Such a shift would mean a substantial deviation from the set-up of LCA, where environmental aspects are specified and evaluated independently from economic aspects. PROSPECTS As outlined, recent work has been concerned not so much with the specifics of allocation as with extending inventory modeling to overcome the traditional limitations of ‘black-box’ processes with fixed input-output coefficients, with subsequent implications for allocation constituting our prime interest here. In the coming years model refinement is likely to continue along the same three main lines of investigation, viz. inclusion of technical relations, aims of process operators, and market relations. Part 3 : Scientific background


Allocation based on economic value can be applied systematically and needs practical application to see how operational problems can be solved. Further specification of technical relations is to be expected, one possible source of impetus being incorporation of some of the modeling tools used in engineering design. Most engineering models are concerned with short-term relations, however, based on short-term physical and other causalities. If such models could be extended to incorporate variations in installed capacities endogenously, they might be extremely useful in LCA. This kind of modeling tool could then effectively be used to examine alternative investment options. It would certainly reinvigorate the still rather open debate on which processes to include in the analysis, a choice essentially between marginal, non-constrained Including current state-ofthe-art) processes

Guinée et al., Handbook on LCA: Operational Guide to the ISO Standards (2002)


Knowledge demands would be enormous, however, and would oblige LCA practitioners to keep abreast of innovations in each and every process involved. In addition, investment functions are hardly a technical affair only, for investment decisions are based primarily on economic considerations. If such models are to be operationalised for use in LCA, therefore, they must be extended to include the aims of process operators. This also holds for combined waste treatment processes. Assuming (steady-state) functioning of such a process at some optimum level, long-term analysis of some additional amount of waste to be processed would indicate a need to install new capacity. Establishing the effects of processing the waste would then involve a comparison of the system flows with and without the functional unit. Economic considerations play a role in waste processing, too, albeit somewhat less well-defined


Economic considerations play a role in waste processing, too, albeit somewhat less well-defined. Here too, then, analysis shifts into the second group, with operators' aims being included in the inventory model. Overall, then, it is to be concluded that purely technical model specifications do not facilitate LCA modeling to any great degree. Even if such types of model became available for long-term analysis of waste management, they would still not resolve the allocation problem. Incorporating the goals of process operators in models, as a first step towards market modeling, seems feasible in situations where one or just a few process operators dominate the system. Given some additional unit of function, the new optimum for the system could be specified, under the simplifying assumption of constant prices. Such models are quite common in operations research but are then concerned with short-term rather than long-term optimisation


Such models are quite common in operations research but are then concerned with short-term rather than long-term optimisation. Such optimisation models could to some extent resolve the multifunctionality problem, in cases of combined but not joint processes. Optimising from a collective point of view, as in minimising social cost, is not empirical modeling but may be useful in specifying attractive alternatives, which should be subsequently investigated by means of inventory modeling. Incorporating the full complexity of market relations interlinking all the processes in the system is not operationally feasible at the level of technological detail required in LCA-type decision situations. If applied to the main processes only, it may become feasible. However, it doubtful whether allocation makes sense as a subsequent step, after modeling, as substantial information on market effects would then be lost. Without allocation we would leave the realm of functional unit-based LCA


Without allocation we would leave the realm of functional unit-based LCA. CONCLUSIONS We now return to our original question: how to model the system and how to resolve the resultant multifunctionality problems? In order to maintain a clear relation to ISO 14041 certain modeling aspects must be included, as can be seen above. The basic principles to be adopted in inventory modeling are summarised in Table 3.9.3. See further explanations in Sections 1.1, 2.3 and 3.1. What starting points are there as requirements for allocation? The first main principle, a basic requirement of the allocation procedure, is that the sum of the allocated values of all product flows equals the total value of the flows of the unallocated process: the ‘100% rule’. Optimisation models do not generally satisfy this requirement


Optimisation models do not generally satisfy this requirement. Allocation does not involve the principle of mass balance, which may hold for the unallocated, multifunctional process, but not for the monofunctional processes resulting from allocation. As most process descriptions do not include the consumption of oxygen or any other form of respiration, even unallocated processes do not in fact massbalance. A second principle is that allocation should be at the level of multifunctional unit processes only. It is at that level that the clearest view is possible on technical relations, and on the values of the coproducts. A third, more methodological principle relates to reasonableness. In particular, results should be unaffected by the sequence of application of the allocation procedure (cf. Sen, 1969), i.e. which of the coproducts is taken as the initial point of departure. Also, if processes are first allocated and then placed in Part 3: Scientific background


the system scaled to the functional unit, the result should be the same as when they are first scaled, prior to the allocation step. The fourth requirement is that the same principles should be applied to all the categories of coproduction: combined and joint production, combined waste processing, re-use and recycling. The definition of the multifunction problem as modeled does not distinguish different types of multifunctionality, as the term ‘product’ covers both goods and services, allocation is necessary if more than one product is produced by a unit process. Waste having a negative value is not a product; nobody buys it. Processing waste is a service provided to the party supplying the waste, to be allocated to the products of the process producing them. That service relieves the producer of his waste; he pays for that service, being a product


That service relieves the producer of his waste; he pays for that service, being a product. Hence, the different situations for combined production, co-production, combined waste handling and recycling do not differ in their basic definition. In each case the same allocation principles should hence apply. A fifth, more practical principle is that multifunctional processes associated with more than one of the product alternatives being compared should be allocated consistently. If different options for allocation are available, the same option should be applied to all the alternatives compared. Otherwise, any resultant differences between alternatives would be due both to real differences and to differences in the allocation methods applied. Of course in that situation a sensitivity analysis on the different options is due, to see how outcomes differ if one option for allocation is applied to all alternatives, or the other. Table 3.9.3: Inventory modeling and allocation principles


Table 3.9.3: Inventory modeling and allocation principles. Resumé of modeling principles use fixed input-output coefficient process definitions wherever possible detail processes until a truly multifunctional core is reached (ISO 1a) include all processes implied in fulfilling the function definesystem boundarieswhereflows are being paidforbyotherproductsystems define system boundaries where environmental interventions enter or leave the system do not cut off flows that cannot be specified further but estimate them, e.g


using environmentally extended input-output analysis do not use market mechanisms in detailed LCA, but possibly in extended LCA make a clear distinction between system modeling and allocation allocation principles ‘100% rule’ allocate at the level of unit processes only independence of sequence of application consistent principles for all varieties of co-production consistent allocation procedure and results for multifunctional processes associated with different product alternatives The guidelines for dealing with the multifunctionality problem have been elaborated primarily on the basis of economic allocation, in both an extended and a simplified version. These guidelines were codeveloped with Lindeijer, and their background and more detailed formulations are elaborated in Appendix C. As the substitution method is regarded by some as a promising approach to the multifunctionality/allocation problem, we here present its best elaboration as an option for extended LCA


As the substitution method is regarded by some as a promising approach to the multifunctionality/allocation problem, we here present its best elaboration as an option for extended LCA. It does not constitute a recommended method, however, for the theoretical reasons discussed above and also because it is not entirely clear how operational choices for long-term change-oriented LCA are to be reasoned and made in the substitution approach. The method is illustrated with an example elaborated by Weidema for this Guide (see Part 2b, Section 3.9.3.2). Economic allocation is defined as partitioning of all non-product inputs and outputs proportionally to the share of each product from the unit process in total proceeds, see Figure 3.9.1 below. Part 3 : Scientific background


The steps in inventory modeling to avoid the multifunctionality problem, and the solutions to the multifunctionality problem through allocation are now surveyed. The framework for this survey is the two levels of LCA distinguished in this Guide, simplified and detailed analysis, with additional options for extensions. The latter are not in line with the starting points adopted here, however. The relationship of these steps to the framework of ISO 14041 is indicated, if they are included there. As ISO does not distinguish between simplified and a detailed LCA, some ISO elements are in either option, or ISO steps occur in different versions. Also, as ISO now does not fully distinguish between modeling as a possible cause of the multifunctionality problem and the available solutions to that problem, some steps are part of modeling in the Inventory analysis and some in the solution, through some sort of allocation


In due course such further differentiation might be incorporated in revisions of ISO 14040 and 14041. The respective steps to be followed are summarised in general terms below, showing their equivalence, or otherwise, to the steps provided in ISO 14041 (Table 3.9.4). The steps ‘economic allocation’, in detailed and simplified LCA, and ‘symmetrical substitution method’ (termed ‘avoiding allocation’ by Weidema, 2001), as an option for extension, are elaborated in greater detail in Volume 2b. Table 3.9.4: Recommended procedure for handling multifunctionality in detailed version inventory dm1. Divide processes modeling that are not really multifunctional into monofunctional unit processes (ISO step 1a) dm2. For waste processing: model longtermtechnical relations, assuming simple aims of process operators (ISO step 2, extended) allocation da1. Apply economic allocation based on market value or constructed marketvalue (ISO step 3, interpreted economically) simplified version sm1


Apply economic allocation based on market value or constructed marketvalue (ISO step 3, interpreted economically) simplified version sm1. Use made-single databases(correspondence with ISOdifferentperdatabase and process) or use environmentally extended inputoutput data sm2.As a proxy, treat openloopco-production asclosedloop co-production (after ISO), with quality adjustment sm3. Use available waste management models based on physical causality (ISO step 2, simplified) sa1. Apply economic allocation where readily feasible, if not at unit process level then at full system level sa2. Use substitution with already made-single cradle-togate database data (ISO step 1b, simplified) sa3.Applythe most readily available physical product parameter related to value, e.g. mass, volume, energy content (not in line with ISO) sa4. For all remaining multifunctional processes: use some 'quick-and-dirty' measure: e.g


mass, volume, energy content (not in line with ISO) sa4. For all remaining multifunctional processes: use some 'quick-and-dirty' measure: e.g. forrecycling,'50% reduction of primary production' ( not generally in line with ISO) inventory modeling and allocation. options for extension em1. Apply market analysis to main co-products, as a realistic system expansion (after ISO step 1 b, extended) em2. Set up a linear programming model of the main processes if operated by a single organisation (no ISO correspondence) ea1. Applythe symmetrical substitution method (further interpretation and extension of ISO step 1b) Part 3: Scientific background


RESEARCH RECOMMENDATIONS Distinguishing more clearly between modeling and allocation opens new perspectives on improved inventory modeling. With this distinction clearer, the respective merits of LCA and other tools and models for empirical analysis of environmental impacts will be easier to clarify and hence improve. Modeling internal relations in processes from a long-term perspective is a challenge. This will require systems modeling to some degree and requires the explicit introduction of the aims of process operators. By incorporating certain economic mechanisms in the inventory model, particularly in cases involving extremely high or low elasticities, inventory modeling might be made more realistic and some of the principal defects of LCA redressed. Developing parallel models for decision support (e.g. LCA, market modeling, SFA, CBA), each indicating a key effect mechanism, is seen as a more fruitful route than incorporating ever more mechanisms in LCA


LCA, market modeling, SFA, CBA), each indicating a key effect mechanism, is seen as a more fruitful route than incorporating ever more mechanisms in LCA. If such parallel modeling options become available, interpretation of results cannot be restricted to LCA but will involve the full set of available analytical tools. In this connection, research is specifically recommended on how LCA relates to Cost Benefit Analysis (CBA), which avoids allocation problems by accepting inter-system differences in function and expressing these in terms of monetary value. The relative merits of these two methods are as yet unclear. It is recommended, furthermore, to deepen understanding of the precise relationship between system boundaries and allocation, the definitions of which are closely related; Research is also recommended on how substitution can be more systematically integrated in inventory modeling than is currently feasible


3.10 Calculation method TOPIC Collection of process data yields a database of processes. The act of quantitatively relating these processes to one another, scaled to the reference flow following from the functional unit, is referred to here as the calculation method. The calculation result is a set of linked and scaled processes, each with scaled environmental interventions, which are usually aggregated. DEVELOPMENTS IN THE LAST DECADE In ISO 14041 (1998E) several steps are distinguished under the heading of calculation procedures: validation of data; relating data to the unit process; relating data to functional unit and data aggregation; refining the system boundaries. Heijungs et al. (1992) In the 1992 guide scaling of process data, there termed “Creating the inventory table”, consisted of two steps (pp. 37–40): “quantification of the environmental interventions” and “representation of the qualitative environmental interventions”


37–40): “quantification of the environmental interventions” and “representation of the qualitative environmental interventions”. Two calculation procedures were discussed in the 1992 Backgrounds document: the sequential method and the matrix method. The first is still widely used. However, with this method there is no quick and accurate way of dealing with processes which are mutually related (recursion). The matrix method allows feedback to be dealt with immediately. The remarks on data validation in ISO 14041 have already been cited in Section 3.5. Relating data to unit process has been discussed in this Guide under the heading of data collection (Section 3.6). Part 3 : Scientific background


On relating data to the functional unit and data aggregation ISO 14041 clause 6.4.4 states:: Based on the flow chart and system boundaries, unit processes are interconnected to allow calculations on the complete system. This is accomplished by normalizing the flows of all unit processes in the system to the functional unit. The calculation should result in all system input and output data being referenced to the functional unit. Care should be taken when aggregating the inputs and outputs in the product system. The level of aggregation should be sufficient to satisfy the goal of the study. Data categories should only be aggregated if they are related to equivalent substances and to similar environmental impacts. If more detailed aggregation rules are required, they should be justified in the goal-and-scope-definition phase of the study or should be left to a subsequent impact-assessment phase


If more detailed aggregation rules are required, they should be justified in the goal-and-scope-definition phase of the study or should be left to a subsequent impact-assessment phase. (Source: ISO 14041, 1998E) ISO 14041 (1998E) clause 6.4.1 states (see textbox): When determining the elementary flows associated with production of electricity, account shall be taken of the production mix and the efficiencies of combustion, conversion, transmission and distribution. The assumptions made shall be clearly stated and justified. Whenever possible, the actual production mix should be used in order to reflect the various types of fuel that are consumed. Inputs and outputs related to a combustible material, e.g. oil, gas or coal, can be transformed into an energy input or output by multiplying it by the relevant heat of combustion. In this case it shall be reported if the higher heating value or the lower value is used


In this case it shall be reported if the higher heating value or the lower value is used. The same calculation procedure should be consistently applied throughout the study. [...]. All calculation procedures shall be explicitly documented. Source: ISO 14041, 1998E. A distinction between aggregated and non-aggregated inventory results may be useful. In the “Contribution analysis” part of the Interpretation phase, one of the recommendation is to determine the contribution of each process to the total inventory results, for instance. This implies that options enabling such analysis must be provided in the calculation step. ISO sets no restrictions on calculation methods. Currently available methods include matrix inversion (Heijungs et al.,1992; Möller, 1992; Heijungs & Frischknecht, 1998), (simultaneous) sequential calculation of the inputs and outputs of each unit process of the system, with or without a number of iterations, and linear programming


As most LCA studies to date have not specified the calculation methods employed, there may be even more methods in use. The main difference between these different calculation methods concerns the handling of loops, e.g. in the (frequently occurring) case that coal is needed for electricity production and electricity is needed for the extraction of coal. Such loops can be handled by matrix inversion techniques, but cannot be dealt with appropriately by most sequential calculation methods. In the latter the simplification is made to cut out loops altogether, by setting certain input flows to zero.1 Although this is a very common approach, it is generally left implicit. LCA studies are generally performed using dedicated LCA software programs. Several programs are available; for a software review see e.g. Rice (1996), Rice et al. (1997), Menke et al. (1996) and Siegenthaler et al. (1997)


Several programs are available; for a software review see e.g. Rice (1996), Rice et al. (1997), Menke et al. (1996) and Siegenthaler et al. (1997). The ISO text on refining the system boundaries focuses on sensitivity analysis and may be regarded as an issue for Interpretation, here treated in Section 5.6. The result of the calculation step is termed the inventory table, which comprises all the environmental interventions related to the reference flow (of reference process) specified in the goal and scope of the study (see Figure 2.4.1). Furthermore, all economic flows not followed to the system boundary should be reported directly below the inventory table. PROSPECTS As mentioned in Section 3.6, LCA software should preferably allow scaling of data, including a time dimension. On this topic no further specific developments are foreseen


PROSPECTS As mentioned in Section 3.6, LCA software should preferably allow scaling of data, including a time dimension. On this topic no further specific developments are foreseen. 1 Sequential calculation techniques can, in principle, handle loops by iterating a given number of times, but most LCA software programs do not include this type of sequential calculation method. Part 3: Scientific background


CONCLUSIONS We conclude that the best available practice with respect to calculation methods is matrix inversion, although this is not included in most software. If a different calculation method is employed, its precise nature should be specified as well as the differences from matrix inversion. RESEARCH RECOMMENDATIONS LCA software should preferably allow scaling of data, including a time dimension. Part 3 : Scientific background


4. Impact assessment 4.1 General introduction According to ISO 14040 (1997E) Life Cycle Impact Assessment (LCIA), the third phase of life cycle assessment, “is aimed at understanding and evaluating the magnitude and significance of the potential environmental impacts of a product system”. In a similar vein, the first SETAC-Europe Working Group on Impact assessment (WIA–1) defines LCIA as “a quantitative and/or qualitative process to identify, characterise and assess the potential impacts of the environmental interventions identified in the Inventory analysis” (Udo de Haes (ed.), 1996). To this end the individual data of the inventory table, or LCI results, are translated into contributions to selected impact categories, such as ‘depletion of abiotic resources’, ‘climate change’ or ‘acidification’. These contributions are calculated using characterisation models, in which relevant environmental processes are modeled to a so-called category endpoint


These contributions are calculated using characterisation models, in which relevant environmental processes are modeled to a so-called category endpoint. To aggregate the results for each category, these are expressed in terms of a common parameter called a category indicator: ‘infrared radiative forcing’ for climate change, for example. ISO 14042 (2000E) puts it as follows: “The LCIA phase models selected environmental issues, called impact categories, and uses category indicators to condense and explain the LCI results. Category indicators are intended to reflect the aggregate emissions or resource use for each impact category. These category indicators represent the ‘potential environmental impacts’ discussed in ISO 14040. In addition, LCIA prepares for the life cycle Interpretation phase”. (See Figure 4.1.1)


These category indicators represent the ‘potential environmental impacts’ discussed in ISO 14040. In addition, LCIA prepares for the life cycle Interpretation phase”. (See Figure 4.1.1). ISO goes on to note that LCIA can assist in various decision situations, but that “parties should recognise that a complete product system assessment is difficult and may require the use of several different environmental Part 3: Scientific background


assessment techniques”. The latter means that it may sometimes be useful to apply other tools in addition to LC(I)A (see also Sections 2.4 and 2.5 of Part 1 of this Guide and Appendix B of this Part). As elaborated in this Guide, Impact assessment comprises eight steps (see Section 1.4): Procedures (no special section in this Part; see Chapter 1); Selection of impact categories (Section 4.2, p. 529); Selection of characterisation methods: category indicators, characterisation models (Section 4.3, p. 538); Classification (Section 4.4, p. 622); Characterisation (Section 4.5, p. 624); Normalisation (Section 4.6, p. 625); Grouping (Section 4.7, p. 628); Weighting (Section 4.8, p. 630). In elaborating these Impact assessment steps a further point of departure was ISO 14042 (2000E), with respect to the methodological framework and the issues for Impact assessment proposed there


In elaborating these Impact assessment steps a further point of departure was ISO 14042 (2000E), with respect to the methodological framework and the issues for Impact assessment proposed there. Here the ISO proposals have been further operationalised taking into account the work of WIA–1 (Udo de Haes (ed.), 1996 and Udo de Haes et al., 1999) and relevant proposals by other authors. Deviations from ISO have been introduced only in cases where there are significant arguments for doing so. In discussing the substance of the eight steps as well as the individual impact categories themselves, a fixed format has generally been employed below: ‘Topic’, ‘Developments in the last decade’, ‘Prospects’, ‘Conclusions’ and ‘Research recommendations’. Where this format was less appropriate, as with the rubrics ‘Interventions ‘ and ‘Economic flows not followed to system boundary’, it has not been strictly adhered to


Where this format was less appropriate, as with the rubrics ‘Interventions ‘ and ‘Economic flows not followed to system boundary’, it has not been strictly adhered to. As explained in Section 1.5 the step dealing with procedures is not discussed separately in this chapter, but in an integrated manner for all the various phases in Section 1.3. We first provide a short resumé of the international organisations involved in work on Impact assessment, the main ISO requirements and the relation between ISO and SETAC, as a springboard for discussing the substance of the individual steps of LCIA. 4.1.1 International organisations involved Since 1992, two organisation in particular have been involved in work on LCIA: ISO; SETAC, including specifically the SETAC-Europe Working Group on Impact assessment (Udo de Haes (ed.), 1996) and the SETAC-US Work Group on Impact assessment (Barnthouse et al., 1997)


In 1993 the International Organization for Standardization (ISO) established a technical committee (TC 207) concerned with standardisation of a number of environmental management tools, including a subcommittee on LCA (SC5). One of the Working Groups formed under SC5 has dealt with LCIA (ISO/TC 207/SC 5 /WG 4). The main tasks were: to define concepts, to define a technical framework for LCIA and to specify general methodological requirements and procedural requirements. These include requirements for ‘comparative assertions disclosed to the public’, which are much stricter than for other (either internal or non-comparative) applications. This work resulted in the International standard on LCIA (ISO 14042, 2000E)


This work resulted in the International standard on LCIA (ISO 14042, 2000E). Through its North American and European branches, the Society of Environmental Toxicology and Chemistry (SETAC) has played a leading role in bringing LCA practitioners, users and methodology developers together to collaborate on the continuous improvement and harmonisation of LCA methodology. SETAC focuses on the scientific development of LCA methodology. In 1993 a ‘Code of Practice’ was published (Consoli et al., 1993). From then on specific activities were taken up by a series of working groups, including the SETAC-Europe Working Group on Impact assessment (WIA) and the SETAC North America Work Group on Impact assessment. Initial reports have meanwhile been published independently (Udo de Haes (ed.), 1996; Barnthouse et al., 1997)


Initial reports have meanwhile been published independently (Udo de Haes (ed.), 1996; Barnthouse et al., 1997). The European report ‘Towards a methodology for Life Cycle Impact assessment’ (Udo de Haes (ed.), 1996) provides an up-to-date, comprehensive review and critical analysis of existing thinking on LCIA. The principal methodological choices made in each phase of LCIA are analysed. The successor to this first WIA (WIA–2) has already started its work, again in Europe. The goal of WIA–2 is to establish a list of recommended impact categories, together with category indicators to be used in LCIA. This list should, as far as possible, be in line with ISO 14042, which, according to WIA–2, means it should have maximum scientific and technical validity. Furthermore, it should be practicable in terms of number of categories, characterisation methods and inventory data requirements (Udo de Haes et al., 1999)


Furthermore, it should be practicable in terms of number of categories, characterisation methods and inventory data requirements (Udo de Haes et al., 1999). The North American report is principally a critical review of the LCIA framework, scrutinising the accuracy of the various methods available and their applicability in various circumstances. Both the WIA and the SETAC Part 3 : Scientific background


North America Work Group have made significant inputs to the ISO process vis-à-vis development of an LCIA standard. A major new task is the cooperation between SETAC and UNEP on establishing ‘best available practice’ in the field of life cycle Impact assessment, as a follow-up to the work of WIA–2. More specifically, such practice is to be established with respect to (a list of) impact categories, and category indicators and characterisation factors for each of these. It is anticipated that an appropriate form of cooperative organisation will be defined for this task in the year 2001 and possibly also established. Besides these international developments and standardisation activities, there have been numerous proposals for defining impact categories, category indicators and models; these are discussed in Sections 4.2 and 4.3


4.1.2 ISO 14042 requirements ISO 14042 (2000E) describes procedures rather than specific methodologies or models for life cycle Impact assessment, implying that any methodology or model is acceptable as long as it satisfies the general ISO criteria. Figure 4.1.2.1 summarises the overall framework of LCIA, showing the relationship between life cycle inventory results, impact categories, category indicators and category endpoint(s), and illustrating these concepts with reference to the impact category ‘Acidification’. ISO 14042 also defines the term ‘environmental mechanism’: “a system of physical, chemical and biological processes for a given impact category, linking the LCI results to category indicators and to category endpoints”


In this Guide we have opted to refine this definition, to distinguish more categorically between the terms of the ‘real world’ and those of the characterisation models used to simulate certain partially underststood environmental processes.1 Thus, the term ‘environmental process’ has been adopted here for the chain of physical, chemical and biological events in the natural environment that link a particular environmental intervention to a particular impact; typical examples include pollutant accumulation or leaching. For a given impact category, these environmental processes then together form the ‘environmental mechanism’, which is modeled to a greater or lesser extent by the characterisation model, up to one or more category endpoints. 1 In preparing this Guide the ISO terminology proved to be rather inconsistent on a variety of points and for this reason we have here refined the definitions of certain ISO terms. In the Glossary these are specifically indicated


In the Glossary these are specifically indicated. Part 3: Scientific background


According to ISO 14042, category indicators may be chosen anywhere along the environmental mechanism between intervention and endpoint. Operationalisation of each of the chosen impact categories then comprises the following elements: identification of one or more category endpoints; definition of a category indicator for each of these endpoints; identification of the LCI results to be assigned to each category indicator, taking into account the selected category endpoint(s); and identification of the characterisation model and characterisation factors to be used. ISO 14042 states that this procedure “facilitates the collection, assignment, and modeling of appropriate LCI results” and “helps to highlight the scientific and technical validity, assumptions, value-choices and degree of accuracy in the model”


With respect to the selection of impact categories, category indicators and models, ISO sets the following requirements (see textbox): the selection of impact categories, indicators and models shall be consistent with the goal and scope of the LCA study the sources for impact categories, indicators and models shall be referenced; the selection of impact categories, indicators and models shall be justified; accurate and descriptive names shall be provided for the impact categories and category indicators; the selection of impact categories shall reflect a comprehensive set of environmental issues related to the product system being studied taking the goal and scope into consideration; the environmental mechanism and model which relate the LCI results and indicator as a basis for characterisation factors shall be described; the appropriateness of the use of the characterisation model for deriving the category indicator in context of the goal and scope of the study shall be described


In addition a number of recommendations is given for the selection of impact categories, indicators and models: the impact categories, indicators and models should be internationally accepted i.e


based on an international agreement or approved by an international body; the impact categories should represent the aggregated emissions or resource use of the product system on the category endpoint(s) through the indicators; value-choices and assumptions made during the selection of impact categories, category indicators and characterisation models should be minimised; the impact categories, category indicators and characterisation models should avoid double counting unless required by the Goal and scope definition, for example when the study includes both human health and carcinogenicity; the characterisation model for each category indicator should be scientifically and technically valid, and based upon a distinct identifiable environmental mechanism and/or reproducible empirical observation; the impact categories and indicators should be environmentally relevant; it should be identified to what extent the characterisation model and the characterisation factors are


empirical observation; the impact categories and indicators should be environmentally relevant; it should be identified to what extent the characterisation model and the characterisation factors are scientifically and technically valid


Depending on the environmental mechanism and the goal and scope, spatial and temporal differentiation of the characterisation model relating the LCI results to the indicator should be considered. The fate and transport of the substances should be part of the model. LCI results other than mass and energy flow data included in an LCA study, e.g. land use, shall be identified and their relationship to corresponding indicators shall be determined


land use, shall be identified and their relationship to corresponding indicators shall be determined. The environmental relevance of the category indicator or characterisation model should be clearly stated in terms of the following criteria: a) the ability of the category indicator to reflect the consequences of the LCI results in the category endpoint(s) at least qualitatively; b) the addition of environmental data or information to the characterisation model with respect to the category endpoint (S), including the condition of the category endpoint(s), the relative magnitude of the assessed change in the category endpoints, the spatial aspects, such as the area and scale, the temporal aspects, such as duration, residence time, persistence, timing, etc., the reversibility of the environmental mechanism, and the uncertainty of the linkages between the characterisation model and the changes in the category endpoints


Part 3 : Scientific background a) b) c) d) e) f) g) a) b) c) d) e) f) g)


4.1.3 Relation between ISO and SETAC Working Group on LCIA Since ISO 14042 describes no specific methodologies or models for use in Impact assessment, there is a need to specify LCIA methodologies and models that satisfy the broader ISO requirements. This is where the work of the SETAC-Europe Working Group on Impact assessment comes in, for WIA–2 is seeking to draw up an authorised, recommended list of impact categories complete with category indicators and characterisation factors. Proceeding from the ISO requirements, WIA–2 aims to establish a best available practical method for each impact category distinguished. At the time of writing, the work had yielded a list of recommendations for individual impact categories as well as an overall framework for these categories (Udo de Haes et al., 1999). As stated in Chapter 1, the aim of this new Guide is to operationalise the ISO standards, and in particular ISO 14042, by updating and expanding the Guide of Heijungs et al


As stated in Chapter 1, the aim of this new Guide is to operationalise the ISO standards, and in particular ISO 14042, by updating and expanding the Guide of Heijungs et al. (1992) to incorporate all relevant developments since publication of the latter, taking the ISO standards as the basic point of departure and with particular reference to ongoing work within the SETAC community. For the LCIA phase, this means that the work of WIA–2 (Udo de Haes et al., 1999) has been taken as a starting point. WIA–2 was established in the knowledge that its mission would be taken over at the global level by the scheduled cooperation between SETAC and UNEP. The WIA–2 work on best available practice has not yet been completed and this Guide can therefore do no more than make recommendations on best available practice as understood at the present time


The WIA–2 work on best available practice has not yet been completed and this Guide can therefore do no more than make recommendations on best available practice as understood at the present time. Thus, the present Guide goes beyond the work of WIA–2, taking into account as far as possible all relevant developments in LCIA during the last decade. 4.2 Selection of impact categories TOPIC In the Impact assessment phase the results of the Inventory analysis are translated into contributions to relevant impact categories, such as depletion of abiotic resources, climate change, acidification, etc. To this end, relevant impact categories must be identified. The text of ISO 14042 might be interpreted as indicating that these impact categories are to be defined anew for each study


To this end, relevant impact categories must be identified. The text of ISO 14042 might be interpreted as indicating that these impact categories are to be defined anew for each study. To facilitate the work of practitioners a default list of impact categories has here been elaborated, thereby distinguishing between 'baseline' impact categories, 'study-specific' impact categories and 'other' impact categories. In this step of the LCA, then, practitioners are still obliged to select those categories relevant to the goal of his or her particular study, supported by the preliminary selection made in this Guide. DEVELOPMENTS IN THE LAST DECADE ISO 14042 does not provide a default list of impact categories for inclusion in LCIA. The starting points of ISO and WIA–2 regarding the selection and definition of impact categories are summarised in Table 4.2.1. For the underlying argumentation the reader is referred to ISO 14042 and Udo de Haes et al


For the underlying argumentation the reader is referred to ISO 14042 and Udo de Haes et al. (1999), respectively.1 Table 4.2.1: Starting points for definition and selection of impact categories in general and for a specific LCA study. General starting point for the framework of impact categories and category indicators: a framework shall be developed which is open to further scientific progress and further detailing of information (WIA–2) General starting points for the total set of impact categories: 1. the categories shall together permit an all-encompassing assessment of relevant impacts, as currently understood (completeness) (ISO/WIA–2) 2. the categories should have minimum overlap and avoid double counting unless so required by the goal and scope (ISO/WIA–2) 3. the categories should be internationally accepted, i.e. based on an international agreement or approved by a competent international body (ISO) 4


the categories should be internationally accepted, i.e. based on an international agreement or approved by a competent international body (ISO) 4. the total number of impact categories should not be too high (WIA–2) Starting points for the selection of categories in a specific LCA study: 1. the selected impact categories shall be consistent with the goal and scope of the LCA study (ISO) 2. the selected impact categories shall form a comprehensive set of environmental issues related to the goal and scope of the LCA study (ISO) 1 Furthermore, ISO sets several requirements regarding the description and documentation of categories. Part 3: Scientific background


Choice of overall impact assesment method Within the ISO framework the environmental profile resulting from the characterisation step is an important LCA result in its own right, with the grouping and weighting steps (which are more value-based) constituting distinct, optional elements. Some existing methods deviate in this respect, however, weighting interventions directly. One example is the Ecopoints method, in which emissions and extractions are weighted using a distance-to-target method, i.e. based on policy targets (Ahbe et al., 1990). Consequently, this method does not include a separate characterisation step. A second major defining aspect of Impact assessment methods is the point in the environmental mechanism at which the category indicators are defined. They may be defined close to the intervention (the midpoint, or problem-oriented approach, e.g. Heijungs et al., 1992; Udo de Haes (ed.), 1996; Haas, 1997; Wenzel et al., 1997; Beetstra, 1998; Udo de Haes et al., 1999)


Heijungs et al., 1992; Udo de Haes (ed.), 1996; Haas, 1997; Wenzel et al., 1997; Beetstra, 1998; Udo de Haes et al., 1999). Alternatively, they may be defined at the level of category endpoints (the endpoint, or damage approach, e.g. EPS: Steen & Ryding, 1992; Steen, 1996; ExternE: EC, 1995a; Eco-indicator 99: Goedkoop & Spriensma, 1999). A cluster of category endpoints of recognisable value to society is referred to as an “area of protection”. Here, we distinguish four: human health, natural resources, the natural environment and the man-made environment. Despite its name, the midpoint approach still allows definition of category indicators anywhere along the environmental mechanism in question (cf. Udo de Haes et al., 1999), including the endpoint level. This permits use of the best indicator available for each impact category, regardless of where it is located in the environmental mechanism


Udo de Haes et al., 1999), including the endpoint level. This permits use of the best indicator available for each impact category, regardless of where it is located in the environmental mechanism. If indicators are chosen at midpoint level their relationship with the category endpoints should be clearly defined. This will generally be in qualitative terms. In ISO 14042 (2000E) the degree of linkage between the chosen category indicator and the category endpoint is referred to as the “environmental relevance” of the indicator. The environmental relevance is to be clearly stated in terms of the following criteria: 1. the ability of the category indicator to reflect the consequences of the LCI results on the category endpoint(s), at least qualitatively; 2


the ability of the category indicator to reflect the consequences of the LCI results on the category endpoint(s), at least qualitatively; 2. the addition of environmental data or information to the characterisation model with respect to the category endpoint(s), including the condition of the category endpoint(s), the relative magnitude of the assessed change in the category endpoint(s), the spatial aspects, such as area and scale, the temporal aspects, such as duration, residence time, persistence, timing, etc., the reversibility of the environmental mechanism, and the uncertainty of the linkages between the characterisation model and the changes in the category endpoints (ISO 14042, 2000E). In the second case category indicators are defined at endpoint level. The advantage of this approach is that the environmental relevance of the category indicators is high: it is this level which ultimately matters to society and which enables a direct link to be made with weighting methods


However, these kinds of endpoint indicators are still under development and the associated models do not yet include all the relevant effects of common interventions (see textbox below). Since 1992 considerable efforts have been devoted to developing models and category indicators based on the endpoint approach (e.g. EC, 1995a; Goedkoop & Spriensma, 1999; Müller-Wenk, 1997; Hofstetter, 1998). The most comprehensive and recent work in this area is the Eco-indicator 99 by Goedkoop & Spriensma (1999). This approach is reviewed in the textbox and compared with the midpoint, i.e. problem-oriented approach. Part 3 : Scientific background


In the Netherlands two Impact assessment methods have been developed in the last decade, both grounded in the ‘environmental themes’ formulated by the Dutch Government in 1989 (VROM, 1989; RIVM, 1991). Both have the same basic structure, with the indicator results obtained by multiplying the inventory results by the appropriate characterisation factor together forming the so-called environmental profile, which is then normalised (see section 4.6), before serving as input for a possible weighting step. Where the two methods vary is with respect to the characterisation models and characterisation factors developed and proposed for the individual themes. In terms of their operationalisation there are also several clear differences between the methods. The first method, often referred to as the problem-oriented approach and first presented by Heijungs et al


The first method, often referred to as the problem-oriented approach and first presented by Heijungs et al. (1992), operationalised models and characterisation factors for a number of impact categories, but did not operationalise the weighting step. The second Dutch method is the Eco-indicator approach, developed primarily for the purposes of ‘eco-design’. Designers were deemed unable to work with 10–20 indicator results, and the Eco-indicator therefore employs only 1 to 3 weighted indices. Thus, there is greater emphasis on weighting than in the approach of Heijungs et al . In the first version of the Eco-indicator (Eco-indicator 95; Goedkoop, 1995) weighting was based partly on a damage approach, partly on a distance-to-target approach (i.e. based on predefined damage targets). Most of the impact categories identified were adopted from Heijungs et al., although the two toxicity themes were defined rather more narrowly


based on predefined damage targets). Most of the impact categories identified were adopted from Heijungs et al., although the two toxicity themes were defined rather more narrowly. Originally conceived as an experiment, the Eco-indicator method has since been improved. In the latest version (Eco-indicator 99; Goedkoop & Spriensma, 1999) a completely different approach to Impact assessment has been adopted in which a limited number of damage categories are weighted (by a panel, for example). Three types of damage are distinguished, for which weighting is taken to be more readily feasible: damage to resources; damage to ecosystem quality; damage to human health. As in the problem-oriented approach, the natural sciences are used to calculate the relation between the impacts of a (product) system’s life cycle and the resultant damages


As in the problem-oriented approach, the natural sciences are used to calculate the relation between the impacts of a (product) system’s life cycle and the resultant damages. The Eco-indicator methodology thus consists of two parts: scientific calculation of the three forms of damage due to the life cycle of the product under study; a valuation procedure to establish the significance of these damages. Part 3: Scientific background


The method has a modular structure (Figure 4.2.1; source: Goedkoop, 1997) in which the building blocks of the natural science component can be modified or replaced to reflect different value systems (viz. Egalitarian, Individualist, Hierarchist). The authors recommend using the Hierarchist version of the model as the default method, with the other two being run as a form of sensitivity analysis (Goedkoop & Spriensma, 1999). In the Eco-indicator 99 approach, ‘damage to health’ is operationalised using the notion of DALYs: DisabilityAdjusted Life Years. This indicator is said to measure “the total amount of ill health, due to disability and premature death, attributable to specific diseases and injuries. The DALY concept thus compares time lived with disability (YLD: Years Lived Disabled) ands time lost due to premature mortality (YLL: Years of Life Lost). Health is simply added across individuals


The DALY concept thus compares time lived with disability (YLD: Years Lived Disabled) ands time lost due to premature mortality (YLL: Years of Life Lost). Health is simply added across individuals. That is, two people each losing 10 years of disability-free life are treated as the same loss as one person losing 20 years” (Goedkoop & Spriensma, 1999). For the technical details of the DALY concept, the reader is referred to (Box 4.1) of this publication. Both the problem-oriented approach and the Eco-indicator approach conform to the ISO 14042 (2000E) framework, as reflected in ISO/TS 14047 in prep.), since both clearly distinguish the characterisation and weighting steps. Although the Eco-indicator 99 approach is very promising and certainly appealing as an avenue for further research, the problem-oriented approach is currently considered the ‘best available practice’ for Impact assessment and has therefore been adopted in this Guide


The Eco-indicator method still has several serious shortcomings. It includes far fewer inventory items and provides only very limited coverage of human-toxic impacts (carcinogeneity only, thus ignoring a wide range of other health impacts). Some of the constituent models are outdated compared with those now used in the problem-oriented approach, while others involve major uncertainties. Thus, the terrestrial acidification and eutrophication models are based on the local, Dutch situation, while the problem-oriented approach now uses a European model (Huijbregts, 1999b); the data and assumptions of the toxicity model can be improved (cf. Huijbregts 1999a); and linkage of GWP and ODP and other universally accepted factors to damage parameters is still very incomplete and uncertain


Huijbregts 1999a); and linkage of GWP and ODP and other universally accepted factors to damage parameters is still very incomplete and uncertain. Finally, aggregation of ecotoxicological impacts with eutrophying, acidifying and land use impacts is still very preliminary, and the ecosystem impacts of climate change, increased UV radiation and photochemical smog are not yet included. The key feature of the problem-oriented approach is that the category indicators are defined at midpoints along the environmental mechanism, congruent with current environmental policy themes, and can therefore be modeled relatively accurately. The approach has the added advantage of permitting flexible choice of characterisation model and position of category indicator in the environmental mechanism, since for many impact categories more than one model is defensible and available


However, midpoints (wherever their precise position in the mechanism) are a difficult input for weighting and in the problem-oriented approach there is thus still no set of weighting factors covering all impact categories (see section 4.8). The main advantage of the Eco-indicator 99 is that category indicators are defined at the endpoint level, giving them greater environmental relevance. As it is this level that ultimately matters to society, the object of the weighting procedure is more immediate. The major uncertainties associated with modeling from midpoints to endpoints constitute a serious drawback, however. It would be very useful to collaborate and examine whether the advantages of the Eco-indicator 99 and the problem-oriented approaches can be combined into a still better and more comprehensive Impact assessment methodology. Part 3 : Scientific background (


Although the Eco-indicator 99 approach is very promising and is certainly appealing as an avenue for further research, the problem-oriented approach is currently deemed the ‘best practice’ for Impact assessment and has therefore been adopted in this Guide. As discussed in the text box, the former approach does not employ reliable endpoint indicators for all relevant impacts and has a number of other shortcomings, and cannot yet therefore be used to generate a comprehensive environmental profile. As things stand, therefore, we here recommend the problem-oriented approach, with impact categories defined at the midpoint level. This allows the best available indicator to be used for each impact category, regardless of where in the environmental mechanism between intervention and endpoint this category indicator is defined


This allows the best available indicator to be used for each impact category, regardless of where in the environmental mechanism between intervention and endpoint this category indicator is defined. In the future, when more complete endpoint indicators are available, an approach based on categories defined at endpoint level, such as the Eco-indicator 99, may well become the preferred approach (see also textbox). In this Guide, the problem-oriented (midpoint) approach will now be elaborated further as a baseline. The Eco-indicator 99 approach can be used for the purpose of sensitivity analysis, since it is presently the most comprehensive method based entirely on endpoint indicators. The models used to calculate the indicator results for the various impact categories are more up to date and complete than comparable foreign approaches such as the ExternE and Environmental Priority Strategies, or EPS, methods (EC, 1995a; Steen, 1993 & 19961)


By using the Eco-indicator 99 approach as a sensitivity analysis, LCA practitioners can familiarise themselves with the kind of results yielded by this approach and compare them with the results obtained with the problem-oriented approach adopted in this Guide as a baseline. Choice of impact categories In elaborating the problem-oriented approach, a default list of impact categories first needs to be defined. Table 4.2.2 presents such a list, based mainly on the work of the WIA–2 Working Group on Impact assessment and earlier work (Udo de Haes (ed.), 1996; Udo de Haes et al., 1999)2, on more recent developments in the LCA field and, of course, on the basic starting points of Table 4.2.1. This default list acknowledges three groups of impact category. Group A: ‘Baseline impact categories’ comprises those of the categories distinguished and discussed in Udo de Haes et al. (1999) for which a baseline characterisation method3 is selected below, in Section 4.3


(1999) for which a baseline characterisation method3 is selected below, in Section 4.3. Group A impact categories are included in almost all LCA studies. Two revisions have been introduced relative to Udo de Haes et al. (1999): ‘extraction’ of abiotic resources is now ‘depletion’ of abiotic resources, the latter term being more impact-oriented; and the category ‘ecotoxicity’ has been broken down into five subcategories, three of which are included in Group A: freshwater aquatic, marine and terrestrial ecotoxicity. Group B: ‘Study-specific impact categories’ comprises categories that may merit inclusion, depending on the Goal and scope of the LCA study and whether appropriate data are available, and for which a baseline and/or alternative characterisation method is proposed in this Guide. Most of these categories are mentioned by Udo de Haes (ed.;1996) or Udo de Haes et al. (1999), although some have been defined more recently (e.g


Most of these categories are mentioned by Udo de Haes (ed.;1996) or Udo de Haes et al. (1999), although some have been defined more recently (e.g. the subcategories ‘freshwater sediment ecotoxicity’ and ‘marine sediment ecotoxicity’). Group C: ‘Other impact categories’ comprises the categories mentioned by Heijungs et al. (1992), Udo de Haes (ed. 1996) and/or Udo de Haes et al. (1999) for which no baseline characterisation method is proposed in this Guide. These impact categories require further elaboration before they can be used in LCA studies, with research still in progress. Desiccation, for instance, is an issue that is receiving considerable attention in the Netherlands (collaboration between KIWA and RIZA) and Australia (CRC)


Desiccation, for instance, is an issue that is receiving considerable attention in the Netherlands (collaboration between KIWA and RIZA) and Australia (CRC). Besides the impact categories distinguished below in Table 4.2.2, this Guide also identifies two additional rubrics: ‘Interventions for which characterisation factors are lacking‘ and ‘Economic flows not followed to system boundary’, discussed in Sections 4.3.17 and 4.3.18, respectively. These two rubrics should be included in the results of every LCA study. 1 A new version of the EPS method has recently been published (Steen, 1999). This version could not be evaluated in the present Guide, however. 2 Historically, the list of impact catepories has its origins in the ‘environmental themes’ defined in the Netherlands’ first National Environmental Policy Plan (VROM, 1989); it has since been revised and refined for the LCA (and other) purposes


3 A ‘characterisation method’ for an impact category comprises a category indicator, a characterisation model and characterisation factors derived from that model. Part 3: Scientific background


Table 4.2.2: Default list of impact categories and subcategories impact category A. Baseline impact categories Depletion of abiotic resources Impacts of land use land competition Climate change Stratospheric ozone depletion Human toxicity Ecotoxicity freshwater aquatic ecotoxicity marine aquatic ecotoxicity terrestrial ecotoxicity Photo-oxidant formation Acidification Eutrophication B. Study-specific impact categories Impacts of land use loss of life support functions loss of biodiversity Ecotoxicity freshwater sediment ecotoxicity marine sediment ecotoxicity Impacts of ionising radiation Odour malodourous air Noise Waste heat Casualties C


Other impact categories Depletion of biotic resources Desiccation Odour malodourous water single baseline characterisation method provided in this Guide? yes yes yes yes yes yes yes yes yes yes yes no no yes yes yes yes yes yes yes no no no other characterisation method(s) available in the Guide? yes yes yes yes yes yes yes yes yes yes yes yes yes yes yes yes no no no no yes no no In some LCA methodologies - for example, that given in the scientific background to the EDIP method (Hauschild & Wenzel, 1998) and the Nordic report (Lindfors et al., 1995c) - other default lists are used. Here, however, we proceed from the list of Table 4.2.2, which is based mainly on the work of WIA–2, because of the international support it enjoys. Currently modeled (midpoint) category indicators for the baseline impact categories recommended in this Guide with an indication of their environmental relevance are illustrated in Figure 4.2.2, freely adapted from Udo de Haes et al.(1999)


Although in certain contexts it may also be relevant to have information about the total energy consumed by a product system, in practice this may prove to be quite a complex issue. What precise energy requirements are to be taken, for example? The total energy used by the unit processes? The energy that can potentially be supplied by the energy resource, as originally extracted? What energy can be supplied by a particular grade of coal? If there is additional assessment of energy resources in MJ, every effort must be made to ensure that these are not subsequently double-counted in the weighting step. Part 3 : Scientific background ... ... ...


Part 3: Scientific background


With respect to the reporting of the various aspects of Impact assessment, the general requirements of If a third party report according to ISO 14040 (1997E), clause 6, is prepared, the report shall include the following items: a) the LCIA procedures, calculations, and results for the study; b) limitations of the LCIA results relative to the defined goal and scope of the study; c) the relationship of the LCIA results to the defined goal and scope, see annex A; d) the relationship of the LCIA to the LCI results, see annex A; e) impact categories considered, including rationale for their selection and a reference to their source; f) descriptions of or reference to all characterisation models, characterisation factors and methods used, including all assumptions and limitations; g) descriptions of or reference to all value-choices used in relation to impact categories, characterisation models, characterisation factors, normalisation, grouping, weighting, and elsewhere in the LCIA a


of or reference to all value-choices used in relation to impact categories, characterisation models, characterisation factors, normalisation, grouping, weighting, and elsewhere in the LCIA a justification for their use and their influence on the results, conclusions and recommendations; h) a statement that the LCIA results are relative expressions and do not predict impacts on category endpoints, exceedence of thresholds, safety margins, or risks


When included as part of the LCA study the following items shall also be included if a third party report according to ISO 14040 (1997E), clause 6, is prepared: a) a description and justification of the definition and description of any new impact categories, category indicators or characterisation models used for the LCIA; b) a statement and justification of any grouping of the impact categories; c) any further procedures that transform the indicator results, and a justification of the selected references, weighting factors, etc.; d) any analysis of the indicator results, for example sensitivity and uncertainty analysis or the use of environmental data including any implication for the results; e) data and indicator results reached prior to any normalisation, grouping or weighting shall be made available together with the normalised, grouped or weighted results


In addition, for comparative assertions disclosed to the public the report shall include the following items: a) an evaluation of the completeness of the LCA; b) a statement as to whether or not international acceptance exists for the selected category indicators and a justification for their use; c) a justification for the scientific and technical validity and environmental relevance of the category indicators used in the study d) the results of the uncertainty and sensitivity analysis; e) an evaluation of the significance of the differences found; f) if included in the LCA study: the procedures and results used for grouping; a statement that conclusions and recommendations derived from grouping are based on value-choices; a justification of the criteria used for the normalisation and grouping (these can be personal, organisational or national value-choices); a statement when grouping is used that “The ISO 14042 standard does not specify any specific methodology or support the


and grouping (these can be personal, organisational or national value-choices); a statement when grouping is used that “The ISO 14042 standard does not specify any specific methodology or support the underlying value-choices used to group the impact categories"” a statement when grouping is used that “The value-choices and judgements within the grouping procedures are the sole responsibilities of the commissioner of the study (e.g., government, community, organisation, etc)"


Where relevant, the items listed in this subclause should also be considered in the elaboration of other kinds of reports where LCIA results are used Note 1 : A graphical presentation of LCIA results as part of the report may be useful but one should consider the fact that this invites implicit comparisons and conclusions. Note 2: Due to the inherent complexity of the LCIA phase, the aforementioned additional documentation beyond the requirements stated in ISO 14040 may be desirable for internal and two party reports. ISO 14040, clause 6, apply here (see Chapter 2). In addition, ISO 14042 (2000E), clause 10, lays down the following requirements for third party study reports on Impact assessment: Lindfors et al. (1995a) list the following reporting issues for Impact assessment: “The list of the impact categories considered in the study shall be reported. Whether the category is handled in a quantitative or qualitative way should also be noted


Whether the category is handled in a quantitative or qualitative way should also be noted. Deviations from the list in Table 7.1 [in Lindfors et al. (1995a)] should be highlighted and justified. This can be regarded as a part of the Goal definition and Scoping component [...]. Under the heading of each impact category all inputs and outputs that cancontribute to the impact shall be noted. Inputs and outputs for which no quantitative information is available shall also be Part 3 : Scientific background


noted. If desired, as an alternative, the result of the classification can be reported together with the characterisation subcomponent. If a ‘red-flag’ classification of chemicals is performed, reference to the relevant list(s) shall be made. A report should contain a brief description of the background to the list(s), the criteria used in the list(s) and the motivation for the choice of the list(s). If a ‘red-flag’ classification of risks of accidents is performed, the criteria used for the flags shall be reported. [...] The reader shall be given sufficient information to be able to reproduce the results and check the data sources. Explanations of all methodological choices should be presented. This may be done by referring to another available document where the choices are justified. In cases where the results are a function of the chosen method(s), an explicit reference to the used method(s) shall be made where results and conclusions are reported


In cases where the results are a function of the chosen method(s), an explicit reference to the used method(s) shall be made where results and conclusions are reported. Not only the total contribution to each impact category should be presented, but also the contribution from each parameter considered in the classification. Results from the characterisation shall be presented in the form of tables and/or matrixes. A discussion on what differences in the results are regarded as significant should be included. As a part of an initial valuation, the characterisation results may be presented in a qualitative manner, using signs to indicate differences and significance of differences in tables and/or matrixes. Justifications for conclusions shall be presented where appropriate. If a normalisation is performed, documentation of methods and data (with adequate references), and explanations for chosen reference areas shall be included in the report


If a normalisation is performed, documentation of methods and data (with adequate references), and explanations for chosen reference areas shall be included in the report. If valuation methods are used, the method(s) shall be described, weighting factors (including data gaps) shall be presented with references [and] arguments for the choice of method(s) should be presented. [..]” More detailed reporting recommendations for the various impact categories are presented in Chapter 7 of Lindfors et al. (1995a). In its report, the SETAC-Europe Case studies Working Group (Meier et aI., 1997) gives the following (minimum) reporting guidelines for the Impact assessment phase: “If an Impact assessment has been carried out then the methodology used should be clearly detailed in the report. The reasons for notcarrying out an Impact assessment should be detailed


The reasons for notcarrying out an Impact assessment should be detailed. If the CML/SETAC approach has been used then details under the following headings should be included: Classification The selection of impact categories should reflect the goals of the study and should be justified. Any specific exclusion or inclusion of impact categories should be clearly detailed and justified. Characterisation The characterisation methodology should be detailed and explained. If any normalisation has been carried out the methodology adopted should be clearly justified including geographic and temporal considerations


If any normalisation has been carried out the methodology adopted should be clearly justified including geographic and temporal considerations. Valuation If a valuation has been undertaken the methodology should be clearly explained and justified, both for quantitative and qualitative approaches.” In this Guide the name and sequence of some of the steps of Impact assessment deviate slightly from those used in the ISO standards (see Section 1.4.3), and the ISO reporting issues have therefore not been adopted precisely as they stand. However, all the ISO issues are covered in the reporting Guidelines provided in Part 2a. Furthermore, where possible and useful the ISO guidelines have been rendered more explicit, based on the guidelines provided by Lindfors et al. (1995a) and Meier et al. (1997). PROSPECTS No specific developments are foreseen in this area. CONCLUSIONS In this Guide the problem-oriented (midpoint) approach has been elaborated as a baseline for Impact assessment


(1997). PROSPECTS No specific developments are foreseen in this area. CONCLUSIONS In this Guide the problem-oriented (midpoint) approach has been elaborated as a baseline for Impact assessment. The Eco-indicator 99 approach may be used as a sensitivity analysis. A default list of impact categories is furthermore proposed, distinguishing between: Part 3: Scientific background


Group A: Baseline impact categories Group B: Study-specific impact categories Group C: Other impact categories baseline characterisation method to be selected in Section 4.3 – to be included in (almost) all LCA studies – to be included if appropriate to the topic of study and if data are available baseline and/or alternative characterisation methods available – no baseline characterisation method available – require further elaboration for inclusion in LCA RESEARCH RECOMMENDATIONS No specific research is recommended. 4.3 Selection of characterisation methods: category indicators, characterisation models and factors TOPIC The interventions recorded in the inventory table are quantified in terms of a common category indicator. To this end characterisation models are used, from which characterisation factors are derived for individual pollutants and so on


To this end characterisation models are used, from which characterisation factors are derived for individual pollutants and so on. For a given impact category, a characterisation method comprises a category indicator, a characterisation model and characterisation factors derived from the model. The impact categories distinguished in this Guide (see Table 4.2.2) are treated individually in Sections 4.3.1 to 4.3.16, below, thereby discussing the models, factors and indicators available for the category in question. Wherever feasible a baseline characterisation method is recommended which in our view represents the current best available practice. In cases where a choice of methods was available, the arguments in favour of the preferred baseline method are presented, designating one or more of these other methods as an alternative or additional method as relevant


These choices are based on selection criteria distilled from the relevant ISO standards and the work of the second SETAC-Europe Working Group on Impact assessment, WIA–2 and these criteria will be discussed in this section. DEVELOPMENTS IN THE LAST DECADE With respect to the selection of characterisation methods, ISO has set out a number of starting points (see earlier textbox) for comparative assertions, as has been done more generally by SETAC WIA–1 (Udo de Haes (ed.; 1996). Combining the input from these two sources and adding further starting points related to the fact that the present Guide goes beyond the work of WIA–2, a list of relevant criteria has been drafted for the selection of the baseline characterisation methods recommended in the present Guide (see Table 4.3.1). The criteria based on the (marginally modified) WIA–1 starting points and the additional criteria adopted in this Guide are discussed immediately after Table 4.3.1


The criteria based on the (marginally modified) WIA–1 starting points and the additional criteria adopted in this Guide are discussed immediately after Table 4.3.1. In the tables evaluating the baseline method recommended for each impact category (Sections 4.3.1 to 4.3.16) these selection criteria are referred to in abbreviated form, as indicated in italics in Table 4.3.1. Part 3 : Scientific background –


Table 4.3.1: Selection criteria for the baseline characterisation method recommended in this Guide, with reference to ISO and WIA–2 starting points. ISO starting points the category indicator should (shall for comparative assertions) be modeled in a scientifically and technically valid way in relation to the environmental interventions, i.e., using a distinct identifiable environmental mechanism and/or reproducible empirical observation the category indicators and models shall be environmentally relevant, i.e. shall be sufficiently clearly related to the category endpoints, at least qualitatively the category indicators and models should be internationally accepted, i.e


based on an international agreement or approved by a competent international body value-choices and assumptions should be minimised category indicators can be chosen anywhere in the environmental mechanism of an impact category, from environmental interventions to category endpoints (focal point in environmental mechanism) not included not included not included not included not included not included not included WIA–2 starting points included included not explicitly included included included it should be possible to multiply characterisation factors by mass or other units indicating the magnitude of the environmental interventions the preferred time span for fate and effects is eternity, with 100 years as a second option; all effects of the emission/ extraction occurring now and in the future should be taken into account the category indicators and models should include the modeling of fate,exposure/intake and effects, as relevant the category indicators and models should include


future should be taken into account the category indicators and models should include the modeling of fate,exposure/intake and effects, as relevant the category indicators and models should include effects below thresholds (‘less is better’ approach) selection criteria for baseline method recommended in this Guide included included included included included modified: the baseline category indicators should be linear (linearity) included included included it should be possible to perform Impact assessment without information on time or location (time- and locationindependent) the method should be operational for a sufficient number of environmental interventions the uncertainty margins of the baseline indicator result should be as small as possible Part 3: Scientific background


Criterion 6: the baseline category indicators should be linear In this context, linearity means that characterisation is based on characterisation factors that are independent of the magnitude of the environmental intervention. This has been adopted as a selection criterion for the baseline category indicators for two reasons: in this way the general structure of LCA is not altered: and the quantity chosen in the definition of the functional unit, which is often arbitrary, does not influence the mutual relationship between the results for different categories. Criterion 10: it should be possible to perform Impact assessment without information on time or location There is broad ongoing debate on whether the location of emissions (or extractions) and receptors should be taken into account in LCIA. One of the classic examples concerns the emission of salt to the sea; the same may hold true for the and emissions of ships sailing in open sea


One of the classic examples concerns the emission of salt to the sea; the same may hold true for the and emissions of ships sailing in open sea. It would seem logical to introduce spatial information to resolve the problem of possible overestimation of impacts. If one knows the location of a given emission, one can determine whether effects are likely to occur, using information on the sensitivity of the location or the region. Similar considerations hold with respect to the fate of emissions. There are four main approaches to location-dependent characterisation: 1. No such characterisation. LCA is regarded as a macroscopic tool of particular value for indicating the overall (potential) environmental impacts of a (product) system. 2. Distinction between sensitive and non-sensitive areas, with emissions being ignored only when nonsensitive areas are clearly involved. No fate modeling is performed. 3. Introduction of an effect-oriented site factor


No fate modeling is performed. 3. Introduction of an effect-oriented site factor. The EDIP programme proposes consistent introduction of a site factor of between 1 and 0, depending on the projected sensitivity of the area in which the substance is emitted (Wenzel et al., 1997). No fate modeling is performed. 4. Introduction of location-dependent characterisation factors based on fate and effect modeling. Potting et al. (1998) and Huijbregts (1999b) propose an approach including both fate and regional sensitivity, based on a dispersion model and a model predicting the sensitivity to acidification. For some LCA applications this may be a useful approach, but the additional inventory data required constitute a significant drawback. These four approaches focus principally on the location of effects, with only Potting et al. (1998) and Huijbregts (1999b) proposing a location-dependent approach that integrates fate and effect


These four approaches focus principally on the location of effects, with only Potting et al. (1998) and Huijbregts (1999b) proposing a location-dependent approach that integrates fate and effect. For a further discussion on integrated fate and effect modeling, see Section 1.2.3.3 and Wegener Sleeswijk (in prep.). Location-dependent characterisation suffers from a number of general problems and limitations. Although methods and models often allow for location-dependent elaboration, such elaboration increases the quantity of inventory data required, making the inventory table more complex, as data on different locations cannot be aggregated. Data requirements may even become prohibitive, for example when different sites are distinguished for each impact category. Until now location-dependent factors have been proposed for different impact categories independently


Until now location-dependent factors have been proposed for different impact categories independently. A comprehensive approach distinguishing the same, limited number of regions for all impact categories would be preferable, to keep the locational differentiation in the inventory manageable. In many LCA studies, however, no information is available on the location of emissions and/or extractions. Our baseline category indicators must therefore be able to assess at least locationindependent inventory data. If, in addition, location-specific information can also be assessed with the same category indicator this is then, of course, an advantage. Besides spatial differentiation, differentiation in the temporal domain may also be relevant for certain impact categories; see the general discussion in Section 1.2.3.3. Examples of impact categories for which temporal differentiation might be useful are day- and nighttime noise and summer and winter smog


Examples of impact categories for which temporal differentiation might be useful are day- and nighttime noise and summer and winter smog. The same reasoning can be applied to temporal as to spatial differentiation. For many interventions and unit processes there will be no temporal information available, and it should therefore at least be possible to apply non-time-specific characterisation factors. It is not only the time of occurrence of interventions and/or impacts that may be of interest, but also their duration; see Section 1.2.2.3. In the case of the Inventory analysis, opting for a cradle-to-grave analysis already implies a basic perspective of infinite time. For impact assessment the situation is arguably different. Ecosystems are slow to adapt to changes in environmental conditions, for instance. It might Part 3 : Scientific background


therefore be argued that impacts with a very long time horizon require form of time-discounting or cut-off beyond, say, 100 or 1000 years (Udo de Haes (ed.), 1996). Criterion 11: The method should be operational for a sufficient number of environmental interventions A qualitatively excellent category indicator operational for only a small fraction of relevant environmental interventions is not a good choice for the baseline method, because the results obtained with the indicator should represent all the interventions contributing to the category in question. Criterion 12 (+2): The uncertainty margins of the baseline indicator result should be as small as possible (while the indicator remains as environmentally relevant as possible) Criterion 12 is often in conflict with criterion 2: endpoint indicators often lead to greater uncertainties in results than midpoint indicators, for example, but they have more environmental relevance


A balance should therefore be sought between these two selection criteria: indicators should be as environmentally relevant as possible without introducing undue uncertainty into the results. Whether a method is based on a proportional or a marginal approach plays no role in the choice of baseline. As current characterisation models employ either linear relationships (e.g. for toxicity, acidification) or non-linear relationships, it is not possible to derive both proportional and marginal characterisation factors. While such a theoretical distinction can be made, then, this choice rarely presents itself in practice and we shall have to work with the (heterogeneous) factors we have


While such a theoretical distinction can be made, then, this choice rarely presents itself in practice and we shall have to work with the (heterogeneous) factors we have. (See Section 1.2.3.4 for a more extensive discussion of this and related topics.) PROSPECTS The new SETAC-Europe Working Group on life cycle Impact assessment (WIA–2) and the forthcoming SETAC/UNEP cooperation have been charged with preparing an authorised, recommended list of impact categories with respective category indicators and characterisation factors. These may differ from the categories, indicators and factors adopted here, in which case these documents will need to be updated in due course (in about 3–4 years). CONCLUSIONS On the basis of the selection criteria presenter above, in this Guide a distinction has been made between: a baseline characterisation method, i.e


CONCLUSIONS On the basis of the selection criteria presenter above, in this Guide a distinction has been made between: a baseline characterisation method, i.e. the method recommended here as the current best available practice for the impact category in question; alternative characterisation methods, which may be adopted instead of the baseline method if duly justified and documented, or may be used in tandem with the baseline method, as a sensitivity analysis; additional characterisation methods, which may be applied similarly to alternative methods, but requiring additional effort (e.g. collection of additional data, development of additional models); variant characterisation methods, which start from entirely different principles


collection of additional data, development of additional models); variant characterisation methods, which start from entirely different principles. RESEARCH RECOMMENDATIONS Short-term research It is recommended to investigate the usefulness and feasibility of category indicators for the impact categories ‘human toxicological impacts’ and ‘ecotoxicological impacts’ calculated using the damage approach. The marginal and proportional category indicators currently used in LCIA should be inventoried. A comprehensive site- or location-dependent approach should be developed (although the relevant level of differentiation might vary among impact categories). Appropriate time horizons for LCIA should be examined in more detail. Long-term research It is recommended to investigate the potential for developing category indicators for impact categories within an overall framework based on a damage approach


Long-term research It is recommended to investigate the potential for developing category indicators for impact categories within an overall framework based on a damage approach. As a start, the scope could be investigated for developing category indicators for human and ecosystem health, including not only toxicological impacts but also such other impacts as casualties, smog, etc. It is recommended to study, for several impact categories, the differences between LCIA results based on proportional and a marginal modeling and the influence of including/excluding background concentrations. The scope for fully integrating fate in the characterisation factor for each impact category should be investigated. Part 3: Scientific background


Because of the wide variation in the dilution volume of substances, the scope for including this parameter in multimedia models should be investigated. In the following sections each of the impact categories listed in Table 4.2.2 is described and documented and a baseline characterisation method recommended. 4.3.1 Depletion of abiotic resources TOPIC ‘Abiotic resources’ are natural resources (including energy resources) such as iron ore, crude oil and wind energy which are regarded as non-living. Abiotic resource depletion is one of the most frequently discussed impact categories and there is consequently a wide variety of methods available for characterising contributions to this category. To a large extent these different methodologies reflect differences in problem definition. Depending on the definition, this impact category has only natural resources, or natural resources, human health and the natural environment as areas of protection (see Figure 4.2.2)


Depending on the definition, this impact category has only natural resources, or natural resources, human health and the natural environment as areas of protection (see Figure 4.2.2). DEVELOPMENTS IN THE LAST DECADE Three types of abiotic resources can be distinguished: deposits, funds and flows. Deposits are resources that are not regenerated within human lifetimes. Examples of deposits are fossil fuels, minerals, sediments, clay, etc. Funds are resources that can be regenerated within human lifetimes. Groundwater and soil are examples of funds. Flows are resources that are constantly regenerated, such as wind, river water and solar energy (Finnveden, 1996a). It is debatable whether all three types of abiotic resources can or should be aggregated into one measure for abiotic depletion


It is debatable whether all three types of abiotic resources can or should be aggregated into one measure for abiotic depletion. It will be difficult to combine Impact assessment for flows, for which there is no reserve to be depleted but rather a maximum utilisable flow, with that for deposits and funds (Guinée & Heijungs, 1995). Heijungs et al. (1992) In Heijungs et al. (1992) for a given resource i, abiotic depletion was defined as the ratio between the quantity of resource extracted and the recoverable reserves of that resource yielding a dimensionless indicator result. The units used for both extractions and reserves could thus be freely selected, as long as this was consistent for a given resource. Ores were normally expressed in kg and natural gas in although MJ could be used as an alternative. “Abiotic depletion” also covered depletion of some energy resources such as fossil fuels. Heijungs et al


“Abiotic depletion” also covered depletion of some energy resources such as fossil fuels. Heijungs et al. (1992) observed that this is a simplified method and that it should ultimately be extended to include the extraction rate, expressed in kg/yr or Heijungs et al. (1997) make a distinction between resources that can be depleted and those that are competitively used, stating that deposits fall into the former category and flows into the latter, while funds may fall into either category. They state that the two categories should be assessed using two different methods: resources that are depleted should be assessed by a method based on depletion, those that are competitively used by a method based on competition. One implication of this is that aggregation of abiotic resources into a single measure is not meaningful. Reviews of existing Impact assessment methods for the depletion of abiotic resources are provided in several publications, in particular Heijungs et al


Reviews of existing Impact assessment methods for the depletion of abiotic resources are provided in several publications, in particular Heijungs et al. (1992), Fava et al. (1993), Guinée & Heijungs (1995), Lindfors et al. (1995a,c), Lindfors (1996), Finnveden (1996a) and Heijungs et al. (1997). Broadly speaking, the available methods fall into six groups, reviewed very briefly below (mainly Finnveden, 1996a, and Lindfors et al., 1995c, supplemented with more recently developed methods). 1. No assessment or aggregation (e.g Lindfors, 1996). 2. Aggregation of natural resource extractions on a mass basis (e.g. Lindfors et al., 1995c). 3. Aggregation and assessment based either on a) ‘ultimate reserves’, i.e. the quantity of resource (as a chemical element or compound) that is ultimately available, estimated by multiplying the average natural concentration of the resource in the primary extraction media (e.g. the earth’s crust) by the mass or volume of these media (e.g


the earth’s crust) by the mass or volume of these media (e.g. the mass of the crust) (Guinée, 1995); or b) on ‘economic Part 3 : Scientific background


reserves’, i.e. that part of the reserve base1 which can be economically extracted at the time of determination (United States Department of the Interior - Bureau of Mines, 1993) and/or current extraction rate. The method of Heijungs et al. (1992) is an example of assessment based on reserves. Alternative assessment methods proceed from resource extraction rates relative to reserves (see Guinée & Heijungs, 1995 and Ekvall et al., 1997), from extraction rates only (Goedkoop, 1995) or from per capita reserves (see Hauschild & Wenzel, 1998) Aggregation and assessment based on the cost of ‘restoring’ the resource to its original, natural state, or on the costs associated with substituting current extraction processes by presumed ‘sustainable’ processes. Pedersen (1991) and Steen (1995) describe such methods. Aggregation and assessment based on energy content or exergy content or consumption (e.g. Finnveden, 1996b; see also Ayres et al., 1996 and Ayres, 1998)


Aggregation and assessment based on energy content or exergy content or consumption (e.g. Finnveden, 1996b; see also Ayres et al., 1996 and Ayres, 1998). Exergy is the amount of energy that can be obtained when matter is brought reversibly into equilibrium with its surroundings. It is the fraction of the energy content that can be used for work (= available energy). The exergy consumed is the exergy of the resources extracted as input minus the exergy of the outputs (Finnveden, 1996b). Finnveden suggests that the potential exergy of an ore might be used as a measure for depletion of abiotic resources in LCA. Aggregation and assessment based on the change in, the anticipated environmental impact of the resource extraction process due to lower-grade deposits having to be mined in the future. This method is described by Blonk et al. (1997a) and Müller-Wenk (1998) among others. It has been operationalised for metal ores and energy resources


This method is described by Blonk et al. (1997a) and Müller-Wenk (1998) among others. It has been operationalised for metal ores and energy resources. In the case of metal ores, the virtual additional energy that will be required for future extraction processes is estimated. This additional energy, and the energy content of energy resources, are converted into the weighting indices of the Eco-indicator 95 approach by performing an LCA for the transformation of 1 kg ‘heavy-grade’ oil into thermal energy. Within the Eco-indicator 99 methodology a method has been developed based on that of Müller-Wenk (1998) (Goedkoop & Spriensma, 1999). The authors differ in their conclusions as to the best method for characterising abiotic resource depletion. Lindfors (1996), for instance, recommends not aggregating abiotic recources at all in the characterisation phase of LCAs conducted under the ecolabeling programme (= method group 1, above)


Lindfors (1996), for instance, recommends not aggregating abiotic recources at all in the characterisation phase of LCAs conducted under the ecolabeling programme (= method group 1, above). He states that all characterisation methods introduce value-based judgement and that there is no consensus on this point. Finnveden (1996a) states that further discussion is necessary before one particular methods can be selected. A major debating point will be problem definition. In our view, however, the current lack of any aggregation of abiotic resources at all has the disadvantage that the outcome of the characterisation phase comprises many separate scores for this impact category, which the LCA practitioner will probably not put to any use. The implication is that the problem is thus neglected. Hauschild & Wenzel (1998) also advise against applying aggregation in the characterisation phase, proposing use of weighting at a later stage, after normalisation


Hauschild & Wenzel (1998) also advise against applying aggregation in the characterisation phase, proposing use of weighting at a later stage, after normalisation. In their weighting method resource depletion is specified as a fraction of known per capita economic reserves in 1990. They eventually therefore recommend using a method from group 3, above. The only difference from other methods in group 3 is that they advise aggregating in a later phase of the LCA, during weighting rather than characterisation. In their review, Heijungs et al. (1997) employ five criteria to assess all the methods then available. They conclude that there is no ‘perfect’ method meeting all their criteria. The main conclusion here is that there is as yet no consensus about what constitutes the best category indicator for ‘abiotic depletion’, the choice depending crucially on the definition of this term


Although in most publications the problem of abiotic depletion is not precisely defined, four groups of definitions can be broadly distinguished, based on what is seen as the key problem: 4. 5. 6. 1 The reserve base is that part of an identified resource that meets specified minimum physical and chemical criteria related to current mining practice (United States Department of the Interior - Bureau of Mines, 1993). the decrease of the resource itself – method groups 2 and 3 the decreasing reserves of useful energy/exergy in the world –method group 5 the contribution of current extraction processes, or possible ‘restoration’ of the resource, to other impact categories (in the first case, this means that resource depletion is not in fact regarded as a separate environmental problem at all, as the environmental impacts of extraction processes are already included in LCAs (cf