) in UK economic analysis for IGCB (see chapter 5, IGCB, 1998; IGCB, 2006). Some of the more robust LCA impact assessment methods have adopted more sophisticated approaches to try and tackle these categories, working with site-dependent characterisation factors for the impact categories acidification, eutrophication and photochemical ozone formation, for most European countries (as with the EDIP 2003 method (Potting & Hauschild 2005). This moves the LCA output to a closer match with the underlying economic valuation literature of these categories.

There are also some impact categories that are extremely problematic for valuation, notably ecosystems and biodiversity. This applies equally to any LCA based approach or any impact pathway approach. In many cases, these categories are only reported in terms of physical impacts22.

There are some studies that try to use LCA output and provide overall externality estimates for all impact categories (e.g. IPTS, 2006), but the results are generally unsatisfactory. 19 Calculated as the number of human life-years affected multiplied by a severity score (quality adjustment) between 0 and 1, where 0 is equal to death and 1 is equal to perfect well-being.20 Method to calculate health related indicators and differences in the burden of disease and in causes of illness, including morbidity and mortality in a weighted indicator.21 We stress, however, that there are issues about how accurately this sort of generic approach captures mortality and morbidity aspects, i.e. capturing all the impacts of illness in terms of resource costs, opportunity costs and dis-utility. There is also considerable uncertainty attached to transferring values between contexts. Contextual differences include: the absolute size of the risks; knowledge of the risks; level of involuntariness associated with the risks; socio-cultural determinants of preferences relating to risk, different causes of death, etc. There are issues according to when impacts occur (and LCA does not discount) for example whether there is a lag between exposure and impact (as in the case of cancer), and the age group affected, for example as different WTP values are usually applied for infants. 22 from a policy perspective, the output from LCA as physical outputs (tonnes of acidifying pollutants) is less useful than output from impact pathway analysis (which in recent studies has been in km2 of area that exceed critical loads).

As a final issue, there are some impacts that are missing in LCA, which are relevant for externalities. This includes: Amenity effects (including noise and nuisance). These are usually covered in

externality studies. Accidents and occupational risk. These are often included in externality studies

(e.g. ExternE, 1998; Sansom et al, 2001), but there are issues whether these are partially of fully internalised through wage premiums.

4.2.2. Using LC Inventory output and monetising using impact pathway

The alternative approach we have considered is to use the LCI data and directly value these burdens using a simplified impact pathway approach with damage costs (e.g. cost per tonne estimates).

This has a major advantage as damage costs are based on an underlying impact based approach, and for many burdens, it is possible to use accepted approaches for the quantification and valuation of impacts, e.g. for the human health impact of air pollution, where clear DoH guidance exists (e.g. COMEAP, 1998; 2001; 2006). This approach also allows a direct match between impacts and valuation endpoints; so for example, it is possible to match WTP values for a morbidity illness (e.g. using values recommended by IGCB, 2006) to the specific health endpoint.

Where this approach falls down is that there are relatively few areas where UK Government has agreed damage cost estimates. These tend to be limited to GHG emissions, air quality (focused on the classical air pollutants (SO2, NOX, VOCs, CO and PM)), and some areas of amenity.

Of course there are values in the literature for many of the LCA burdens identified (and impacts in the LCA) such as for trace metal pollutants, acidification, and even for some categories that are difficult to quantify (e.g. ecosystems and biodiversity) but these are not agreed. Note this problem, however, applies equally to the use of LCA impact assessment output above.

There is one potential variation on this approach that is also worth consideration. This is to use the damage cost approach as above, but to explore the potential gaps using sensitivity analysis. Two potential approaches have been considered for this. The first uses typical damage cost from the literature, to test whether any potential emissions are significant compared to the quantified impacts. The second is to use the estimates of specific burdens (e.g. valuation of PM), and then use these values with the LCA impact assessment data on the relative importance to back-calculate all other burdens within each impact category. For the scoping analysis here, we have focused on the first of these approaches, but recommend the second approach is investigated in any follow-up work.

4.3. Scoping of Externalities for the Three Material Life Cycles

The study has investigated the potential externalities for the three material life cycles. It was hoped that both approaches above could be investigated in the study, but due

to time constraints, it was decided to pursue the most promising approach – using LCI data and existing damage costs based on impact pathway analysis (the second option above). The approach adopted is outlined by pollutant below.

4.3.1. Greenhouse Gas Emissions

We have used the Government recommended values for the Social Cost of Carbon (SCC) (GES, 2002). This suggested an illustrative central value of £70/tC for use in appraisal, within a range of £35 to £140/tC, as an illustrative estimate for the global damage cost of carbon emissions23. It was recommended that these (2000) values are increased in real terms at £1/tC per year. The GES paper also recommended periodic review of these values. A review was recently completed (Downing et al, 2005; Watkiss et al, 2006), which recommended new estimates, though this has not yet led to an update of the values24.

There are two potential approaches for using these social damage cost estimates: Firstly, to take the global warming potential output (GWP100) from the LC impact

assessment data, and apply the damage costs for carbon to these. Secondly, to keep the different GHG emissions separate, and apply specific

social cost estimates for each gas (i.e. for CO2, for methane, for N2O etc). This leads to different relative values than with the use of GWP above, because each GHG has a different atmospheric lifetime, and so has a different (relative) net present value of damage costs because of discounting. To illustrate, methane has a social cost that is far less than implied by its GWP because it has a short atmospheric lifetime25.

The choice of approach is potentially important for the wood life cycle, as this material will be in buildings for 40 years, prior to disposal. It is also important in looking at the profile of emissions from landfill, which are spread over many years.

The most correct approach would be to use the second of the above options, with the important qualification that any analysis should using rising values over time. However, the current guidance only has estimates for carbon and so the first approach has been used here.

We highlight that the SCC values are highly uncertain (see Downing et al, 2005), and a wide range of estimates exist around these values, not least over the choice of key variables such as discount rate, equity weighting, etc. Given the importance of CO2 externalities in the overall results (see later section), a more comprehensive analysis of uncertainty and sensitivity analysis would be recommended for any follow-up studies.

23 The SCC is the marginal global damage cost of carbon emissions. It is usually estimated as the net present value of the impact over the next 100 years (or longer) of one additional tonne of carbon emitted to the atmosphere today. 24 An indicative set was produced in the final study reports. These were used to help inform the analysis of potential benefits in the climate change programme review, however, any changes to the guidance are unlikely until after the Stern review has been published in late 2006..25 and so lower marginal costs as any methane emitted today will have disappeared from the atmosphere before the most severe climate change impacts occur.

4.3.2. Air Pollutants and Health/Non-health for the Main Combustion Pollutants

We have used the latest damage costs from the recent work undertaken for the Air Quality Strategy Review (Watkiss et al, 2006b). This provides an update on earlier damage costs from the Air Quality Evaluation, and follows the guidance from COMEAP on health quantification (1998: 2001, and 2006), combined with guidance on valuation from IGCB (2006), including the use of the estimates from Chilton et al (2004). The values include both mortality and morbidity (including chronic mortality)26.

Damage cost estimates are given for PM10, SO2, NOx, and VOCs. The damage costs include health impacts from primary pollutant impacts (e.g. from PM, NOx or SO2) and secondary impacts (from ozone and secondary particulates or acidification, formed from these primary pollutants). The values also include damage to crops and materials, but exclude impacts on ecosystems and cultural heritage. These two omissions have been the focus of recent research for Defra (Eftec, 2005: Eftec and EFL, 2006), but there is no agreed approach for applying such effects within the impact pathway framework, mostly due to issue of benefit transfer.

The impacts considered are summarised in the table below.

Table 4.1. Impacts included in the Damage Costs for Combustion Pollutants (IGCB, 2001; Watkiss et al, 2006b)

Burden EffectHuman exposure to PM10/PM2.5 (emitted directly or formed indirectly from NO2 or SO2)

Chronic effects on MortalityAcute effects on Morbidity (Respiratory and Cardiovascular hospital admissions)

Human exposure to ozone (formed indirectly from VOCs and NO2)

Acute effects on Mortality and Morbidity (Respiratory hospital admissions)

Human exposure to SO2 (emitted directly)

Acute effects on Mortality and Morbidity (Respiratory hospital admissions)

Exposure of crops to ozone Yield loss for barley, cotton, fruit, grape, hops, millet, maize, oats, olive, potato, pulses, rapeseed, rice, rye, seed cotton, soybean, sugar beet, sunflower seed, tobacco, wheat

Damage to materials Acidic depositionOzone damage to polymeric materialsBuilding soiling

We have applied the AQSR damage costs (Watkiss et al, 2006b) directly to the Life Cycle Inventory, i.e. to the burdens in tonnes of pollutant emitted to air.

4.3.3. Air Pollutants and Health – Other Toxic Air Pollutants

26 We highlight that while there are recommended UK values in place, and a general consensus on impacts, there remains some debate on the exact impacts and valuation estimates. To illustrate, different values (and damage costs) are used in Europe (e.g. Holland et al, 2005). Given the importance of these externalities in the overall results (see later section), a more comprehensive analysis of uncertainty and sensitivity analysis would be recommended for any follow-up studies.

The LCI data has a much wider list of air emissions than covered by the damage costs above, numbering several hundred compounds. Many of these are trace pollutants which are emitted in small quantities, but have potentially high toxicity.

There are European damage costs for some of the main toxic pollutants (e.g. Bickel and Friedrich, 2001, Watkiss et al, 2005, ExternE, 1998 and 2005) and these estimates have been used here. This includes estimates for CO, many of the heavy metals (e.g. arsenic, cadmium, lead, nickel, chromium) and some particularly toxic organics (e.g. benzene, 1,3-butadiene, dioxins, PAHs, benzo[a]pyrene). It is stressed that these impact pathways and valuation derived numbers are not accepted by the Department of Health (DoH)27.

We have applied these damage costs directly to the Life Cycle Inventory, i.e. to the burdens in tonnes of pollutant emitted to air.

However, while this covers the most important pollutants, it does not cover all potential burdens to air. To address this, a screening analysis has been used with the LCI data, applying a damage cost £1 million or £100,000 or £10,000 per tonne of emissions to the entire inventory of ‘air’ emissions, to see if any additional materials are significant relative to the main air pollutants above. We highlight that an alterative approach for screening other toxics would be to use the life cycle impact assessment toxicity results, and weight these relative to the damage costs (from Watkiss et al, 2006b) for one or two of the key combustion pollutants.

4.3.4. Emissions to Water (and effects including Eutrophication)

There has been less application of the impact pathway approach to water pollutants, and many external cost estimates are derived from ‘top-down’ approaches (i.e. based on generic burden data), such as in Pretty et al (2000, 2002), Enviros and Eftec (2004), Eftec and IEEP (2004), Markandya et al (2005).

These studies only partially cover the categories included in the LCA, with respect to human toxicity (from water pollution), fresh water aquatic ecotoxicity, and acidification and eutrophication. Moreover, they tend to work towards valuation of aggregate levels of water quality for surface water (which include an overall rating based on all pollutants and water pollution levels), or the costs of water pollution treatment as a proxy for damage costs for surface water and groundwater.

However, the studies do cover effects that are not included in the LCAs, such as capturing the amenity value of water resources (e.g. from leisure, fishing, etc). These are potentially important with respect to valuation, because the existing legalisation on water treatment (for drinking water)28 reduces the actual impacts from this burden.

27 DoH have low confidence in the values, as they are usually based on US dose-response functions that extrapolate high dose exposure down to very low concentrations linearly, without thresholds. We highlight that given this uncertainty, a more comprehensive analysis of review, uncertainty and sensitivity analysis would be recommended for any follow-up studies.28 The existing legislation in place for drinking water reduce the potential impacts from water pollution on human health to very low levels. Note, however, that pollution can potentially increase the costs or degree of water treatment needed.

We have also considered emerging impact pathway values and damage costs from the MethodEx study for valuation of nitrate and phosphate pollution (based on WTP studies to avoid eutrophication).

For trace water pollutants, we have again adopted a screening analysis with the LCI data, applying a damage cost of £1 million or £100,000 per tonne to all ‘water burdens’, to see if any additional emissions are significant relative to the main pollutants. As highlighted above, it would also be possible to screen other water emissions (for health toxicity) using the life cycle impact assessment toxicity results and damage costs.

Of course, the actual impacts of emissions to water vary enormously according to the receiving environment, and exposure pathways, which are very site specific. It is this issue that complicates the impact pathway approach application, and it also raises the issue of uncertainty and lower confidence in the LCA values.

A large number of the emissions to water are generated from the waste disposal pathways (especially for landfill). Recent studies have concluded that it is not currently possible to quantify the external costs of leachate pollution from landfill sites in the UK, due to a lack of relevant data (e.g. Enviros and Eftec (2004)29.

4.3.5. Emissions to the Marine Environment and Marine Toxicity

The emissions to the marine environment involve a different set of pathways to freshwater above. These are less focused on the potential health effects of pollution, and more focused on marine ecosystems. The area is generally less extensively covered in the analysis of externalities, though there has been some work on the potential arising from offshore oil extraction (including drill cuttings, drilling muds, and chemicals) and oil spills (e.g. ExternE, 1995b; Watkiss and Lewis, 1995) and work on ant-fouling agents. It has not been possible to apply damage cost estimates for this study, and so sensitivity analysis has been used here to examine the potential importance of these burdens.

4.3.6. Emissions to soil

Emissions to soil are more difficult to capture through an impact pathway approach, and have often been valued using abatement costs associated with land remediation (e.g. Markandya et al, 2006). Sensitivity analysis has been used here to examine the potential importance of these burdens.

4.3.7. Ozone layer depletion

This area has not been considered because none of the material life cycles have significant ozone depleting substances involved (see the results of the three impact assessments in the previous chapter).

29 A COWI (2000) study recommended that if it was essential to use a value, a figure which is equivalent to around £0.5 per tonne of waste could be used.

4.3.8. Resource Depletion (Abiotic depletion)

For externalities, the area of resource depletion is different to the other areas above. The burdens from emissions to air, water and soil lead to environmental impacts and environment costs (externalities)30. Resource depletion involves a different concept, and as indicated by the initial peer review, resource depletion is a category where there is a user cost / scarcity rent already charged (though there is an important issue whether the extent the observed path of resource prices can be interpreted as an indication of the change in resource scarcity over time, see Markandya et al, 2005). We have therefore not considered valuation here. However, we highlight the wider issues of economic theories of efficient and sustainable resource depletion (for a discussion see Markandya et al, 2001 - the SAUNER project). This is an area that is highly relevant for the Sustainable Consumption and Production research programme, but warrants separate study.

4.3.9. Amenity

Amenity effects are not captured in the LCA, but they are important for waste disposal, i.e. the local amenity dis-benefits from noise, nuisance, odour, litter and vehicles for landfill and incineration. We have used the Cambridge Econometrics report (Cambridge Econometrics in association with EFTEC and WRc, 2003) to value amenity impacts per tonne of waste for landfilling. This estimated the average reduction in house prices near each landfill site equated to an average of £1.52 to £2.18 per tonne of landfill waste (at 1995 prices updating by consumer inflation to 2002 prices, for an assumed average flow of 100 million tonnes pa for 28 years at a 6% discount rate). We have also used initial data from the MethodEx study to assess the potential amenity cost for incinerators31.

4.3.10. Accidents

There is also another potential group of effects that are excluded in LCA. This includes accidents associated with the resource flows.

Occupational (worker) health effects include accidents and occupational diseases. Occupational accidents involve an immediate direct physical impact on the worker, with an obvious relationship between cause and effect. Occupational diseases generally occur as a more or less delayed response to a continuous, often long-term, exposure of an external burden. The calculation of occupational health impacts is usually based on (historical) statistics.

30 Or at least only partially internalised (where regulation or taxes are already in place). 31 MethodEx is an ongoing 6th Framework Programme Study, investigating the externalities from the waste, extraction and agricultural sectors. It will report in December 2006. http://www.methodex.org

There is also an additional complication when analysing the ‘external costs’ of occupational effects, because of the possibility that occupational health impacts are internalised by the action of labour or insurance markets. For this reason we do not consider occupational effects further here (see ExternE, 1995).

The one residual area that is potentially important is public accidents, notably arising from transport activity. The quantification and valuation of traffic accidents has long been included in Government appraisal and there is current guidance32. With a more detailed LCA, it would be possible to estimate the increase in transport activity, and quantify and value these potential effects, though it has not been possible within this scoping study.

4.3.11. Land use and other effects on Ecosystems and Biodiversity

As highlighted above, the quantification and valuation of all burdens on ecosystems has been extremely problematic. While it is possible to apply environmental economic techniques to value specific areas, there are issues with the transferability of these values, and also problems deriving unit values for marginal changes in ecosystem quality or biodiversity from environmental burdens of air, water and soil pollution.

There is also the issue of land-use change, and the impact of economic activity across the resource flow (but particularly from extraction) changing the available habitat area. The LCI data does quantify the land area across the life-cycle, but it is difficult to translate this into values, not least because of the issue of the site-specific nature of land type (i.e. whether the land used has any natural resource value, and if so, what specific type).

Within this scoping study, we have not attempted to assess the additional effects on ecosystems (as outlined above), but highlight that this will be important in future detailed work.

4.4. Specific issues on using the LCI data

A number of other issues have been identified which are relevant when using the LCI data, in relation to the use for estimating externalities.

Time specific issues. Discounting: LCA studies aggregate impacts across time (i.e. it uses static

weighting and normalisation factors). For example, the burdens that arise from the production and use of timber in housing occur now, whilst the burdens from waste disposal will occur in the future in 40 years or so at the end of the material life, but both are added together in the LCA. This is a problem for valuation because of discounting, though this is not an issue for LCA as it is not an economically driven approach. The impact pathway approach (as an economic method) specifically considers the time dimension, so that future impacts can be discounted.

32 http://www.webtag.org.uk/

This is an important for this study, and we have considered the effects that this can have in the analysis. It is recommended that a time dimension is considered for future studies using LCI to estimate externalities.

Dynamic values. Related to the point above, the importance of some burdens change over time. To illustrate, the Government SCC values increase over time because of the time profile of climate change. This leads to different SCC values being used in different time periods, i.e. it involves dynamic damage cost values that change over time. The same mechanism is used for the analysis of future air quality and health impacts (see IGCB, 2006). Policy appraisal often uses different values in different time periods, and so knowledge is needed on when burdens occur. It is recommended that the potential use of dynamic values is considered for future studies using LCI for externalities.

Country of origin.

UK vs. non UK impacts. The impacts from some burdens are not local – notably air pollution which is trans-boundary. LCA does not differentiate the burdens that apply within the country of origin, and overseas. Externality studies – especially in relation to policy applications – usually separate UK only effects33.

Policy coverage. The country of origin is also relevant for the study here, due to the degree of policy coverage. The level of policy coverage will vary according to country. Again, we recommend that the potential dis-aggregation of LCI burdens by country are considered for future studies using LCI for externalities.

The following sections apply the damage cost estimates to the LCI data for each of the three materials.

4.5. PET

The analysis of PET has focused on two specific stages:

The life cycle burdens to produce 1 kg of PET as plastic bottles, and deliver those to the consumer.

The burdens arising from disposal of the plastic bottles, looking at landfill, incineration and recycling.

For the first phase (up to production of PET bottles and delivery to the consumer), there are 625 burdens identified in the LCI: 189 of these are associated with raw materials; 206 are air emission burdens; 168 are water emissions; 8 are waste burdens; 54 are soil burdens.

33 even though this is counter to most international policy agreements and is also inconsistent with the social cost of carbon approach (see Watkiss et al, 2005).

Up to the stage of delivery to the consumer, the LCA (see previous chapter) scored ‘Abiotic depletion’ highest, due to the use of crude oil as a feedstock and (together with other fossil fuels) as an energy source in production and in electricity generation. As outlined above, we have not considered resource depletion here. Global warming and acidification also scored high, reflecting energy used in the PET manufacture and the electricity used in the bottle blowing process. We have focused on the externalities associated with these two processes, concentrating on the global warming impact, human toxicity and acidification effects.

A consideration of the LCI data reveals the following ‘air’ burdens as most important by weight: Carbon dioxide (CO); Carbon monoxide (CO2) Methane, fossil (CH4) Nitrogen oxides (NOX) NMVOC, non-methane volatile organic compounds, unspecified origin (NMVOC) Particulates, > 10 um (PM10) Sulphur dioxide (SO2)

The burdens of these pollutants to air, to produce 1 kg of PET as a product, are shown below.

Table 4.2. Burdens (in kg emitted to air) for 1 kg of PET material in the product (note equivalent to tonnes of burden per tonne of product)

CO2, fossil Methane NOx NMVOC PM10 SO2

Total kg 4.43 0.0099 0.0115 0.0047 0.001 0.0123

Cardboard (PET bottle manufacture) 0.0001 0

HDPE (PET bottle manufacture) 0.0001 0.0002

Production of PET granules for PET manufacture

2.51 0.0059 0.007 0.0026 0.001 0.0079

Electricity used in bottle manufacture 1.92 0.004 0.0035 0.0002 0.0004 0.0044

Transport PET granules to bottle manufacturer 0 0

Transport (bottle to retailer) 0 0

Transport (retailer to consumer) 0.0008 0.0018

Using the damage cost estimates (see earlier section) for greenhouse gases and health/non-health impacts of air pollutants, the externalities have been calculated for the production of a tonne of PET in the product. These are shown below, split by pollutant and life cycle stage. They confirm that the highest externalities are associated with the energy use for PET production and the electricity used on bottle manufacture (as indicated by the LCA). All other stages are low in comparison.

The key externalities are primarily associated with CO2 emissions, though NOx emissions are also important, and to a lesser extent, SO2.

Table 4.3. Externalities (£) for 1 tonne of PET material in the product

CO2, fossil CH4 NOxNMVO

C PM10 SO2

TOTAL

Total £ 85 4 33 3 5 14 141

Cardboard (PET bottle manufacture) 0 0 0 0 0 0 0

HDPE (PET bottle manufacture) 0 0 0 0 0 0 0

Production of PET granules 48 2 20 2 4 5 79

Electricity used in bottle manufacture 37 2 10 0 1 9 58

Transport PET granules to bottle manufacturer 0 0 0 0 0 0 0

Transport (bottle to retailer) 0 0 0 0 0 0 0

Transport (retailer to consumer) 0 0 2 1 0 0 3

Notes:CO2: current SCC value £70/tCCO, fossil value from Air Quality EvaluationMethane, SCC value, adjusted for GWPNOx – AQSR value, 6% no lagNMVOC – AQSR valuesPM, < 2.5 um - AQSR value, 6% no lagPM, > 10 umSO2- AQSR value, 6% no lag

The values are surprisingly large at £141 for every tonne of PET. Note, however, that 1 tonne of PET would produce around 25,000 plastic bottles, so the unit externality per bottle is extremely low. The value is also consistent with the embodied energy values in plastics which are two to three times more energy intensive per kg than steel (see later case study).

To check the potential gaps in the analysis, a number of additional sensitivities have been undertaken.

Firstly, the damage costs for key trace pollutants have been applied (e.g. heavy metals and key organic toxics). This has shown that none of these key trace pollutants are significant (in this case that they have externalities within two orders of magnitude of the total value above, i.e. above £1/tPET). This is not surprising, as due to the nature of these pollutants, they have been subject to tight regulations, reducing any residual impacts.

The analysis has also gone through the inventory to see if there are any other potential pollutants that could be important.

To do this, we have undertaken a screening analysis using hypothetical damage costs of £1 million, £100,000, and £10,000 per tonne, and applied this to all the emissions in the inventory. This highlights whether any other air emissions are potentially of concern. With the use of the damage cost of £10,000, none of the other air emissions are significant. With a damage cost of £100,000, the following are potentially significant in relation to the values above:

Acetic acid Aluminum Ammonia Benzene Boron Butane Ethane Ethene Hexane Hydrocarbons, aliphatic, alkanes,

unspecified

Hydrocarbons, aromatic Hydrogen chloride Hydrogen fluoride Methanol Pentane Potassium Propane Silicon Xylene

However, looking at this list, the only pollutants likely to have damage costs of £100,000 per tonne, would be highly toxic carcinogens. This could include aromatic hydrocarbons, though we stress that even benzene only has estimated damage costs around £1000 per tonne (source: Air Quality Evaluation – Watkiss et al, 2005). The inventory shows that the aromatic hydrocarbons are associated with the production of PET granules. Interestingly, the analysis did not pick up nickel as significant (though this had been identified as the major driver in the LCIA score for freshwater toxicity), nor mercury (important in the LCIA terrestrial ecotoxicity score).

A similar approach was adopted for water emissions. The analysis tested the potential importance of major emissions to water (e.g. N and P), but as the loads of these are relatively low, they are not important when externality damage costs are applied. The analysis has also tested the importance of key trace pollutants (e.g. heavy metals) across the life cycle. This shows that these are insignificant in relation to air emission above (below two orders of magnitude). This result differs from the LCA, which indicates a number of trace metals (e.g. vanadium) are potentially important. Finally, the screening analysis with indicative damage costs (e.g. £ 1 million, £100,000, £10,000) does not identify additional trace pollutants, and testing larger flows (e.g. biological oxygen demand) with realistic damage costs shows that these are not significant compared to the main analysis above.

There is, however, a potential issue over whether the externality assessment captures all the major burdens associated with extraction of fossil fuels (associated with oil extraction for PET feedstock, and all fossil fuels for energy and electricity generation) for example in relation to the marine environment. This is highlighted as an area for future investigation, and is relevant in respect of the policy coverage. We also raise the issue of sustainable resource consumption, for subsequent consideration against the policy coverage34. 34 There is a growing literature on the economic theory of efficient resource use and sustainable use of resources. This addresses the likelihood that desirable paths of economic growth, and the non-renewable resource depletion paths that support these growth paths, may require the application of certain rules for sustainability, rather than simply a reliance on market forces. Note sustainable extraction of a non-renewable resource is, in practical terms, impossible, but there are proposed rules for sustainability and natural resource depletion, e.g. where

The analysis has also included the disposal of PET, investigating the burdens from the alternative options of landfill, incineration and recycling (note the problem of non-authorised disposal has not been covered here). The previous chapter showed that recycling reduces all environmental impacts significantly. The results indicated that landfill (and to a lesser extent incineration) significantly increased the score for aquatic toxicity.

The three waste disposal options are summarised below, showing the main pollutants from the LCI analysis for the main GHG and air quality pollutants.

Table 4.4. Burdens (in kg emitted to air) for 1 kg disposed (or tonnes per tonne)

CO2, fossil Methane NOx NMVOC PM10 SO2

Landfill +0.0264 +0.00216 +0.000169+0.000054

1 +0.000011+0.000021

7Incineration +1.36 -0.00133 -0.000598

+0.0000397 -0.00013 -0.0015

Recycling -2.23 -0.00546 -0.00525 -0.00194 -0.00096 -0.00678

Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits (for consistency with earlier sections).

The externalities are shown below - note these include both positive and negative externalities.

Table 4.5. Externalities for 1 tonne disposed

CO2, fossil

Methane NOx

NMVOC PM10 SO2

Amenity

Total

Landfill +0.5 +0.9 +0.5 +0.0 +0.0 +0.1 +3.0 +5.0Incineration +26.0 -0.5 -1.7 0.0 -0.4 -8.8

n.q.+14.5

Recycling -42.6 -2.2 -14.8 -1.3 -3.3 -40.0 -104.1

Notes:Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits (for consistency with earlier sections).CO2: current SCC value £70/tCCO, fossil value from Air Quality EvaluationMethane, SCC value, adjusted for GWPNOx – AQSR value, 6% no lagNMVOC – AQSR valuesPM, < 2.5 um - AQSR value, 6% no lagPM, > 10 umSO2- AQSR value, 6% no lagAmenity: central estimate of £3/tonne waste (Enviros, 2004, in £, 2003).n.q. not quantified

The results show the positive benefits of recycling over the other disposal options35. Landfilling leads to dis-benefits, though the impacts are relatively low, as plastic does not generate methane (as it does not decompose in situ in the landfill).

resource rents (the decrease in value of a natural resource stock due to depletion) are invested in alternative capital stocks to maintain the overall value of an economy’s capital (Markandya et al, 2001).35 the beneficial impact of recycling may be slightly overstated as the collection (and any sorting) required has not been modelled.

Incineration leads to the largest net dis-benefit. This mostly arises from the CO2 and other air pollutants that are emitted when the plastic is burnt. Even though there are benefits from energy recovery from incineration, these are not sufficient to offset the negative effects. This issue is discussed again in the policy coverage section36.

Two additional points are raised on the values above. Firstly, the amenity dis-benefits of incineration are not included. Some of the literature implies much higher amenity impacts per tonne incinerated than landfilled (e.g. Enviros, 2004 and Eunomia/ECOTEC 2002) reflecting the urban location of most incinerators. Secondly, the damage costs values we have used do not take location into account (i.e. they assume equal values for all emissions in all areas). As most incinerators are located in, or near, urban areas, they have potentially higher impacts in relation to some primary pollutants. Both of these points would further raise the negative values for incineration as a disposal option.

This does raise an issue with the current waste hierarchy, at least with respect to incineration of plastics as a suitable waste disposal option.

It is also highlighted that the assumption about the displaced electricity is important (i.e. whether the electricity from incineration displaces the average mix, the marginal plant, etc). Previous work (e.g. ExternE, 2005) has shown that very large changes can occur in the externalities of incineration, according to the choice of generating technology assumed for displacement, both now and in future years. This issue should be explored as a priority in future work, and would also apply to the assumptions about landfill and energy recovery for options that involved degradable waste.

The analysis has also tested the LCI data for key trace pollutants (e.g. heavy metals, toxic organics) and screened using indicative damage costs against all possible trace emissions. This has not shown additional pollutants of concern for landfill or incineration, relative to the values above. Indeed, it shows that for some of the potential problem pollutants (e.g. dioxins from incinerators, mercury, etc) the actual significance of these pollutants is actually low in contrast to public perception.

The LCA has also used an alternative impact assessment approach – the Eco-Indicator99 IA methodology. This method matches the damage cost analysis above. Fossil fuel use (in the manufacture of PET) is the most significant impact using this methodology. The next two most significant impacts are the impacts on human health which result from respiratory inorganics and climate change. The main contributors to these impacts are emissions of nitrogen oxides (NOx), sulphur dioxide (SO2) and particulates (for respiratory organics) and CO2 (climate change), from fossil fuel combustion, firstly to provide heat in the production of PET and its constituent chemicals (e.g. xylene) and secondly to generate electricity for PET blowing. The use of this alternative approach does show that the choice of impact assessment method in the LCA can influence the apparent importance of different burdens. For any future work, we therefore recommend the use of alternative assessment methods.

36 While this shows that burning plastic is a less efficient way of generating electricity than conventional generation, the process does recover some useful resource (energy) from the waste material – so the relevant comparison is against other waste streams.

4.6. Iron and Steel

The same approach as above has been applied to calculate the potential externalities associated with steel production and recycling.

Steel production is known to be an energy intensive process, and so global warming and acidification impacts (which are primarily associated with pollutants from fossil fuel combustion) score highly with the CML methodology. The CML method also indicates a ‘high’ score for fresh water aquatic toxicity, due to nickel and cobalt emissions to water from the disposal to landfill of smelter slags from ferroalloy production (particularly ferronickel) and the disposal of slag from the electric arc furnace. Disposal to landfill of slags from ferroalloys (particularly ferrochromium) is also the main contributor to human toxicity (through emissions of chromium VI). In the case of terrestrial ecotoxicity, emissions to air of mercury from the electric arc furnace process are the main contributor. The burdens to produce 1 kg of Iron and steel are shown below.

Table 4.6. Burdens (in kg emitted to air) for 1 kg of Iron and Steel in the product (or tonnes per tonne)

CO2 Methan

e NOxNMVOC

s PM10 SO2

Total 1.793 0.00344 0.00822 0.00091 0.00796 0.00751Production of steel in O2 Furnace 0.083 0.00000 0.00001 0.00000 0.00005 0.00000Production Pig Iron in Blast Furnace 1.060 0.00218 0.00536 0.00046 0.00460 0.00518Ferronickel production 0.454 0.00098 0.00096 0.00010 0.00062 0.00173Ferrochromium production 0.000 0.00006 0.00019 0.00004 0.00060 0.00016Ferromanganese production 0.000 0.00000 0.00000 0.00003 0.00020 0.00000Molybdenite production 0.045 0.00008 0.00139 0.00025 0.00177 0.00017Quicklime production 0.045 0.00000 0.00000 0.00002 0.00000 0.00000Liquid oxygen production 0.000 0.00004 0.00000 0.00000 0.00001 0.00011Disposal oxygen furnace wastes 0.000 0.00000 0.00000 0.00000 0.00000 0.00000

Table 4.7. Externalities (£) for 1 tonne of Iron and Steel material in the product

CO2 Methan

e NOxNMVOC

s PM10 SO2 Total

Total34.2

3 1.38 23.23 0.61 27.28 44.28 131.01Production of steel 1.59 0.00 0.04 0.00 0.18 0.00 1.81

Production of Pig Iron 20.2

4 0.87 15.14 0.31 15.75 30.53 82.86Ferronickel production 8.67 0.39 2.71 0.07 2.11 10.18 24.14Ferrochromium production 0.00 0.03 0.54 0.02 2.06 0.96 3.61Ferromanganese production 0.00 0.00 0.00 0.02 0.67 0.00 0.69Molybdenite production 0.86 0.03 3.92 0.17 6.07 0.99 12.03Quicklime production 0.86 0.00 0.00 0.01 0.00 0.00 0.87Liquid oxygen production 0.00 0.02 0.00 0.00 0.02 0.65 0.70Disposal O2 furnace 0.00 0.00 0.00 0.00 0.00 0.00 0.00

wastes

Note ferrochromium and ferromanganese production both occur overseas. The damage costs are dominated by production of pig iron, and by greenhouse gas emissions, and the effects of PM and SO2 emissions. This matches the findings of the LCA.

The analysis has also assessed the potential trace burdens (e.g. nickel, chromium VI, mercury) with indicative damage costs. This does show some potential issues for chromium (IV) with externalities that are around £5/tproduct. For other trace metals (nickel, arsenic, lead, cadmium) the estimated externalities are lower (£0.2 to 2/tproduct). Whilst these values are lower than the main impacts of GHG and energy related air quality above, they are potentially important, and reinforce the LCA findings. However, some care must be taken in interpreting these results for subsequent policy analysis. The compounds are of potential concern because of their persistence and their potential to accumulate through the environment. However, in this case, the burdens all arise from disposal to landfill and subsequent emissions. The burden data on emissions does not differentiate the time of release (as outlined in an earlier section). The LCI assumes full release of these burdens over long time scales (hundreds of years) rather than on current annual emissions, which would in fact be very low for fully lined landfill sites with leachate collection.

The same approach as for PET has been used to screen for other potential omissions from air and water emissions. This has shown no significant additional burdens.

The values for car production are shown below. These are additional to the values above to reflect the total externalities associated with one tonne of material in the product. It is interesting to note however, that car production makes a significant contribution compared to the iron and steel production.

CO2 Methan

e NOxNMVOC

s PM10 SO2 TotalEmissions 1.74 0.0038 0.0044 0.0004 0.0030

Externalities 33.3 0.0 10.7 3.0 0.0 17.8 64.8

The recycling data are shown below. This shows high benefits per tonne of steel recycled (around £50/t)

Table 4.8. Burdens (in kg emitted to air) for 1 kg prompt scrap recycling and end of life shredding and recycling (or tonnes per tonne), and externalities

CO2 Methan

e NOxNMVOC

s PM10 SO2 Total

Emissions-

0.393 -0.0008 -0.0025 -0.0004 -0.0062 -0.0021

Externalities -7.5 -0.3 -7.1 -0.3 -21.2 -12.2 -48.6

Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits.

There is a time dimension that warrants consideration on disposal. This relates to the time between vehicle manufacture and recycling – based around the typical life of a vehicle (e.g. 10 years). There are issues related to discounting future benefits from recycling, and also considering the use of different values for different time periods (especially in relation to carbon). This is discussed in more detail for wood (below).4.7. Wood

Finally, a similar analysis has been made for material flow and disposal for wood.

The LCA identifies a significant benefit in global warming due to the carbon dioxide which is sequestered in the wood. Production of wood is non-energy intensive (compared to other construction materials such as steel), and therefore the contribution of transport to impacts such as acidification and eutrophication is relatively high.

One issue that is different for wood is that 39% of the total timber that goes into the construction processes is lost as waste, and goes immediately for recycling, incineration or landfill. Another issue is that the time period between production and disposal (from buildings) is very long, e.g. 40 years or so. When the wood comes out of the building on demolition, it is assumed to go down the same three routes; but after a significant time delay.

The main burdens are shown below. The analysis has assessed timber sourced from Scandinavia and the UK. The values exclude the waste disposal of waste wood, and the CO2 embodied in the wood. The values take account of the wood waste.

Table 4.9. Burdens (in kg emitted to air) for 1 kg of Wood in the product (or tonnes per tonne)

CO2 Methane NOx NMVOCs PM10 SO2

Total -2.77Sequestration of Cin wood -3.15Total excluding sequestration 0.381 0.00051 0.00293 0.00234 0.000481 0.000794Sawing of timber at building site 0.004 1.6E-05 9.6E-06 5.4E-04 1.0E-06 3.0E-07Transport UK timber (mill to building site) 0.029 2.7E-05 2.4E-04 5.0E-05 2.3E-05 6.3E-06Transport forest to sawmill 0.018 1.7E-05 1.5E-04 3.1E-05 1.4E-05 4.0E-06UK Electricity used for planing at mill 0.031 6.5E-05 5.8E-05 2.7E-06 6.7E-06 1.2E-06UK Electricity used For drying wood 0.029 6.0E-05 5.3E-05 2.5E-06 6.2E-06 1.1E-06UK Electricity at sawmill 0.040 8.2E-05 7.3E-05 3.5E-06 8.6E-06 1.5E-06Wood chips burnt for drying 0.008 1.0E-05 2.5E-04 2.9E-05 9.8E-05 2.7E-06Capital burdens of UK planing mill 0.007 1.1E-05 2.6E-05 5.5E-06 1.4E-05 9.7E-06Capital burdens of drying infrastructure 0.000 6.2E-07 9.1E-07 2.0E-07 3.7E-07 2.1E-07

Capital burdens of sawmill 0.007 1.1E-05 2.1E-05 4.3E-06 1.6E-05 1.1E-05Roundwood production in forest 0.056 5.4E-05 5.9E-04 8.2E-04 6.2E-05 7.0E-06

Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits (for consistency with earlier sections).The externalities are shown below.

Table 4.10. Externalities (£) for 1 tonne of Wood in the product

CO2 Methane NOx NMVOCs PM10 SO2 TotalTotal -52.86 0.21 8.28 1.56 1.65 4.68 -36.49Sequestration of Cin wood -60.14 -60.14Total excl. sequestration 7.27 0.205 8.283 1.556 1.647 4.679 23.64Sawing of timber at building site 0.08 0.006 0.027 0.362 0.004 0.002 0.49Transport UK timber (mill to building site) 0.55 0.011 0.678 0.033 0.078 0.037 1.39Transport forest to sawmill 0.35 0.007 0.426 0.021 0.049 0.023 0.87UK Electricity used for planing at mill 0.60 0.026 0.163 0.002 0.023 0.007 0.82UK Electricity used For drying wood 0.55 0.024 0.150 0.002 0.021 0.007 0.76UK Electricity at sawmill 0.76 0.033 0.207 0.002 0.029 0.009 1.04Wood chips burnt for drying 0.15 0.004 0.720 0.019 0.337 0.016 1.25Capital burdens of UK planing mill 0.14 0.004 0.073 0.004 0.050 0.057 0.33Capital burdens of drying infrastructure 0.01 0.000 0.003 0.000 0.001 0.001 0.01Capital burdens of sawmill 0.12 0.005 0.060 0.003 0.054 0.063 0.31Roundwood production in forest 1.06 0.022 1.660 0.545 0.211 0.041 3.54

Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits (for consistency with earlier sections).

The same approach has been adopted for other air and water emissions, testing for trace pollutants and screening with indicative damage costs. This has shown no significant additional burdens.

The externalities for the disposal phase have been considered for recycling, incineration and landfill. This proved difficult to set up correctly in the model, and so externality values are not shown below, but the analysis indicates that as with the LCIA analysis, there is a net benefit from landfill and incineration due to energy recovery from these processes which means that conventional fossil fuel generation is avoided.

There are some methodological issues for the waste disposal of wood, as these activities occur some 40 years or so in the future. Firstly, the carbon (and air quality) externalities should be adjusted to account for

the different social cost of carbon values in future years (using the £1/tC per year uplift) and increases in health valuation (a 2% annual uplift is generally applied – see IGCB, 2006). This would have the effect of increasing the values. For example, the illustrative central value for the social cost of carbon recommended for 2040 is £110/tC (rather than the value of £70 in the year 2000). If the carbon is released from the wood through incineration, this negates the benefits that occur from the product use (i.e. above), but should be valued using this higher value.

Secondly, there is a need to consider discounting. If a discount rate of 3.5% was applied to the results, they would be reduced by about 70%.

We highlight that neither of the impact assessment methodologies capture a number of environmental impacts, which can be associated with forestry. These include reduction of biological diversity and its associated values, water resources, soils, and unique and fragile ecosystems and landscapes. These are not assessed in the externality assessment here either.

4.8. Findings

The chapter outlines the approach for externality assessment. There is now a growing consensus on the approach for quantifying external costs, using a bottom-up assessment called the impact pathway approach. This is an impact driven approach, quantifying physical impacts, and then assessing the monetary valuation of these.

A review has highlighted the difference in objectives between life cycle analysis (options appraisal) and externalities (economic appraisal), summarised below.

For the analysis here, we have explored two broad approaches to use the information in the previous chapter, and to estimate the externalities for the three material life-cycles. These are:

1. Take the LC impact assessment methodology output (in terms of the traditional output of global warming potential, human toxicity, acidification, fresh water aquatic ecotoxicity, etc.) and convert these values to monetary estimates directly. The major advantage of using the LCIA output directly is the comprehensive and holistic coverage. The major dis-advantage is that the LCIA approaches are not consistent with UK specified impact assessment methods (e.g. for health) and do not link that usefully with the valuation endpoints. There are also contextual issues in applying valuation estimates across the impacts.

2. Take the LCI output (i.e. the inventory data on burdens), and apply damage cost values to these numbers. This is a more rigorous approach, and is more consistent towards answering the study objectives, but raises issues over coverage of impact categories.

Essentially there is a trade off between a high coverage of all potential pollutants with a lower level of confidence in (quantification and) valuation, versus a focus on a smaller number of more robust valuation estimates but with gaps.

Table 4.13. Comparison of LCA and Externality (impact pathway) approaches

LCA/LCIA Externalities (impact pathway)Analysis approach

Focus on average estimates (e.g. average of the sector)

Focus on marginal estimates (e.g. additional plant)

Quantification approach for effects

Historic focus on burdens (pressures, e.g. emissions), though most LCIA now incorporating impact driven approach

Focus on impacts, linking changes in state of environment with stock at risk and exposure response relationships

Aggregation Burdens are aggregated across steps and across impact categories

Impacts are kept separate, until final aggregation of monetary values

Location Historically no location specified, but many LCA now incorporating options for location specific inputs

Site-specific, in order to quantify physical impacts

Time period Focus on current burdens Usual focus on future policy and time-scales

Time dimension

Time of burden not specified Time of impact captured, to allow discounting of future versus current effects

Units Static single value for all periods Dynamic values, e.g. rising values for future time periods

Omissions Amenity, accidents, direct impacts on ecosystems / biodiversity

Major omissions for areas which difficult to value, notably ecosystems and biodiversity.

Coverage All burdens Screening of all burdens, followed by detailed analysis for priority impacts only

The approach we have considered is to use the LCI data and directly value these burdens using a simplified impact pathway approach with damage costs (e.g. cost per tonne estimates). This has a major advantage as damage costs are based on an underlying impact based approach, and for many burdens, it is possible to use accepted approaches for the quantification and valuation of impacts. As there are relatively few areas where Government recommended approaches for quantification and valuation are in place (at least compared to the hundreds of potential burdens identified in LCA), this has been complemented with the use of European estimates for important trace pollutants (heavy metals and toxic organics), and the use of a screening approach to filter the full inventory to identify any potential omissions.

The approach has therefore been: To start with dis-aggregated LCI data output; To use the LCA as a screening tool to identify the important steps and burdens; To select these priority burdens and assess in more detail using an impact

pathway approach (in this case a simplified IA approach) to estimate externalities, agreed impact and valuation approaches for UK policy.

To check the LCI data to see if there are any potential major burdens that have been overlooked.

We highlight that for any subsequent work, an alternative approach would be possible to investigate gaps. This would be to use the estimates of specific burdens (e.g. valuation of PM), and then use these values with the LCA impact assessment data on the relative importance to back-calculate all other burdens within each impact category.

The chapter has also reviewed the potential methods for quantifying and valuing burdens. It has then applied these to quantify the externalities for the three materials. The results are summarised below.

PET Steel WoodExternality per tonne of material in productExternality/tonne +141 +131 (steel

production) +65 (car production)= +196

-36.5

Key impacts CO2, NOX and SO2 from energy use and electricity

CO2, NOX SO2 and PM from energy use

Benefit from carbon sequestered in wood

Key omissions Marine toxicity from crude oil/gas extractionResource depletion (sustainable resource consumption)Ecosystems (generally)

Externalities associated with extraction.Some potential trace pollutants may be important (but issue time-scale)Resource depletionEcosystems (generally)

Ecosystems and biological diversity, water resources, soils.

Externality from waste disposal - recycling -104.1 -48.6 Not quantified - incineration +14.5 - landfill +5.0

Positive benefits of recycling. Landfilling leads to dis-benefits, though low as plastic not decompose. Incineration leads to

Positive benefits of recycling.

Benefits do not capture additional benefits from

Positive benefits for landfill and incineration due to energy recovery.

the largest net dis-benefit, due to burning of plastic (despite energy recovery)

avoided primary extraction

Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits. The values broadly reflect the findings of the LCIA, though there is less importance attached to trace pollutants (heavy metals and persistent organics) in the externality results.

The analysis has shown significant externalities arise from these three materials, though these appear to be dominated by greenhouse gas emissions and air quality pollutants, associated with energy/material manufacture. We highlight one associated issue with this point. The major burdens from energy use and electricity will change over time. This should be considered in setting up future life cycle analysis, for any subsequent analysis, to ensure results are applicable.

We also highlight a related point in respect of energy recovery. It is clear that the assumptions about waste disposal, e.g. energy recovery assumed, type of energy used in recycling, have a big impact on determining the externalities. This is one area where further consideration is needed for transferability, and further work is needed to improve the analysis.

5. Policy Coverage

5.1. Introduction

This chapter reviews the policy coverage against the burdens identified for the three material resource life cycles.

To address these different policy aspects, the policy coverage analysis has covered: Existing UK economic instruments for the different life cycle steps. Potential economic instruments in place in other countries. Other UK or EU policies which are targeted at environmental issues, e.g.

standards (and possibly voluntary agreements). Other policies which impact on the material, even though the main purpose is not

to internalise a given environmental externality (examples would include fuel taxes, energy taxes).

These are set out below, and include an assessment of the policy coverage of each of the three materials.

5.2. Input to policy development

There are a number of potential applications for the use of externality estimates across Government. These include:• Project appraisal (project cost-benefit analysis);• Regulatory Impact Assessment (policy cost-benefit analysis); • Input to policy development (including economic instruments);

For this study, it is the last of these that is most relevant (though it is also important to look at command and control legislation in determining the full policy coverage, especially where the latter has been justified by an economic cost-benefit analysis that includes assessment of these externalities).

As set out by Defra, ‘economic instruments are capable of influencing the behaviour of consumers and manufacturers in ways that are more subtle, yet potentially more powerful, than conventional regulatory controls, and which are capable of achieving results at lower cost’37.

In the UK, a number of economic instruments have been introduced to address environmental issues through successive budgets.

The underlying driver for these is explained in ‘Tax and the Environment’ (HMT, 2001). A summary of the key current policies that involve taxes, charges or subsidies are summarised below.

The landfill tax was initially based on an externality study (CSERGE et al, 1993). It encourages efforts to minimise the amount of waste generated and to develop

37 http://www.defra.gov.uk/environment/economics/index.htm

more sustainable waste management techniques. In line with the five-year escalator announced in 1999, the standard rate was increased by £1 per tonne to £15 per tonne on 1 April 2004. As announced in Budget 2003, this rate was increased by £3 per tonne in 2005-06, and by at least £3 per tonne in the following years to reach a medium- to long-term rate of £35 per tonne. The lower rate of landfill tax remains at £2 per tonne. The externalities from waste disposal were recently reviewed (Enviros et al, 2004). Budget 2006 announced that standard rate of landfill tax will increase by £3 per tonne to £21 per tonne from 1 April 2006. The lower rate, covering inactive or inert waste, will remain unchanged.

The aggregates tax was another economic instrument set on the basis of externality review, recognising that the extraction of aggregates imposes a range of environmental costs. Introduced in April 2002, the aggregates levy seeks to incorporate these costs into the price of virgin aggregate and encourages the use of alternative materials, such as wastes from construction and demolition that would otherwise be disposed of in landfill. It also promotes greater efficiency in the use of virgin aggregate. The levy was initially set at a rate of £1.60 per tonne of virgin aggregate commercially exploited in the UK. Budget 2004 froze this rate. Revenue from the aggregates levy is recycled to businesses through a cut in employer NICs, and through the aggregates levy sustainability fund, which supports projects seeking to minimise the impact of quarrying on local communities. Budget 2006 announced that that in 2006-07 the rate of the levy will continue to be frozen at £1.60a tonne.

The climate change levy is a 0.43 pence/kWh charge applied to industry, though the revenues generated are recycled through climate change agreements. It should be noted that the levy is an energy tax rather than a carbon tax (as it does not apply the same rate to carbon emissions, and instead varies so that carbon emissions from coal are taxed at a lower rate than gas). Budget 2006 announced am increase in the climate change levy, in line with inflation, from 1 April 2007;

The emissions trading scheme, now superseded by the EU Emission Trading Scheme (ETS) to reduce the EU emissions of greenhouse gases cost-effectively and meet its obligations under the United Nations Framework Convention on Climate Change and the Kyoto Protocol. The first phase of the proposed scheme is taking place between 2005 and 2007 across the enlarged European Union (EU25). The EU ETS works through Member States setting emissions allowances at a national level (national allocation plans) for each trading phase. After the first phase, the EU foresees 5 year trading phases. Grandfathering (i.e. free allocation on the basis of historic emissions) has been the predominant allocation method. Member states are allowed to use auctioning for up to 5% (first phase) and 10% (second phase) of the total allowances.

Duty level of fuel. There has been a series of duty differentials in the UK – starting with unleaded petrol in the early 1990s – and more recently for early introduction (in 2000 to 2001) of ultra-low sulphur fuel for petrol and diesel. This has been extended to alternative fuels.

The budget 2004 set a Government commitment to a three-year rolling guarantee on the fuel duty differentials for all alternative fuels. This include the 20 pence per litre duty incentive in favour of biodiesel that will be maintained until at least 2007 (recognising biofuels offer significant environmental benefits in terms of a reduction in the emissions of greenhouse gases and local air quality improvements). Budget 2006 announced further detail on the Renewable Transport Fuel Obligation to increase the use of biofuels – with the obligation set at 2.5 per cent in 2008-09 and 3.75 per cent in 2009-10, and the biofuels duty incentive maintained at 20 pence per litre in 2008-09. Note however, in the 2003 Pre-Budget Report, the Government announced a gradual reduction of the duty differential in favour of liquefied petroleum gas (LPG) over the next three years towards a level commensurate with the environmental benefits of the fuel. Budget 2004 confirmed that the differential for LPG will be reduced by 1 penny per litre in 2004-05, and by a further 1 penny per litre in both 2005-06 and 2006-07. The differential for natural gas (NG) will remain at its current rate, which is equivalent to 41 pence per litre, until 2007. Budget 2004 increased the duty on red diesel and fuel oil by 2.42 pence per litre from 1 September 2004. This reduces the differential between rebated oils and the main road fuels by 1 penny per litre to 40.88 pence per litre38.

Vehicle excise duty and company car taxation. Vehicle excise duty (VED) for cars first registered from March 2001 is based on the car’s carbon dioxide emissions and fuel type. This has offered motorists the opportunity to reduce their VED by up to £95 a year by choosing less polluting cars. The Government continues to evaluate the carbon dioxide-based company car tax system, introduced in April 2002, to gauge its impact on consumer behaviour. Following three years of increases in the rates, the level of emissions qualifying for the minimum charge in 2006-7 will be frozen at 140 grams per kilometre. Budget 2006 announced reforms to vehicle excise duty (VED) to sharpen environmental incentives including reducing the rate to zero for cars with the very lowest carbon emissions and introducing a new top band for the most polluting new cars.

The fuel duty escalator. This was introduced in 1993, and was an annual increase in the real level of motor fuel duty by 3% per annum. The escalator was increased to 5% in November 1993. The July 1997 budget included a commitment to annual increases of 6% in real terms in the duty on road fuels, except road fuel gases. This was ended in 1999. This escalator has some residual effects on the price of fuel in the UK, though this is difficult to disentangle (on its introduction it added three pence to a litre of fuel, and by 1997, when the escalator was at 5%, it contributed a 11.1 pence rise to the cost of unleaded fuel).

The Renewables obligation. Introduced in 2002. This requires all licensed electricity suppliers in England & Wales to supply a specified and growing proportion of their electricity sales from a choice of eligible renewable sources (DTI, 2003). Suppliers are responsible for demonstrating that compliance to OFGEM through a system of Renewables Obligation Certificates (ROCs).

38 This follows the 2003, HM Customs and Excise consultation document, Duty differentials for more environmentally friendly rebated oils, which examines whether preferential duty rates for rebated oils with a lower sulphur content would deliver worthwhile environmental benefits.

The Obligation is the key policy mechanism by which the Government is encouraging the growth necessary to reach the UK's renewable energy targets - by 2010, 10% of the UK's electricity should be supplied from renewable sources .

The Government has also considered a number of areas for potential economic instruments39, these include: Economic instruments for water pollution charges; Economic instruments for water pollution; Economic instruments in relation to water abstraction; Design of a tax or charge scheme for pesticides; Economic instruments for Aviation (Aviation and the Environment, using

economic instruments, 2003); Economic instruments for diffuse water pollution (associated with agriculture).

Most of the above instruments apply to all three material life cycles, as they affect the generic activities of energy/electricity, transport and waste disposal.

5.3. Policy Development Outside the UK

A large number of economic instruments are in place internationally. These are collated in a OECD database of environmental taxes40 which has been used here to investigate other potential economic instruments in place, i.e. which could apply to the 3 material fuel cycles.

5.3.1. CO2 taxes

A large number of countries have some form of energy tax or a specific CO2 tax. As similar taxes exist in the UK, these are of less interest per se, but are of interest in respect to the relevant level – to compare to the UK. It is stressed that not all of these declared CO2 taxes are direct and unambiguous CO2 taxes - in many cases the level of the tax is not set in an attempt to reflect each tonne of CO2 emissions. Direct CO2 taxes have been implemented in Norway, Denmark, Finland and Sweden. The systems in place in Finland and Norway are complicated in that the tax is only charged for a limited number of sectors (significantly not including the ESI). Most of these taxes use fuel as the tax base – using this as a proxy for CO2 emissions – based upon a stated value of CO2 emissions. The highest value for CO2 is that implicit in the Swedish tax system. Finland and Sweden have CO2 taxes that replace previous general energy taxes and so have high levels of revenue in comparison to other taxes (whereas other taxes are additional to existing energy taxes).

5.3.2. SO2 and NOx taxes

There are also examples of specific taxes on air pollutants. While the UK has used duty differentials to introduce low sulphur fuels, it does not have a system of emission taxes.

39 See http://www.defra.gov.uk/environment/economics/index.htm40 www.oecd.org/EN/document/0,,EN-document-8-nodirectorate-no-1-3016-8,00.html

Other countries have specific taxes in place. For SO2, these are often set as direct pollution taxes – i.e. they are set per unit of SO2 emissions – usually per tonne emissions. However “general” sulphur taxes are also set by means of variation in the taxation of fuels, according to the sulphur content (usually by means of bands of fuels – over and under 0.5% sulphur, for example, rather than by means of specific sulphur content). In Europe, a number of countries have introduced air emissions charges or taxes for SO2. Norway, Sweden, France, Italy and Denmark tax SO2 emissions (or the sulphur content of the fuel). The Czech Republic also have emission-based charges or fees. Denmark and Italy have a direct SO2 tax, charged per tonne SO2 with values of Euro 1.3/tSO2 (Denmark) and Euro 53/tSO2 (Italy). Similarly Australia (New South Wales) has a load based licensing pollution charge of Euro 2.5/tSO2 (below a threshold), rising to Euro 5/tSO2 (for emissions above a threshold). Finally, the Czech republic has a charge of 29 Euro/tSO2.

NOx taxes have been implemented in France, Italy, Sweden, Australia and the Czech republic. The values are summarised in the Table below. All of these are direct pollution taxes charged per tonne or kilogramme of NOx, but some have varied rates for different areas and different levels of emissions (i.e. the sophisticated Australian system). Many of the taxes implemented are limited to specific types of plant, most commonly to large combustion plants. The Swedish NOx tax (introduced in 1992) is noticeably higher (at Euro 4329/tonne NOx) than anywhere else – the next highest tax (in Italy) being around Euro 100/tNOx. The tax is revenue neutral with plants receiving rebates in proportion to the final production of energy (plants with high emission relative to energy output are net payers). However, this has meant that certain industrial sectors with a substantial capacity to reduce emissions gain from the tax industry are net winners, and sectors such as the pulp and paper industry, which have little potential to reduce emissions, are net losers. The French NOx tax has been implemented recently – the revenues are used to subsidise abatement equipment.

5.3.3. Other Atmospheric pollutants, waste and water

There are some taxes for other atmospheric pollutants in place, though these are not so common. Switzerland has a direct charge on volatile organic compounds (VOCs), the Czech Republic has charges on particulates, carbon monoxide (CO), hydrocarbons (HCs) and “Other polluting substances”, and the load-based licensing scheme in NSW, Australia has rates for an extensive variety of atmospheric pollutants. France has a tax system similar to SO2 and NOx above for NMHC/VOC. A few countries, including the Czech Republic and Denmark have charges on Ozone depleting substances (such as CFCs).

Taxes on waste and water abstraction are more common. However, not all of the latter will have a significant effect upon the power-generating sector. Waste taxes have been documented elsewhere (e.g. OECD) and the UK has led the way in the implementation of the landfill tax.

Note, some waste and water taxes have been excluded from the list where they would not impact on the ESI, for example, taxes on domestic consumers and taxes on defined types of waste from specialised industries.

Table 5.1 Other Taxes Relevant to the Material Cycles.

Country Particulates tax VOC/HC emission tax

Waste tax Water abstraction/ consumption tax

Trace elements tax (ash constituents)

EuropeAustria No No Yes Yes NoBelgium No No No No NoDenmark No No Hazardous waste Yes NoFinland No No Landfill only Yes NoFrance No No Yes Yes NoGermany No No Yes Yes NoGreece No No Yes No NoNetherlands No No Yes No NoIreland No No No No NoItaly No No Municipal waste

onlyManagement of water resources

No

Luxembourg No No No No NoNorway No Emission permit

application feeHazardous waste only

No No

Portugal No No No No NoSpain No No No No NoSweden No No Landfill only No NoOther Australia Yes Yes Yes No Yes, NSW onlyCanada Quebec charge

on air pollutionQuebec charge on air pollution

No Quebec only No

Czech Rep. Yes Yes Yes No NoJapan No No No No NoNew Zealand No No No No NoSlovenia No No Yes Consumption tax

for industryNo

Switzerland No Yes No No NoUnited States No No Yes, vary by

stateYes, national water use charges

No

The only country of the above to operate a trace elements tax is Australia, as part of the New South Wales load-based licensing scheme. This imposes charges for emissions of a variety of trace elements to land, air and water.

Table 5.2. Particulate taxes.

Country Tax description Tax rateAustralia New South Wales load-based licensing

scheme – levels for coarse and fine particulates. Applies to stationary plant

Unit cost varies, depending on whether coarse or fine particulates, area-specific threshold, and level of emissions relative to this threshold

Canada Measured emissions to air – annual levels for several types of pollutants. Applies to stationary plant

CAD$2 (EUR 1.4) per ton

Czech Rep.

Tax on particulate emissions. Applies to stationary plant,

CZK 3000 (EUR 88.1) per tonne

Table 5.3 VOC/HC emission taxes.

Country Tax description VOC/HC tax rateAustralia New South Wales load-based

licensing scheme – stationary sources

Unit cost varies, depending on area-specific threshold and level of emissions relative to this threshold

Canada Measured emissions to air – annual levels for several types of pollutants. Applies to stationary plant

CAD$ 2 (EUR 1.4) per ton

Czech Rep. Applies to stationery plant CZK 2000 (EUR 58.8) per tonneSwitzerland Incentive tax on VOCs CHF 2 (EUR 1.32) per kg

Table5.4. Water abstraction/consumption taxes.

Country Water abstraction/ consumption tax

Tax description Water abstraction/ consumption tax rate

Austria Yes Based on volumetric water consumption

N/A

Denmark Yes Tax on consumption of water, all sectors.

DKK 5 (EUR 0.83) per m3

Finland Yes Tax on consumption of water, all sectors.

Total charge FIM 8.4 (EUR 1.28) per m3

France Yes Charge on water abstraction

Rates vary according to water origin and sensitivity of the area

Germany Yes Charge on water extraction DEM 0.01-0.4 (EUR 0.0051-0.2046) per m3

Italy Yes Management of water resources

Tariff formula decided by monitoring committee

Slovenia Yes Industrial water charge SIT 108.8 (EUR 0.34) per m3

United States Yes Metered water charges Vary (and only charged to 49% of users in 1995)

5.4. Other Policies

There are a number of other policies, which address the environmental impacts from the life cycles of the material and products considered.

5.4.1. Raw Material Extraction

PETThe raw materials for PET are crude oil and natural gas, which could be sourced from the UK or overseas. In the UK there are a number of policies regulating oil and gas extraction, including Offshore Petroleum Activities (Oil Pollution Prevention and Control) Regulations 2005, to control the discharge of oils, Offshore Chemical Regulations 2002 to control the use and discharge of chemicals and the Offshore Combustion Installations (Prevention and Control of Pollution) Regulations 2001 to control atmospheric emissions.

Iron and SteelAlmost all the main raw materials for steel production (coal, iron ore and ferro-alloys) are imported into the UK form a variety of countries (e.g. Australia, Brazil, South Africa, Venezuela and Canada for iron ore). Policies and regulations to control the environmental impacts of mining, disposal of wastes and reclamation of land are likely to vary significantly between countries.

WoodThere are a number of voluntary certification schemes for timber production. The main schemes relevant to the production of softwood are: Forest Stewardship Council (FSC) – international in scope; Programme for the Endorsement of Forest Certification Schemes (PEFC)

international in scope; Canadian Standards Association – applicable in Canada; Sustainable Forestry Initiative – applicable in US and Canada;

The FSC scheme sets performance based standards which must be met; the PEFC scheme endorses national forestry certification scheme. So for example in Scandinavia (the source for much softwood used in the UK), The PEFC scheme endorses the Finnish Forest Certification Scheme, the Forest Owners Associations (Family Forest Certification ) system in Sweden and the Living Forests Standards for Sustainable Forest Management in Norway.

The schemes all provide assurances that the timber is legally harvested, but assessments of the schemes (e.g. FERN, 2004) suggest that they vary in the standards they set for sustainable forest management.

5.4.2. Production of material

All of the production processes are subject to pollution control regulations. Iron and steel production and foundries, and PET production (including the production of its constituent chemicals from crude oil in e.g. refineries) are regulated by the Environment Agency under the Integrated Pollution Prevention and Control (IPPC) regime (previously Integrated Pollution Control (IPC). Operators are required to use Best Available Techniques (BAT) to limit emissions. BAT is defined in a series of guidance notes, and are used to set limits for emissions to all media in authorisations for individual plant.

The manufacture of timber is regulated by Local Authorities under the local air pollution control (LAPC) system established under Part I of the Environmental Protection Act 1990. The main pollutant of concern is particulates (dust) arising from wood machining.

5.4.3. Waste Management/End of Life

PETA significant proportion of PT bottles will end up in the household waste stream; some (discarded as litter) will end up the remainder are likely to end up in the commercial waste stream (e.g. through catering establishments). In the Waste Strategy 2000 the government set statutory performance standards for Local Authorities for household waste recycling and composting; these are set at 30% for 2005/6.

As packaging, PET bottles fall under the (revised) Packaging Directive (2004/12/EC) implemented in the UK as the Producer Responsibility Obligations (Packaging Waste) Regulations 2005. This sets mandatory recovery and recycling targets for packaging waste, which must be met by 31 December 2008. The material specific recycling rate for plastics is 22.5%.

Iron and Steel Wastes from iron and steel production, and wastes from vehicle production would be subject to normal waste regulation. Disposal of a vehicle at the end of its life is subject to the End of Life Vehicle Directive (implemented in England and Wales as the End of Life Vehicle Regulations 2003 (SI 2635)). By 2007 manufacturers of cars must take back vehicles free of charge. There are also targets for recycling and recovery of materials from vehicles.

TimberWood waste from construction and demolition sites would be subject to waste disposal regulations e.g. the use of registered waste carriers. As wood is not an inert material it would be subject to the biological component of the landfill tax. There are no specific targets or incentives for the recycling of construction and demolition wood waste.

LandfillingLandfills in England and Wales are regulated by the Environment Agency under the Landfill (England and Wales) Regulations. These implement the Landfill Directive (Council Directive 1999/31/EC), which aims to prevent, or to reduce as far as possible, the negative environmental effects of landfill. They set engineering and operating standards (e.g. leachate treatment and recovery of landfill gas. The directive also requires biodegradable waste to be progressively diverted away from landfill. This is implemented via the Waste and Emissions Trading Act (2003) which provides for the allocation of allowances, which may be tradable, to waste disposal authorities.

Waste incinerationIncineration of waste is subject to the Waste Incineration Directive (WID) through permits issued by the Environment Agency under the Pollution Prevention and Control Regime. Operating conditions, emission limit values for a range of substance to air and water, and emission monitoring requirements are all set out.

RecyclingWhile there are no specific policies on recycling apart from those mentioned for PET above, wood and plastics are two (of the six) materials which are a focus for WRAP (the Waste & Resources Action Programme). The aim of WRAP's Plastics Programme is to increase the capacity for plastic recycling in the UK, fund development of new applications for recycled plastics and raise awareness of the range and quality of recycled products already available. A specific aim is to provide support for the establishment of a 20,000 tonnes plastic bottle sorting/recycling facility; The priority of the wood programme is to stimulate demand for recycled wood products through a marketing and education initiative targeted at the construction industry and local government. This highlights the opportunities and benefits of using recycled products such as coloured mulches and extrusions, and is supported by a range of R&D activities. The programme also seeks to stimulate increased investment in the panel board industry to utilise greater amounts of recycled wood and is developing a web-based wood map, "RecycleWood", to match producers of waste wood with recycling services.

5.5. Internalisation

In assessing the economic instruments above, or other polices, it is useful to consider the degree of internalisation, i.e. whether the taxes in place cover the externalities or whether policy is optimal41.

Looking at the policy coverage of the three life-cycles, the review above shows that there are policies in place (in most cases with economic instruments) to address the externalities of the main burdens: energy/electricity use, transport and waste.

There appears less focus on some of the other areas identified, e.g. toxic trace pollutants, and resource extraction, and these are highlighted as potential gaps, worth further investigation.

The next question is whether the current levels of taxes or charges are sufficient (or whether policy is optimal). To answer this properly, a very detailed analysis would be needed. However, using information from previous studies, we make the following conclusions: Sansom et al (2001) have shown that a national aggregate level, the social costs

of transport are accounted for in taxes and charges (e.g. fuel duty and road tax). However, the study found that the pattern varied at a dis-aggregated level across the network, such that the social costs for urban and motorway driving, especially at peak times, were not fully internalised.

Watkiss et al (2006) has shown that the social cost of carbon (climate change) are not completely internalised in current policy. This can be shown with the levy of the climate change levy, which has a lower implicit tax per tonne of carbon than the Government SCC value for electricity, and a very much lower value for coal use.

41 For command and control legislation, whether the policy put in place leads to the optimal policy position where the marginal costs of abatement are equal to the marginal damage costs.

ExternE (2006) and Watkiss (2005) have shown that the external costs of electricity generation and energy and not fully internalised. As well as the issue of carbon (above) this arises from the external costs of the other air pollutants, i.e. PM10, NOX and SO2. Note the EC has indicated it might consider NOX and SO2 trading as part of future air quality policy42. There has been previous consideration of NOX and SO2 trading in the UK, with previous research in this area.

Enviros and Eftec (2004) investigated the potential externalities associated with waste disposal. This has been further studied within the MethodEx project. For landfill, the general finding is that the level of tax in place does cover most of the externalities – though this finding is heavily influenced by the assumed social cost of methane (i.e. a higher value for methane implies externalities are only partially internalised). However, these studies and others in the literature (e.g. ExternE, 2005; Eunomia/Ecotec, 2002) imply that the external costs of incineration are not fully captured, not least because there are no economic instruments in place. This position has reflected the recommended waste hierarchy, which has sought to use landfill as a last resort. The analysis in the previous chapter has shown that there are likely to be significant externalities from this technology (though interestingly these arise from conventional combustion pollutants, rather than the heavy metal and dioxin emissions that have dominated public perception of impacts).

Resource depletion is a category where a user cost / scarcity rent already charged. However, there is an important issue whether the observed resource prices can be interpreted as an indication of the change in resource scarcity over time (see Markandya et al, 2005). We highlight the wider issues of economic theories of efficient and sustainable resource depletion.

It is also worth considering a number of other potential policy questions: Whether the current practice of using instruments for selected processes or

individual burdens may miss some important environmental externalities up- or down-stream, and how large these may be.

Whether there is policy overlap that actually duplicates the correct pricing regime. Whether there are any perverse effects, where action to address one isolated

problem actually increases overall environmental burdens (and externalities) elsewhere in the system (e.g. where action to control one process leads to a change in practice that increases overall pollution levels elsewhere in the life cycle).

These questions are beyond this scoping study, but are highlighted for consideration in any follow-up work.

One related issue that is raised here is the international dimension. The material flows here involve import of primary material (and will also involve exports of products, as well as import of finished product).

42 Within the adopted Thematic Strategy on Air Pollution, the Commission has indicated it will pursue the option of introducing regional (including regional trans-boundary) emissions trading for NOX and SO2 when revising the NEC Directive in 2006. This would permit individual plant to trade emissions reductions that go beyond current LCPD limits.

In considering the policy coverage, and any overlap or duplication of policies, it is important to recognise the international nature of resource flows. Where a resource flow occurs entirely in the UK, then the appropriate policy would be to tackle the burdens directly with policy. However, this situation is rarely the case, and action to address specific burdens with across the resource cycle in one country may lead to differences between countries, for example, with different levels of environmental tax on resource extraction, and so different extraction prices. There are potential ways to address this issue. One might be to consider import tariffs, though this is a hugely complex (and political) issue43. An alternative is to introduce policy instruments downstream, e.g. for the aggregate burdens of extraction and primary processing. While this may address the issue of imports, it does not provide the necessary targeted incentives for reducing environmental burdens for extraction. A third alternative might be through the use of supply chain evaluation/procurement policy – where producers ask for some form of environmental declaration of impacts from suppliers of raw materials (though it is more difficult to see how this could be applied to some of the bulk raw materials and chemicals). We highlight that these international issues should be investigated for any subsequent detailed analysis.

5.6. Findings

The policy coverage by lifecycle stage for each of the three material/product uses is summarised in Table 5.5. In addition there are the common policies on transport and energy use.

Table 5.5 Summary of Policy Coverage

Lifecycle stage Wood in construction

Iron and steel in car

PET in drinks bottle

Raw material extraction

Forestry certification schemes

Various regulations on offshore extraction of oil

Material Production LAPC IPPC IPPC

Incorporation into product

LAPC

Disposal of product/material

Landfill Regulations and Landfill taxWaste incineration directive

End of life vehicle regulations

Packaging regulationsLandfill regulations and taxWaste Incineration Directive

It is clear that economic instruments in the UK have been focused on a number of discrete areas – notably energy/electricity use, transport and waste. However, the comparison with the LCA and externalities assessment shows these are the major burdens for the three materials.

43 For example, the possibility of border adjustment tariffs, on energy intensive products imported from countries without CO2 targets, has been raised in the climate change arena.

For some of the secondary impacts, there does appear to be less policy coverage. This includes toxic trace pollutants, material extraction, and resource depletion, and these are highlighted as potential gaps. A comparison of other policies in place suggests that other policies partially address these gaps, and extend the policy coverage to other parts of the life-cycle, but a number of gaps remain. The most important of these appears to be with raw material extraction. While there are certification schemes for forestry which can help to ensure that wood is produced sustainably, there is no equivalent policy for ensuring that environmental impacts from e.g. mining of raw materials for steel, are addressed adequately.

The analysis has also included a review of the degree to which externalities may be internalised by the existing policy coverage. For transport, the existing economic instruments seem to cover the potential

externalities (i.e. the degree of internalisation suggests prices reflect the appropriate signals).

For energy/electricity, it suggests that important externalities remain (i.e. prices do not fully reflect environmental burdens).

For the waste disposal options, the policy coverage for landfill seems broadly appropriate, but there are gaps with respect to incineration for the materials studied.

There is also a clear message that the benefits of recycling are not fully captured, indicating there must be remaining externalities upstream in the fuel cycles (including from extraction, use and disposal).

There appear to be potential gaps for resource extraction, particularly of non-renewable materials.

From this we suggest there might be a need for: Full policy coverage to address the total externalities of CO2 emissions, and the

conventional air pollutants (SO2, NOX and PM10). Market signals to recognise the benefits of recycling (i.e. to reflect avoided

externalities). The analysis suggests that recycling avoids significant externalities across the life cycle (i.e. from avoided extraction, processing and waste disposal).

Policies to tackle burdens with resource extraction, e.g. mining of raw materials, oil and gas extraction.

Policies to ensure optimal natural resource depletion (implying reliance on market forces is insufficient).

In addition, the results of this study suggest that not all the externalities of incinerating plastic in an energy from waste plant would be directly addressed by current policies. However, plastic would not normally be burned on its own, but as part of a mixed household/commercial waste stream. It would be useful to examine whether for the incineration of mixed wastes in energy from waste plant, there are externalities which are not addressed by polices (such as the Waste Incineration Directive and local planning regulation), which impact upon waste incineration facilities at the moment.

6. Recommendations and Research Gaps

6.1. Key Findings

This report presents the findings from the SCP study on the Policy Coverage of Environmental Impacts of Materials. The project aims were to: Investigate the environmental burdens associated with 3 materials at each stage

of their life cycle, including their use in products and final disposal; Assess the potential externalities associated with these burdens; Review the policies in place to address these burdens (the policy coverage),

considering both economic instruments and other forms of legislation.

The first phase of the study identified a large number of materials for consideration. These included primary materials (e.g. metals, aggregates, wood) and secondary materials and intermediate products (e.g. plastics, paper). Following an initial screening, five materials were chosen for a more detailed assessment. After further review, and consultation with the study steering group, the study selected the following three materials: Plastics, specifically PET (PolyEthylene Terephthalate). This material was

chosen to capture the growing use of plastic materials, and to investigate the emerging plastic recycling industry.

Iron and steel. This material was chosen as it is an energy intensive material, used in construction and products, and an existing recycling market exists.

Wood. This material was chosen because it is a renewable material and is a major construction material.

Existing life cycle data was found for the extraction and processing of all three materials (although not always for the UK situation), though there were less useful data found for each material product. However, one of the major findings of the study has been the existing LCA studies are not useful for subsequent policy externality analysis. This is because: The life-cycles of existing studies are set up to assess a particular problem or

question. They do not easily lend themselves to an examination of life cycle burdens for resource flows, per unit of material.

Data are often aggregated together across the life-cycle (or across several lifecycle stages). Furthermore, life cycle impact assessment methodologies are often used to characterise and weight emissions of individual burdens into a few categories to allow comparison of options. This makes it difficult to assess the effect of individual stages or burdens in detail, which is needed to examine subsequent externalities and policy coverage.

The LCA studies are aimed at answering a different question to externality analysis. LCAs are primarily undertaken as a comparative (relative) way of looking at different options, using burdens which are calculated on average effects.

In contrast, detailed externality assessment has a focus on the estimation of (absolute) impacts (not burdens), calculated based on marginal (not average) effects.

Most LCA studies do not contain the detail on the location of burdens. This information is important for externality studies, both in relation to UK versus non UK effects, and because externalities vary significantly between locations.

This finding is one of the main policy conclusions of the study. It suggests that any detailed analysis following this study would need dis-aggregated life cycle inventory data i.e. of emissions of individual pollutants, split by individual life cycle stage.

The study initially planned to use existing LCAs to identify burdens for the three material flows. Due to the reasons above, this was not possible. Instead, the study used an LCA tool (SimaPro) to model simplified life-cycles for each material. This included raw material extraction and processing, one main product use for each material (wood in construction, steel in a car and PET in a drinks bottle) and end of life management of the product. This allowed the output of disaggregated life cycle inventory (LCI) data, split by life cycle stage. Two recognised impact assessment methodologies were then applied to the inventory to identify the most important environmental impacts (as a screening tool).

Using this inventory burden data, the study scoped the externalities for the three material life cycles. Two potential approaches were considered. Firstly, to use life cycle impact assessment (LCIA) output and try and monetise this directly. Secondly, to use the LCI (inventory) output, and apply existing externality estimates using damage costs (simplified £ per tonne values). Regarding the two approaches, the study has found: It is difficult to apply valuation estimates directly to LCIA output, as most units do

not relate directly to the economic valuation estimates in the literature. There are also problems of consistency, because LCIA approaches do not follow standard UK guidance for impact assessment – an example being for health impact assessment of air pollutants. However, the use of LCIA data has a major advantage, as it has complete coverage across all life cycle steps and burdens.

The use of the LCI data and an impact pathway approach (or simplified impact approach using damage costs) is much more satisfactory in terms of policy consistency, and is methodologically more robust. It also allows a direct match between impacts and valuation endpoints. The main disadvantage is that there are not agreed approaches for analysis and valuation of all burdens from a LCI (though this problem also applies to the use of LCA output). In addition, applying this approach to all the burdens from an LCI is time consuming.

The two points reflect the trade-off between trying to value everything (with inconsistent methods and high uncertainty) versus only valuing a few areas robustly, but leaving potential gaps.

This study adopted the second approach, using LCI data on burdens. This provided the most robust and evidence based approach for subsequent externality analysis. Our approach has been: To produce dis-aggregated LCI data output;

To use the LCIA as a screening tool to identify important steps and burdens; To select the priority burdens and assess these in detail using an impact pathway

approach to estimate externalities (in this case a simplified approach using damage costs), using agreed impact and valuation approaches for UK policy.

To screen the LCI data to see if there are any potential major burdens that have been overlooked. We also highlight that it would be possible to use the priority burden estimates, in combination with the LC impact assessment data, to calculate the relative values for all burdens (i.e. by using the weighting factors in LCIA to weight the other burdens relative to the quantified estimates). This approach should be investigated in any subsequent studies.

The analysis revealed significant externalities for each of the three material flows. The externalities are dominated by greenhouse gas emissions and air quality pollutants arising from energy and electricity use, though this reflects the choice of materials (high embodied energy or energy intensive materials). For iron and steel, some trace pollutants (heavy metals) may be potentially significant. The analysis is summarised below in Table 6.1. Note transport was not identified as a major burden for any of the three material flows.

During this analysis, a number of specific methodological issues were identified for externality assessment using LCI data. These were: LCI data does not (generally) specify the location of burdens. It does not specify when burdens occur (there is no time dimension). Note this is

important for economic analysis and discounting. A number of potential externalities are not included in LCA frameworks (as LCA is

focussed on quantification of resource and emission flows). These include impacts on ecosystems and biodiversity, social impacts, accidents, and amenity.

A large proportion of burdens are associated with transport and electricity, and the burdens from these sectors change over time (e.g. LCI may be using current information to address a future policy question). The assumptions on the benefits of energy recovery from waste, or recycling are also dependant on energy mixes assumed, and the same issue applies.

The study assessed the policy coverage of the three life-cycles. These are summarised below in Table 6.2. In addition there are the common policies on transport and energy use.

The review has found that economic instruments in the UK have been focused on a number of discrete areas – energy/electricity use, transport and waste. However, given the findings above, this focus address the major burdens of the three materials assessed here.

The study also assessed whether these polices internalised the impacts. For landfill as a waste option, the existing economic instruments seem to cover the potential externalities (i.e. the degree of internalisation suggests prices reflect the appropriate signals).

For energy/electricity, and for incineration as a waste option, the analysis suggested that important externalities remain (i.e. prices do not fully reflect environmental burdens), particularly for CO2 emissions, and the conventional air pollutants (SO2, NOX and PM10).

Table 6.1. Summary of Externality Analysis

PET Steel WoodExternality per tonne of material in productExternality/tonne +141 +131 (steel

production) +65 (car production)= +196

-36.5

Key impacts CO2, NOX and SO2 from energy use and electricity

CO2, NOX SO2 and PM from energy use Energy use.

Benefit from carbon sequestered in wood

Key omissions Marine toxicity from crude oil/gas extractionResource depletion (sustainable resource consumption)Ecosystems (generally)

Externalities associated with extraction.Some potential trace pollutants may be important (but issue time-scale)Resource depletionEcosystems (generally)

Ecosystems and biological diversity, water resources, soils.

Externality from waste disposal - recycling -104.1 -48.6 Not quantified - incineration +14.5 - landfill +5.0

Positive benefits of recycling. Landfilling leads to dis-benefits, though low as plastic not decompose. Incineration leads to the largest net dis-benefit, due to burning of plastic (despite energy recovery)

Positive benefits of recycling.

Benefits do not capture additional benefits from avoided primary extraction

Positive benefits for landfill and incineration due to energy recovery.

Negative numbers (-) are benefits. Positive numbers (+) are dis-benefits.

Table 6.2 Summary of Policy Coverage

Lifecycle stage Wood in construction

Iron and steel in car PET in drinks bottle

Raw material extraction

Forestry certification schemes

Various regulations on offshore extraction of oil

Material Production LAPC IPPC IPPC

Incorporation into product

LAPC

Disposal of product/material

Landfill Regulations and Landfill taxWaste incineration directive

End of life vehicle regulations

Packaging regulationsLandfill regulations and taxWaste Incineration Directive

There were also a number of gaps identified. A comparison of other policies in place suggested that other policies partially address these gaps, and extend the policy coverage to other parts of the life-cycle, but a number of gaps remain. Following from this, and the analysis of wider policies, we suggest there might be a need for: Full policy coverage to address the total externalities of CO2 emissions, and the

conventional air pollutants (SO2, NOX and PM10). Market signals to recognise the benefits of recycling (i.e. to reflect avoided

externalities). The analysis suggests that recycling avoids significant externalities across the life cycle (i.e. from avoided extraction, processing and waste disposal).

Policies to tackle burdens associated with resource extraction, e.g. mining of raw materials for mining, and oil/gas extraction.

Policies to ensure optimal natural resource depletion (implying reliance on market forces is insufficient).

In addition, the results of this study suggest that not all the externalities of incinerating plastic in an energy from waste plant would be directly addressed by current policies. However, plastic would not normally be burned on its own, but as part of a mixed household/commercial waste stream. It would be useful to examine whether for the incineration of mixed wastes in energy from waste plant, there are externalities which are not addressed by polices (such as the Waste Incineration Directive and local planning regulation), which impact upon waste incineration facilities at the moment.

The study findings are summarised by material below in Table 6.3.

Table 6.3 Key environmental Impacts from lifecycle of PET used in a bottle

From Lifecycle IA Methodologies From externalities UK Policy Coverage of Key Impacts

Comments

Main ImpactsEmissions of CO2, leading to global warming/climate change impacts

Emissions of SO2, NOx and particulates leading to damage to human heath and acidification and to a lesser extent eutrophication impacts on ecosystems

Emissions arise from fossil fuel combustion in 1) production of constituent chemicals, and 2) production of electricity used in bottle blowing

Secondary Impacts

Freshwater ecotoxicity (nickel from oil combustion, vanadium from steel use and disposal of chemical wastes, and disposal of PET in landfill or incinerator)

Terrestrial ecotoxicity (mercury from chemicals production, chromium water borne emissions from disposal of chemical wastes, and vanadium emissions from oil combustion)

Main ImpactsCO2 emissions (60%)

NOx emissions (23%)SO2 emissions (10%)

Emissions arise from fossil fuel combustion in 1) production of constituent chemicals, and 2) production of electricity used in bottle blowing

Major difference between waste disposal and recycling options with incineration largest impacts, recycling large benefits.

Secondary Impacts

Other emissions/impacts identified as relevant in LCA work do not appear as significant in the externalities work, particularly the importance of trace pollutants (heavy metals)

CO2 emissions from energy use – Climate Change Levy EU ETS,

SO2 and NOx from energy use – Large combustion plant Directive (for electricity generation), IPPC

IPPC for incineration

All methodologies agree that emissions of CO2, SO2, NOx from combustion of fossil fuels used in production of constituent chemicals, PET itself, and in electricity used for bottle blowing and leading to global warming, acidification and impacts on human health are main sources.

Policy coverage for these emissions (mostly UK, and if not probably within Europe) is fairly good.

Externalities partly internalised, through existing economic instruments but do not fully account for carbon, or for electricity.

More minor impacts on ecotoxicity

Externality assessment shows that large impacts from incineration, not fully covered by policy

Note assumption on displaced electricity key in determining exact effects of incineration.

Issues with coverage for resource extraction and resource depletion.

Table 6.4 Key environmental Impacts from lifecycle of steel use in a car

From Lifecycle IA Methodologies From externalities UK Policy Coverage

Comments

Main ImpactsEmissions of CO2, leading to global warming/climate change impacts

Emissions of SO2, NOx and particulates leading to damage to human heath and acidification and to a lesser extent eutrophication impacts on ecosystems

Emissions arise from fossil fuel combustion in 1) blast furnace process, 2) sintering of iron 3) car production, and in the case of particulates from iron ore extraction

Secondary ImpactsFreshwater ecotoxicity (nickel and cobalt emissions to water from disposal of smelter slags from ferroalloy production and disposal of slag from recycling of steel)

Human toxicity (chromium emissions from disposal of ferrochromium slags)

Damage to human health from carcinogens (arsenic and cadmium to water from iron ore extraction and pelletisation)

Terrestrial ecotoxicity (mercury from electric arc furnace in steel recycling (CML) and emissions of heavy metals from ferrochromium and ferronickel production and disposal of smelter slag)

Main ImpactsSO2 emissions (36%)Particulates (22%)CO2 emissions (21%)NOx emissions (19%)

Emissions arise from fossil fuel combustion in 1) production of constituent chemicals, and 2) production of electricity used in bottle blowing

Secondary Impacts

Other emissions/impacts identified as relevant in LCA work also potentially important (nickel, chromium, arsenic and cadmium). However, issue over the time-scales of release of these compounds, as inventory quantifies emission from landfill disposal, and does not specify time period (assumes hundreds of year). Within policy time-scales, assuming modern sites, emissions much lower. This highlights an issue on time dimension for burden.

CO2 emissions from energy use – Climate Change Levy EU ETS,

SO2 and NOx from energy use – IPPC

Other emissions from iron and steel production, (including recycling in electric arc furnaces and in slag handling) are covered under IPCC.

End of Life Vehicle Directive encourages recycling which reduces most impacts

All methodologies agree that emissions of CO2, SO2, NOx from combustion of fossil fuels and leading to global warming, acidification and impacts on human health are main sources.

Policy coverage for these emissions where steel and car production are within the UK is fairly good. For imported cars, policy coverage is likely to be similar within other EU countries. Externalities partly internalised, through existing economic instruments but do not fully account for carbon, or for electricity.

All mining (and some preparation) of iron ore and most mining and production of ferroalloys occurs outside the UK (and EU). There are a number of secondary impacts from this these stages (to ecosystems and human health) due to emissions and disposal of slags from production of ferroalloys. The extent of policies to address these issues in the country of origin was not examined in this study, but could vary considerably.

The generation of over burden and other mining waste in iron ore and ferroalloy production, in terms of volume of waste produced and the amenity impacts of mining are significant potential environmental impact which are not identified by the life cycle impact assessment methodologies.

Issues with coverage for resource extraction and resource depletion.

More minor impacts on ecotoxicity, though there is an issue of the time dimension (see column to left)

Table 6.5 Key environmental Impacts from lifecycle of wood used in a building

From Lifecycle IA Methodologies From externalities UK Policy Coverage of Key Impacts

Comments

Main impactPositive environmental impact from reduction of global warming impacts due to CO2 stored in wood when wood recycled) and offset fossil fuel generation when wood disposed of to landfill or incineration

Secondary impactsEmissions of SO2, NOx from transport, and operations in forest and at sawmill leading to acidification and eutrophication.

Eutrophication impact increases significantly if waste wood is landfilled.

Damage to human health from carcinogens (cadmium, arsenic, and respiratory inorganics (particulates and NOx) from transport of the wood (by lorry), combustion of wood chips for kiln drying and fuels used during cultivation and felling.

Damage to ecosystems from land-use

Main ImpactsKey impacts are from CO2, SO2 and NOx emissions from use of fossil fuels in transport and operations in forest and sawmill, and in avoided emissions due to offset fossil fuel based electricity production from energy generation if wood is landfilled or incinerated.

Net benefits for CO2 emissions and SO2 emissions if waste wood is incinerated in energy from waste plant .

Issue raised of the time dimension, and disposal occurs in the future (+40 years) – this should involve the use of different values (e.g. for CO2) and potential consideration of discounting.

Other emissions/impacts identified as of secondary importance in LCA work do not appear as significant in the externalities work

Forestry labelling and certification schemes

Transport – standards for vehicle emissions

Landfill regulations, and landfill tax, waste incineration directive.

Combustion related pollutant of CO2, SO2 and NOx leading to acidification and eutrophication are important Unlike the other two materials considered, production of this resource is not very energy intensive, and transport is therefore a greater source of these fuel related emissions.

A number of impacts on forestry (e.g. on biological diversity, water resources, soils etc) not captured by life cycle impact assessment methodologies.

Note: abiotic depletion and mineral and fossil use are not included in the above table as from an economic perspective these are dealt with by the price mechanism.

6.2. Transferability and Further Work

In relation to the transferability of this approach to other examples or materials, we have found a major issue in transferring previous LCA studies into a form suitable for externality assessment (for the project aims here).

The use of a detailed material flow - identifying sources of raw materials, locations of processes and combining this with a very detailed ‘step by step’ life cycle model - would provide the necessary approach to undertake the analysis here properly.

However, even with relatively simple lifecycles, the volume of data generated by such an approach would be large and such an approach would be data and resource intensive. For this reason, the priority approach adopted here will be needed to make the analysis possible.

The overall approach developed here is, we believe, transferable to other studies, and the analysis of the externalities associated with materials (e.g. PET, iron and steel, wood) would be transferable to other products provided a proper in-depth analysis was undertaken (rather than the scoping analysis here).

This approach might require some additional work for each product, to set up the life cycle stages from material production through to disposal for the specific product. It would also require work to set up for the actual disposal route (e.g. for the specific material to landfill, incineration, and recycling).

There are some issues also to resolve – including the issue of incorporation into products more thoroughly. This includes coping with the complex patterns of product use, and the necessary data for incorporation into products. This includes the issue of additives, and how they are included within materials, and the issue of produce categorisation in terms of the activity associated with formation (e.g. the heating, shaping, etc). It might be possible to tackle the second of these through an analysis of the magnitude of impacts.

There is one final issue on transferability. This is over the year of the study and the assumed energy / electricity mix in place, and whether this should reflect current or future baselines. Different policy questions or analysis might want to work with different policy baselines, which would require the adjustment of the life cycle set up. Similarly, it is clear that the assumptions about waste disposal, e.g. energy recovery assumed, type of energy used in recycling, etc. have a big impact on determining the externalities, and again these might change for different studies. This is one area where further consideration is needed for transferability, and further work is needed to improve the analysis.

In relation to research priorities, the study recommendations are: Further testing with other LC impact assessment methods would be useful,

particularly some of the more advanced tools that include a more impact driven approach.

The externalities and life cycle impact assessment are generally closely correlated in identifying the most important impacts, but do differ in some areas (notably over the importance of trace pollutants). It would be useful to reconcile these differences in future material flow analysis.

More work is needed to attribute damage cost values (impact pathway approach) to all LCI burdens, or to develop approaches to screen all burdens more robustly. Similarly, work to progress the valuation of LCA output directly would be useful. We stress that these gaps are not a barrier to the successful implementation of a follow on phase of this study.

Sensitivity analysis would be useful to examine the potential effect on the results of uncertainty (e.g. on impacts and values) and over the choice of assumptions (e.g. on which energy source is displaced by waste to energy schemes, on future rather than current technologies/policies, etc).

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8. Glossary

CBA: Cost-benefit analysis. An economic technique for assessing policies or projects. CBA quantifies costs and benefits of options in monetary terms, including values not captured by markets

CFCs. Ozone depleting substance.

CO. Carbon monoxide. An air pollutant.

CO2. Carbon dioxide. A greenhouse gas.

COMEAP: UK Department of Health’s Committee on the Medical Effects of Air Pollutants

CCL Climate change levy.

CML: An life cycle assessment methodology (CML IA) that assesses the contribution of emissions to environmental problems by weighting them to allow their aggregation.

CV: Contingent valuation. Valuation technique which elicits the WTP of non-market goods through direct questionnaire

DALYs: Disability Affected Life Years. Method/unit to directly compare differences in the burden of disease and in causes of illness, including morbidity and mortality in a weighted indicator.

Damage cost: simplified externality estimates for use in scoping analysis, that link burden directly to valuation, usually expressed as e.g. £ per tonne of emission.

Defra: The Department for Environment, Food and Rural Affairs

DfT. Department for Transport.

DoH: The Department for Health.

Discounting: technique used to compare costs and benefits that occur in different time periods, recognising people prefer to receive goods and services now rather than later. The rate at which future goods or services are discounted to today requires the use of a discount rate.

EC: European Commission.

Euro standard. European Commission emission standard legislation, relating to Euro standards I to V. These have the effect of reducing emissions from new vehicles.

External Costs/Externalities: economic activities have environmental and social impacts (e.g. air pollution) which have economic costs. These are typically not paid

for by providers or users (of the goods or services in question) and are known as external costs or externalities.

GHG: Greenhouse gas.

GIS: Geographical Information System.

GWP: Global warming potential. A conversion factor to allow the expression of different GHG in equivalent terms (CO2 equivalent).

IGCB: Inter-departmental group on costs and benefits.

IPPC: Integrated Pollution Prevention and Control. Integrated approach to the environmental regulation of major industrial activities.

LCA: Life Cycle Analysis.

LCI: Life Cycle Inventory.

LCPD: Large Combustion Plant Directive. EU legislation setting limits for emissions of SO2, dust and NOX for new power stations burning solid, liquid or gaseous fuels with a thermal input of over 50 MW

MCA: Multi-criteria assessment (or analysis). Non-monetary technique for option appraisal.

NO. Nitric oxide. An air pollutant.

NOx. Oxides of nitrogen (includes NO and NO2). An air pollutant.

NO2. Nitrogen dioxide. An air pollutant.

O3: Ozone. An air pollutant.

PET: Polyethylene Terephthalate

PM10. Particulate matter less than 10µm aerodynamic diameter. An air pollutant.

QALY: Quality Adjusted Life Years. Unit for assessing relative health impacts, where number of human life-years affected is multiplied by a severity score (quality adjustment).

PM2.5. Fine particles less than 2.5 µm in diameter

RIA. Regulatory Impact Assessment. Option appraisal, mostly centred on cost-benefit analysis, for assessing proposed policies in the UK.

ROCs. Renewables Obligation Certificates. In April 2002, the Renewables Obligation was introduced. This requires all licensed electricity suppliers in England

& Wales to supply a specified and growing proportion of their electricity sales from a choice of eligible renewable sources. Suppliers are responsible for demonstrating that compliance to OFGEM through a system of Renewables Obligation Certificates (ROCs).

SCC: Social Cost of Carbon. The social costs of climate change occurring.

SCP: Sustainable Consumption and Production.

Social Costs. The wider costs to society from economic activity. See external costs above.

SO2: Sulphur dioxide. An air pollutant.

VED Vehicle excise duty (i.e. road tax).

VOC: Volatile organic compounds. An air pollutant.

VOLY: Value of life year. Valuation unit used for changes in life expectancy.

WTP Willingness to pay. The willingness to pay of the affected individual to avoid a negative impact.

Appendices

CONTENTS

Appendix 1 Screening of MaterialsAppendix 2 Life Cycle Studies for Candidate Materials

Appendix 1:Screening of Materials

Initial Screening of Materials

A large number of potential ‘materials’ were identified for initial screening in the study:

Aluminium; Iron and Steel; Heavy metals (e.g. cadmium, mercury); Aggregates; Quarry materials e.g. limestone, slate; Cement; Bricks; Glass; Wood/timber; Paper; Natural textiles (cotton and wool); Man-made textiles; Rubber; Plastics; Fertilisers; ‘toxic chemicals’ e.g. pesticides.

Materials which it was agreed to exclude at the kick-off meeting were: Primary fuels, Transport fuels, Electricity and Agricultural products/food.

The environmental impacts of primary fuels, transport fuels and electricity over their lifecycles have been studied in considerable detail previously, and the externalities assessed. It was therefore felt their inclusion would provide little added value to this study. Agricultural products and food were excluded as they are the focus of other parts of the SCP research programme

The materials were screened by assessing them against the following selection criteria identified in the Terms of Reference for the study.

Volume; Hazardous nature; % of total volume of material disposed of by landfill; Evidence of environmental concern (key envioronemtnal impacts as identified by

other studies); Sufficient data, especially LCA data; Transferability of learning.

Assessment of these criteria was important to enable the study:

To obtain a good coverage against major areas of environmental concern, and try to select materials that have major environmental burdens in each area, e.g. in

each of climate change, air quality, water quality, resource depletion, and biodiversity.

To pick specific materials with known environmental burdens, e.g. with high greenhouse gas emissions, others with high air pollutants, others with toxic or persistent chemicals / metals, others with waste arising (to landfill), etc.

To review the quality of LCA information. This is an extremely important factor, because it will dictate the potential for later analytical steps in the project.

Other aspects which were evaluated and the reason for inclusion were:

Trends in use (growth/decline): it would not be useful to study a material whose use is already declining as a result of policies in place

Level of recycling % of material imported : Number of product uses; this is an important factor in determining the

complexity of analysis needed.

The results of the screening are shown in Table A1.1

Table A1.1 Screening matrix for materials

Material Renewable material

Growth rate % imported No of product uses Recycling Policy coverage Key Lifecycle

Stages (provisional assessment)

Data availability and quality (provisional assessment)

Other comments

Raw materials

Processed material

Aluminium X HIGH ALL MEDHIGH (mainly in

buildings, packaging, vehicles

50%(1998)

Energy taxes/CCL. IPPC, Landfill tax

Extraction and processing

European data on production, LCAs for some products

Iron and Steel X MED ALL LOW17.7%

18,500tonnesHIGH 40%

(1998)Energy taxes/CCL. IPPC, Landfill tax

Extraction and processing

European data on production, LCAs for some products

Heavy metals (e.g.

cadmium, mercury)

X

LOW – but

RoHS may

cause a decline

Cad = HIGHMercury = HIGH

Cad = MedMercury = High

Hazardous Waste Directive

RoHSIPCC, WID

Extraction?, processingDisposal

Some studies on NiCd batteries, use of Cd in electronic

goods and PV panels – may not be

UK based

major intentional uses of cadmium are Ni-Cd

batteries (65%), pigments (17%), stabilisers 10%) ,

coatings, cadmium alloys and cadmium

electronic compounds

Aggregates X

LOW1%

Growth rate

VERY LOW0.69% (2003) LOW 24%

Aggregates levy, planning process

Landfill tax

ExtractionDisposal

Some data available

Quarry materials e.g.

limestone, slate

X MED LOW LOW LOW

All

have

pot

entia

l to

be

recy

cled

as a

ggre

gate

s.

No

figur

es fo

und.

LAPC, planning process

ExtractionDisposal

Data available for production

Cement X MED VERY LOW1.49%

LOW(Primarily

construction)

Energy taxes/CCL IPPC Landfill taxAggregates levy

Extraction. Production

Data available for production

Bricks X MED VERY LOW(0%)

VERY LOW(0%)

LOW)

Energy taxes/CCL IPPC/LAPC Landfill tax

Extraction, production

Data available for production

Glass X MED LOW (22%)

MED (over 90% used for containers,

and glazing in buildings and

vehicles)

46% (2004)

Energy taxes/CCL IPPC Landfill tax Production

Data for production and some products

available

Table 1 Screening matrix for materials (Continued)

Material Renewable material

Growth rate % imported No of product uses Recycling Policy coverage Key Lifecycle

Stages (provisional assessment)

Data availability and quality (provisional assessment)

Other comments

Raw materials

Processed material

Wood/timber HIGH HIGH ?MED

Primarily furniture, paper, construction

11-16%

Voluntary labellingLAPC. Landfill

Directive.Packaging Directive

Extraction, production,

disposal

Data for European production

available, also for some products

Additional materials added (preservatives etc)

can have significant impact

May need to define type of timber (e.g.

softwood/hardwood as sourced differently

Paper HIGHHIGH(over 70% of

pulp)

HIGH(60% of paper)

HIGH 37% (2004)

IPCC, Landfill tax, Landfill Directive

Extraction, production,

disposal

Production data available but generally at

European not UK level; some product

LCAs as well

Large number of grades and uses, but cold focus on some example uses

(e.g. a packaging use or finished product such as

newspaperNatural textiles

(cotton and wool)

LOWHIGH HIGH

HIGH (clothing, household soft

furnishings, industrial textiles)

(25%) LAPC, Landfill taxLandfill directive

CultivationProduction

Some production data Processing (e.g. dying

can have significant impacts)

Man-made textiles X LOW HIGH (as above) ? IPCC, LAPC Production

Processing (e.g. dying can have significant

impacts)

Rubber MED ALL (natural)

HIGH60%

(Synthetic)

MED (tyres account for 70%)

19% (tyres – 1999)

Landfill directive (Tyres)

Plastics X HIGH HIGH LOW HIGH 7% (2001)IPCC.

Packaging and Packaging Waste

Production, disposal

Data on plastics manufacture at

European level; data on products likely to

be available but need to define first

May need to pick particular plastic with

more limited no of uses e.g. EPS

Fertilisers X MED

HIGH70-80% of

raw materials

HIGH (51%) LOW X IPPC, Nitrates Directive Production, use

Production phase well characterised, some data on use

phase‘toxic

chemicals’ e.g. pesticides

X MED ? HIGH HIGH (about 60^) MED X

Pesticide levy,Groundwater

Directive

Production, use, disposal

Likely be limited May be difficult to identify ‘generic’ product to study

Definition of imports High, >50%, Medium, 25-50%, Low 25%, Very Low <5%)

Table 1 Screening matrix for materials (continued)

Emissions to Air Water Waste Other issuesClimate change*

Ozone layer depletion*

Acidification

Eutrophication

‘Smog’ (Photochemical ozone0

Eco-toxicity/human toxicity

Water use

Water quality

Waste (volume)

Waste(hazardousness)

Energy use

Abiotic resource depletion

Biotic resource depletion

Land use/biodiversity

Amenity Social impacts

Aluminium (26) (E) (25)

(E) (E)

Iron and Steel (10) (E) (16)

(E) (E)

Heavy metals (e.g. Cd, Hg)

(E) ?

Aggregates c Quarry materials e.g. limestone, slate

(E) (E)

Cement (2) (E) (4)

(E) (E)

Bricks (13) (E) (20)

(E) (E)

Glass (30) (E) (21)

(E) (E)

Wood/timber Paper (18) (E)

(17)(E) (E)

Natural textiles (cotton)

Man-made textiles

Rubber Plastics 25 (E)

24(E)

Fertilisers 24 (E) 23

(E)

‘toxic chemicals’ e.g. pesticides

(agrochemicals 34)

(E) (agrochemicals 33)

(E) (E)

Notes: (E) Any process with significant energy use will have emissions of combustion related pollutants (e.g.NOx, SO2, particulates and VOCs) which will contribute to acidification, eutrophication, smog and will have an impact on human health.No. in brackets (for climate change and acidification) indicate ranking in UKs environmentally extended accounts emissions per £M. Should be considered indicative only due to age of data. Not all materials correspond with sector classification used in account

The analysis undertook a rapid review of other studies that have prioritised products/sectors/materials by examining their environmental impacts on a life cycle basis. These include relevant key consumption sectors, materials, or processes and are shown in Table A2.2 below.

Table A1.2 Areas with key environmental impacts identified in other studies

EIPRO (examined final consumer consumption)

Van de Voet et al, 2004 (examined materials)

CEC, 2004 (examined, resource use/activities)

Housing (construction of building)

Clothing and footwear (particularly clothing)

Recreation (household audio/video equipment)

Pesticides

Plastics Concrete Iron and steel Paper and board Glass, clay, lead,

nickel, zinc

Solvent use Metal extraction and

refining Dissipative uses of

heavy metals Housing and

infrastructure Marine activities Chemical industry

Following analysis of the screening matrix, and discussion with the steering group at the kick-off meeting, six broad themes of particular interest were taken forward for more consideration. These were: Wood or paper, as a renewable material; Materials with local amenity impacts that are mined or quarried, e.g. aggregates. Energy intensive materials (e.g. iron and steel, aluminium, glass); Construction and building material (e.g. bricks or concrete) due to the high

quantities involved; Plastics, due to the growing use of these materials. Materials with high specific toxicity or major impacts during the use phase (heavy

metals, toxic pollutants (e.g. pesticides), and fertilisers (important for water pollution).

Further consideration of these materials, and their pros and cons, is shown in Table A1.3.

Table A1.3 Choice of five materials for more in-depth analysis in the screening phase

Material characteristics Suggested Candidates

Pros/Cons

Renewable material WoodPaper

Wood is a much broader category than paper (would include paper), Flows could be complex, but could be interesting to examine labelling schemes

Materials with local amenity impacts, and upstream policy coverage and possibilities for recycling

Aggregates, Other quarried products

Inclusion of aggregates would allow examination of aggregates levy, which was specifically designed to address externalities.

Energy intensive materials (so have impacts in several categories), also recyclable

Iron and steel,AluminiumGlass

EEA recently produced LCA for iron and steel in Europe – could be good source of dataGlass – mostly UK production

Plastics Various Use is increasing. Interesting from the perspective of recycling. While acknowledging the complexity of plastics, this was not felt appropriate as a reason to exclude them from the analysis. Need to pick plastic with small number of products.

Construction/building product Cement, brick May be more recycling routes for cement

Impacts during use/disposal phase; impacts on water bodies

Heavy metals – probably Cd,Pesticides,Fertilisers

Heavy metals and pesticides score highly on toxicity/hazardousnessPesticides/fertilisers may be interesting from policy coverage point of view (e.g. taxes in other countries)Best data set is probably for fertilisers

Following further discussion and guidance from the steering group, five specific materials were considered in more detail. These are: Wood/paper; Aggregates; Iron and steel; Fertilisers; Plastic (specifically PET);

These are described in the next section.

Assessment of Candidate Materials

The suitability of the five candidate materials for detailed study was examined by looking at: The flows of the materials through the economy in more detail, to establish the

diversity of ‘routes’ through the economy and the potential to construct representative lifecycles which capture the majority of the material use

The potential availability of life cycle data and studies, both for the raw materials, intermediate products and final products. This was assessed by looking at the life cycle literature, and also at publicly available life cycle inventory data and life cycle inventory data (for materials not products) to which the project team has access through its use of LCA software tools (the Ecoinvent database).

Wood

The material flow for wood in the UK economy is complex (TRADA, 2005). As well as domestic production, there are imports and exports of wood, sawn timbers, intermediate products such as plywood, veneer sheets, particle boards and fibre boards, and of finished products such as furniture and of packaging (mainly as pallets). A simplified view of the material flows is given in Box 1, together with details of the main uses of wood in the UK: construction, accounts for just over a third, and packaging and pallets, almost one-fifth each. About 14% is transformed into paper, and the remainder is used for joinery. About 4 million tonnes of wood are estimated to enter the waste stream from construction and demolition sites44, together with over 1 million tonnes from packaging. Estimated wood wastes from the furniture industry are lower (about 0.3Mt). About 10% of waste wood in the UK is estimated to be recycled via a number of routes including use as animal/poultry bedding, mulching/composting, use in panel manufacture and combustion to produce heat and/or power.

LCA studies are available for timber, particle and fibreboard, some wood building components, and for pulp and for paper products. No specific UK studies were found;, most are based in the main timber producing regions (Scandinavia, Canada) and Germany. Life cycle inventory data is available from Ecoinvent for timber, wooden boards, particle boards and fibreboards.

44 http://www.globaltrees.org/proj.asp?id=4

Box 1 Material Flows in the UK for the Production and Use of Timber

Flow of Materials for the Production and Use of Timber in the UK

Softwood: 9070 kt (green) produced in the UK in 2004**

Raw Materials Intermediate Products Final Products

* FAOSTAT

** Forestry Commission

Hardwood: 510 kt (green) produced in

UK in 2004**

Timber Production in UK with UK wood**

5000 kt to sawmills

1500 kt to wood panel mills

660 kt to other mills

700 kt to pulp millsProcessed Pulp: 500 kt produced in the UK in 2003*

Paper & Board: 6226 kt produced in 2003*

1500 kt pulp imported 2003*

1700 kt paper & board exported 2003*

7500 kt paper & board imported 2003*

12026 kt consumed in the UK in 2003930 kt (green)

exported **

2780 sawn timber produced; 370 exported; 8647 imported (in thousands of cubic metres)**

3533 panels produced ; 511 exported; 3786 imported (in thousands of cubic metres)**

6,806,000 m3 panels consumed in 2004 in the UK

11,059,000 m3 timber consumed in 2004 in the UK

In addition, UK sawmills processed 9,798,000 m3

imported softwood and hardwood in 2004. UK Panel mills processed 4,259,000 m3 imported wood.**

Processed timber is used primarily to make fencing and packaging or is used in construction. In 2004, 34% of softwood production was used to make fencing; 33% was used in construction; 31% was used to make packaging and pallets; 2% was used to make other products, primarily wood chips and sawdust, or was burnt for heat. In 2004, 81% of hardwood production was used in construction; 4% was used to make packaging and pallets; 15% was used to make other products, primarily wood chips. (UK Timber Statistics 2004).

The finished products from the wood panel mills were primarily particleboard or OSB (oriented strand board). In 2004, 2.6 million cubic metres of the finished panels were particleboard or OSB and 880 thousand cubic metres were fibreboard (MDF) (UK Timber Statistics 2004).

Consumption of Wood and Wood Based Products in the UK

Source: TRADA Technology Ltd, 2005

Aggregates

In contrast to wood, aggregates are produced and used almost entirely in the UK (Box 2). Their flow through the economy is much simpler. Recycled materials account for about a fifth of consumption.

Key environmental impacts are in the extraction phase (where the aggregates levy was introduced to account for amenity impacts), transport and to some extent disposal (although they are an inert material, large volumes are involved). Lifecycle inventory data is available for sand and gravel (for the Swiss situation). A UK LCA study of construction and demolition waste (Craighill and Powell, 1999) could provide some information on the disposal and recycling of aggregates. Work to support the introduction of the aggregates levy could provide data on amenity impacts associated with extraction.

Box 2 Material Flows in the UK for Aggregates:

IMPORTS

3.1Mt/Yr (2004)

Life Cycle of Aggregates in the UK

UK PRODUCTION

214.6 MT/Yr

Recycled Materials* = 18%

Secondary Aggregates** = 4%

Rail

3Mt/Yr

RAW MATERIALS END OF LIFE

http://www.mineralsuk.com/britmin/import_aggs_uk.pdf

http://www.mineralsuk.com/britmin/mm6.pdf

http://www.mineralsuk.com/free_downloads.html

http://www.environment-agency.gov.uk/commondata/103196/waste2?referrer=/yourenv/eff/1190084/resources_waste/213982/152399/

EXPORTS

12.2MT/Yr (2004)

Construction

(202.5 Mt/Yr)

38% Sand & Gravel

62% Crushed Rock

104 Mt/Yr

Crushed Rock

4.1

Sand & Gravel

8.1

Crushed Rock

0.634

Sand & Gravel

0.942

47.2 Mt/Yr

Crude Granite

1.558

*Recycled Materials = C&D Waste, Railway Ballasts, Asphalt Planning

**Secondary Aggregates = Power station ash, Iron & Steelworks Slag, China Clay Waste, Colliery Spoil, Slate Waste

USE

1.2 Mt/Yr (spent ballasts)

205.5 Mt/Yr

(2004)

Iron and Steel

While almost all the raw materials for iron and steel production are imported into the UK, relatively little (less than 5%) of primary or secondary steel is imported. While four sectors (engineering, automotive industry, construction and metal goods) account for almost 80% of steel use; there are a wide range of uses within these sectors.

Life cycle inventory data is available for iron and steel production, although the initial literature review revealed few studies of products. The European Topic Centre on Resource and Waste Management (Moll et al, 2005) have recently looked at resource flows of iron and steel in the EU, and this contains much useful lifecycle data and information on environmental impacts at the EU level, mostly for the extraction of raw materials and production of steel. The study highlighted several difficulties (both conceptual and data wise) in considering the environmental impacts of processing of steel into goods, and of the ‘use stage’ of steel containing goods. The latter were not estimated in the study, neither were the impacts of recycling. scrap steel.

Box 3 Material Flows in the UK for the Production and Use of Iron and Steel

Flow of Materials for the Production and Use of Iron and Steel in the UK

Iron orePrimary Steel: 10700 kt produced in 2004**

Semi-finished steel products

12,114 kt of finished steel products were produced in the UK in 2003 *

Raw Materials Intermediate Products Final Products

*British Geological Survey, 2005.

** UK Iron and Steel Statistics Bureau

24% Engineering

14% Metal goods

17% Automotive

22% Construction

23% Other industries **

A number of minerals are needed to produce crude steel, including:

Limestone: 2018 kt used in the steel industry, primarily produced in the UK (in 2003 a total of 79000 kt of limestone was produced in the UK)

Chromium: 55.4 kt consumed in steel industry, primarily imported (107 kt imported in 2003)

Ferro-manganese: 99.3 kt consumed in steel industry, primarily imported (79 kt imported in 2003)

Nickel: 15.8 kt consumed in steel industry, (26.8 kt produced in UK in 2003; 146.4 kt imported)*

Coke

Scrap steel

5,200 kt consumed 2004**, 139 kt of

which was imported*

Crude Steel

Secondary Steel: 3500 kt produced in 2004** 1306 kt

exported 2003*

540 kt imported 2003*

Iron ore is primarily imported into the UK. In 2004, 16500 kt of iron ore were imported for use in the iron & steel industry. Coke was either imported directly (600 kt in 2004) or made in the UK from imported coal (6000 kt) (UK Iron and Steel Statistics Bureau). Scrap steel from semi-finished and finished products is re-used to make secondary steel. In 2003 the UK exported 7000 kt of scrap steel. (British Geological Survey, 2003).

The markets for steel can be broken down further: Engineering: mechanical machinery and equipment (9%), DEA’s (2%), other

electrical products (3%), and other mechanical engineering (10%) Metal goods: packaging (5%) and furniture and other goods (9%). Construction category: structural steelwork (13%) and building and civil

engineering (9%). Other industries: wire drawing (6%), forging and stamping (5%), cold forming

(3%), oil and gas (3%).

Plastics

The complexity of the plastics sector – both in terms of types of plastics and in terms of end uses (Figure 2) and in relation to imports and exports of plastics components and plastics in products - means that it is necessary to consider the plastic type separately in order to be able to map the resource flows. The example chosen was PET (Polyethylene Terephthalate) as it is manufactured in significant quantities in the UK, has a single predominant use – drinks bottles, and is a plastic that is recycled as bottles are an easily segregated waste, with relatively low levels of contaminants. Lifecycle inventory data for the manufacture of PET is available (although based on European rather than UK data) and there are LCA studies which include recycling of PET, although not for the UK situation. For example the LCA study comparing PET and HDPE bottles by Lehmann et al (2005) is for Sweden.

Figure 2 UK Consumption of Plastics by Sector and Polymer Type.

Source: Plastics in the UK Economy – A report from Wastewatch

Box 4 Material Flows in the UK for the Production and Consumption of Polyethylene Terephthalate (PET Plastic)

Flow of Materials for the Production and Use of PET plastic in the UK

Crude oil and Natural Gas

Ethylene glycol (EG) and TerephthalicAcid (PTA)****

235kt of PET plastic products manufactured in the UK in 2000**

Total supply of PET plastic products in 2000: 261 kt**

Raw Materials

Intermediate products

Production Consumption

*British Geological Survey, 2005.

** Plastics in the UK Economy

*** Digest of UK Energy Statistics 2005

**** Lehmann 2005.

Synthesised PET

60% clear plastic bottles

12% coloured bottles

18% other, including textile fabrics and food packaging **

A number of minerals needed to produce synthesised PET, including:

Limestone (6799 kt produced in UK in 2003 for industry)

NaCl (6000 kt produced in UK in 2003; 217 kt imported, 537 kt exported)

Bauxite (mainly imported; UK imported 271 kt in 2001) *

Net import of 26 kt**

In 2004 the UK produced 95,374 kt (kilo tonnes) of crude oil, imported 62,516 kt and exported 64,504 kt. The UK also produced 1,115,744 GWh of natural gas, imported 133,035 GWh, and exported 114,111 GWh (DUKES 2005).

Fertilisers

Box 5 shows the material flows for production of inorganic nitrate, phosphate and potassium fertilisers in the UK. The greatest use is of nitrogen fertilisers (1.1 million tonnes, and just under half of this is produced in the UK. The main raw material is natural gas, from which ammonia is produced. Much ammonia production for fertiliser production is UK based, and is often directly linked to or integrated with fertiliser production plant45. The product flow for fertilisers is very straightforward, with virtually all fertilisers being used in agriculture. If this material is chosen for further study, it is proposed that the system boundary is drawn at the ‘farm field’, i.e. impacts from the use phase (emissions and leaching) are considered, but that the fate of nitrogen which is taken up by the crops is not.

45 Although recently higher gas prices have seen some UK ammonia production plant shut down as it has become cheaper to import ammonia..

The main environmental impacts arise with the production phase, (as this is very energy intensive and a fossil fuel is used as a feedstock) and with the use phase – nitrate leaching into water bodies and emissions of the greenhouse gas nitrous oxide. Life cycle inventory data is available (e.g. from Ecoinvent) for production of several types of fertiliser (European average data). Estimates of average leaching rates from fertiliser use are available (e.g. as reported in Lillywhite and Rahn (undated) and emissions of N2O could be estimated using recognised emission factors (i.e. as in the IPCC methodology for compiling national greenhouse gas emission inventories).

Box 5 Material Flows in the UK for Nitrogenous, Phosphorous, and Potassium/Potash Fertilisers:

Note: Mt = metric tonne

Flow of Materials for the Production and Use of Nitrogenous Fertilisers in the UK

Nitrogen (from air)

+

Hydrogen (from natural gas)

AmmoniaTotal Nitrogenous fertilisers produced in 2002 in the UK: 540,000 Mt*

Total Nitrogenous fertiliser consumption in 2002 in the UK: 1,142,000 Mt*

Approximately 99% of fertilisers in the UK are used for agricultural purposes. The remaining 1% are used for domestic purposes.**

236,000 Mt imported in 2000*

182,800 Mt exported in 2000*

636,000 Mt imported in 2002*

30,000 Mt exported in 2002*

Raw Materials

Intermediate products

Fertiliser production

Fertiliser use

* FAOSTAT Data, 2004.

** Estimate from personal communication with the Agricultural Industries Confederation (AIC)

Flow of Materials for the Production and Use of Phosphate Fertilisers in the UK

Rock Phosphate or Basic Slag

(both imported***)

Phosphoric Acid, single superphosphate, or NP slurries

Total Phosphate fertilisers produced in 2002 in the UK: 50,000 Mt*

Total Phosphate fertiliser consumption in 2002 in the UK: 283,000 Mt*

Approximately 99% of fertilisers in the UK are used for agricultural purposes. The remaining 1% are used for domestic purposes.**

186,299 Mt phosphoric acid imported in 2000*

282,000 Mt imported in 2002*

0 Mt exported in 2002*

Raw Materials

Intermediate products

Fertiliser production

Fertiliser use

* FAOSTAT Data, 2004.

** Estimate from personal communication with the Agricultural Industries Confederation (AIC).

*** British Geological Survey, 2005.

Flow of Materials for the Production and Use of Potassium/Potash Fertilisers in the UK

Mineral deposits +

Chloride or sulphate

Purified and concentrated material

Total Potassium fertiliser produced in 2002 in the UK: 540,000 Mt*

Total Potassium fertiliser consumption in 2002 in the UK: 376,000 Mt*

Approximately 99% of fertilisers in the UK are used for agricultural purposes. The remaining 1% are used for domestic purposes.**

279,111 Mt imported in 2002*

323,186 Mt exported in 2002*

Raw Materials

Intermediate products

Fertiliser production

Fertiliser use

*FAOSTAT Data, 2004.

** Estimate from personal communication with the Agricultural Industries Confederation (AIC)

*** British Geological Survey, 2005.

Note: Boulby mine in North Yorkshire is the only potash source in UK. Chloride and sulphates must be imported. ***

Bibliography of Sources Used for Initial Screening

Aluminium www.alupro.org.uk)(http://www.foe.co.uk/pdf/sustainable_development/tworld/metals.pdf)http://www.psi.org.uk/publications/archivepdfs/environment/economicdimensionsExecSum.pdfhttp://www.aluminum.org/template.cfm?Section=The_Industry)

Iron and Steel

http://www.psi.org.uk/publications/archivepdfs/environment/economicdimensionsExecSum.pdf

Heavy metals

http://www.cadmium.org/)http://www.bartleby.com/65/me/mercury.html) http://www.wasteonline.org.uk/resources/InformationSheets/Batteries.htm

Aggregates http://www.aggregain.org.uk/sustainable.html)Page 22 of http://www.mineralsuk.com/britmin/ukmy2004.pdf.

Cement http://www.publications.parliament.uk/pa/ld200304/ldselect/ldeucom/179/4050502.htm)Bricks http://www.mineralsuk.com/britmin/brickclay.pdfGlass http://www.defra.gov.uk/environment/waste/topics/packaging/pwreg97/09.htm)

http://www.britglass.org.uk/Industry/IndustryHome.htmlPlastics http://www.manufacturingtalk.com/news/fro/fro168.html)

http://www.bpf.co.uk/bpfindustry/An_Introduction_to_Plastics.cfm)Wood http://www.forestry.gov.uk/forestry/ahen-5gbls4

http://www.wasteonline.org.uk/resources/InformationSheets/Wood.htmhttp://www.wwf.org.uk/filelibrary/pdf/uk_imports_wood-based.pdf) http://www.wrap.org.uk/downloads/Wood0504i.38e10805.pdf

Rubber http://www.statistics.gov.uk/downloads/theme_commerce/PRQ-20044/PRQ24170-20044.pdfhttp://www.iom3.org/divisions/plastics_rubber/news_files/rubber_tech_roadmap.PDFhttp://www.lancashire.gov.uk/environment/lancashireprofile/sectors/rubber.asphttp://www.dti.gov.uk/sustainability/downloads/tyre.pdf

Fertilisers http://www.statistics.gov.uk/downloads/theme_commerce/PRA-20030/PRA14300_20030.pdfOther quarried materials

http://www.mineralsuk.com/britmin/ukmy2004.pdf

Paper http://www.paper.org.uk/info/statistics/statshome.htmCEPIhttp://www.poptel.org.uk/iied/smg/pubs/rethink3.html

Textiles http://www.wasteonline.org.uk/resources/InformationSheets/Textiles.htmhttp://www.e4s.org.uk/textilesonline/content/6library/fr_library.htm

Recycling http://www.defra.gov.uk/environment/statistics/waste/

Appendix 2:Lifecycle Studies for Candidate Materials

Wood

Title: A Life Cycle Assessment of the production of a daily newspaper and a weekly magazine

Products: Newspapers & weekly MagazinesCountries: Sweden, Germany, CanadaLife Cycle Stages: Cradle to GraveYear of Study: 1998Availability of report:

Free on Web

Quality of data: Good

Title: End of Use and End of Life Aspects in LCA of Wood Products – Selection of Waste Management Options and LCA Integration

Products: Wood ProductsCountries: EuropeLife Cycle Stages: End of Life onlyYear of Study: Circa 2001Availability of report:

Free on the wed

Quality of data: Good

Title: Energy, Carbon and Other Material Flows in the Life Cycle Assessment of Forestry and Forest Products

Products: From the forestry industryCountries: FinlandLife Cycle Stages: Not a full LCA – reviews case studies.Year of Study: 2001Availability of report:

Free on the Web

Quality of data: OKTitle: LCA for Paper & Packaging Waste management

Scenarios in VictoriaProducts: Paper and Packaging WasteCountries: AustraliaLife Cycle Stages: End of lifeYear of Study: 2001Availability of report:

Free on the web

Quality of data: Good

Title: Life Cycle Assessment of Particleboards and Fibreboards

Products: Particleboard & FibreboardCountries: Germany & PolandLife Cycle Stages: Cradle to ManufactureYear of Study: 1996/1997Availability of report:

Free on the web

Quality of data: GoodTitle: Life Cycle Assessment Studies in the Timber IndustryProducts: TimberCountries: Canada & GermanyLife Cycle Stages:Year of Study: 1995Availability of report:

Free on the web

Quality of data: Speech made by authors.

Title: Life Cycle Environmental Impact Analysis for Forest Products

Products: Forest ProductsCountries: USALife Cycle Stages: Not a full LCA more of a presented paper.Year of Study: 1995Availability of report:

Free on the web

Quality of data: Not GoodTitle: Life cycle assessment of paper and plastic checkout

carrier bagsProducts: Carrier Bags (Plastic & Paper)Countries: USALife Cycle Stages: Cradle to GraveYear of Study:Availability of report:

Free on the Web

Quality of data: Good

Title: A Portrait of the Québec Pulp and Paper IndustryProducts: Pulp & PaperCountries: CanadaLife Cycle Stages: Not an LCA – covers environmental impactsYear of Study: 2001Availability of report:

Free on the web

Quality of data: OK

Title: Study and Assessment of Available Information for a Pilot Project on a Teak Garden Chair

Products: Teak Garden ChairsCountries: Europe & IndonesiaLife Cycle Stages: Cradle to GraveYear of Study: 2005Availability of report:

Free on the web

Quality of data: Good LCI

Title: Timber and the circle of lifeProducts: Tropical TimberCountries:Life Cycle Stages: Cradle to GraveYear of Study: 2004Availability of report:

Free on the web

Quality of data: OK – summary of report.

Title: Wood components in steel and concrete buildings – In-fill exterior wall panels

Products: WoodCountries: UK, the Netherlands, France, Germany, Poland, China and

the Nordic countries (Sweden, Norway and Finland).Life Cycle Stages: Not an LCAYear of Study: 2003Availability of report:

Free on the web

Quality of data: Gives details of environmental impactsTitle: Comparative LCA:s for wood construction and other

construction methods – Energy use and GHG emissions.

Products: Wood frame structuresCountries: SwedenLife Cycle Stages: Cradle to GraveYear of Study: 2003Availability of report:

Free on the web

Quality of data: Good

Title: LIFE CYCLE OF WINDOW MATERIALS - A COMPARATIVE ASSESSMENT

Products: Wood, Plastic and metal framesCountries: UKLife Cycle Stages: Cradle to GraveYear of Study: ?Availability of report:

Free on the Web

Quality of data: OKTitle: Green by DesignProducts: WoodCountries: CanadaLife Cycle Stages: Limited Cradle to Grave dataYear of Study: ?Availability of report:

Free on the Web

Quality of data: Limited

Title: Wood Processing and Furniture Making: Cleaner Production Fact Sheet and Resource Guide

Products: WoodCountries: ?Life Cycle Stages: Offer limited Environmental impact data.Year of Study: USAAvailability of report:

Free on the Web

Quality of data: Limited

Plastics

Title: COMPARATIVE LCA ON PLASTIC PACKAGING Products: PET & HDPE used in bottlesCountries: SwedenLife Cycle Stages: Cradle to GraveYear of Study: 2005Availability of report:

Free on the web

Quality of data: GoodTitle: LCA recycled Plastic steel and wood blocksProducts: Recycled PlasticCountries: USALife Cycle Stages: Cradle to graveYear of Study: 2000Availability of report:

Free on the web

Quality of data: Primarily a cost assessment

Title: Life cycle assessment of paper and plastic checkout carrier bags

Products: Carrier Bags (Plastic & Paper)Countries: USALife Cycle Stages: Cradle to GraveYear of Study:Availability of report:

Free on the Web

Quality of data: GoodTitle: Life Cycle Assessment of PVC and of principal

competing materialsProducts: PVC + Other competing materialsCountries: EULife Cycle Stages: Cradle to GraveYear of Study: 2004Availability of report:

Free on the Web

Quality of data: Good

Title: Solid Waste and Greenhouse gases - LCA of Emissions and Sinks

Products: Plastics including LDPE & PETCountries: USALife Cycle Stages: Cradle to GraveYear of Study: 2002Availability of report:

Free on the web

Quality of data: Good

Title: LCA of the Industrial Use of Expanded Polystyrene Packaging in Europe

Products: Packaging systems for TV setsCountries: EULife Cycle Stages: Cradle to GraveYear of Study: 2001Availability of report:

Free on the web

Quality of data: Good

Title: LIFE CYCLE INVENTORY OF PACKAGING OPTIONS FOR SHIPMENT OF RETAIL MAIL-ORDER SOFT GOODS

Products: PackagingCountries: USALife Cycle Stages: LCIYear of Study: 2004Availability of report:

Free on the web

Quality of data: Good

Aggregates

Title: Life Cycle Assessment of Road - A Pilot Study for Inventory Analysis

Products: Asphalt, Concrete, Stone etcCountries: SwedenLife Cycle Stages: Cradel to GraveYear of Study: 2001Availability of report:

Free on the web

Quality of data: GoodTitle: A LIFECYCLE ASSESSMENT AND EVALUATION OF

CONSTRUCTION AND DEMOLITION WASTEProducts: C&D materialsCountries: UKLife Cycle Stages: Life cycle InventoryYear of Study: ?Availability of report:

Free on the web

Quality of data: Good

Title: Background Document for Life-Cycle Greenhouse Gas Emission Factors for Clay Brick Reuse and Concrete Recycling

Products: Bricks and ConcreteCountries:Life Cycle Stages:Year of Study: 2003Availability report:Quality of data:Title: LCA of ConcreteProducts: Ordinary & frost resistant concreteCountries: SwedenLife Cycle Stages: Whole life cycle except use phase.Year of Study: 2005Availability of report:

Free on the Internet

Quality of data: Masters Thesis

Title: Life Cycle Analysis of High Quality Recycled AggregateProduced by Heating and Rubbing Method

Products: AggregatesCountries: JapanLife Cycle Stages: Full cycle assessmentYear of Study: ?Availability report: Need to search for full reportQuality of data: Good

Title: LC environmental and economic assessment of using recycled materials for Asphalt Pavements

Products: AsphaltCountries: USALife Cycle Stages: Full life cycleYear of Study: 2003Availability of report:

Free on the internet

Quality of data: Good

Fertiliser

Title: Damage Costs of Nitrogen Fertiliser and Their Internalization

Products: FertilizerCountries: Bulgaria/UKLife Cycle Stages: ?Year of Study: 2004Availability of report:

Free on the internet

Quality of data: Good

Iron & Steel

Title: The industrial ecology of steelProducts: SteelCountries: USALife Cycle Stages: Cradle to GraveYear of Study: 2000Availability report: Free on the InternetQuality of data: Good

Title: IISI Life Cycle Inventory Study for Steel Industry Products

Products: Hot rolled coil (with and without pickling), cold rolled coil (with and without finishing), hot dip and electrically galvanised sheet, painted sheet, tinplate and tin-free sheet, tubes, sections, plate, rebar/wire rod, and engineering steels.

Countries: ?Life Cycle Stages: Cradle to GraveYear of Study: 2000Availability report: From World steel organisation

http://www.worldsteel.org/?action=storypages&id=112 Quality of data: Good

Nobel House17 Smith SquareLondon SW1P 3JR

www.defra.gov.uk