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Lehmann, A., Bach, V., & Finkbeiner, M. (2015). Product environmental footprint in policy and market

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Lehmann, A.; Bach, V.; Finkbeiner, M.

Product environmental footprint in policy

and market decisions: Applicability and

impact assessment

Accepted manuscript (Postprint)Journal article |

Product Environmental Footprint in Policy

and Market Decisions: Applicability and

Impact Assessment

Annekatrin Lehmann1*,Vanessa Bach1* and Matthias Finkbeiner1

1Chair of Sustainable Engineering, Department of Environmental Technology, Technische Universität Berlin, Berlin, Germany

Lehmann and Bach contributed equally to the work.

* Address correspondence to: vanes[email protected]

EDITOR'S NOTE:

This paper represents 1 of 7 articles in the special series “LCA Case Study Symposium 2013,” which

was generated from the 19th SETACLCA Case Study Symposium “LCA in market research and policy:

Harmonization beyond standardization, ”heldin November 2013, in Rome, Italy. This collection of

invited papers reflect the purpose of the symposium and focus on how LCA can support the decision-

making process at all levels, that is, industry and policy contexts, and how LCA results can be

efficiently communicated and be used to support market strategies.

ABSTRACT

In April 2013, the European Commission published the Product and Organisation Environmental Footprint

(PEF/OEF) methodology—a life cycle-based multicriteria measure of the environmental performance of

products, services, and organizations. With its approach of “comparability over flexibility,” the PEF/OEF

methodology aims at harmonizing existing methods, while decreasing the flexibility provided by the

International Organization for Standardization (ISO) standards regarding methodological choices. Currently, a

3-y pilot phase is running, aiming at testing the methodology and developing product category and

organization sector rules (PEFCR/OEFSR). Although a harmonized method is in theory a good idea, the PEF/OEF

methodology presents challenges, including a risk of confusion and limitations in applicability to practice. The

paper discusses the main differences between the PEF and ISO methodologies and highlights challenges

regarding PEF applicability, with a focus on impact assessment. Some methodological aspects of the PEF and

PEFCR Guides are found to contradict the ISO 14044 (2006) and ISO 14025 (2006). Others, such as prohibition

of inventory cutoffs, are impractical. The evaluation of the impact assessment methods proposed in the

PEF/OEF Guide showed that the predefined methods for water consumption, land use, and abiotic resources

are not adequate because of modeling artefacts, missing inventory data, or incomplete characterization

factors. However, the methods for global warming and ozone depletion perform very well. The results of this

study are relevant for the PEF (and OEF) pilot phase, which aims at testing the PEF (OEF) methodology (and

potentially adapting it) as well as addressing challenges and coping with them.

Keywords: Comparability; Harmonization; Impact assessment; ISO 14044; ISO 14025; OEF; PEF

INTRODUCTION

International Organization for Standardization (ISO) 14044 (2006) is currently the only standardized

method to assess a range of potential environmental impacts of products and services. It is

transparent, widely used, and accepted throughout policy and in the market. Life Cycle Assessment

(LCA) is applied both internally for environmentally oriented product and process development as

well as externally to document environmental performance in environmental product declarations or

in the development of type I label criteria. In 2013, the European Commission published the Product

and Organisation Environmental Footprint (PEF/OEF) methodology—a life cycle-based multicriteria

measure of the environmental performance of products, services, and organizations (European

Union 2013a) (PEF/OEFGuide), developed as part of the 2020 European strategy “A resource efficient

Europe” (European Union 2013b). With its approach of “comparability over flexibility,” the PEF/OEF

methodology aims at harmonizing existing methods, while decreasing the flexibility provided by the

ISO standards regarding methodological choices. This is reflected, for example, in predefined life

cycle impact assessment (LCIA) methods.

The PEF/OEF methodology is currently widely discussed amongst stakeholders, and skepticism exists

in industry and consumer organizations, as well as in the scientific community. For example, concerns

expressed by the European Automobiles Manufacturers Association (ACEA), Orgalime (European

Engineering Industries Association), and the European consumer voice in standardization (ANEC)

(Orgalime, ACEA, ANEC 2013) refer to potential unfair competition and market distortion at the

consumer level. Concerns from a scientific perspective are described in Finkbeiner (2014) including a

lack of conformity with international standards, predefined and partly untested impact assessment

methods, and objections to the PEF/OEF “comparability over flexibility” approach.

Therefore, the methodology is also a challenge for existing environmental labeling schemes such as

The Blue Angel (https://www.blauer-engel.de/) or the EU-Flower (http://ec.europa.eu/environment/

ecolabel/index_en.htm). Altogether, this reveals risks of impeding a life cycle-based environmental

analysis of products and services.

The first wave of a 3-y PEF/OEF pilot phase started in November 2013 with the aim of testing the

methodology in practice and developing product category and organization sector rules

(PEFCR/OEFSR) for selected products and sectors; a second wave started in June 2014. PEFCRs and

OEFSRs are life cycle-based rules that complement the general PEF/OEF Guide by providing further

specification at the level of a specific product or sector category (European Commission 2014a). In

the pilot phase, challenges regarding its applicability—known as well as new ones, likely to appear

during the pilot phase—need to be addressed and coped with.

The goal of this article is to present and discuss the main differences between the PEF and the ISO

methodologies (ISO 14044 [2006] and ISO 14025 [ISO 14025 2006]), identify and highlight related

challenges regarding the PEF methodology and its applicability in practice, with a focus laid on impact

assessment. Our report complements the previous analyses in the PEF Guide (European Union

2013a, Annex X) and Finkbeiner (2014) by further discussing challenges and the applicability of the

PEF requirements. Regarding the impact assessment, 14 methods on midpoint level that were

adopted by the PEF/OEF methodology and have to be applied when carrying out a PEF/OEF study

were evaluated. Even though an analysis of these methods exists, conducted within the ILCD

(International Reference Life Cycle Data System) Handbook (European Commission—Joint Research

Centre—Institute for Environment and Sustainability 2011) in a comprehensive way, important

criteria regarding their applicability are underrepresented in that analysis and thus picked up in this

paper. Hence, our evaluation complements the existing assessment and highlights challenges of the

proposed impact assessment methods. Additionally, selected challenges of the PEF methodology

identified in our analyses are illustrated by using a theoretical example.

METHODOLOGICAL APPROACH

This study followed a 2-step approach: 1) summary and discussion of differences between the ISO

and the PEF and highlighting challenges of the methodology, and 2) a thorough analysis of the

PEF/OEF impact assessment methods and description of their challenges. The methodological

approach of the latter is further described in the next subsection. A theoretical example (comparing

biobased with conventional plastics) was used to test selected methodological requirements of the

PEF methodology and to illustrate the identified challenges. As an illustrative example, we compare a

thermoplastic cellulose acetate and a thermoplastic polymethylmethacrylate for biobased and

conventional plastics, respectively (with a functional unit of 1 kg plastic).

This example addresses all challenges related to the impact assessment and the following selected

other challenges: identification of a suitable CPA/NACE code (Classification of Products by

Activity/Nomenclature statistique des activites economiques dans la Communaute europeenne) and

the ban of cutoffs.

Criteria-based evaluation of impact assessment methods

To assess the suitability of the methods, a new evaluation scheme based on the evaluations from the

European Commission Joint Research Centre (2011) and Reimann et al. (2010) was developed.

Overall, 4 main criteria are considered for evaluation: “stakeholder acceptance,” “documentation

and review,” “environmental relevance,” and “applicability.” As shown in Figure 1, each criterion is

specified by 2 to 5 subcriteria.

Figure 1. Assessment scheme for evaluating 14 impact assessment methods considering 4 main criteria each with 2–5

subcriteria.

For the criterion “stakeholder acceptance,” 3 subcriteria were considered. They were chosen to

represent the following requirements: To communicate the results to the broad public or layperson,

politics, and industry, the impact indicator as well as the corresponding unit has to be easy to

understand. Moreover, when being supported by an authorized body, it is more likely that the

method will be accepted by other stakeholders as well. For the criterion “usable results for policy are

provided,” all considered impact assessment categories are assumed to be relevant for politics or

policy. Thus, the assessment focuses on whether the models are able to provide results that are

useful on a policy level.

The criterion “documentation and review” has been evaluated based on whether the method is

documented and reviewed and uncertainties are addressed. A thorough documentation of the

method itself, background models, and resulting available characterization factors other than a

reviewed journal paper is required for a holistic understanding. Furthermore, the scientific quality of

the method is assumed to be higher when scientists dealing with the same or related topics have

reviewed the method thoroughly. Addressing uncertainties regarding, for example, model

assumptions— qualitatively as well as quantitatively (as done within the USEtox model [Rosenbaum

et al. 2011])—improves the understanding of potential uncertainties of the impact assessment

results, which therefore supports the interpretation.

The third criterion, “environmental relevance,” evaluates the ability of a method to reflect the

impacts on the environment adequately. Therefore, it is evaluated in the sub criteria if relevant

substances and chemical reactions are considered, if effect mechanisms are modeled adequately,

and if relevant distribution pathways and accumulation are considered. Furthermore, in a method

properly reflecting the environmental impacts, characterization factors are assumed to be provided

for all relevant substances. Missing factors would lead to a neglect of substances relevant for the

impact category. Thus, as a first step, a classification of all relevant substances has to take place.

Relevant substances can be identified based on the current knowledge in this field. Moreover,

existing impact assessment methods should be included, because the elementary flows considered

up to now are assumed to present relevant substances for this category. Based on the identified

relevant substances, related chemical reactions and distribution pathways should be identified

considering state-of-the-art research in the individual science disciplines. The work of the

Intergovernmental Panel on Climate Change (IPCC 2007), for example, recently highlighted the

growing problem of ocean acidification, which is so far not covered in LCA. Based on this research, an

LCIA method can be developed considering relevant elementary flows such as carbon dioxide as well

as relevant reactions such as the formation of hydrogen ions in water. Because the impact categories

land use, depletion of water, and depletion of abiotic resources are input based, they have to be

evaluated with a different set of criteria. For a comprehensive assessment of land use, the different

land use types have to be considered, because their contribution to biodiversity or soil quality varies.

Also, several effect mechanisms attributable to different land use types have to be modeled

adequately. For water depletion, the resource availability and differentiation of water bodies need to

be considered within the model, and regional aspects have to be reflected. To determine abiotic

resource depletion, both the physical availability as well as socioeconomic aspects have to be

considered.

The fourth criterion assesses the applicability of the impact assessment method. A sufficient amount

of inventory data has to be available to use the method in practice. Global validity is also important,

because supply chains have a global, not a European, scope.

As a result, the methods are ranked between levels 1 and 3 for each criterion. A 3-level approach was

chosen for the classification, following an easy-to-understand traffic light rating system. Such

systems are already widely applied in literature, for example, for impact assessment in Social LCA

(e.g., Franze and Ciroth 2011; Martınez-Blanco et al. 2014). A classification of 1 indicates that the

method fulfils all requirements of the criterion, and a ranking of 3 means that the demands could not

be reached. Because many methods can only partly achieve the criteria requirements, an

intermediate ranking was included (level 2).

The concluding recommendation is based on the results of this assessment and supported by an

example comparing bio-based and conventional plastics.

RESULTS AND DISCUSSION

In the following sections, we summarize the differences between ISO and PEF methodologies and the

identified challenges as well as the results from the evaluation of impact assessment methods, using

the previously mentioned example for illustration.

Differences between ISO and PEF and challenges of the PEF methodology

The PEF differs regarding goal and scope, mainly because it introduces a set of new terminology for

globally accepted terms (e.g., unit of analysis instead of functional unit), it is communication driven,

whereas ISO is not, and it specifies how the unit of analysis shall be defined (amongst other methods,

linked with CPA/NACE codes) (see also Finkbeiner [2014]). The introduction of (in our opinion

unnecessary) new terms was already reviewed by the European Commission, noting that renaming

common terms found limited support by the stakeholders (Galatola and Pant 2014). With regard to

communication, appropriate methods or vehicles are yet missing. In the context of communication

(and intended application), in the PEF overview on key requirements of PEF methodology (European

Union 2013a, Annex X), business-to-business and business-to-consumers are claimed as

communication targets of ISO 14044, whereas this is not directly stated in ISO 14044 (2006). The PEF

requirement of linking the unit of analysis with CPA/NACE codes appears not to be a straightforward

task: Taking the theoretical example on bio-based and conventional plastics, the category

“manufacture of plastic products” (CPA/NACE code 25.2) would apply most. However, more than 1

category would be needed to cover the product system adequately—the production of raw materials

for bio-based plastic (here mainly cellulose from recycled paper and pulpwood), for example, differs

from raw material production of conventional plastic, but it is not reflected in the previously

mentioned category and should thus be described in an additional category (maybe CPA/NACE code

17.12, “manufacture of paper and paperboard”).

Regarding the inventory, one main difference is that the PEF Guide does not allow cutoffs. This would

mean that all processes have to be modeled—in the plastics example, all machinery, even the

extruder, which has a negligible contribution to the overall environmental impact. This is hardly

feasible in practice, because it causes extra effort and costs without providing additional information.

However, this requirement is relativized in the PEFCR Guide: although not using the term “cutoff,” it

states that processes identified (in the screening study) as being not relevant—thus neglible—do not

need to be considered in the subsequent PEF study. As in ISO 14044 (2006), fixed criteria defining

when a process is considered as relevant or not (e.g., contribution of less than 5% to the impact

category) are not provided in the PEF/PEFCR Guides. As a result of the screening study during the

PEFCR development, one may exclude certain impact categories in the PEFCR, if identified as not

being relevant (European Commission 2014b), without knowing yet how relevance should be

measured. Methodologically, excluding impact categories, however, means that elementary flows

are cut off. Moreover, the PEF methodology provides numerous specifications, for example, on how

to deal with fossil and biogenic carbon (C), land use change, C storage, and so forth, or on

quantitative minimum data quality requirements and data sources as well as regarding allocation by

proposing a new “recycling formula.”

Regarding impact assessment and interpretation, the PEF methodology differs because it provides a

default set of 14 impact assessment methods (further explained in the following section).

Normalization is a recommended step and has to be tested within the PEF pilot phase. Therefore,

European normalization factors are provided by the European Commission (European Commission

2014c). Considering only European emission data disregards the fact that most raw materials are

produced outside of Europe. Furthermore, except for some aspects such as greenhouse gas

emissions, global as well as regional emission data necessary to determine normalization factors are

not yet available. Hence, the provided normalization factors are so far based on estimations, and

normalized results should be interpreted with care, such as the significance of impact assessment

categories for a specific product system (Heijungs et al. 2006; Benini et al. 2014). To calculate credits

or burdens for the end of life (EoL), a specific formula to deal with multifunctionality in recycling

situations is given (European Commission 2014b), which requires the collection of 18 factors

regarding the EoL product system. Generic data should be used if specific data are not available. For

bio-based plastic, in which so far generic processes are lacking, not all requested factors can be

provided by the modeled system, and generic data from conventional plastic have to be considered.

Thus, in this example, a distinction between the impacts of bio-based and conventional plastic in the

EoL stage is not possible. Moreover, the PEF methodology allows weighting as an optional step

(results provided as “additional environmental information”) for business-to-business or business-to-

consumers communication with comparisons and comparative assertions. This clearly contradicts ISO

14044 and ISO 14025 as well as the fact that a PEF study can be used as claim about the

environmental superiority or equivalence of a product compared with another product (European

Commission 2014b), whereas ISO states that LCIA shall not be the sole basis for interpretation in

comparative assertions (ISO 14044 2006).

Criteria-based evaluation of impact assessment methods

In this section, the evaluation results for all 14 impact assessment methods are presented. Because

of the complexity of the analysis, not all criteria can be discussed in detail. Key points that reflect the

overall maturity of the method are highlighted.

Table 1 shows that stakeholder acceptance has been rated very well for most methods.

Environmental topics discussed in the public for years such as global warming (Barker 2007; IPCC

2007) have the advantage of being understood even by laymen, although underlying indicators might

be complex. Stakeholder acceptance of methods for ionizing radiation, toxicity, water, and abiotic

resource depletion is lacking as several sub criteria are evaluated with level 3. Because the methods

for assessing abiotic resource depletion based on reserves instead of ultimate reserves (Guinée et al.

2001; Oers et al. 2002) and ionizing radiation (Frischknecht et al. 2000) have never been used in LCAs

so far and therefore have never been tested, their stakeholder acceptance is rated as minor;

however, for policy the abiotic depletion potential (ADP) method (Guinee et al. 2001) might be of

interest as the economic resource availability is assessed. Because of its complexity and high

uncertainties, the UseTOX method for assessing human and ecosystem toxicity (Rosenbaum et al.

2008; Rosenbaum et al. 2011) is limited in its comprehensibility and applicability for policy. Even

though it becomes more reliable by addressing the uncertainties within a method, when these

uncertainties are very high as is the case for USEtox, where the result varies by 1 or 2 orders of

magnitude, policy cannot rely on the results. Furthermore, because USEtox has never been applied in

a policy context before—contrary to, for example, the accumulated exceedance method for

acidification and eutrophication, which has been used in mitigation strategies in Europe for years—

real life experience using and interpreting this method is missing. The ecological scarcity method

(Frischknecht et al. 2009) is easy to understand, but because of its severe flaws, such as modeling

Table 1. Evaluation results of impact assessment methods shown for main criteria and subcriteria

aSubcriteria: indicators and unit are easily understandable, an authorized body supports the method, usable results for policy are provided

bSubcriteria: method is documented, method is reviewed, model uncertainties are assessed

cSubcriteria: relevant substances are considered, relevant chemical reactions are considered, effect mechanism is modelled adequately, relevant distribution pathways are considered, accumulation is considered

dSubcriteria: method is globally valid, inventory data are adequate

eSubcriteria: different land use types are considered, effect mechanism is modelled adequately

fSubcriteria: resource availability is considered, different water bodies are considered, regional differences are reflected

gSubcriteria: resource availability is considered, socio-economic factors are considered

Impact Assessment Categories

Stakeholder Acceptancea

Documentation & Reviewb

Environmental Relevancec

Applicabilityd

Climate change

1 # 1 # 1

1 # 1 # 1 # 1 # 1

1 # 1

Ozone depletion

1 # 1 # 1

1 # 1 # 1 # 1 # 1

1 # 1

Particulate matter

1 # 2 # 2

2 # 1.5 # 3

2 # 1 # 1 # 1 # 1

2 # 2

Ionizing radiation

1 # 3 # 3

1 # 1 # 2

1 # 1 # 2 # 2 # 2

3 # 2

Photochemical ozone creation

1 # 2 # 2

2 # 1 # 3

1 # 1 # 1 # 1 # 1

2 # 1

Eutrophication, terrestrial

1 # 2 # 1

2 # 1 # 3

3 # 1 # 1 # 1 # 1

2 # 1

Eutrophication, aquatic

1 # 2 # 2

2 # 1 # 3

3 # 1 # 1 # 1 # 1

2 # 1

Acidification, aquatic

1 # 2 # 1

2 # 1 # 3

3 # 1 # 1 # 1 # 1

2 # 1

Human toxicity, carcinogen

3 # 1 # 2

1 # 1 # 1

3 # 2 # 1 # 1 # 2

1 # 2

Human toxicity, noncarcinogen

3 # 1 # 2

1 # 1 # 1

3 # 2 # 1 # 1 # 2

1 # 2

Eco toxicity

3 # 1 # 2

1 # 1 # 1

3 # 2 # 2 # 2 # 2

1 # 2

Land use

1 # 2 # 2

2 # 1 # 3

2 # 3e

2 # 2

Water depletion

1 # 3 # 3

1 # 1 # 3

2 # 3 # 3f

2 # 3

Abiotic resource depletion

1 # 3 # 2

1 # 1 # 3

2 # 3g

1 # 3

artefacts that lead to a higher water scarcity in Germany than in Australia (Finkbeiner 2014), it is not

suitable for use in policy in its current form.

The documentation is evaluated as very well for most methods. However, some methods are using

background models, which are not publicly accessible and are therefore ranked at level 2. All

methods besides the Riskpoll model for particulate matter (Rabl and Spadaro 2004) are published in

scientific journals. Furthermore, only 5 of the 14 methods provide uncertainty assessments.

Many methods such as the World Meterological Organization method for ozone depletion (WMO

2010) or the IPCC (2007) for climate change perform very well regarding their environmental

relevance. Because both methods are very mature, they could be used, not exclusively though, for a

comparative assertion such as bio-based and conventional plastics. The ionizing radiation method

(Frischknecht et al. 2000) only considers effect mechanism regarding nuclear power production and

only within France. Thus, its reflection of the environmental relevance is not satisfying. The plastics

example shows that ionizing radiation is predominantly related to electricity production. Because the

method has not been tested sufficiently so far, that other pathways and related inventory data for,

for example, fertilizer production (Scholten and Timmermans 1996), are considered adequately as

well cannot be ensured. For the example chosen, this means that excluding this aspect could lead to

an inequitable advantage for the bio-based material. The USEtox model (Rosenbaum et al. 2008) for

assessing human toxicity and ecotoxicity only covers a small amount of the potentially toxic

chemicals (although more than 3000 characterization factors are included in the latest version

[Rosenbaum et al. 2011] and cumulative toxic effects cannot be assessed so far [Finkbeiner et al.

2014]).

Thus, its environmental relevance is not satisfying. Some characterization factors such as those for

metals are marked as interim; for them, proceeding with caution is advised (Rosenbaum et al. 2011).

Hence, this method should be used with care in comparative LCA studies. The accumulated

exceedance method (Seppälä et al. 2006) for acidification and eutrophication only provides few

characterization factors (Seppälä et al. 2006; Finkbeiner 2014). Within the plastics example, the

accumulated exceedance method (Seppälä et al. 2006) covers 7 elementary flows less for the

acidification potential than the impact assessment method developed under the lead of the Center

of Environmental Science of Leiden University (CML) (Guinee et al. 2001). For the products compared

here, these neglected emissions only sum up to 2% of the overall impact, but for other product

systems their contribution may be more significant, especially when metals are considered. The land

use method (MilaiCanalsetal.2007) is the third evaluated as not being satisfactory. Soil organic

matter only serves as a headline indicator for soil fertility and climate regulation; other aspects such

as loss of biodiversity have to be considered by applying additional impact assessment methods (Mila

i Canals et al. 2007; Brandao et al. 2011). Thus, this method should be applied with care within

comparative studies as well, especially when the assessed systems differ significantly (such as for bio-

based vs conventional plastics). The ecological scarcity method (Frischknecht et al. 2009) used to

assess water depletion does not adequately consider all environmental relevant factors for the

resource freshwater, because it is only based on relative scarcity (Finkbeiner 2014). Because no

regional inventory data are available in LCA databases so far, water footprint case studies are very

time consuming and difficult to carry out, especially for complex industry products (Berger et al.

2012). For PEF studies, European weighted characterization factors are provided, which consider

different water bodies but are not geographically explicit. Especially for the evaluation of water

depletion, regional differences are of upmost importance (Pfister et al. 2009; Berger et al. 2014;

Finkbeiner et al. 2014). Thus, applying results of average characterization factors cannot be used to

support the decision making for a product system. This is shown in the example of bio-based plastics

compared with conventional plastic, where regional differences are not considered and only the

amount of water consumed is the deciding factor. Conventional plastic is produced from crude oil,

which is extracted from countries all over the world. High amounts are obtained in Saudi Arabia

(Central Intelligence Agency 2012), which is a water-scarce country. The bio-based plastic example

cellulose acetate is made mainly from wood pulp (Sustainable Forestry Initiative 2014; Balser et al.

2015). Current trade statistics (United Nations Economic Commission for Europe 2013) show that low

amounts of industrial round wood used for wood pulp are exported within Europe. Thus, the 30%

cellulose acetate produced in Europe (Dammer et al. 2013) can be assumed to be manufactured from

the wood of European trees. European countries have on average less water scarcity than most oil-

producing countries outside of Europe. Thus, a regionalized impact assessment method should be

recommended. The method for ADP (Guinee et al. 2001) is also not sufficient regarding its

environmental performance, because it evaluates resources only from an economic point of view

considering reserves. Neither physical availability nor proper socioeconomic factors addressing

scarcity are considered. Furthermore, regarding the example, ADP indicator results for conventional

plastic would probably be worse than for bio-based plastic because ADP only considers abiotic

resources. So far, no impact assessment methods are available for evaluating biotic resource

depletion (European Commission—Joint Research Centre—Institute for Environment and

Sustainability 2011; Finkbeiner et al. 2014). Thus, a comparison of abiotic and biotic resource

depletion is not possible.

The biggest challenge regarding the applicability of these methods is the availability of inventory

data. Most of the methods including accumulated exceedance method (Seppälä et al. 2006) are not

valid worldwide because only European background data are considered. When assessing silvicultural

systems such as forestry for bio-based plastic, regional aspects might be of importance for the overall

environmental profile. Acidifying substances, which are caused by transports, might be considered

less or more critical when emitted, for example, in Canada instead of Europe. Because supply chains

are spread out around the world, a uniform world model as used within RiskPoll (Rabl and Spadaro

2004) or USEtox (Rosenbaum et al. 2008) has clear advantages. Because eutrophication and

acidification are assessed within LCA for many years, the inventory data can be assumed as more or

less complete.

RECOMMENDATIONS AND OUTLOOK

In summary, half of the impact assessment methods evaluated here cannot be seen as adequate to

be used solely for decision support such as comparative assertions. However, methods for global

warming and ozone depletion are very good, as evaluated in all subcategories.

The idea of the PEF/OEF to harmonize existing life cycle-based methods is good, but although all

methods reference ISO 14044 as their basis (except for the C footprint standards with limited focus

of climate change) (Finkbeiner 2014), the PEF/ OEF methodology shows several differences and even

contradiction, such as allowing for weighting of LCA results to single scores. Besides the fact that the

new methodology may cause confusion amongst stakeholders and bears the risk of impeding a life

cycle-based environmental analysis of products and services as such, the applicability (as well as the

acceptance) of the method is not yet proven. Questions remain, for example, regarding the

identification of suitable NACE/CPA codes, dealing with the ban of cutoffs, and impact assessment

methods, especially for water depletion and land use. These questions and related challenges were

discussed in this article and illustrated using a theoretical example comparing bio-based and

conventional plastic. Identifying a suitable NACE/ CPA code is not a straightforward task, because

more than one category would be needed to cover the product system adequately.

The evaluation of the PEF/OEF impact assessment methods and their application in the example

showed that using the proposed methods for land use and water depletion is not recommended in

general, but especially not for comparing different product systems. For water depletion, more

commonly used and better tested methods are available (Pfister et al. 2009; Berger et al. 2014;

Finkbeiner 2014). Other relatively new methods such as ionizing radiation (Frischknecht et al. 2000)

or accumulated exceedance (Seppälä et al. 2006; Posch et al. 2008) should be tested regarding their

applicability and necessity within the pilot phase. Applying USEtox (Rosenbaum et al. 2008) presents

a challenge because of several issues mentioned within this paper. The pilot phase should be used to

determine which product systems should use this method. The IPCC method (IPCC 2007) as well as

the World Meterological Organization method (WMO 2010) can be applied without any difficulty.

The results of this evaluation of impact assessment methods can be taken into account within the

PEF/OEF pilot phase, together with the original evaluation and classification of the European

Commission (2011).

Furthermore, important environmental effects such as loss of biodiversity, which cannot be

addressed adequately within LCA so far (Finkbeiner et al. 2014), might have a big influence on the

result and should be considered qualitatively, maybe within the section Additional Environmental

Information in the PEFCRs/OEFSRs.

Further challenges of the PEF methodology beyond the scope of this paper include the applicability

of the data quality assessment scheme, the suitability of the provided allocation approach for

recycling, and weighting or the identification of appropriate measures to communicate PEF/OEF

results. Benchmarks and performance classes shall be defined during PEFCR development to enable

comparisons between products, but for now it is not clear how they should be defined practically nor

theoretically if a benchmark system is to be meaningful.

The European Commission promotes the use of PEF and OEF for measuring and communicating

environmental life cycle performance of products and organizations. Thus, it provides opportunities

to support decision-making processes at all levels, that is, industry and policy. However, for this,

information obtained must be based on a solid methodology, and appropriate communication tools

must be used, but both of these conditions are not yet ensured.

The ongoing pilot phase intends to test the PEF/OEF methodology, to deal with remaining questions,

and to solve problems. This includes further clarifying the goal of PEF/OEF (PEFCR/OEFSR), testing

normalization factors, reconsidering whether weighting is indeed an appropriate way for

communicating LCA/PEF results, and adapting some of the predefined impact assessment methods

(Finkbeiner 2014). The results of this paper are relevant for the pilot phase, because they highlight

theoretical and practical challenges and propose possibilities to cope with them.

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