At the moment the EPA is working on a list of so-called ‘PBT’ chemicals that are persistent, bioaccumulative and toxic, to be used to focus source-reduction and recycling measures. To compile the list some 4000 chemicals are being screened and the resultant PBT information stored in a database. This database might perhaps in the future be used in combination with the data of Huijbregts (1999a) and the USES-LCA data to gain an approximate indication of human and ecotoxicological characterisation factors for a large group of substances. CONCLUSIONS Given the foregoing considerations, for the same reasons as for impact category human toxicity we recommend using Huijbregts (1999a), based on fate modeling with USES-LCA, as the baseline characterisation method for ecotoxicity. USES-LCA treats intermedia transport in most realistically and


Part 3: Scientific background comprehensively, using continuous variables. The fact that the method of Huijbregts also comprises factors for human toxicity (see Section 4.3.7) is an additional advantage, allowing the two impact categories to be assessed with a similar fate model. An infinite time horizon is thereby adopted (criterion 7), and the global scale as spatial horizon (criterion 10). As with human toxicity, however, it should be borne in mind that the list of 180 substances provided by Huijbregts (1999a) represents merely a very small subset of all known and unknown toxic substances. If the case study involves substances suspected of contributing to ecotoxicity for which no characterisation factors are provided in this Guide, such factors should be calculated. If this is not feasible, an attempt should be made to estimate such factors based on similar or related substances for which they are available (by extrapolation, for example; cf. Section 4.3.17)

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


If this is not feasible, an attempt should be made to estimate such factors based on similar or related substances for which they are available (by extrapolation, for example; cf. Section 4.3.17). It is not advised to use ‘old’ characterisation factors that ignore fate, such as those developed by Heijungs et al. (1992), even if these are available for more substances than the new, fatebased factors. Fate is a particularly important consideration in the context of the ecotoxicity of chemical pollutants and its exclusion might lead to misleading results. The baseline characterisation method for ecotoxicity, using the characterisation factor is evaluated with respect to the (ISO-based) criteria in Table 4.3.8.1. Table 4.3.8.1: Evaluation of the baseline characterisation method for ecotoxicity, using the characterisation factor with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12


criterion scientifically and technically valid environmentally relevant internationally accepted value-choicesandassumptions focal point in environrmental mechanism linear time span fate, exposure/intake and effects less is better time- and location-independent operational uncertainty margins evaluation partly: although fate calculations are based on well understood environmental mechanisms, actual toxicological effects are assessed via a very crude aggregation of very different effects (from reduced fertility to mortality) among species and ecosystems yes, the category indicator represents risks to ecosystems (near-endpoint level) no; however, USES-LCA is very similar to the model EUSES, supported by the European Union (EUSES is in fact based on USES) yes; the choice of valuing all effects and species equally is particularly debatable (nearly) endpoint yes eternity fate included, exposure/intake not relevant, effects included yes; below-threshold effects are also included yes,


effects and species equally is particularly debatable (nearly) endpoint yes eternity fate included, exposure/intake not relevant, effects included yes; below-threshold effects are also included yes, although in fact the ETPs are European factors yes, for 180 substances emitted to air, water and soil presumably about the same as for the other category indicators that include fate (method group 2) The following methods are included in this guide as options for sensitivity analysis, particularly for LCA studies in which results for this impact category are dominated by metals and volatile, persistent halogenated organics: The FAETPs, MAETPs, FSETPs, MSETPs and TETPs for the time horizons 20, 100 and 500 years at the global scale: Huijbregts et al


(2000). The FAETPs, MAETPs, FSETPs, MSETPs and TETPs for an infinite time horizon at the continental scale (i.e. excluding global effects): Huijbregts et al. (2000). Recommendations for extended LCAs: If the case study involves substances suspected of contributing to ecotoxicity for which FAETPs, MAETPs, FSETPs, MSETPs and TETPs are not provided in this Guide, such factors should be calculated. If this is not feasible, an attempt should be made to estimate ETPs for these substances based on similar or related substances for which these ETPs are available (by extrapolation, for example; cf. Section 4.3.17).


Part 3 : Scientific background Freshwater aquatic ecotoxicity: method status baseline alternative 1 alternative 2 alternative 3 alternative 4 additional variant characterisation method/factor – – Marine ecotoxicity: method status baseline alternative 1 alternative 2 alternative 3 alternative 4 additional variant characterisation method/factor – – Terrestrial ecotoxicity: method status baseline alternative additional variant characterisation method/factor – – Freshwater sediment ecotoxicity: method status baseline alternative 1 alternative 2 alternative 3 alternative 4 additional variant characterisation method/factor – – Marine sediment ecotoxicity: method status baseline alternative 1 alternative 2 alternative 3 alternative 4 additional variant characterisation method/factor – – reference Huijbregts, 1999a Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 – Hauschild & Wenzel, 1998; Guinée et al., 1996; Jolliet & Crettaz, 1997 reference


Huijbregts, 1999a Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 – Hauschild & Wenzel, 1998; Guinée et al., 1996; Jolliet & Crettaz, 1997 reference Huijbregts, 1999a Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 – – reference Huijbregts, 1999a Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 – Hauschild & Wenzel, 1998; Guinée et al., 1996; Jolliet & Crettaz, 1997 reference Huijbregts, 1999a Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 – – reference Huijbregts, 1999a Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 Huijbregts et al., 2000 – – RESEARCH RECOMMENDATIONS Long-term research: The potential for integrating the method of Huijbregts (1999a) with the so-called PAF approach should be investigated


The potential for using the PAF approach to aggregate ecotoxicological with such other effects as eutrophication and acidification should be further investigated.


Part 3: Scientific background This database might perhaps in the future be used in combination with the data of Huijbregts (1999a) and the USES-LCA data to calculate approximate human and ecotoxicological characterisation factors for a larger group of substances. Photo-oxidant formation TOPIC Photo-oxidant formation is the formation of reactive chemical compounds such as ozone by the action of sunlight on certain primary air pollutants. These reactive compounds may be injurious to human health and ecosystems and may also damage crops. The relevant areas of protection are human health, the man-made environment, the natural environment and natural resources (Udo de Haes et al., 1999)


The relevant areas of protection are human health, the man-made environment, the natural environment and natural resources (Udo de Haes et al., 1999). Photo-oxidants may be formed in the troposphere under the influence of ultraviolet light, through photochemical oxidation of Volatile Organic Compounds (VOCs) and carbon monoxide (CO) in the presence of nitrogen oxides Ozone is considered the most important of these oxidising compounds, along with peroxyacetylnitrate (PAN). Photo-oxidant formation, also known as summer smog, Los Angeles smog or secondary air pollution, contrasts with winter smog, or London smog, which is characterised by high levels of inorganic compounds, mainly particles, carbon monoxide and sulphur compounds. This latter type of smog causes bronchial irritation, coughing, etc. Winter smog, as far as considered in this Guide, is part of human toxicity


This latter type of smog causes bronchial irritation, coughing, etc. Winter smog, as far as considered in this Guide, is part of human toxicity. Today, three methods are available for comparing the ozone creation potential of different species of VOC, based on: POCPs (e.g. Derwent & Jenkin, 1990, Derwent et al., 1998) fate factors (Hofstetter, 1998) Incremental Reactivity (e.g. Carter, 1994) 1. 2. 3. DEVELOPMENTS IN THE LAST DECADE Heijungs et al. (1992) The numerous atmospheric species of VOC vary widely in their contribution to photo-oxidant formation. In Heijungs et al. (1992) Photochemical Ozone Creation Potentials (POCPs) were used as a characterisation factor to assess and aggregate the interventions for the impact category photo-oxidant formation: where (kg) is the mass of substance i released, the photochemical ozone creation potential of the substance and Photo-oxidant formation is the indicator result, which is expressed in kg ethylene-equivalents. In Heijungs et al


In Heijungs et al. (1992) it was proposed to flag emissions to remind practitioners of their crucial relevance to photo-oxidant formation. In most current LCA studies the available data set will not include specific VOC species, but only groups of hydrocarbons and methane. In such cases Heijungs et al. (1992) recommended deriving average POCPs for these groups from the arithmetical average of the data for individual VOCs. The POCP is calculated as the estimated quantity of ozone formed photochemically by a given VOC, using a model to calculate ozone formation in the presence and absence of this compound (UNECE, 1990). This means that, unlike ODPs and GWPs, the POCPs calculated by Derwent & Jenkin (1990) are based on an ‘average’ rather than a ‘marginal’ approach. Heijungs et al. (1992) suggested that it would be more attractive to calculate POCPs on the basis of a marginal change in emissions, as with GWPs and ODPs


Heijungs et al. (1992) suggested that it would be more attractive to calculate POCPs on the basis of a marginal change in emissions, as with GWPs and ODPs. One of the advantages of a marginal approach is that would also be included, as explained below, in contrast to Derwent & Jenkin’s method, which covers only VOCs. Although acts as a catalyst in the chemical reactions involved in photochemical smog formation and is therefore not consumed, its background concentration affects the course of ozone production. As is also removed from the atmosphere in the form of (acidic, eutrophying) precipitation and deposition, photochemical reactions will only be maintained if there is a continuous supply thereof. In the methods currently used to estimate POCPs the emission-effect curve is assumed to be a straight line through the origin


In the methods currently used to estimate POCPs the emission-effect curve is assumed to be a straight line through the origin. This is an extremely coarse approximation of the actual situation and may introduce errors for certain substances, including In principle, it would be possible to calculate a POCP for However, setting emissions to zero for the reference part of the calculation outlined above would lead to absurd results, as there would be no ozone production at all without The emission-effect curve of is far from linear. If POCPs were calculated on the basis of marginal changes in emissions, a POCP could be derived for 4.3.9


Part 3 : Scientific background Methods based on POCPs Photochemical Ozone Creation Potentials (POCPs) were originally developed to assess various emission scenarios for volatile organic compounds (Derwent & Jenkin, 1990). A UN protocol defined the POCP of a VOC as the ratio between the change in ozone concentration due to a change in the emission of that VOC and the change in the ozone concentration due to a change in the emission of ethylene (UNECE, 1990). Expressed as a formula: where is the change in ozone concentration due to a change in the emission of VOC i and the integrated emission of VOC i up to that time, with the denominator containing these parameters for ethylene, the reference substance. The POCPs of Derwent & Jenkin (1990), used as characterisation factors in Heijungs et al. (1992), were updated in 1996, 1998 and 1999 (Derwent et al., 1996, 1998; Jenkin & Hayman, 1999). These figures are based on a 5-day trajectory model of VOC transportation above Europe


(1992), were updated in 1996, 1998 and 1999 (Derwent et al., 1996, 1998; Jenkin & Hayman, 1999). These figures are based on a 5-day trajectory model of VOC transportation above Europe. In the first update POCPs for NO, and several other inorganic substances were also included, in contrast to the Derwent & Jenkin POCPs of 1990. However, it is not clearly explained how these were calculated. In the last two updates only the POCPs for VOCs were revised, using a new calculation method. The original POCPs for these compounds were based on the difference in ozone formation with and without the VOC in question, which, as already mentioned, is an ‘average’ approach (see text box). The new POCPs (Derwent et al., 1998; Jenkin & Hayman, 1999) are based on a ‘marginal’ approach, being calculated from the 5-dayintegrated incremental ozone production due to an incremental emission of the VOC on top of a basic emission scenario. Andersson-Sköld et al


Andersson-Sköld et al. (1992) have calculated POCPs for a different trajectory with a lower background concentration of POCPs have thus been calculated for two scenarios: a scenario with a relatively high background concentration (i.e. with a surplus of and VOC as limiting factor for ozone formation), yielding POCPs’ (Derwent et al., 1998); and a scenario with a relatively low background concentration (i.e. with a surplus of VOC and as limiting factor for ozone formation), yielding POCPs’ (Andersson-Sköld et al., 1992). Methods based on fate factors Hofstetter (1998) has developed characterisation factors based on Disability Adjusted Life Years (DALYs) for respiratory diseases due to air pollution (see also Section 4.3.7)


These factors, which are used in the Eco-indicator 99 (Goedkoop & Spriensma, 1999), take account of ozone-induced respiratory disease for a number of VOCs and based on a fate factor and the DALY for A fate factor for and for Non-Methane VOC (NMVOC) was derived using the EMEP model1, with a fate factor for other VOCs then being derived by multiplying the ratio between the POCP of the VOC in question and that of NMVOC by the fate factor for NMVOC derived with the EMEP model: The only difference between the characterisation values developed by Hofstetter (1998) and the POCPs discussed above is located in two constants: the DALY for and (Note that DALYs do not cover effects on ecosystems or crops, and that these effects should be assessed separately.) Hofstetter (1998) has also developed characterisation factors for respiratory diseases for the main inorganic chemicals causing winter smog: particles, CO and sulphur compounds


With this method summer and winter smog can therefore be amalgamated into one category indicator. Methods based on Incremental Reactivity An alternative approach to describing the potential contribution of VOCs to photo-oxidant formation is thyat based on the notion of Incremental Reactivity (IR; Carter 1994, Carter, 1997). The incremental 1 European Monitoring and Evaluation Programme (Barret & Berge, 1996).


Part 3: Scientific background reactivity of a VOC in a pollution scenario is defined as the change in ozone caused by adding a small amount of the VOC to the emissions in the scenario, divided by the amount of VOC added (Carter, 1994): with: the Incremental Reactivity of substance I; the change in ozone (g); the incremental mass of VOC added (g). IRs are calculated using a so-called ‘base case scenario’ (i.e. specific pollution scenario) that represents a specific ozone exceedence episode in an area of the United States


The base case scenario is subsequently adjusted, resulting in three derived scenarios and three associated IRs: MIR scenario: Maximum Incremental Reactivity scenario, in which the emissions in the base case scenario are adjusted to yield the highest incremental reactivity of the initially present VOC mixture MOR scenario: Maximum Ozone Reactivity scenario, in which the emissions in the base case scenario are adjusted to yield the highest peak ozone concentration EBIR scenario: Equal Benefit Incremental Reactivity scenario, in which the emissions in the base case scenario are adjusted such that VOC and NOX reductions are equally effective in reducing These three scenarios provide three different IRs: the MIR scenario yields MIRs (Maximum Incremental Reactivity), the MOR scenario MOIRs (Maximum Ozone Incremental Reactivity) and the EBIR scenario EBIRs (Equal Benefit Incremental Reactivity (Carter, 1994) There are three main differences between the approach based on IRs and that


(Maximum Ozone Incremental Reactivity) and the EBIR scenario EBIRs (Equal Benefit Incremental Reactivity (Carter, 1994) There are three main differences between the approach based on IRs and that employing POCPs


First, while recent POCPs were developed on the basis of regional European scenarios (Derwent et al., 1998; Andersson-Sköld et al., 1992), IRs are grounded in scenarios for urban areas in North America (Carter 1994, Carter, 1997). Second, POCPs are based on a trajectory model of VOC transport over Europe, IRs on a single-cell box model. Third. POCPs are based on a time span of 5 days, IRs on a time span of at most one day. In other respects, too, the scenarios used for the respective calculations are at variance and are therefore difficult to compare. At the same time, though, there is a reasonable correlation between the POCP and MIR values, which generally predict the same relative importance of different classes of VOC (Hahn & Will, 1995; Jenkin & Hayman, 1999)


MIRs consider ozone formation on a much shorter time scale, however, and therefore give greater weight to those VOCs that are rapidly oxidised, whereas POCPs provide better resolution of less reactive VOCs (Jenkin & Hayman, 1999). Moving back to the wider discussion of photo-oxidant formation, several important issues have been raised in the recent literature. In the first place, a number of authors consider it preferable to characterise emissions under the heading of photo-oxidant formation and several attempts have been made to do so. Frischknecht (1998) bases a POCP for on the assumption that and VOCs are equally responsible for photo-oxidant formation. He then uses the quotient of annual Swiss emissions of and non-methane VOCs (NMVOCs) to calculate a POCP of 0.645 kg ethylene-equivalents per kg Apart from this being a very crude assumption, these calculations in fact appear to be incorrect1


Hoffstetter (1998) calculates a fate factor for from data generated by the aforementioned EMEP model (Barret & Berge, 1996). This factor, which he himself characterises as very uncertain, is equal to that of all NMVOC combined (kg ethylene-equivalents). Finnveden et al. (1992), Lindfors (1996) and Nichols et al. (1996) propose dividing the category into two, one for and one for VOCs (and CO). Lindfors et al. (1995c) even add a separate subcategory for CO. However, they do not elaborate these proposals into characterisation factors for Derwent et al. have calculated POCPs voor NO and It has also been proposed that POCPs should better reflect the fact that differences in background concentration lead to different relative contributions of individual VOCs to ozone formation. To tackle this 1 Frischknecht assumes that all VOC is ethylene and probably confused these numbers with those for emissions of NMVOC and


Part 3 : Scientific background problem, Nichols et al. (1996) and Hauschild & Wenzel (1998) propose using two different lists of POCPs for VOCs: one for situations with a high background concentration (data from Derwent & Jenkins, 1990, recently updated in Derwent et al. 1998) and one for situations with a low background concentration (data from Andersson-Sköld et al., 1992). Another option would be to use MOIRs and EBIRs in situations with high or low concentrations, respectively. Many authors hold that a distinction should be made between long-term and short-term processes (e.g. Finlayson-Pits & Pits, 1993; Nichols et al., 1996). Short-term smog episodes characterised by high ozone concentrations (e.g. >120 ppb) are best understood in terms of the effects of oxidant formation on human health. However, longer periods with lower ozone levels may also cause damage to human health, crops and forest ecosystems


However, longer periods with lower ozone levels may also cause damage to human health, crops and forest ecosystems. Both POCPs and MIRs (MOIRS and EBIRS) are concerned specifically with ozone formation potential in peak situations and it is unclear whether these factors are also valid for extended periods of low ozone concentration. Finally inventory data are often insufficiently differentiated, mentioning only ‘total VOC emission’, or similar. One way of tackling this problem is to use average POCPs for the total VOC emissions of different types of emission source. (For a description of other approaches to the problem of group parameters, see Section 3.6) This approach has been followed by Hauschild & Wenzel (1998) and by Derwent et al. (1996)


(For a description of other approaches to the problem of group parameters, see Section 3.6) This approach has been followed by Hauschild & Wenzel (1998) and by Derwent et al. (1996). A second option is to solve the problem in the inventory phase, using sourcespecific emission profiles sources (see Section 3.6) PROSPECTS In the future, further progress is anticipated on Impact assessment modeling to the various endpoints, viz. damage to human health, damage to ecosystem health, etc. (cf. Figure 4.2.2); see Section 4.2 for a more extensive discussion of this subject). CONCLUSIONS As a baseline it is recommended to use the most recent POCPs of Derwent et al. (1998) and Jenkin & Hayman (1999), supplemented with the POCPs for inorganic substances1 of Derwent et al. (1996), which include NO and Although MIRs, MOIRS and EBIRs also focus on ozone formation potential, these indicators consider ozone formation over a much shorter time scale


(1996), which include NO and Although MIRs, MOIRS and EBIRs also focus on ozone formation potential, these indicators consider ozone formation over a much shorter time scale. Because the longer term is relevant in the LCA context, POCPs are to be preferred over MIRs (criterion 7). The POCPs of Andersson-sköld et al. (1992) are based on the same model as the POCPs of Derwent et al. (1998) and Jenkin & Hayman (1999). Hauschild & Wenzel (1998) advise using the POCPs for Scandinavia and other situations with low concentrations and the POCPs for the rest of Europe and other regions with high concentrations. Given the fact that the latter are more frequently updated, however, the difference between the two appears to be accrue not only from regional atmospheric variation, but also from growing scientific knowledge. In our view, therefore, it is inappropriate to use both types of POCP in one and the same LCA


In our view, therefore, it is inappropriate to use both types of POCP in one and the same LCA. As a baseline we recommend using the POCPs, because these are more up to date and because we anticipate that most emissions covered by LCAs will occur in regions with high concentrations. It is advised to use emission profiles to disaggregate aggregated VOC emissions before Impact assessment, rather than using POCPs for specific VOC mixtures (see Section 3.6 and Part 2b, Section 3.6). Such disaggregation is also relevant for other impact categories like human toxicity and ecotoxicity. The evaluation of the baseline characterisation method using the characterisation factor high NOx POCP with respect to the (ISO-based) criteria is shown in Table 4.3.9.1. 1 When consulted on this topic, D. Derwent confirmed that the old (average) POCPs for inorganic substances can be combined with the new (marginal) POCPs for VOCs.


Part 3: Scientific background Table 4.3.9.1: Evaluation of the baseline characterisation method for photo-oxidant formation, using the characterisation factor high POCP, with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. criterion scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linear time span fate, exposure/intake and effects less is better time- and location-independent operational uncertainty margins evaluation yes; based on a widely supported trajectory model yes; the adverse effects of high ozone concentrations (especially those on human health) are well understood not officially, but the POCPs of Derwent et al


(1998) are widely used all over the world yes midpoint yes 5 days fate included, exposure/intake not relevant, effects included in terms of effects on photo-oxidant formation yes; below-threshold effects are also included yes, although in fact POCPs are representative for European countrieswithhighbackground concentrations yes, for 120 VOCs it is not known how the uncertainties in POCPs compare with those associated with other methods for this impactcategory The following methods are included in the guide as options for sensitivity analysis: MIRs, MOIRs and EBIRs (Carter, 1997), when there is interest in assessing VOCs on a shorter time scale. The POCPs of Andersson-Sköld et al. (1992) for LCA studies in which most emissions take place in regions with low concentrations. No additional recommendations are provided for extended LCAs


The POCPs of Andersson-Sköld et al. (1992) for LCA studies in which most emissions take place in regions with low concentrations. No additional recommendations are provided for extended LCAs. method status baseline alternative 1 alternative 2 alternative 3 alternative 4 additional variant characterisation method/factor high POCP MIR MOIR EBIR low POCP – – reference Derwent et al., 1996; Derwent et al., 1998; Jenkin & Hayman, 1999 Carter, 1997 Carter, 1997 Carter, 1997 Andersson-Sköld et al., 1992 – – RESEARCH RECOMMENDATIONS Long-term research: The old POCPs are based on an average approach, the new POCPs on a marginal approach. This makes them a good starting point for a general analysis of the differences between using a marginal and an average approach for the same impact category. 4.3.10 Acidification TOPIC Acidifying pollutants have a wide variety of impacts on soil, groundwater, surface waters, biological organisms, ecosystems and materials (buildings)


4.3.10 Acidification TOPIC Acidifying pollutants have a wide variety of impacts on soil, groundwater, surface waters, biological organisms, ecosystems and materials (buildings). Examples include fish mortality in Scandinavian lakes, forest decline and the crumbling of building materials. The major acidifying pollutants are and Areas of protection are the natural environment, the man-made environment, human health and natural resources (see Figure 4.2.2). NOx


Part 3 : Scientific background DEVELOPMENTS IN THE LAST DECADE Acidification is one of the impact categories in which local sensitivity plays an important role and the possibility of including regional differences in the LCA model has been a key issue in recent years. It is Heijungs et al. (1992) In Heijungs et al. (1992) acidification potentials (APs) were used as a characterisation factor to assess and aggregate the interventions for the impact category acidification: where (kg) is the mass of substance i released, the Acidification Potential of the substance and Acidification is the indicator result, which is expressed in kg Potential acid deposition can be expressed in terms of potential allowing potentially acidifying emissions to be aggregated on the basis of their capacity to form It is assumed, for instance, that one molecule of yields two ions, with one molecule of nitrogen oxides and one molecule of reduced nitrogen compound each yielding a single ion


The acidification potential (AP) of substance i is defined as the number of ions produced per kg substance relative to where represents the number of ions that can potentially be produced per kg substance i, and the number of ions produced per kg The emission-effect curve is a straight line through zero. therefore discussed here as an extension of the general discussion on regionalisation earlier in this chapter (see discussion on criteria below Table 4.3.1). As stated, APs reflect the maximum acidification potential of a substance. The actual impact will be governed by local processes and circumstances, and will be reduced as mineralisation and denitrification rates increase. The acidification caused by a particular substance may also be reduced if the anions accompanying the hydrogen ions become bound to the impacted system (for a certain period, for it is not an infinite buffer) or absorbed and removed by biomass (Hauschild & Wenzel, 1998; Lindfors, 1996)


This is particularly relevant for and where actual acidification may vary between 0% and 100% of the potential value. Several methods have been proposed to deal with local differences in sensitivity to acidification: 1. 2. 3. 4. neglecting emissions in non-sensitive areas (e.g. Hogan et al., 1996); weighting emissions according to the sensitivity of the area in which they are emitted (e.g. Hauschild & Wenzel, 1998). assessing a maximum and a minimum scenario (e.g. Lindfors, 1996; Nichols et al., 1996); extending models to include regional sensitivity and fate (e.g. Potting et al, 1998; Huijbregts, 1999b) . These four options are considered in turn. 1. Neglecting emissions in non-sensitive areas The simplest solution is to neglect all emissions occurring in non-sensitive areas, as exemplified in the method presented by Hogan et al. (1996). As this requires knowledge of the geographical location of the acidifying emissions (i.e


(1996). As this requires knowledge of the geographical location of the acidifying emissions (i.e. the country of origin), additional information must be collected in the Inventory phase. 2. Weighting emissions according to local sensitivity A more sophisticated method is to weight emissions according to the sensitivity of the area in which they are emitted. Examples of this approach are Blau & Seneviratne (1995), Tolle (1997) and Hauschild & Wenzel (1998). Blau & Seneviratne (1995) propose three different sensitivity classes for Europe and for the World, while Tolle (1997) proposes state-specific scaling factors for the USA. Hauschild & Wenzel (1998) propose the same characterisation factors as Heijungs et al. (1992). In addition, they propose using a ‘site factor’ to account for local circumstances that reduce the acidifying impact of certain substances in particular areas


(1992). In addition, they propose using a ‘site factor’ to account for local circumstances that reduce the acidifying impact of certain substances in particular areas. For instance, they report that 25% of the annual nitrogen input to forest ecosystems in Denmark is removed in harvested trees. This leads to a site factor of 0.75 for and in Danish forests. These methods also require the location of the acidifying emissions to be known


Part 3: Scientific background (Blau & Seneviratne: more or less at national level; Tolle: at state level; Hauschild & Wenzel: at ecosystem level, e.g. forest, natural area1). 3. Assessing a maximum and a minimum scenario Lindfors (1996) and Nichols et al. (1996) propose an entirely different approach. They suggest that two acidification scenarios should be studied: a maximum scenario including the contribution of and and a minimum scenario excluding the contribution of these compounds, which may vary widely depending on soil conditions (influencing anion leaching) and ecosystem management (biomass removal). 4. Modeling regional sensitivity and fate The above methods all take account of the sensitivity of the region in which the acidifying substances are emitted but not for the subsequent fate of the substance. Potting et al


Potting et al. (1998) propose an approach that includes both fate and regional sensitivity based on a dispersion model developed by EMEP, the Cooperative Program for Monitoring and Evaluation of the long-range transmission of air pollutants in Europe (Amann et al., 1996; Barret et al., 1996) and on the acidification model RAINS (Posch et al., 1997): with with the total acidification; ecosystem e in grid cell j; the critical load for ecosystem e in grid cell j; the deposition on grid cell j; a step function which is 0 if the deposition is below the CL of ecosystem e and 1 if it is above; a transport factor: the fraction of deposited on j; the emission of substance x in region r. Although it is an advantage that this method takes account of regional differences in sensitivity and fate, there is one concern: the use of a step function. For each type of ecosystem distinguished by RAINS a critical acidification load for sulphur and nitrogen is calculated based on ecosystem properties


For each type of ecosystem distinguished by RAINS a critical acidification load for sulphur and nitrogen is calculated based on ecosystem properties. Europe is divided into a very large number of grid cells, and for each of these a cumulative distribution can be made of the critical loads of all the ecosystems it comprises; see Figure 4.3.10.12. This step function is used to find the change in unprotected ecosystem area as a result of a change in deposition (which is in turn the result of a change in acidifying emissions). 1 As emissions do not frequently occur in such natural areas, however, this ecosystem-based method appears to require a fate component. 2 In reality this is rather more complicated because Rains-Europe now uses critical load functions represented by isolines for combined sulphur and nitrogen loads. However, the principle of calculating critical load exceedance remains the same.


Part 3 : Scientific background Because of this step function a marginal increase in emission will not generally lead to a change in the ecosystem area that is unprotected, although occasionally there will be a relatively large change. In other words, the derivative of is either zero or infinity. This is certainly not a suitable basis for a characterisation factor, which is designed to describe the effects of marginal changes in emissions. This becomes less problematical if the number of kinds of ecosystem in a grid cell is sufficiently large and a sufficiently large ‘marginal’ increase in emissions is used to derive acidification factors. Potting therefore takes a 10% increase in regional emission as a ‘marginal’ change, which he claims is sufficient to overcome this problem. If a smaller change had been chosen, however, the resulting characterisation factors would have been different. The equivalency factors are thus highly dependent on the choice of ‘marginal’ emission change


The equivalency factors are thus highly dependent on the choice of ‘marginal’ emission change. Huijbregts (1999b) has therefore proposed an improvement. He suggests using not the change in unprotected ecosystem area as the effect, but the change in relative risk. This relative risk is characterised by the ratio between deposition and critical load. Huijbregts thus replaces: by This solves the problem by changing the step function into a continuous function, with the characterisation factor for acidification now based on a deposition/critical load ratio simular to the PEC/PNEC ratio used for ecotoxicity. Based on this relative risk and some other minor changes to RAINS, Huijbregts has adapted RAINS to the purposes of LCA, calling it RAINS-LCA, and using it to calculate regional characterisation factors for acidification. A second concern is the need for global characterisation factors in LCA. Potting et al


A second concern is the need for global characterisation factors in LCA. Potting et al. (1998) have calculated regional characterisation factors for approximately 40 regions in Europe, but no average European or global factors. Huijbregts (1999b) has calculated average European factors, by weighted summation of the regional factors for each acidifying substance: with: the regional acidification potential of substance x in region r; ecosystem e (in grid cell j); the critical load for ecosystem e (in grid cell j); a transport factor: the fraction of deposited on j; the emission of substance x in region r. In the ExternE project (EC, 1995a) three different types of effects of acid rain are modeled separately: effects on ecosystems, effects on buildings, and effects on recreational fishery. Substance fate is also


Part 3: Scientific background PROSPECTS modeled using two different models: a Gaussian plume model and a trajectory model. However, ExternE does not provide straightforward characterisation factors that can be used directly in LCA. For the Eco-indicator 99 (see Section 4.2), damage factors were calculated using the so-called ‘Nature Planner’, a fate and effect model developed for the Netherlands (Goedkoop & Spriensma, 1999). The authors have expressed doubts as to whether the model is still valid at the European or global scale, however. CONCLUSIONS In the future, further progress is anticipated on Impact assessment modeling to the various endpoints, viz. damage to human health, damage to ecosystem health, etc. cf. Figure 4.2.2); see Section 4.2 for a more extensive discussion of this subject). Concluding, the average European characterisation factors of Huijbregts (1999b) are recommended for the time being as the best available practice


Concluding, the average European characterisation factors of Huijbregts (1999b) are recommended for the time being as the best available practice. Regional factors have not been adopted as the baseline, because it is not always possible, nor desirable, to consider differences between emission sites in LCA. It is therefore important that emission site-independent characterisation factors become available, even for those impact categories for which local sensitivity is important (criterion 10 for selection of baseline characterisation method). The average European factors of Huijbregts (1999b) are better linked to the category endpoint natural environment than the old factors of Heijungs et al


The average European factors of Huijbregts (1999b) are better linked to the category endpoint natural environment than the old factors of Heijungs et al. (1992), even when used together with first three options above: ‘maximum and minimum scenarios for and (Lindfors, 1996; Nichols et al., 1996), ‘neglecting emissions in non-sensitive areas’ (Hogan et al., 1996) and ‘weighting of emissions according to the sensitivity of the area in which they are emitted’ (Hauschild & Wenzel, 1998), because they account for the fate and sensitivity of the receiving ecosystem (criteria 2 and 8). The ExternE method is not operational in terms of characterisation factors (criterion 11) and the method of Goedkoop & Spriensma has an important source of uncertainty (criterion 12): “the very crude assumption that the average sensitivity of Dutch natural areas is representative for the average sensitivity of other natural areas“ (Goedkoop & Spriensma, 1999)


The baseline characterisation method for acidification, using the AP based on RAINS-LCA, is evaluated with respect to the (ISO-based) criteria in Table 4.3.10.1. Table 4.3.10.1: Evaluation of the baseline characterisation method for acidification, using the AP based on RAINS-LCA, with respect to the (ISO-based) criteria of Table 4.3.1 . 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12


1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. criterion scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty margins evaluation yes, based on the RAINS model, supported by United Nations Economic Commission for Europe (UN/ECE) yes, near-endpoint level not officially, but RAINS is supported by UN/ECE present, but accepted by an international community near endpoint yes eternity all included yes, effects below threshold are also included yes, although in fact they are representative for Europe yes, for the three main acidifying substances: and greater than for the method using the APs of Heijungs et al. (1992) The following method is included in the guide as an option for sensitivity analysis: The method using the acidification potentials of Heijungs et al. (1992)


(1992) The following method is included in the guide as an option for sensitivity analysis: The method using the acidification potentials of Heijungs et al. (1992). Several substances have been added to the 1992 list (source: Hauschild & Wenzel, 1998). This data set is useful when substances otherthan or are involved. Recommendations for extended LCAs: If most emissions take place in Europe, the regional characterisation factors of Huijbregts (1999b) should be used. This means gathering extra information in the Inventory phase on the region of origin of the acidifying emissions. (


Part 3 : Scientific background method status baseline alternative additional variant characterisation method/factor average European AP generic AP region (site) dependent AP – reference Huijbregts, 1999b Heijungs et al. , 1992 (updated with 1998) Huijbregts, 1999b Hogan et al., 1996; regional factors of 1998; Potting et al.(1998) Hauschild & Wenzel, Hauschild & Wenzel, RESEARCH RECOMMENDATIONS Long-term research: It is desirable to develop a standard method to account for fate and regional differences for all impact categories. To this end a research programme should be initiated focusing on all impact categories for which fate and regionalisation may be important, including acidification. 4.3.11 Eutrophication Eutrophication covers all potential impacts of excessively high environmental levels of macronutrients, the most important of which are nitrogen (N) and phosphorus (P)


4.3.11 Eutrophication Eutrophication covers all potential impacts of excessively high environmental levels of macronutrients, the most important of which are nitrogen (N) and phosphorus (P). Nutrient enrichment may cause an undesirable shift in species composition and elevated biomass production in both aquatic and terrestrial ecosystems. In addition, high nutrient concentrations may also render surface waters unacceptable as a source of drinking water. In aquatic ecosystems increased biomass production may lead to a depressed oxygen levels, because of the additional consumption of oxygen in biomass decomposition (measured as BOD, biological oxygen demand). As emissions of degradable organic matter have a similar impact, such emissions are also treated under the impact category ‘eutrophication’. The areas of protection are the natural environment, natural resources and the man-made environment (see Figure 4.2.2)


The areas of protection are the natural environment, natural resources and the man-made environment (see Figure 4.2.2). SETAC-Europe has also assigned the emission of waste heat to this impact category. In this guide, however, waste heat is treated as a separate category (see Section 4.3.12). TOPIC


Part 3: Scientific background DEVELOPMENTS IN THE LAST DECADE Hauschild & Wenzel (1998) take more or less the same approach as Heijungs (1992). They calculate nutrient enrichment potentials relative to this being one of the key nutrifying agents. They also base Heijungs et al. (1992) In Heijungs et al. (1992) the eutrophication (there termed ‘nutrification’) due to N, P and C (measured in terms of Chemical Oxygen Demand, COD) was aggregated by quantifying their potential contribution to biomass formation. Eutrophication potentials (EPs) were used as a characterisation factor to assess and aggregate the interventions for the impact category eutrophication: where is the mass of substance i released to the air, water or soil, the Eutrophication Potential of the substance and Eutrophication the indicator result, which is expressed in kg -equivalents. (dimensionless) reflects a substance’s potential contribution to biomass formation


(dimensionless) reflects a substance’s potential contribution to biomass formation. As a formula: where and are the potential contributions to eutrophication of one mole of substance i and ref (i.e. respectively,and and are the mass of i and ref (i.e. EPs are based on the ‘average’ chemical composition of aquatic organisms: (Stumm & Morgan, 1981), assumed to be representative of average biomass. The only emissions considered to be eutrophying by Heijungs et al. (1992) were emissions of nitrogen and phosphorus compounds. They determined the contribution of each of these nutrients to biomass formation, assuming unlimited supply of other nutrients. In this approach one mole of biomass requires 16 moles of N and 1 mole of P. When degradable organic matter is emitted, the required forits degradation can be measured as chemical oxygen demand (COD)


When degradable organic matter is emitted, the required forits degradation can be measured as chemical oxygen demand (COD). It is assumed that degradation of one mole of biomass requires 138 moles of Therefore if the contribution toeutrophication, of one mole of P is 1, the contribution of one mole of N, is 1/16 and the contribution of COD (as is 1/138. The contribution of one mole is then expressed as the contribution of one gram by dividing by the molecular weight. In the formula for is the total contribution of one mole of substance i and is the molecular weight of i. The reference substance is used to create eutrophication potentials (EPs), similar to ODPs, GWPs, APs and POCPs. Hence, the characterisation factor for eutrophication is independent of whatever substance happens to be the limiting factor in a particular location. This approach was adopted for two reasons: to obtain universal, i.e


This approach was adopted for two reasons: to obtain universal, i.e. global characterisation factors, independent of local differences, and because it is unknown which medium (freshwater, salt water, groundwater or soil) an emitted substance will eventually enter. these potentials on the assumption that one mole of P contributes as much to the formation of biomass as 16 moles of N. Besides this overall nutrient enrichment potential they also propose two separate potentials for nitrogen and phosphorus, expressed as gram N and gram P per gram substance, which can be used if it is known on the basis of site-dependent information which of the nutrients is limiting. Only biologically available nitrogen is considered is not classified). Hauschild & Wenzel (1998) do not assign BOD or waste heat to the impact category eutrophication. A general comment in the literature is that the approach of Heijungs et al


Hauschild & Wenzel (1998) do not assign BOD or waste heat to the impact category eutrophication. A general comment in the literature is that the approach of Heijungs et al. (1992) disregards the media of emission as well as the sensitivity of the receiving environment and the limiting nutrient. Two suggestions have been made for overcoming these limitations: by distinguishing ecosystem subcategories, and by including fate and site- or region-dependent effect modeling. The first solution, proposed by Nichols et al. (1996) and by Lindfors et al. (1995c), involves distinguishing ecosystem subcategories. Nichols et al. distinguish 3 subcategories: i) ii) iii) terrestrial ecosystems, aquatic ecosystems: smaller inland surface waters, and aquatic ecosystems: larger inland surface waters and seas. Lindfors et al. (1995c) propose 5 subcategories representing 5 scenarios: 1. 2. 3. 4


Lindfors et al. (1995c) propose 5 subcategories representing 5 scenarios: 1. 2. 3. 4. terrestrial ecosystems, emissions of N to air (because most terrestrial ecosystems are N-limited) aquatic ecosystems, P-limited; emissions of P and organic matter to water aquatic ecosystems, N-limited; emissions of N and organic matter to water aquatic ecosystems, N-limited; emissions of N to air and water and organic matter to water


Part 3 : Scientific background 5. aquatic ecosystems, emissions of N, P and organic matter to water and N and P to air. The emissions in these scenarios can be characterised in terms of either - or -equivalents or oxygen demand. In the latter case the characterisation factor of substance i, in g oxygen/kg i, is the oxygen required for the mineralisation of the organic matter (average composition) produced from one kg of i when i is the limiting nutrient, with one mole of N and P corresponding respectively to 8.6 and 138 moles of consumed The two types of units for the characterisation factor boil down to the same and are in fact interchangeable. The second option suggested is to include fate and site- or region-dependent effect modeling. Besides acidification factors, Huijbregts (1999b) calculated eutrophication factors for air emissions of and using RAINS-LCA, which also contains critical loads for eutrophication. He calculated both regional and average European factors


He calculated both regional and average European factors. These factors account only for eutrophication of terrestrial ecosystems due to air emissions. Direct emissions to soil are not included, nor are emissions to water. Effects on aquatic ecosystems are likewise not included. For the Eco-indicator 99, damage factors were calculated using the so-called ‘Nature Planner’, a fate and effect model developed for the Netherlands (Goedkoop & Spriensma, 1999). The authors have expressed doubts as to whether the model is still valid at the European or global scale, however. PROSPECTS In the future, further progress is anticipated on Impact assessment modeling to the various endpoints, viz. damage to human health, damage to ecosystem health, etc. (cf. Figure 4.2.2); see Section 4.2 for a more extensive discussion of this subject). Huijbregts & Seppälä (2000) recently proposed a method for LCA assessment of the effects of eutrofication on aquatic ecosystems


Huijbregts & Seppälä (2000) recently proposed a method for LCA assessment of the effects of eutrofication on aquatic ecosystems. This method combines the fate part of RAINS-LCA for air emissions with the characterisation factors of Heijungs et al. (1992). Water emissions are also included, using a fate factor of one (i.e. all emissions to water remain in water). CONCLUSIONS Concluding, our preference would be for assessment of eutrophication using a method that includes multi-media fate and exposure as well as the varying sensitivity of the ecosystems exposed. At present, however, such a method is only available for the terrestrial effects of air emissions of and (criterion 11)


At present, however, such a method is only available for the terrestrial effects of air emissions of and (criterion 11). Huijbregts & Seppälä (2000) propose using this method for terrestrial eutrophication in conjunction with a separate method for aquatic eutrophication, combining a fate factor (calculated with RAINS-LCA for air emissions of and and set to 1 for water emissions) with the old characterisation factors of Heijungs et al. (1992). However, this proposal has two drawbacks: i) emissions to soil are not included, although these may be particularly relevant in LCA studies covering agricultural products (criterion 11), and ii) using different methods for terrestial and aquatic eutrophication would yield two subcategories that cannot simply be summed. The environmental profile would thus contain two scores instead of one for eutrophication


The environmental profile would thus contain two scores instead of one for eutrophication. The method of Goedkoop & Spriensma has an important source of uncertainty (criterion 12): “the very crude assumption that the average sensitivity of Dutch natural areas is representative for the average sensitivity of other natural areas“ (Goedkoop & Spriensma, 1999). We therefore propose adopting as a baseline the method described in Heijungs et al. (1992), in which all emissions of N and P to air, water and soil and of organic matter to water are aggregated into a single measure, because this method allows both terrestrial and aquatic eutrophication to be assessed. The characterisation factors in -equivalents, - equivalents and are all interchangeable, and in this Guide we opt to use the same concept and factors as Heijungs et al., in -equivalents. Several substances have been added to the original 1992 list


Several substances have been added to the original 1992 list. The baseline characterisation method for eutrophication, using the eutrophication potentials of Heijungs et al. (1992) as a characterisation factor, is evaluated with respect to the (ISO-based) criteria adopted in this Guide in Table 4.3.11.1.


Part 3: Scientific background Table 4.3.11.1: Evaluation of the baseline characterisation method for eutrophication, using the eutrophication potentials of Heijungs et al. (1992) as a characterisation factor, with respect to the (ISObased) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. criterion scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty margins evaluation not relevant category indicator is close to intervention no, but widely used present (e.g


fixed C÷N÷P ratio is representative for biomass, terrestrial and aquatic) midpoint yes eternity fate and exposure not included yes, effects below threshold also included yes yes, for many substances containing bioavailable N P and COD low all or The following method is included in the Guide as an option for sensitivity analysis: The method using the average European eutrophication potentials for terrestrial ecosystems (Huijbregts, 1999b). Recommendations for extended LCAs: If most emissions take place in Europe, the regional characterisation factors of Huijbregts (1999b) should be used. This means gathering extra information in the Inventory phase on the region of origin of the eutrophying emissions. If there are compounds containing bioavailable N or P that are relevant for the study but are not included in the list, calculate characterisation factors using the formula specified above


If there are compounds containing bioavailable N or P that are relevant for the study but are not included in the list, calculate characterisation factors using the formula specified above. method status baseline alternative additional variant characterisation method/factor generic EP average European EP region (site) dependent EP – reference Heijungs et al. , 1992 (updated) Huijbregts, 1999b Huijbregts, 1999b RESEARCH RECOMMENDATIONS Long-term research: It is desirable to develop a standard method to account for fate and regional differences for all impact categories. To this end a research programme should be initiated focusing on all impact categories for which fate and regionalisation may be important, including eutrophication. 4.3.12 Waste heat TOPIC Emissions of waste heat may increase temperatures on a local scale: in a city or lake, for example. They cannot contribute to global warming on a scale such as that associated with emissions of greenhouse gases


They cannot contribute to global warming on a scale such as that associated with emissions of greenhouse gases. The effects on ecosystems of waste heat emissions to the air are negligible. Depending on local conditions, the discharge of waste heat into surface waters may result in a substantial temperature rise, with a consequent impact on local aquatic ecosystems. SETAC-Europe includes waste heat under the impact category ‘eutrophication’, because it may likewise lead to lower oxygen concentrations (measured as COD, for example). In this Guide waste heat is treated as a separate impact category, although it covers only aquatic emissions of waste heat such as cooling water emissions from power stations. The areas of protection are the natural environment and natural resources.


Part 3 : Scientific background DEVELOPMENTS IN THE LAST DECADE Heijungs et al. (1992) Waste heat is treated as a separate impact category in Heijungs et al. (1992). The local character of the problem is abstracted by expressing waste heat emissions into water in megajoules (MJ). These emissions can be quantified by estimating the proportions of waste heat discharged via the cooling water and via the air (based on the energy balance of the plant). As this data is available from the Inventory analysis, no further action is required in the classification step. In essence, the characterisation factor for all waste heat emissions to water is 1, and 0 for all waste heat emissions to the atmosphere. There are no new developments relating to the topic of heat emissions to surface water. PROSPECTS No specific developments are foreseen in this area. CONCLUSIONS We propose using the method described in Heijungs et al. (1992)


PROSPECTS No specific developments are foreseen in this area. CONCLUSIONS We propose using the method described in Heijungs et al. (1992). This means that all heat emissions to water are multiplied by a charaterisation factor of 1. This baseline characterisation method is evaluated with respect to the (ISO-based) criteria adopted in this Guide in Table 4.3.12.1. Table 4.3.12.1: Evaluation of the baseline characterisation method for waste heat, using characterisation factor = 1, with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12


1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. criteria scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty evaluation not relevant category indicator is close to intervention no no close to intervention yes not relevant fate and exposure not included yes yes yes very low No methods are included in this Guide as options for sensitivity analysis. No additional recommendations are given for extendend LCAs. method status baseline alternative additional variant characterisation method/factor unweighted aggregation of energy – – – reference Heijungs et al., 1992 – – – RESEARCH RECOMMENDATIONS No research is foreseen.


Part 3: Scientific background 4.3.13 Odour TOPIC Odour becomes a problem when a given concentration of odorous substances is experienced as unpleasant. Whether an odour is experienced as stench will depend on the particular individual exposed. Above a certain emission level, however, every individual will experience it as such. Here, the term odour will be used for effects. The area of protection is human health. Odour may be defined as the observed difference between a sample of clean air and a sample of contaminated air. The concentration at which such a difference cannot quite be observed varies from substance to substance, and depends on the physical and chemical properties of the substance (Brasser et al., 1985)


The concentration at which such a difference cannot quite be observed varies from substance to substance, and depends on the physical and chemical properties of the substance (Brasser et al., 1985). The odour threshold value of a substance is defined as the concentration of that substance under defined standard conditions at which 50% of a representative sample of the population can just detect the difference between a sample of air mixed with that substance and a sample of clean air. Odour can be measured fairly objectively, while odour nuisance is more a matter of individual sensitivity. DEVELOPMENTS IN THE LAST DECADE We are not aware of any new developments in LCIA relating to the topic of odour. Indeed, most LCIA methodologies do not even have an impact category ‘odour’. Lindfors et al. (1995c) propose that odour be assigned to their impact categories ‘human health’ and ‘habitat alterations’


Lindfors et al. (1995c) propose that odour be assigned to their impact categories ‘human health’ and ‘habitat alterations’. Within the former category they suggest using the same method for odour as Heijungs et al. (1992). The only difference is the name of the impact category. Within the category ‘habitat alteration’ no category indicator has yet been operationalised for odour. Heijungs et al. (1992) In Heijungs et al. (1992) emissions of odorous substances were classified using a method similar to the critical volumes approach, by dividing the emission of a potentially malodorous substance by the odour threshold value of that substance


A distinction must be made between emissions of potentially malodorous substances to the atmosphere and to water, for each is associated with a different odour threshold value, as expressed in the following formulae: where malodorous air is the quantity of air contaminated to the odour threshold value the emission of substance i into the atmosphere (kg) and the odour threshold value in air of substance i and where malodourous water and represent the same parameters for water. This provisional, critical volumes approach does not account for processes of dispersal, degradation or transformation, which are highly dependent on the particular substance involved. This also means there is no assessment of indirect odour emissions such as ozone. Direct ozone emissions are negligible, particularly in comparison with the ozone formed by photochemical reactions involving volatile organic compounds and (see also Section 4.3.9)


Direct ozone emissions are negligible, particularly in comparison with the ozone formed by photochemical reactions involving volatile organic compounds and (see also Section 4.3.9). This provisional approach was only partially adopted in the old Guide, the main reason being that no uniform odour threshold values had yet been agreed for many substances. For atmospheric emissions a comprehensive list of odour thresholds was calculated, however, using a uniform method. No such list was yet available for water. PROSPECTS Because odour is not generally considered to be as significant a category as, say, human health or climate change, the proposal of Lindfors et al. (1995c) to include odour in an impact category ‘human health’ seems reasonable. If and when such an impact category is distinguished aggregating all human health impacts, in the form of DALYs (see Section 4.2) for example, odour might be assigned to this new category


If and when such an impact category is distinguished aggregating all human health impacts, in the form of DALYs (see Section 4.2) for example, odour might be assigned to this new category. Until then, it should be possible to develop Odour potentials (OPs) in the same way as the human toxicity potentials and aquatic ecotoxicity potentials developed by Guinée et al. (1996) or Huijbregts (1999a). A multi-media model might then be used to calculate the fate of the substance and the resulting i


Part 3 : Scientific background environmental concentrations divided by the odour threshold. The resulting values could then, finally, be compared with the corresponding values for a reference substance. Although the analogy between OPs and HTPs may be an advantage if odour is assigned to human health, the effects on which the threshold values are based are nonetheless very different: adverse health effects for HTPs and annoyance or maybe headaches for OPs. CONCLUSIONS For the present it is recommended to use the critical volumes method described by Heijungs et al. (1992) for malodorous air. No method is yet available for malodorous water. It is proposed to include fate at some time in the future, possibly using a method analogous to that developed for toxic substances by Guinée et al. (1996) or by Huijbregts (1999a). The score of this approach with respect to the (ISO) requirements is shown in Table 4.3.13.1


(1996) or by Huijbregts (1999a). The score of this approach with respect to the (ISO) requirements is shown in Table 4.3.13.1. Table 4.3.13.1: Evaluation of the baseline category indicator for malodorous air, the critical volumes approach of Heijungs et al. (1992), with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12


(1992), with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. criteria scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty margins evaluation no more or less; odour is not the same as odour nuisance, which may affect human well-being no by adoptng the odour threshold level as a reference, a value choice is made to value every odour as malodorous midpoint yes not relevant fate and exposure not included yes, effects below threshold are also included yes yes unknown No methods are included in this Guide as options for sensitivity analysis. Recommendations for extended LCAs: Develop OPs by using USES-LCA for the fate part and 1/OTV for the effect part


Recommendations for extended LCAs: Develop OPs by using USES-LCA for the fate part and 1/OTV for the effect part. Malodorous air method status baseline alternative additional variant characterisation method/factor inverse OTV _ based on fate model and odour threshold values – reference Heijungs et al., 1992 – – – RESEARCH RECOMMENDATIONS Short-term research: It is recommended to investigate the scope for developing OPs that include fate. The potential for integrating odour in the categories ‘human health’ and possibly ‘ecosystem health’ should also be researched. In the case of ‘human health’ weighting factors might, for example, be developed to weight odour nuisance with reference to toxic effects. Malodorous water method status characterisation method/factor reference baseline alternative – – additional variant – – – – – –


Part 3: Scientific background 4.3.14 Noise TOPIC Noise, or noise nuisance, refers to the environmental impacts of sound. In principle, these impacts could cover at least human health and ecosystem health, but the environmental mechanisms are complex, non-linear and highly dependent upon local circumstances. Moreover, noise is similar to odour in that a given level of exposure is experienced differently by different individuals. Something considered a nuisance by one person might be appreciated by another, as exemplified by the case of loud music. Hence, whether or not sound waves will lead to ‘nuisance’ depends partly on the actual situation and partly on the person interviewed. DEVELOPMENTS IN THE LAST DECADE Heijungs et al. (1992) Heijungs et al. (1992) aggregated the sound production (in obtained in the inventory to yield an abstract sound level that is non-location- and non-person-specific and referred to it as “potential noise nuisance”


(1992) aggregated the sound production (in obtained in the inventory to yield an abstract sound level that is non-location- and non-person-specific and referred to it as “potential noise nuisance”. In this case the characterisation factor is essentially 1. This ignores the fact that some sound emissions may not cause any nuisance at all (e.g. source remote from those exposed) while others may cause a great deal of nuisance in certain environments (e.g. in a road tunnel os street canyon): where G is the sound production (in Most LCIA methodologies do not have an impact category ‘noise’. This runs counter to the observed fact that most people deem noise to be a major environmental problem. Lindfors et al. (1995c) propose assigning noise to the impact categories ‘human health’, ‘human health in the working environment’ and ‘habitat alterations’. Lafleche & Sacchetto (1999) describe a methodology that attempts to include noise in LCA


Lafleche & Sacchetto (1999) describe a methodology that attempts to include noise in LCA. This methodology embodies a regional approach which is operationalised for one specific case: transport by car or truck along the Bologna-Milan highway. This methodology is still insufficiently developed for general use in LCA. Müller-Wenk (1999) propose a method for expressing road traffic noise in terms of DALYs (see Section 4.2). This method links transport kilometres directly to DALYs, thereby more or less skipping the inventory phase. As a consequence, this method is useful for road traffic noise only and not for aircraft noise and so on. PROSPECTS In the future, further progress is anticipated on Impact assessment modeling to the various endpoints, viz. damage to human health, damage to ecosystem health, etc. (cf. Figure 4.2.2); see Section 4.2 for a more extensive discussion of this subject). The method of Müller-Wenk (1999) might provide a starting point for endpoint modeling of noise


(cf. Figure 4.2.2); see Section 4.2 for a more extensive discussion of this subject). The method of Müller-Wenk (1999) might provide a starting point for endpoint modeling of noise. CONCLUSIONS For the present it is recommended to use the method described by Heijungs et al. (1992) as the baseline characterisation method, thus mutiplying all sound produced by a characterisation factor of 1. This method is evaluated with respect to the (ISO-based) criteria adopted in this Guide in Table 4.3.14.1.


Part 3 : Scientific background Table 4.3.14.1: Evaluation of the baseline characterisation method for noise, using characterisation factor = 1, with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. No criteria scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty margins evaluation not relevant category indicator is close to intervention no no close to intervention yes not relevant fate and exposure not included yes yes yes, verylow methods are included in this Guide as options for sensitivity analysis. Recommendations for extended LCAs: – The method of Müller-Wenk (1999) is recommended as a supplementary method for extended LCAs in which road traffic noise is an important item


Recommendations for extended LCAs: – The method of Müller-Wenk (1999) is recommended as a supplementary method for extended LCAs in which road traffic noise is an important item. method status baseline alternative additional variant characterisation method/factor unweighted aggregation of sound _ based on DALY – reference Heijungs et al., 1992 _ Müller-Wenk, 1999 – RESEARCH RECOMMENDATIONS Long-term research: It is recommended to investigate the scope for future assignment of noise to the categories ‘human health’ (in terms of DALYs) and perhaps ‘ecosystem health’. It is also recommended to develop DALY-based characterisation factors for the intervention noise, or sound. 4.3.15 Impacts of ionising radiation TOPIC The impact category ‘impacts of ionising radiation’ covers the impacts arising from releases of radioactive substances as well as direct exposure to radiation, in building materials for example. Exposure to ionising radiation is harmful to both human beings and animals


Exposure to ionising radiation is harmful to both human beings and animals. The areas of protection are therefore human health, the natural environment and natural resources (see Figure 4.2.2). Ionising radiation is expressed in terms of the number of atoms disintegrating (or decaying) per unit time. The SI unit of radioactivity is the becquerel (Bq), one Bq corresponding to one disintegration per second. The radioactivity of a substance is expressed in or Radioactivity always declines in the course of time and the time taken for the radioactivity of a given substance to decline by half is known as the half-life of the substance. The amount of radioactive material is generally expressed in mass terms, i.e. in kg


The amount of radioactive material is generally expressed in mass terms, i.e. in kg. The relationship between emitted radiation and mass m is given by: where is the radiation in Bq, m the mass in kg, the half-life in s, M the molecular mass in and is Avogadro’s number in Different forms of radiation may be released in the process of radioactive disintegration, viz. alpha, beta, gamma and neutron radiation and X-rays. These forms of radiation may add or remove electrons to or from the atoms they encounter, i.e. ionise them. The degree of ionisation depends on the type of


Part 3: Scientific background radiation and the material irradiated. During ionisation of a material, energy is transferred to that material. As a result the energy of the radiation is reduced and after a number of ionisations will have been fully absorbed by the irradiated material or tissue. The energy absorbed per unit mass material or tissue is known as the absorbed dose (Jonker et al., 1988). The different types of ionising radiation impinge very differently on living tissue. The effects of a given dose of alpha radiation will differ from those resulting from the same dose of gamma radiation, for example (Jonker et al., 1988). This difference in harmfulness is expressed by means of a so-called ‘quality factor’: 20 for alpha radiation, 1 for beta and gamma radiation and X-rays, and 10 for neutron radiation (these types of radiation are abbreviated to X and n, respectively). Multiplying the absorbed dose by this quality factor yields the ‘dose equivalent’, expressed in sievert (Sv)


Multiplying the absorbed dose by this quality factor yields the ‘dose equivalent’, expressed in sievert (Sv). Thus, one dose equivalent of alpha radiation is assumed to have the same aggregate effect as twenty dose equivalents of gamma radiation. It is to be noted that characterisation of the impacts of ionising radiation will remain unfeasible until such time as data on emissions of radioactive substances are included in the Inventory analysis. DEVELOPMENTS IN THE LAST DECADE Most LCIA-methodologies do not have an impact category ‘impacts of ionising radiation’. Lindfors et al. (1995c) propose assigning ionising radiation to the impact categories ‘human health’ and ‘ecotoxicological impacts’. They do not operationalise a category indicator, however. In connection with the impact category ‘impacts of ionising radiation’ Heijungs (1994b) distinguishes two Heijungs et al. (1992) Heijungs et al. (1992) discussed the scope for including ionising radiation as an LCA impact category


(1992) Heijungs et al. (1992) discussed the scope for including ionising radiation as an LCA impact category. They saw scope for developing ionisation creating potentials (ICPS) based on models used to calculate the ratio between the absorbed and equivalent doses of a substance and the absorbed and equivalent doses of a reference substance. They also discuss assessment of the impacts of ionising radiation by means of an adapted critical volumes approach, by dividing emissions of a radionuclide by some form of ‘radiation standard’ for the substance in question. For lack of a better alternative they propose using ALIs (Annual Limits of Intake: defined for workplace exposure by the International Commission on Radiological Protection (ICRP, 1979, 1991)). It should be noted that the critical volumes approach is a worst-case approach and ignores differences between substances in terms of half-lives and absorbed doses. However, Heijungs et al


It should be noted that the critical volumes approach is a worst-case approach and ignores differences between substances in terms of half-lives and absorbed doses. However, Heijungs et al. (1992) do not ultimately include a characterisation method for (the impacts of) ionising radiation. types of emission: emission of radioactive substances to the environment, with subsequent release of radiation during decay; direct emission of radiation to the environment, without actual emission of a radionuclide (e.g. from building materials); and two types of radiation impact: impacts resulting from internal exposure to radiation, due to intake of air, food or drinking water; impacts resulting from external exposure to radiation, due to the presence of radioactive substances in the human environment or in ecosystems. Internal exposure can result only from physical emissions of radionuclides


Internal exposure can result only from physical emissions of radionuclides. For this kind of exposure Heijungs proposes a method including fate, broadly similar to the method for human toxicity adopted by Guinée et al (1996). Instead of ADI (Acceptable Daily Intake) the aforementioned ALI (Annual Limit of Intake) might be used. A possible provisional method is the critical volumes approach: where is the emission of substance i and the Annual Limit of Intake for that substance in Bq/kg of body weight. External exposure can result from both types of emission: emission of radionuclides or direct emission of radiation. Adequate characterisation of this kind of exposure requires inclusion in the inventory of at least three technical parameters:


Part 3 : Scientific background the magnitude of the source in Bq; the characteristic decay energy of the radionuclide concerned in eV; the type of particle emitted during decay n). If all these properties are known, one can sum overall external exposure to emitted radiation and to emitted radioactive substances, as expressed in the following formula: where is the quality factor of type i radiation (1 for and 0.1 for n, 0.05 for the magnitude of the source (Bq) and the average decay energy of the radionuclide involved (eV; electronvolt, a unit of energy used in most tables of radionuclides). Heijungs sees two severe limitations to this approach, however: it does not include the relationship between radiation and absorbed dose, and the assumed linear additivity of low and high energy radiation is unrealistic


Solberg-Johansen (1998) suggests a different approach for characterising the risk to humans of emissions of radioactive substances which allows the probability of human exposure to be included. It is based on the following scheme (Figure 4.3.15.1): In thefigure, is the activity of an emission pulse of radionuclide i (Bq), a ‘dispersion factor’ based on the dispersion ‘models’ used ‘exposure concentration’ of i in medium j and is a ‘screening factor’ for i in j This screening factor is based on the exposure pathways for human beings (via food, air and external radiation) and the dose per unit intake then represents the dose incurred due to emission of i to compartment j (Sv). F is a probability coefficient expressing the probability of occurrence of detrimental health effects like cancer or hereditary disease is then the contribution to human irradiation by radionuclide i, defined as the annual risk of detrimental health effects


Screening factors and F-values for a range of radionuclides have been published by the International Commission on Radiological Protection (ICRP, 1979, 1990 and 1991). The different screening levels in Figure 4.3.15.1 stand for different levels of complexity in the dispersion models employed. Level I embodies a very simple and conservative approach which assumes that the concentration in the receiving environment is equal to the concentration in the stream emitted: the


Part 3: Scientific background where is the activity of an emission pulse of radionuclide i (Bq), the dispersion factor for radionuclide i in environmental medium j the screening factor for i in j the volume flow rate of the discharge to j F the probability coefficient Level II accounts for dispersion in the atmosphere and surface waters by using a Gaussian plume model for emissions to air and an advection-diffusion model for emissions to water. Intermedia transport is not accounted for. Level III, developed for air emissions only, uses a Gaussian plume model in combination with individual screening factors for the various exposure pathways. Levels II and III require site-specific data for the dispersion models. Solberg-Johansen recommends using the level II procedure in LCA


Levels II and III require site-specific data for the dispersion models. Solberg-Johansen recommends using the level II procedure in LCA. However, this would mean that instead of using substance-specific characterisation factors, for each local process emitting radionuclides a specific dispersion model would have to be used to estimate the fate of these substances. A regionalised inventory is also needed, including information on weather conditions and/or conditions in the surface waters concerned. Solberg-Johansen states that the level I procedure might be used as a global approach. In this case the only ‘site-specific’ information required is the volume flow rate of the discharge in which the radionuclide was emitted. Both the level I and the level II procedure are operational. provided that the appropriate information is collected in the Inventory phase


Both the level I and the level II procedure are operational. provided that the appropriate information is collected in the Inventory phase. Besides the risk to human health posed by radioactive emissions, Solberg-Johansen (1998) also introduces two other topics: the human risk posed by storage of radioactive solid waste and the general risk to the environment of radionuclides. For both of these she has also developed methods, which we briefly describe. The risk posed by radionuclides in solid waste was estimated on the basis of on a specifically British waste management scenario: Low Level Waste (LLW): characterisation factors based on a site-specific study on the Driggs disposal site by Smith et al. (1998); Intermediate Level Waste (ILW): characterisation factors based on a study by PAGIS (PAGIS, 1988); High Level Waste (HLW): characterisation factors based on a study by PACOMA (Mobbs et al., 1991). Characterisation factors are expressed in for LLW, and in for ILW and HLW


Characterisation factors are expressed in for LLW, and in for ILW and HLW. Inventory data should therefore be in Bq.yr or Bq per nuclide. As the method is based on the UK situation, it is probably not directly applicable to other waste management scenarios. Although characterisation factors are presented for each scenario, no full description is given of how these were calculated. The method adopted by Solberg-Johansen to assess the environmental impacts of radionuclide emissions is based on the same fate models as those used for human impacts. It is only the effect component that differs, the factor for human exposure in the previous equation being replaced by: where is the Environmental Increment factor for radionucide i in medium j, defined as one standard deviation of the mean background concentration, and is the half-life of radionuclide i


This method for calculating environmental impacts of exposure is operational for emissions of a number of radionuclides to water and soil, but not to air. and


Part 3 : Scientific background Frischknecht et al. (2000) have also developed a method to assess human radiation impacts (see Figure 4.3.15.2): with: radiation the indicator result (yr); the characterisation factor for substance i emitted to compartment ecomp based on DALYs the activity of substance i emitted to compartment ecomp (kBq).1 The method accounts for fate, exposure and effect. Fate modeling is based on EC (1995b), using a Gaussian plume model (for France) for air releases and a simple (regional) box model for water releases. For several globally dispersed radionuclides, global models are used in addition to the regional models. Exposure modeling is based on EC (1995b) and UNSCEAR (1993), and exposure calculated as absorbed dose. Effect modeling uses information on carcinogenic and hereditary effects (sources: Ron & Muirhead, 1998; ICRP, 1990) and the DALY approach (Murray & Lopez, 1996 and Hofstetter, 1998; cf. text box in Section 4.2)


text box in Section 4.2). The Frischknecht method yields characterisation factors incorporating fate, exposure and effect. Two sets of factors are available: for a time horizon of 100,000 years, without ageweighting of DALYs, and for a horizon of 100 years, with age-weighting. 1 We have written the unit for the indicator result and for the characterisation factor, where Frischknecht et al. (2000)write’DALYs’and This is in agreementwith Section 2.4, where itwasconcludedthat SIunits are to be used. In fact, ’DALY’ can be seen as the name of the quantity (like ’length’), and ’yr’ as one possible unit for measurement (like ’metre’). ’yr’


Part 3: Scientific background PROSPECTS In the future, further progress is anticipated on Impact assessment modeling to the various endpoints, viz. damage to human health, damage to ecosystem health, etc. (cf. Figure 4.2.2); see Section 4.2 for a more extensive discussion of this subject). No further new developments are known to be in progress. CONCLUSIONS We recommend using the characterisation method of Frischknecht et al. (2000) as the currently best available practice for assessing the human impact of radioactive emissions to air and water. This method covers fate and exposure (criterion 8 for selection of baseline characterisation method). As a baseline we recommend using the characterisation factors without DALY age-weighting (criterion 4) based on the longest time horizon (=100,000 years; criterion 7). The method of Solberg-Johansen (1998) is operational for more radionuclides than that of Frischknecht et al. (2000)


The method of Solberg-Johansen (1998) is operational for more radionuclides than that of Frischknecht et al. (2000). However, the level I procedure of Solberg-Johansen (1998) includes only exposure and not fate. Although the more refined level II procedure does include fate, it is not yet operational in the form of characterisation factors. This level also requires detailed information on the local situation (including the distance between emission source and ‘receptor’). In most LCA studies this information will not be available, and the level II method therefore seems more suitable for local risk analysis than for LCA. The characterisation factors developed by Solberg-Johansen (1998) for the risks of radioactive solid waste are not fully described in her thesis. Consequently, they could not be properly assessed. Moreover, they require information on the composition of the nuclear waste to be collected in the inventory, and this is often not available


Consequently, they could not be properly assessed. Moreover, they require information on the composition of the nuclear waste to be collected in the inventory, and this is often not available. The baseline characterisation method for impacts of ionising radiation, the method of Frischknecht et al. (2000), is evaluated with respect to the (ISO-based) criteria adopted in this Guide in Table 4.3.15.1. Table 4.3.15.1: Evaluation of the baseline characterisation method for impacts of ionising radiation, the method of Frischknecht et al. (2000), with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12


(2000), with respect to the (ISO-based) criteria of Table 4.3.1. 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. criterion scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions focal point in environmental mechanism linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty margins evaluation yes, the characterisation factors are based on common dispersion models and pathways of human exposure yes, endpoint level no present endpoint yes 100,000 years all included yes, effects below threshold are also included representative for France yes, for 49 radioniclides (including Radon–222) unknown The following method is included in the guide as an option for sensitivity analysis. If there are substantial emissions of radionuclides that are not included in Frischknecht et al


If there are substantial emissions of radionuclides that are not included in Frischknecht et al. (2000), the level I procedure of Solberg-Johanson (1998) for emissions may provide valuable additional information, as it covers more nuclides. In such cases it is recommended to use this level I procedure for emissions. Recomendations for extended LCAs The level II procedure of Solberg-Johansen (1998) can be used to augment the baseline method if radionuclide emissions prove to be important in the results of the Impact assessment step and detailed local information on these emissions is available. method status baseline alternative additional variant characterisation method/factor ionising radiation damage factors screeningfactors–level I screening factors – level II – reference Frischknecht et al., 2000 Solberg-Johansen, 1998 Solberg-Johansen, 1998 –


Part 3 : Scientific background RESEARCH RECOMMENDATIONS 4.3.16 No method has yet been developed to assess direct external exposure to radiation, from building materials, for example. It is recommended to investigate the scope for integrating assessment of such exposure into the methods for exposure to releases of radioactive substances. Further research is necessary to assess whether current emissions of radionuclides have a significant impact on the natural environment. If this is the case, current methods for assessing the impacts of ionising radiation should be refined. Further research is also required to determine whether the method of Solberg-Johansen (1998) for assessing the risks of radioactive waste can be integrated with the method of Frischknecht et al. (2000). Moreover, due efforts should be made to obtain a more detailed picture of how both the short- and the long-term risks of both solid and liquid waste can be incorporated in the method


(2000). Moreover, due efforts should be made to obtain a more detailed picture of how both the short- and the long-term risks of both solid and liquid waste can be incorporated in the method. Long-term research: Casualties TOPIC This impact category refers to casualties resulting from accidents. The area of protection is human health (see Figure 4.2.2). In Heijungs et al. (1992) this impact category was referred to as ‘direct victims’, but below this unfortunate term has been replaced by ‘casualty’. DEVELOPMENTS IN THE LAST DECADE Heijungs et al. (1992) In the classification step, various categories of casualty would have to be weighted if casualties other than fatalities (slightly injured, seriously injured and so on) were to be included in the Inventory analysis. At the time the 1992 guide was being prepared there were no methods for such weighting, nor indeed was such a method needed, as graded data on accident victims were lacking. For this reason the Heijungs et al


For this reason the Heijungs et al. (1992) assessed and aggregated casualties by using a characterisation factor 1 : where C is the number of casualties (dimensionless) listed in the inventory table. Most LCIA-methodologies do not include an impact category ‘casualties’. Schmidt & Brunn Rasmussen (1999) describe a very useful method for including the working environment in LCA which encompasses casualties. It is based on a database developed by EDIP in which the working environment impacts per kilo of produced goods are listed for a number of economic activities. PROSPECTS If and when an impact category is distinguished aggregating all human health impacts, in the form of DALYs (see textbox in Section 4.2), casualties might be assigned to this new category. The method of Schmidt & Brunn Rasmussen (1999) might also be useful for including various impacts on human health in the working environment


The method of Schmidt & Brunn Rasmussen (1999) might also be useful for including various impacts on human health in the working environment. CONCLUSIONS Although the method of Schmidt & Brunn Rasmussen (1999) seems a promising starting point for developing a characterisation method for casualties, the scope of the present project precluded an indepth evaluation. We therefore recommend using the method of Heijungs et al. (1992), in which the number of fatal casualties is simply multiplied by a charaterisation factor of 1. In the future it is recommended to assign casualties to the impact category human health. The method of Heijungs et al. (1992) is evaluated with respect to the (ISO-based) criteria adopted in this Guide in Table 4.3.16.1.


Table 4.3.16.1: Evaluation of the baseline characterisation method for casualties, using characterisation factor = 1, with respect to the (ISO-based) criteria of Table 4.3.1. 2. 3. 4. 5. 6. 7. 8. 9. No criterion scientifically and technically valid environmentally relevant internationally accepted value-choices and assumptions evaluation not relevant category indicator is close to intervention no no focal point in environmental mechanism close to intervention linearity time span fate, exposure and effects less is better time- and location-independent operational uncertainty margins methods are included in this Guide as yes not relevant not applicable yes yes yes very low options for sensitivity analysis. No recommendations are given for extended LCA


No recommendations are given for extended LCA. method status baseline alternative additional variant characterisation method/factor unweighted aggregation of victims – – – reference Heijungs et al., 1992 – – – RESEARCH RECOMMENDATIONS Short-term research: Research should be initiated to investigate the scope for including the method of Schmidt & Brunn Rasmussen (1999) for the working environment. Long-term research: Research is required to develop a method to assign casualties to a future impact category ‘human health’. 4.3.17 Interventions for which characterisation factors are lacking TOPIC Many practical cases will involve emissions of toxic chemicals for which no toxicity potentials are listed in the tables with characterisation factors. The same applies to acidifying substances, ionising substances, depletable resources and so on. A general guideline given for extended LCAs is to calculate, estimate or extrapolate missing characterisation factors


A general guideline given for extended LCAs is to calculate, estimate or extrapolate missing characterisation factors. This will often be unfeasible, however, for lack of time or knowledge, for instance. In such cases these overlooked items should be discussed in a separate part of the impact assessment A distinction should be made between: interventions known to contribute to an impact category and for which no characterisation factor is available but for which a factor can be calculated, estimated or extrapolated; interventions known to contribute to an impact category but for which no characterisation factor can be found, calculated, estimated or extrapolated; interventions assumed to be environmentally relevant but not contributing to any of the selected impact categories; interventions assumed not to be environmentally relevant


As far as possible, all interventions for which no characterisation factors are available should be assigned to all impact categories to which they are known to contribute, based on the best available knowledge. These interventions shall than be listed in the environmental profile under the appropriate impact category. If possible, a qualitative or quantitative estimate of their potential environmental impact should be given. Quantitative estimation can, for example, be accomplished by finding an intervention similar in terms of chemical structure for which (a) characterisation factor(s) is/are available. Part 3: Scientific background 1.


Part 3 : Scientific background If it is not known to which impact category an intervention should be assigned it shall be listed in a separate category ‘Interventions for which characterisation factors are lacking‘. Interventions anticipated to be environmentally irrelevant may be excluded from the environmental profile, but this should be transparently justified in the LCA study report. CONCLUSIONS Interventions for which characterisation factors are lacking should be dealt with in the following manner: Interventions for which a characterisation factor can be calculated, estimated or extrapolated should be included in the environmental profile under the relevant impact category, accompanied by a clear explanation of the divergent status of the characterisation factor and the method used to obtain it


Interventions for which no characterisation factor can be calculated, estimated or extrapolated but which are known to contribute to one or more impact categories should be included in a separate part of the environmental profile labeled ‘Interventions for which characterisation factors are lacking‘, accompanied by all relevant additional information such as: substance name; emission compartment; amount emitted; impact category to which a contribution is suspected; if possible, an indication of the significance of the suspected impact. Interventions known to be of environmental relevance but contributing to an impact category that is not selected should be included in the environmental profile in the same way. Interventions expected to be environmentally irrelevant can be excluded from the environmental profile, but this should be transparently justified in the LCA study report