Strategic environmental assessment methodologies—applications within the energy sector

Strategic environmental assessment

methodologies—applications within the

energy sector

Goran Finnvedena,*, Mans Nilssonb, Jessica Johanssona,Asa Perssonb, Asa Moberga,c, Tomas Carlssond

aEnvironmental Strategies Research Group (fms), Swedish Defence Research Agency, PO Box 2142,

SE 103 14 Stockholm, SwedenbStockholm Environment Institute (SEI), P.O. Box 2142, SE 103 14 Stockholm, Sweden

cEnvironmental Strategies Research Group (fms), Department of Systems Ecology,

Stockholm University, P.O. Box 2142, 103 14 Stockholm, SwedendDepartment of Energetic Materials, Swedish Defence Research Agency, SE 147 25 Tumba, Sweden

Received 1 February 2002; received in revised form 1 August 2002; accepted 1 September 2002

Abstract

Strategic Environmental Assessment (SEA) is a procedural tool and within the

framework of SEA, several different types of analytical tools can be used in the

assessment. Several analytical tools are presented and their relation to SEA is discussed

including methods for future studies, Life Cycle Assessment, Risk Assessment, Economic

Valuation and Multi-Attribute Approaches. A framework for the integration of some

analytical tools in the SEA process is suggested. It is noted that the available analytical

tools primarily cover some types of environmental impacts related to emissions of

pollutants. Tools covering impacts on ecosystems and landscapes are more limited. The

relation between application and choice of analytical tools is discussed. It is suggested that

SEAs used to support a choice between different alternatives require more quantitative

methods, whereas SEAs used to identify critical aspects and suggest mitigation strategies

can suffice with more qualitative methods. The possible and desired degree of site-

specificity in the assessment can also influence the choice of methods. It is also suggested

that values and world views can be of importance for judging whether different types of

tools and results are meaningful and useful. Since values and world views differ between

0195-9255/02/$ – see front matter D 2002 Elsevier Science Inc. All rights reserved.

PII: S0195 -9255 (02 )00089 -6

* Corresponding author. Tel.: +46-8-402-38-27; fax: +46-8-402-38-01.

E-mail address: [email protected] (G. Finnveden).

www.elsevier.com/locate/eiar

Environmental Impact Assessment Review

23 (2003) 91–123

different stakeholders, consultation and understanding are important to ensure credibility

and relevance.

D 2002 Elsevier Science Inc. All rights reserved.

Keywords: Strategic environmental assessment; Applications; Energy sector

1. Introduction

1.1. Background

The main purpose of strategic environmental assessment (SEA) is to facilitate

early and systematic consideration of potential environmental impacts in strategic

decision-making (Therivel and Partidario, 1996; Partidario, 1999). It is intended

to be used on policies, plans and programmes. The growing significance of SEA

as a form of support to decision-making is manifested by the recent EC directive

(2001/42/EC) on the assessment of environmental effects from certain plans and

programmes (Feldmann et al., 2001). However, a number of challenges need to

be overcome for SEA to be an effective tool. In order to be effective, a number of

criteria need to be met. The International Association for Impact Assessment

(IAIA) has published the IAIA principles that stipulate best practice for EIA

(IAIA, 1999).

The principles are: rigorous, practical, relevant, cost-effective, efficient,

focused, adaptive, participative, interdisciplinary, credible, integrated, transpar-

ent, and systematic. While established for EIA, they are of key relevance also for

SEA and in a workshop hosted by the Federal Ministry for the Environment

Nature Conservation and Nuclear Safety (2001), these principles were adapted

towards SEA.

A number of publications have been concerned with how to design an SEA

process that can be integrated with the decision-making process (e.g. European

Commission, 1994; Therivel and Brown, 1999; Naturvardsverket, 2000; ANSEA

Project, 2002). Slightly different steps are defined in different sources, although

the main features remain the same. The following steps are identified here, also

based on (Nilsson et al., 2001):

1. Definition of objectives.

2. Formulation of alternatives.

3. Scenario analysis.

4. Environmental analysis (including the use of objective and acceptable

aggregated indicators, based on more traditional natural sciences).

5. Valuation (including the use of controversial aggregation methods, and

political and ethical values).

6. Conclusions, review of quality/follow up measures, etc.

G. Finnveden et al. / Environmental Impact Assessment Review 23 (2003) 91–12392

In addition, consultation and public participation are important aspects of the

SEA process and should take place at several occasions in the process. Similarly,

careful analysis of the uncertainties involved in the assessment, through methods

such as sensitivity analysis should be applied throughout various stages of the

SEA.

A legal basis for undertaking SEA will be developed in Sweden within the

next few years when the EC directive is incorporated into national legislation.

The directive is, however, limited in scope. It applies to plans and programmes,

and modifications of them, that are subject to preparation and/or adoption by an

authority at national, regional and local level, and that are required by legislative,

regulatory or administrative provisions (Article 2). Thus, it does not apply to

policies, to the private sector or to plans and programmes that are not formally

required. It is currently not clear which, if any, applications within the energy

sectors will require an SEA according to the directive (Adolfsson, Swedish EPA,

personal communication) and there is also a lack of methodological guidelines

for this application.

SEAs can be useful and effective for a number of applications where SEAs

are not formally required. For example, SEA could be a useful tool for energy

policies developed at national level (Nilsson et al., 2001; Noble and Storey,

2001). The national energy policy programmes established by the Swedish

Parliament are normally based on major impact analyses prepared by the

Government Commissions (for example SOU, 1995, p. 139, preceding the

1997 policy programme). However, some decisions are also taken in between

these major programmes. At the local level the Swedish municipalities are

required by law to carry out energy plans (SFS, 1977, p. 439) as a means to

promote efficient use of energy and take action for a secure and sufficient

energy supply. The plan shall concern the supply, distribution and use of

energy within the municipality and it shall contain an analysis of what impacts

these activities will have. The municipal energy plans are not legally binding,

however. On a local level, SEA could also be useful for other purposes. The

municipalities still have many important roles in the energy arena, for example

as owners of energy companies and a large real estate stock, as environmental

and planning authorities and as providers of information to the public. Within

companies, SEA can support internal decision-making in several ways,

including to minimise future environmentally related risk and associated

economic costs and to gain competitive advantage. There is currently a lack

of methods for this purpose (Bardouille, 2001) and SEA can possibly be

useful within this context. Another possible application area is to use results

from SEA as arguments in a public debate. A third application area can be to

use a non-site-specific SEA in a tiered approach with a site-specific project

EIA at a lower level. Also NGOs can use SEAs to develop and support

sustainability arguments and positions in their campaigns and other activities.

However, above all it should be used to improve decisions towards sustainable

solutions.

G. Finnveden et al. / Environmental Impact Assessment Review 23 (2003) 91–123 93

SEA can be viewed as consisting of three components: institutional arrange-

ments, procedure, and methods (Kørnøv, personal communication). While much

of the SEA literature is focused on issues surrounding institutional arrangements

(ANSEA Project, 2002) and process aspects (see, for instance, Therivel and

Partidario, 1996), challenges related to what methods and analytical tools to use

in SEA remain and need more attention. Noble and Storey (2001) address this

question by developing a framework focused on multi-criteria decision-making.

Questions remain however, when it comes to the environmental analysis that can

support those methods.

As a procedural tool, SEA can include a number of different analytical tools

(Wrisberg et al., 2000). (The focus of procedural tools is on procedures to guide

the process to reach and implement environmental decisions whereas analytical

tools are modelling the system in a quantitative or qualitative way aiming at

providing technical information for a better decision (ibid.). In this paper we use

the words ‘‘tools’’ and ‘‘methods’’ as synonyms.) However, appropriate methods

need to be established. For instance, SEA guidance often refers to Environmental

Impact Assessment (EIA)-type analyses but it is often difficult to use the methods

associated with project EIA in SEA because they are adjusted for site-specific

information and local impacts whereas SEA often is not site-specific and can

often be primarily concerned with cumulative and indirect impacts (e.g. Petts,

1999b). The lack of methodological guidance for SEA also acts as a barrier to the

implementation of SEA in general and the European directive on SEA (European

Parliament, 2001) in particular.

1.2. Aim

The aim of this study is to examine how various analytical tools can be used

within the SEA process, especially in the following steps: Scenario Analysis,

Environmental Analysis and Valuation. Examples include Economic Valuation

methods, Life Cycle Assessment (LCA) and Risk Assessment (RA) (see also

Petts, 1999a). Relations between SEA applications and choice of methods are

also discussed. The applications of particular interest in this study are within the

energy sector in Sweden. The results and the discussion are however of relevance

also for other applications.

It should be noted that the study is not comprehensive in scope in so far as not

all possible tools are discussed. It goes through a limited set of quantitative tools.

There are, in addition to this, several quantitative tools that have not been

discussed. There are also several qualitative tools that could be useful in an SEA

context, such as various types of group decision making and preference-

indifference models (Noble and Storey, 2001). The tools that are discussed have

been selected because they are established methodologies that are well-known,

they have been developed, tested and applied in the energy sector, at policy, plan,

and programme levels, and there is data available from these tools for testing the

methods in a pilot study.


2. Describing methods

A large number of methods and tools for assessing environmental impacts are

available (Baumann and Cowell, 1999; Moberg et al., 1999; Petts, 1999b;

Wrisberg et al., 2000). We need to characterise different methods in order to

better understand their interrelationships and the appropriateness of different

methods in different applications. In this section some frameworks for describing

methods and their relations are presented.

Firstly, there are several analytical features of the methods that need to be

considered when several methods are used in combination (Finnveden and

Moberg, submitted for publication):

� Degree of site-specificity. Some methods are generally site-specific (e.g. local

air quality models), whereas others are generally site-independent (e.g.

traditional Life Cycle Assessment methodology further described below).

Between these two extremes, there may be a continuum with different types of

site-dependency (Potting, 2000).� Degree of time-specificity. In parallel to site-specificity, we can distinguish

between time-specific, time-independent and different types of time-depend-

ency (Potting, 2000).� Type of comparison. Most methods include some sort of comparison, either

between different alternatives, or within a studied system or against a

reference.� Degree of quantification.� System boundaries, which are largely determined by the object of study.

Distinctions can be made between different types of objects, for example:

chemical substances, products and functions, companies and organisations,

nations and regions, sectors, projects, and policies, plans and programmes

(Finnveden and Moberg, submitted for publication).� Types of impacts and effects considered.

Differences between methods with regard to these aspects can determine if and

how different methods can be used in the context of an SEA.

Secondly, the information produced by different methods should be clas-

sified. A useful categorisation developed for environmental and sustainable

development indicators is the DPSIR (Driving forces, Pressure, State, Impact,

Response) model. This model represents a systems analysis view of the

interaction between the human and environmental systems (Smeets and

Weterings, 1999). Example of indicators from the energy sector in this model

are:

� Driving force-space used in residential living.� Pressure-emissions from electricity production.� State-ambient concentrations of pollutants.


� Impact-cases of cancer due to pollutants.� Response-fraction of renewable fuels.

(In this paper, environmental changes or effects are used as encompassing

terms covering changes of pressure, state and impacts.)

The following sections will introduce and discuss the methods in relation to

these criteria.

3. Methods

3.1. Future studies

Since SEA is a process for facilitating and improving strategic decision-

making processes much of what is handled concerns the future. Therefore it is

often necessary to include some idea of the future in SEA. Future scenarios are

often used in SEAs (see e.g. case studies in Therivel and Partidaro, 1996; Noble,

2000; Noble and Storey, 2001). However, much less is written on the appropriate

methods for future studies within SEA. Scenario analysis is mentioned in Dom

(1999) as an often-used tool for evaluating future impacts within SEA, in

Naturvardsverket (2000) scenario analysis is used as an example of methods

for SEA and the use of forecast and backcasts (defined below) in SEA is

discussed by Noble (2000).

There are several different approaches for studying the future (Dreborg, 2001).

Forecasts try to indicate a probable future and they are often based on trends and

mechanisms that can be seen in past years. Such trends and mechanisms are then,

more or less directly, extrapolated into the future-giving a forecast. As devel-

opment and change are a constant part of society, reliable forecasts are useful

mainly for the shorter term and for well-defined areas (Hojer and Mattsson, 2000;

Rescher, 1998; Makridakis et al., 1998). Conditional forecasts are based partly on

extrapolations of historical data or mechanisms and partly on assumptions or

results from scenarios. Conditional forecasts can then result in a number of

scenarios based on some conditional assumptions. Scenarios are helpful when

there is a significant qualitative uncertainty about the future. External scenarios

means that the studied scenarios are dependent on factors which cannot be

controlled by the user of the scenarios, but which are still relevant for the same.

Scenarios present possible futures. External scenarios may be used in scenario

planning to find strategies that are robust across a range of possible futures (see

e.g. van der Heijden, 1996). If the future is considered to be possible to influence

in a significant way by the user of the scenarios, policy scenarios may be suitable

to use. In this case the user will act, not only react. In back-casting the future goal

is first set and then images of what this preferable future could be/look like is

made. Then in a second step, the ways to get there are described. For a more

extensive presentation of back-casting, see e.g. Dreborg (1996).


Modelling can be used to support the different approaches to future studies.

Modelling is often based on assumptions made from historical trends, but new

mechanisms may also be used. An advantage of modelling is that quantitative

information on, for example, future energy use and fuel mix can often be

produced. A disadvantage with complex models is the possible loss of transpar-

ency. There is also a risk that models based on assumptions and mechanisms that

describe the current situation are used for long-term future studies, where these

assumptions and mechanisms cannot be presumed. Dynamic modelling is

frequently used in energy systems studies. One example of such models is

MARKAL which is a macro-level model. It has for example been used to study

the effects of increased cooperation and cross-border energy-related trade in the

Nordic countries (Unger, 2000). Another example is MARTES, a model used for

local district heating systems (Olofsson, 2001).

3.2. Life cycle assessment

Life Cycle Assessment (LCA) is a tool to assess the environmental impacts

and resources used throughout a product’s life from raw material acquisition

through production use and disposal. An ISO standard has been developed for

LCA providing a framework, terminology and some methodological choices

(ISO, 1997, 1998, 1999). Initiatives have also been taken to develop best

available practice (Udo de Haes et al., 1999a,b, in press). The basis for the

calculations is the functional unit to which all inputs and outputs are related. An

example of a functional unit is 1 kW h electricity or 1 MJ heat. When different

alternatives are compared, the functional unit is the basis for the comparison.

According to the ISO-standard, an LCA is divided into four phases:

1. Goal and scope definition.

2. Inventory analysis, where inputs and outputs to and from the systems are

identified and quantified.

3. Life cycle impact assessment (LCIA), aimed at understanding and evaluating

the magnitude and significance of the potential environmental impacts of a

product system. This phase is further divided into three mandatory elements.

3.1. Selection of impact categories, indicators for the categories and models

to quantify the contributions of different inputs and emissions to the

impact categories.

3.2. Assignment of the inventory data to the impact categories (classifica-

tion).

3.3. Quantification of the contributions from the product system to the

chosen impact categories (characterisation).

4. Interpretation, where the findings of either the inventory analysis or both the

inventory analysis and the life cycle impact assessment phases are combined

in line with the defined goal and scope of the study.


In principle, LCA is a comprehensive environmental assessment. In practice,

not all types of environmental effects are equally well covered (Finnveden,

2000). Effects associated with land use are traditionally difficult to assess,

although there has been a considerable methodological development during the

last years (Lindeijer et al., in press). Toxicological effects are often only included

with data gaps. Effects associated with radiation, accidents and disamenities are

typically not covered at all. The impacts typically best covered in a traditional

LCA are environmental impacts from emissions to air, such as global warming

and acidification, and use of energy resources.

LCA is traditionally a site- and time-independent tool. In a traditional LCA, no

consideration is given to when and where emissions are taking place (Udo de

Haes, 1996) and characterisation factors used for quantifying the contribution to

different impact categories do not have any site-dependent information resulting

in a site-generic result. This is mainly for two reasons. The first is a practical

reason; it is not practically possible to gather site-specific information for all

places included in an LCA. The second is more theoretical. In an LCA, not all

emissions are considered. Only the emissions that are allocated to the functional

unit are considered. There is however a trend towards making LCA more site-

dependent if not site-specific (Huijbregts, 2001; Krewitt et al., 2001; Nigge,

2001a,b; Potting, 2000; Spadaro and Rabl, 1999). Introducing some typical

environments and emissions situations does this. For example, emissions at low

height in an urban environment are differentiated from emissions at high

elevation in rural areas. Impact assessment factors are then calculated for these

typical environments. Site-dependent characterisation factors can thus be calcu-

lated and used, resulting in a site-dependent characterisation.

Besides the mandatory elements within LCIA, there are also some elements

described as optional (ISO, 1999), e.g. weighting that aims at converting and

possibly aggregating indicator results across impact categories resulting in a

single result. Methods for weighting include both economic valuation methods

and multi-attribute approaches, further discussed below (Finnveden et al., 2001).

Within the DPSIR-framework for indicators discussed above, LCA can

produce information on pressure and impact levels. Results on emissions are

on a pressure-level and depending on how far in the environmental mechanism

the impact assessment modelling goes, the resulting indicators may be described

as pressure- or impact-indicators. LCA does however not describe a state. This is

because only the fraction of the emissions which are related to the functional unit

is considered, not the whole picture. In order to calculate a state, the whole

picture including background emissions needs to be considered which is not done

in an LCA.

3.3. Environmentally extended input/output analysis

Input–Output Analysis (IOA) is a well-established analytical tool within

economics and systems of national accounts (Miller and Blair, 1985) using a


nation or a region as the object of the study. The input–output matrices describe

trade between industries. By performing an input/output analysis a calculation of

the sectors or industries involved in the production of a product or service going

to final demand can be calculated. IOA can be applied to include environmental

impacts by adding emissions coefficients to the monetary IOAs (Lave et al.,

1995; Joshi, 2000). IOA are typically applied using retrospective data and

methodology for accounting purposes.

Environmentally extended IOA is a quantitative method. It presents results for

broadly defined sectors or products groups within a nation or a region. It is site-

independent within the nation or region. The emissions most often included are

the traditional air pollutants, but in some applications more pollutants have been

included giving results similar to LCAs.

3.4. Risk assessment of chemicals and accidents

Risk assessment is a broad term covering many different types of assessments.

Already the word ‘‘risk’’ is problematic (Hansson, 1999). Here, a distinction is

made between risk assessment of chemical substances and risk assessment of

accidents. The latter may include environmental aspects. Risk assessment of

accidents concerns unplanned incidents, e.g. explosions or fires. This is typically

in contrast to risk assessment of chemicals, where dispersion of chemicals is often

planned and forms part of its use.

Methods and protocols for risk assessment of chemicals have been developed

in several international fora, e.g. EU and OECD (Eduljee, 1999). In the risk

assessment of chemicals, an exposure assessment including a description of the

nature and size of exposed targets, as well as magnitude and duration of

exposure, is combined with an effect assessment (Eduljee, 1999; KemI, 1995;

Olsen et al., 2001). The exposure assessment is done using some kind of model.

Broadly speaking, two classes of models can be distinguished for toxic chemicals

(Hertwich et al., in press): multi-media fate and exposure models that take into

account the fate of pollutants across medium boundaries and model multi-

pathway exposure routes, and spatially explicit single-medium models which

can take into account dispersion and reactions. Multi-media models are for

example used in the EUSES, the software which in practice is used for carrying

out risk assessment of chemicals within the European regulation (Olsen et al.,

2001). A single-medium model, Ecosense, is further described below in the

section on the Impact Pathway Approach.

In accident risk assessment, accident consequences and their frequency are

estimated. The assessment is usually divided into three parts: hazard identifica-

tion, consequence analysis and frequency estimation. For the hazard identifica-

tion various methodologies have been developed to aid system experts to identify

hazards in their expert domain, e.g. HAZOP (among others reviewed by Khan

and Abbasi, 1998) and Safety Function Analysis (Harms-Ringdal, 2002). For the

consequence analysis, several methods to calculate consequences due to explo-


sions, fires and releases of toxic chemicals are available (AIChemE, 1998). For

the frequency estimation, historical records, fault-trees and event-trees are used.

Risk assessment of accidents is typically done prospectively for different types

of projects (Raddningsverket, 2000), and it is typically site-specific. Risk

assessment of chemical substances can be site-specific but also more site-

independent for a region or a nation. It typically includes all emissions of the

substance within the geographical boundary or from a particular project or plant.

Comparisons can either be made between different alternatives (which alternative

poses the greatest risk?) or against a standard (is the risk acceptable or not?).

Comparisons can also be made internally within a system to identify the greatest

risk. Risk assessment of chemicals is typically done quantitatively. Risk assess-

ment of accidents can be made both quantitatively or qualitatively.

3.5. Impact pathway approach

The Impact Pathway Approach (IPA) can be regarded as a special case of a

risk assessment approach that is of particular importance for the environmental

assessment of different energy systems. In the IPA the analytical sequence

‘economic activities>emissions>dispersion>concentrations>dose>impact’ is

handled systematically. The ExternE project (Commission for the European

Communities, 1995, 1999) and the URBAIR project (Jitendra et al., 1997)

provides a methodology. Fig. 1 outlines the general stages in the impact pathway

approach.

Emissions factors for various energy systems are used to model the pollution

load. Ambient concentrations depend not only on emissions but on background

concentrations and topography and meteorology. Exposure assessment is of

particular importance when studying health impacts. It serves to estimate how

large a share of the population is exposed to different concentrations of

pollutants. Impacts can be divided into impacts on health, materials (corrosion),

ecological systems, forests, and agriculture (Commission for the European

Communities, 1999). The impact assessment is based on combining information

on the exposed receptor population and the concentration with dose–response

Fig. 1. The impact pathway approach.


relationships for various impacts and pollutants. The theoretical and statistical

underpinnings of dose–response relationships are described in Calthrop and

Maddison (1996) and health impacts of air pollution for the most common

pollutants are thoroughly documented in the literature (Ostro, 1994; Commis-

sion for the European Communities, 1995; Dockery and Pope, 1994; Dockery

et al., 1993).

IPA is a site- and time-specific approach. It can however also be used for more

generic applications by performing calculations for typical conditions. Also by

applying typical conditions, site-dependent characterisation factors can be calcu-

lated which could be used in LCA as discussed above. IPA is a quantitative

method that can be used for comparisons, either between different alternatives or

within a studied system. Calculated ambient concentrations can also be compared

to reference values. IPA is data demanding and therefore mostly applicable to

conventional air pollutants (such as particulates, SOx, NOx, ozone, and CO).

Within the DPSIR framework, IPA can be used to calculate indicators on

Pressure, State and Impact levels.

3.6. Ecological impact assessment

For ecological impact assessment we have not found any clear-cut tools that

would be obvious to use within an SEA for the energy sector. Tools for

assessing ecological impacts are currently being developed within the context of

the Convention on Biodiversity (SBSTTA, 2001). Some guidance on how to

deal with nature conservation in SEA is provided by Therivel and Thompson

(1996). For impacts occurring at specified sites methods may be adopted from

project EIA (see e.g. Wathern, 1999). In SEA the areas affected will typically be

larger and the detailed assessment methods may have to be adjusted to a coarser

resolution. In SEA more emphasis may be put on landscape ecological issues.

For many impacts in SEA for the energy sector the specific site will not be

known and here the assessment will have to rely on some sort of classification

of affected landscape types and a general estimation of what effect a certain

activity will have in different ecosystems. A promising approach is to use

indicator species (see e.g. Treweek et al., 1998; Mortberg and Balfors, 2000;

Dıaz et al., 2001).

The issue of assessing impacts resulting from land use in a non-site specific

context has also been dealt with within the field of LCA. For a recent review

of the methods proposed, see Lindeijer et al. (in press). Weidema and Lindeijer

(2001) suggest a method for quantified assessment of the physical impacts of

land use in terms of indicators for biogeochemical substance and energy

cycles, ecosystem productivity, biodiversity, cultural value and migration and

dispersal. The data presented for the method are at biome level and the results

from using it would accordingly be very crude. Supplying more refined data

for the areas where most of the impacts occur could enhance the method. One

problem that this method displays is the difficulty in determining how different


dimensions of land use impacts can be aggregated. This may be a common

problem to all quantified ecological assessments as long as more than one

indicator is chosen.

3.7. Multiple attribute analysis

Multiple attribute analysis (MAA) (also referred to as multicriteria, multi-

objective or multiattribute trade-off or utility analysis in the literature) aims to

improve decision-making by making choices about conflicting or multiple

objectives explicit, rational and efficient. It can help to structure the decision-

process, display trade-offs among criteria, help people apply value judgments

concerning trade-offs, help people make consistent evaluations of risk and

uncertainty, and facilitate negotiation (Hobbs and Meier, 2000). The application

of MAA usually follows a sequence of steps and hence shows many similarities

with the SEA as a whole. It includes pre-decision issues such as problem

definition, selection of attributes, alternatives selection, and quantification of

impacts to various attributes. These steps emphasize the generation of trade-off

information. These elements of MAA and experiences gained in the field can be

useful in several stages of the SEA. Later stages are more specific for MAA, such

as the trade-off analysis and evaluation steps that lead to the actual choices and

decisions (Hobbs and Meier, 2000) and as such match closely with the valuation

stage of the typical SEA.

Decision research has accumulated a wide array of methods and concepts for

MAA, including, for instance, ‘Analytical Hierarchy Process’ (Saaty and

Vargas, 1994) and ‘Value-Focused Thinking’ (Keeney, 1992). The choice of

MAA method is in itself a multiple attribute problem (Patton and Sawicki,

1993). One has to decide on the use of weights, rating systems and aggregation

measures. The choice of attributes, or criteria, is a crucial aspect and also a

difficulty in the approach that has great importance in the SEA method. As

Keeney (1992) points out, the key to a successful model is the sets of

objectives and the attributes that measure them. Most decision literature pays

significant attention to this problem (Miser and Quade, 1985; Keeney, 1992;

Vries, 1999).

A key question with particular importance in the context of aggregation

regards the weighting or non-weighting of criteria (Vries, 1999). People and

decision-makers often value certain attributes more than others. The introduction

of weights to criteria recognizes the relative importance of different criteria.

Weights can be arrived at through observer-derived indirect methods, based on

observed choices, and direct methods, where decision-makers are asked directly

to assign numerical values. (Keeney and Raiffa, 1976; Hobbs and Meier, 2000).

Matrix display systems are important for the MAA problem and representa-

tion. The common approach in such a system is to indicate the criteria on one axis

(such as environmental quality objectives) and the alternatives on the other (such

as different policy options) and create a scorecard or goals achievement matrix.


This is not just applicable in expert analysis. It has also been applied in

stakeholder participation settings by, for instance, Bots and Hulshof (2000), in

which an impact matrix is established as a first stage in the group decision-

making process. This particular work was done in the health sector, but the

approach could just as well be applied in the energy sector. The matrices can

display data at different levels of aggregation. In value-laden public decision-

making that is typical for SEA processes, a more disaggregated approach tends to

be more acceptable and useful (Patton and Sawicki, 1993). In the SEA context

matrix display systems can help in avoiding a need for producing single summary

values or indices that are supposed to capture different environmental dimen-

sions. Aside from matrix display systems, multiple attributes can be displayed in

a diamond model (Nilsson, 1997), or in a value path (Hobbs and Meier, 2000).

MAA is a very flexible family of methods. It can be applied for all kinds of

impacts, be made site-/time-specific or not, and quantitatively as well as

qualitatively. The relationship to the criteria in terms of pros and cons of the

alternatives might not be possible to quantify, but the assessor or decision-maker

would still be able to rank the alternatives in terms of achievement or in other

ways display the best or worst alternative and hence provide a graspable imagery

of the results of the MAA. One should be aware that there can be significant

disagreement between methods as well as disagreements among individuals and

one would normally expect great differences in results from the MAA. However,

the purpose of the MAA is not to come up with one answer but also to be a

learning process that exposes interest groups to different views, and forces people

to think about the problems at hand (Hobbs and Meier, 2000).

3.8. Environmental objectives

An SEA should be conducted as a means to achieve environmental and other

objectives through the PPP proposal. Keeney (1992) suggests that objectives,

which essentially articulate values, can be identified for example by using

existing sets of strategic and generic objectives. In Sweden, a set of 15 national

environmental objectives have been established by Parliament (Table 1) as

guiding principles for environmental policy (Government Bill, 2000). In the

relatively few regional and sectoral SEAs that have been undertaken to date in

Sweden, the use of the national objectives has in some way been a common

feature (Boverket och Naturvardsverket, 2000). Although plans and policies in

the energy sector have a more profound influence on some of these objectives, all

of them can be affected by different energy policies.

The 15 national objectives are quality objectives that specify the conditions, or

states, which actions should be directed towards and the general time frame is one

generation (Government Bill, 2000). Since they are of a qualitative rather than

quantitative character, a valuation against these could only give indications

whether the PPP would contribute positively or negatively towards achievement,

and possibly the strength of that contribution (in qualitative terms).


The Government has also proposed a more detailed framework with interim

targets and action strategies for each objective (e.g. Government Bill, 2000).

While the objectives specify desirable states only, the targets concern pressures,

states, impacts and responses. Targets for environmental pressures and responses

are by far the most common ones. Examples of targets for some of the objectives

are presented in Table 2. It can be noted that there is a mixture of quantitative and

qualitative targets. In general, targets related to objectives 1–6 are more

quantitative.

Having accepted the objectives as guiding principles or criteria for the SEA,

they can be used in several places in the process:

1. Definition of objectives in the PPP—the objectives establish the general

environmental standards which all PPPs should be directed towards.

2. Identification of impacts—the objectives can be used as a checklist when

doing an inventory of potential impacts (see e.g. Nilsson et al., 2001).

3. Choice of indicators—several sets of indicators have been developed

specifically for or with the environmental objectives in mind, and these can

provide baseline data and measurement methodologies (Government Bill,

2000, pp. 223–227; SOU, 2000, pp. 723–742).

4. Valuation—assuming that the parliamentary established objectives represent

public values, they provide a comprehensive and solid base for valuation.

Having chosen a set of objectives and/or targets and assembled the required

information, the actual valuation, in terms of measuring attributes, can then take

place in a simple matrix with more or less space for comments. In the matrix, the

objectives are on one axis and the different options on the other. Valuation results

Table 1

National environmental objectives in Sweden

1. Reduced climate impact

2. Clean air

3. Natural acidification only

4. A non-toxic environment

5. A protective ozone layer

6. A safe radiation environment

7. Zero eutrophication

8. Flourishing lakes and streams

9. Good quality groundwater

10. A balanced marine environment, flourishing coastal areas and archipelagos

11. Thriving wetlands

12. Healthy forests

13. A varied agricultural landscape

14. A magnificent mountain landscape

15. A good built environment

Source: Government Bill (2000), The Swedish Environmental Objectives—Interim Targets and

Action Strategies.


can be reported per target or objective, and, in the first case, a summarising

assessment can also be made for the headline objectives if appropriate. The result

is thus a valuation per objective. If a further aggregation is required, this has to be

done using some MAA method, as discussed above. A valuation against

environmental objectives can work with qualitative information as well as

quantitative data. It can be site-/time-specific or unspecific, and provides for

the assessment of a comprehensive set of impacts.

3.9. Economic valuation

In the valuation stage of the SEA, the achievement of multiple objectives and

the trade-offs between different objectives need to be analysed, processed and

interpreted. At this stage, economic valuation methods can be useful. There is an

extensive literature on economic valuation methods, e.g. Markandya and Richar-

don (1992), Viscusi (1997), Leksell (1998) and Commission for the European

Communities (1995). However, economic valuation is controversial and has been

criticized both conceptually and in applications and this is further discussed

below.

At the core of economic valuation in the SEA context is the divergence between

social and private cost in the market as a result of a certain activity, i.e. a negative

environmental externality (Bojo et al., 1992; Naturvardsverket, 1997; Begg et al.,

1987). Economists distinguish between use values and non-use values. Use values

are those values that are related to human production and consumption patterns

and they can be direct, indirect or option values. Sometimes market prices exist for

Table 2

Some environmental objectives and targets of relevance for the energy sector (examples only)

Objectives Interim targets

Clean air 2005: annual average of 5 mg SO2 m� 3 in municipalities;

2010: annual average of 20 mg NO2 m� 3 and hourly average of

100 mg NO2 m� 3 in most places; 2010: ground-level ozone not

exceeding 120 mg m� 3 as an 8-h average; 2010: national emission

of VOCs, excluding CH4, reduced to 241,000 tons

Natural acidification only 2010: not more than 5% of all lakes and 15% of the total length of

running water affected by anthropogenic acidification; 2010: reversed

trend of increased acidification of forests and recovery under way;

2010: national atmospheric emission of SO2 reduced to 60,000 tons;

2010: national atmospheric emission of NOx reduced to 148,000 tons

A good built environment 2010: spatial and community planning based on programmes and

strategies for promoting more efficient energy use, the use of

renewable energy sources and the development of production plants

for district heating, solar energy, bio fuels and wind power; 2010:

environmental impact made by energy use in residential and

commercial buildings decreasing and lower than in 1995, through

improving energy efficiency and eventually reducing use

Source: Government Bill (2000).


externalities relating to use values. For example, the effect of reduced crop

productivity due to air pollution can be estimated by the market value of crops

lost. For other such externalities no market prices exist, however, for example

cough episodes and uneasiness as a result of urban air pollution (Commission for

the European Communities, 1995). Non-use values are of a more philosophical

kind and they can be either existence values or bequest values. Non-use values are

usually only possible to estimate through artificial market methods.

Valuation approaches can thus be grouped into those that use conventional

markets, implicit markets or artificial markets (Bojo et al., 1992). Conventional

market valuations include analysis of changes in production, changes in earnings,

replacement cost and defensive expenditure. Implicit market valuations study the

revealed preferences from actual consumer behaviour and choices. These include

wage-risk approaches, travel cost approaches, land and property value or hedonic

pricing approach. Artificial market valuations include measurements of consumer

preferences in hypothetical situations, with Willingness-to-pay (WTP) or Will-

ingness-to-accept (WTA) measures. These are sometimes referred to as direct

methods (Asian Development Bank, 1996).

In the SEA context, it is doubtful whether resources will be available to carry

out valuation studies. Instead, the assessor will probably have to work with the

benefits transfer approach, where one utilises results and data from existing

studies and adjusts them to the decision situation. One can adjust at the impact

quantification stage or in the valuation stage for differences in receptor qualifiers

such as structure, density, income, and behaviour.

3.10. Surveys

In many cases it will not be possible to model or calculate the envir-

onmental consequences in quantitative terms, or even to say something

meaningful in qualitative terms. This situation is expected to arise when we

are assessing effects on landscapes and cultural values that are highly related to

people’s perceptions and preferences, i.e. non-use values, such as for instance

changes in the visual situation in archipelagic or mountainous regions and in

cultural rural landscapes. In those cases, it can be meaningful to apply

expressed-preference survey methods to get an idea of people’s perception of

these consequences.

An example of such a method that is commonly used in economic research is

the contingent valuation (CV) method, a stated preference method (or artificial

market valuation method) where people are asked to express their willingness to

pay for certain environmental features (see Section 3.9). There are also methods

that do not directly aim to get economic values, as CV methods do. Opinion and

attitudinal surveys is one such family of tools (Gregory, 1998). They can be used

to compare the relative importance of different environmental values or to

examine how people balance environmental protection with their economic and

social needs and wants.


There are numerous survey techniques that draw upon various disciplinary

bases. These include public-value surveys or constructive multi-attribute

methods, where the problem is deconstructed, analysed and recomposed;

and decision-pathway surveys, where we draw out attitudes through a set

of linked questions (Gregory, 1998; Keeney and Raiffa, 1976). Another set of

survey methods emphasise small group elicitations, in-depth interviews and

narratives of individuals (Dahinden et al., 2000). Apart from public values

and attitudes, surveys have also been used to solicit expert opinion on

different alternatives. The Delphi technique consists of an iterative question-

naire process, in which the expert panel’s first round of responses are

summarized statistically and fed back in order to give the opportunity for a

second round of revised responses (see Dalkey, 1969; Ott, 1978; Sobral et al.,

1981).

Apart from the above-mentioned substantive argument for eliciting public

opinions, i.e. to obtain better information on consequences, there are several good

reasons to elicit public information in the assessment. The normative argument is

that the decision-makers should obtain a meaningful input and participation from

the affected population. This argument is embraced in legislation. The argument

is that a meaningful participation is likely to increase acceptance for decisions

and reduce the risk of conflict as a result of the process (Stern and Fineberg,

1996). Similar to MAA, surveys in general are a broadly applicable method that

can be used for any type of impact.

3.11. Valuation methods based on mass, energy and area

Partly as reactions to the use of risk assessment and economic assessments,

alternative approaches for valuation have been developed focusing on the inputs

of the studied system. This can be motivated because inputs are often known with

greater certainty, whereas outputs and impacts of those, which are the subject of

risk assessment and economic valuations, are normally more uncertain. Although

these methods have not been primarily developed for SEA, they could be used as

valuation methods within SEA, providing an alternative weighting system to the

economic valuation.

Material Flow Analysis (MFA) is a family of different methods (Bringezu et

al., 1997). A common feature is the focus on material flows, especially on the

input side. Two types are briefly mentioned here: Total Material Requirement

(TMR) and Material Intensity Per Unit Service (MIPS). TMR and related

concepts such as Direct Material Input (DMI) and Direct Material Consumption

(DMC) normally have a nation as their principal object of study. TMR aims at

calculating all material inputs to the society, including both direct and hidden

inputs (Adriaansee et al., 1997) whereas DMI and DMC focus on the direct

inputs, excluding hidden flows such as the overburden from mining operations.

The approach has developed during the 1990s and has mainly been used in

retrospective studies. Instrumental MIPS is similar to TMR, but in this case the


object is a product or a service (Spangenberg et al., 1999). MIPS is similar to

LCA but focuses on the material inputs.

Energy analysis has many similarities with bulk-MFA methods. They focus on

the inputs in physical measures and they may be used as evaluation methods for

different types of objects. There are several different types of energy measures

such as exergy (which can be defined as a measure of available energy; Szargut et

al., 1988) and emergy (which is a historic energy measure describing assimilated

energy from energy, material, information and labour; Odum, 1996).

Ecological footprint is also an evaluation method which in principle can be

applied to different types of objects, although it has mainly been used on regions,

nations and projects such as aquaculture. The results are presented in terms of

area used. The focus is on the area necessary for different types of activities, but

the indirect area which could be used for assimilating different types of emissions

is also included (Wackernagel and Rees, 1996).

4. Integrating tools in SEA

4.1. Some aspects of SEA as a tool

4.1.1. System boundaries

The choices regarding system boundaries can influence the choice of methods

and are therefore discussed briefly here. System boundaries can exist in several

dimensions:

1. Geographical (and time) boundaries for the activity.

2. Geographical (and time) boundaries for emissions and use of resources.

3. Geographical (and time) boundaries for impacts.

4. Geographical (and time) boundaries for other activities.

In the scoping stage of SEA, choices have to be made in all these dimensions.

As an example, consider an energy policy in Sweden during a certain time period

(1). This policy will influence emissions not only in Sweden but also in other

countries, where for example different fuels or materials are produced. The policy

can also lead to emissions far in the future. For example, emissions from landfills

and contaminated ground can prevail for very long time periods after the activity

has finished (2). The impacts will also occur in different areas because of regional

and global transports of pollutants. If the pollutants are persistent, the impacts can

occur long after the emissions occurred (3). A new energy policy in Sweden may

also influence the energy system in other Nordic countries as well as in Europe.

In parallel, even after an activity has finished, it can have an impact on other

activities (4).

Another type of system boundary relates to the life-cycle of products, fuels

and materials. If a life-cycle perspective is used, environmental changes and


effects from raw material acquisition, via production and use, to waste disposal

should be considered. In relation to energy applications, this means for example

that not only the production of electricity should be included but also the

production of fuels, and the disposal of waste materials.

The first of the geographical and time boundaries is necessary as a part of the

definition of the PPP under study. Choices about the other boundaries are not

analytical but rather political choices. If certain geographical and time boundaries

are used for the impacts considered, it means that impacts occurring outside the

geographical boundary and impacts on future generations can be neglected. How-

ever, in relation to the basic recognition of the international and intergenerational

importance of environmental protection we argue for broad system boundaries and

a life-cycle perspective to activities and pressures at least as a starting point.

4.1.2. Types of environmental changes and effects and assessment objects

Another important choice has to be made concerning which environmental

changes and effects to consider in the assessment. One starting point can be a

comprehensive list of environmental objectives. Another choice has also to be

made concerning how to define the environmental effects and changes within the

DPSIR model. This is partly a choice influenced by world views and values and

further discussed below.

We suggest that a comprehensive list of environmental effects and objectives

should be considered at the start of the assessment. An example is the list of

Swedish environmental objectives that can be used as a checklist in the early

stages of the assessment, and later be narrowed down in relation to what is

considered most important. The narrowing down should however be made in a

structured way and the reasons should be documented.

4.1.3. Site-specificity

The degree of site- and time-specificity may vary in different applications of

SEA. Often, the sites of relevance may not be determined at the strategic level

studied in SEA, in which case a site-specific analysis will be difficult. Further-

more, if a life-cycle perspective is used, a site-specific assessment of all the sites

involved often proves unfeasible. Therefore, many SEAs may find themselves in

a non-site-specific context. However, this may vary with the level of decision-

making. In local or regional level processes, the site context can be of high

relevance and more site-dependent information can be used compared to a

national level. The degree of site-specificity is also connected to the choice of

indicators. In general, indicators on an impact level require a higher degree of

site-specificity than indicators on a pressure level.

4.2. A framework of analytical tools

The examination of the different methods in Section 3 reveals different

qualities. Table 3 summarises some of the findings in a few key dimensions.


Fig. 2 shows in which steps of the SEA process (as defined in Section 1) the

different tools discussed above can be used. The environmental analysis step has

been further divided into three substeps:

– A detailed description of the systems that the PPP may affect. (For the energy

sector this may require a detailed description of energy system including fuel

mixes, but also the use of other products and materials, for example in relation

to energy efficiency measures.)

– Identification of environmental interventions (emissions, extractions of resour-

ces, land use, etc.).

– Analysis of environmental change due to the interventions.

Future studies can be used at several steps in the SEA process. The outcome of

future studies in SEA can either be a set of alternatives which all lead towards the

predefined goal (as in back-casting studies) or a set of future scenarios in which

the different alternatives can be placed. If a back-casting approach is used, a more

comprehensive environmental assessment should be performed since normally

only some environmental aspects are used in the goal description. As shown in

Fig. 2, back-casting can be used in the formulation of alternatives. Other types of

future studies will be useful primarily in the scenario analysis.

Concerning when to use which future study approach, only tentative guide-

lines can be presented here. In general it can be concluded that if the decision

process concerns short term issues, in areas where trends can be assumed to be

stable and not desirable or possible to change, forecasts (or conditional forecasts)

may be preferable. If this is not the case then some kind of scenario approach may

be more appropriate. An important aspect in the choice of scenario approach is

the extent to which the decision-maker can influence the future. If the decision-

maker has limited power external scenarios are appropriate, while normative

Table 3

Key qualities of different methods in relation to SEA

Site- and time-specificity Degree of quantification DPSIR

Future studies variable variable

Life cycle assessment low high PI

Risk assessment variable high I

Input/output analysis low high P

Impact pathway approach high high PSI

Ecological impacts high low SI

Multiple attribute analysis variable variable PSI

Environmental objectives variable variable PSI

Economic valuation variable high PSI

Surveys variable variable PSI

Mass/energy valuation variable high P

Ecological footprints variable high P


policy scenarios are more appropriate when the future can be influenced through

strategic action.

Futures studies may also be used within the environmental analysis step of the

SEA (not shown in Fig. 2). The consequences of assessed alternatives depend on

future environmental baseline situations. Future studies could also be of rel-

evance in the valuation step. This is because future societal values may be

different from current ones.

Several analytical tools can be used for the identification and description of

environmental interventions, including life-cycle inventory analysis data and

methodology, different types of checklists (possibly based on environmental

objectives), accident-related risk assessment and environmentally extended input/

output analysis (IOA). Since the industries and product groups are defined rather

broadly in IOA, it may be too blunt a tool for different fuels or fuel mixes. It may

however be useful to provide information on indirect interventions from other

industries as a consequence of changes within the energy sector.

Fig. 2. Relationship between steps of the SEA process and different tools.


For the analysis of environmental change, several tools can be used. The

differences between the tools concern both the degree of site-specificity and what

types of environmental effects and changes are considered in the DPSIR model.

Characterisation methods developed and used for LCA can be used both with

generic (site-independent) characterisation factors and with site-dependent char-

acterisation factors if such are available. Furthermore, risk assessment methods

in general and the Impact Pathway Approach in particular may be used at this

stage.

Concerning the valuation of environmental impacts, a valuation against

environmental objectives producing a multi-dimensional result can be used. If

a one-dimensional result is desired several types of aggregation methods may be

used: multi-attribute approaches, economic valuation, surveys, LCA weighting

methods and valuation using energy, mass or area.

Fig. 2 gives a presentation of which tools can be used at which step in the SEA

process. It is also of interest to consider what type of information different tools

provide and how they can interact. This is illustrated in Fig. 3.

Future studies will, together with the formulation of alternatives, result in a

description of the technical system in each alternative. This information will

typically have some site-specificity, e.g. concerning which nation or region the

PPP is developed for. The information can be qualitative or quantitative. In the

former case, only a qualitative environmental analysis is feasible. If the

description is quantitative, a choice can be made to take either a qualitative or

a quantitative path. Different paths can be chosen for different types of

environmental impacts.

Along the qualitative path, checklists can be used resulting in qualitative

information about environmental impacts. A qualitative multi-dimensional valu-

ation against environmental objectives can then be made.

Along the quantitative path, a switch to the qualitative path can be made at

any time, as indicated at some places in Fig. 3. It is more complicated to go

from the qualitative to the quantitative path. By using survey methods it is

however possible to ask people (laymen, experts or some other group) to value

the qualitative descriptions of the environmental impacts. If people are asked to

give monetary values, it will be an economic valuation. People can also be asked

to give ‘‘points’’ resulting in a non-monetary measure as a part of a MAA

method.

If the quantitative path is taken, Life Cycle Inventory data can be used. It

provides data on environmental emissions, resource extractions, etc., per func-

tional unit, e.g. MJ fuel or kg of material. In a PPP, different alternatives can

possibly provide different functions and this is a difference between SEA and

LCA. When the life cycle inventory data is combined with the description of the

technical system, the result is quantitative information about emissions, resource

extraction etc. for each alternative, often with some site-specific information.

The information available at this stage can either be taken directly to a

valuation step (not shown in Fig. 3) or as indicated in Fig. 3 it can be used for


further processing in the environmental analysis. In Fig. 3, three different paths

are suggested: traditional LCA characterisation, site-dependent LCA-character-

isation or risk assessment including air quality modelling as in the Impact

Pathway Approach. Traditional LCA characterisation will result in quantitative

Fig. 3. Information provided by different tools and their possible interaction.


information without site-specific information. The site-dependent LCA character-

isation and the risk assessment approaches can result in quantitative information

with some site-dependent information. The LCA-approaches can calculate results

on a pressure- or impact-level within the DPSIR-model. They cannot, however,

calculate environmental states (such as a concentration level). This is in contrast

to the risk assessment approach which can produce information at both pressure,

state and impact levels. After this step, a choice can be made as to whether a

valuation against environmental objectives is to be made, if the information

should be taken to another type of valuation or if the process is stopped at this

stage and conclusions are drawn with the available information.

If a valuation against environmental objectives is made, the LCA character-

isation approaches cannot be used if the objectives are expressed as desirable

states. If environmental state indicators are of interest, a risk assessment approach

is required in the earlier steps.

As an alternative to a valuation against environmental objectives, methods

such as multi-attribute approaches, economic valuation, and valuation using

energy, mass or area, can be used. All these methods can produce a one-

dimensional quantitative result, as opposed to a multi-dimensional qualitative

result.

The last step in Fig. 3 is to draw conclusions and formulate recommendations.

This should be made in relation to the aims of the SEA as formulated earlier in

the process. Not shown in Fig. 3 is the possibility to stop at almost any step in the

process and try to draw conclusions. It is therefore not necessary to follow a path

all the way. At this stage it is important to use all types of information produced

in the process. Thus, even if a quantitative, one-dimensional valuation method is

used, conclusions should also be based on results earlier in the process to ensure

that no relevant information is lost.

4.3. Some comments on the framework

The framework as outlined above is open to many choices, which are largely

determined by what is regarded as possible and what is regarded as a meaningful

result. Although partly scientific issues, the answers to such questions are partly

determined by values and world views. A few examples are used to illustrate

these discussions.

One example concerns the choice of which types of impacts to consider. This

can be illustrated by controversies around toxic compounds (Tukker, 1999).

Different stakeholders (e.g. industry, governmental agencies and NGOs) may act

in different frames or paradigms which determine what is considered as a

meaningful object to study. For example, industries may act within a ‘‘risk

assessment frame’’ where risk assessments are considered useful and meaningful

and management should be based on results from such studies (ibid.). NGOs may

on the other hand act within a ‘‘phase-out frame’’ where risk assessments are

considered too uncertain because science does not have enough knowledge and


management must be based on a precautionary principle (ibid.). Analysis based

on material and substance flows may be regarded as more appropriate and

meaningful under a ‘‘phase-out frame’’. Within a ‘‘risk assessment frame’’

indicators on an impact level are typically regarded as relevant and meaningful.

The precautionary principle is by some looked upon as unscientific, while others

see it as no less scientific than other principles (Sandin et al., 2002). Within a

‘‘phase-out frame’’, indicators on a pressure level, or even a driving force level

may be regarded as more meaningful.

Another example is whether it is meaningful and useful to produce a one-

dimensional result using a valuation method. It is clear that the use of valuation

methods involves different types of values, not only in relation to how different

aspects are valued against each other, but also in relation to what type of

valuation method should be used and also whether a valuation method should be

used at all (Finnveden, 1997). An illustration is the controversies around

economic valuation (Georgescu-Roegen, 1971; Daly and Cobb, 1990; Hobbs

and Meier, 2000). Another illustration is whether methods based on mass (Kleijn,

2001), energy or area produce meaningful results.

The framework suggests that several types of tools are to be used. It is

however important to note that this does not imply that the analyses are made

specifically for the SEA. It may be the case that results from earlier studies can be

used in the SEA. For example, LCAs can be time-consuming to perform, but

especially in the energy sector there are several studies that have already been

made. There is also software available with databases that can be used (e.g. Rice

et al., 1997; Jonbrink et al., 2000). Other tools in the framework, for example

surveys, are generally case-specific.

It is important to note that the quantitative road cannot be used for all types of

environmental impacts. In relation to the Swedish environmental objectives, it

can be noted that typical LCAs, and similar tools, can provide information mainly

related to objectives 1–5 and 7 in Table 1. The other objectives are more related

to impacts on ecosystems and landscapes and therefore more difficult to handle.

The lack of appropriate methods for Ecological Impact Assessment within the

framework of SEA was pointed out above.

4.4. The framework in relation to application

It is clear that the application can and should determine the SEA process. It is

however not clear how the application influences the choice of analytical tools

and methods. This is partly because the applications can be described in many

dimensions. In this section we will tentatively discuss some possible influences

on the choice of analytical tools.

4.4.1. Applications in different sectors

Although the focus of this study is on the Swedish energy sector, it can be

noted that large parts of the framework are also valid for other sectors and other


countries. In other sectors, the focus of the assessment may change from air

emissions. It is however suggested that for most sectors, most types of

environmental impacts are of importance. For example, in the energy sector,

not only air emissions are of relevance, but also water emissions (e.g. from waste

water from air pollution control devices, landfilling of combustion residues and

mining residues, and oil leakage) and direct impacts on ecosystems and landscape

(e.g. from production of bio fuels, mining operations, hydro power and wind

power). One possible difference concerns the site-specificity. For some sectors,

for example traffic or physical planning, already at a PPP level, a higher degree of

site-specificity may be possible and desirable. The influence of the site-specificity

is further discussed below.

4.4.2. Applications by different actors

SEA can be applied at national, regional and local levels. The degree of

site-specificity may change at these different levels. On a more local level,

both the possibilities and need for site-dependent assessments may increase.

Values and world views may change between different actors, e.g. govern-

mental agencies, industries and NGOs and this may also influence the choice

of methods. Resources in terms of time and funding may also change between

different actors. Besides these aspects (further discussed below), it is suggested

that the framework and choice of analytical methods is not influenced by the

actor.

4.4.3. Functions of the SEA

An SEA can have several functions, such as supporting a choice between two

or several alternatives or identifying critical aspects of studied alternative(s) and

suggest mitigation strategies. Within the same SEA process, both these functions

can be relevant. These different functions are related to the required degree of

quantification. If the intended application is an identification of critical aspects

in order to suggest a mitigation strategy, a qualitative approach is often

sufficient. Qualitatively it is often possible to determine if something is

‘‘critical’’ or ‘‘significant’’. However, if the objective is to support a choice

between two or several alternatives, the quantitative requirements typically

increase. This is because if a trade-off has to be made between two important

and critical aspects, a quantification of how severe the critical aspects are is

often necessary.

4.4.4. Possible and desirable degree of site-specificity

As noted above, the possible degree of site-specificity may vary between

different applications. It is typically higher at local levels and in applications

where land use is at the centre of the assessment. The desirable degree of site-

specificity is also related to world views and scientific paradigms. In the ‘‘risk

assessment’’ paradigm it is generally considered important to do a site-specific

assessment in order to assess the risks. In a ‘‘phase-out’’, or ‘‘strict-control


frame’’, it may be enough to consider the environmental pressures as discussed

above. Related to this topic is also the discussion on choice of indicators within

the DPSIR-model which is also connected to values and world views as discussed

above.

4.4.5. Values and world views

It is interesting to note that several of the application dependencies discussed

above boil down to a dependency on values and world views which can have a

decisive influence on the choice of analytical tools and methods. Since different

stakeholders have different values and world views, consultation and understand-

ing are important to ensure credibility and relevance. In some cases it may be

useful to use several different approaches and produce several different types of

results. These can then form a basis for further discussions.

In summary it is suggested that SEAs used to support a choice between

different alternatives require more quantitative methods, whereas SEAs used to

identify critical aspects and suggest mitigation strategies can be made with more

qualitative methods. The possible and desirable degree of site-specificity in the

assessment can also influence the choice of methods. It is also suggested that

values and world views can be of importance for judging whether different types

of tools and results are meaningful and useful.

5. Conclusions

The aim of this article is to analyse how various analytical tools can facilitate

and enhance the SEA process, particularly in relation to energy sector SEA. It is

found that several existing tools can contribute, either by focusing mainly on the

identification and modelling of environmental change (e.g. LCA, RA, future

studies) or by focusing mainly on the valuation stage (e.g. MAA methods,

economic valuation methods, surveys). Based on the tools examined here, it

appears that finding useful tools for analysing ecosystem and landscape impacts

is more challenging than tools for analysing emissions of pollutants, at least in the

energy sector context.

Based on the examination of the selected analytical tools according to a set of

analytical features, an integrative framework of methods for SEA is proposed

(see Figs. 2 and 3). Three conclusions can be drawn from this exercise. First, the

key factors influencing the choice of analytical tools are the definition of system

boundaries, the amount and types of environmental changes included in the

assessment, the degree of site-specificity desired, the degree of quantification

desired, the degree of aggregation of results desired, and the preference of

information type according to the DPSIR-model. Second, and in addition to these

factors, the preferred function of the SEA also influences the choice. It is argued

that to support a choice between two or more alternatives quantitative results may

be needed, while a SEA with the purpose to identify critical aspects of


alternative(s) and suggest mitigation strategies might do with qualitative results.

Lastly, it was suggested that underlying the choices that shape the use of methods

in a SEA is the world view and assumptions of the assessor, i.e. considered

relevant information.

The next step will be to test the framework of analytical tools in a SEA on a

Swedish energy sector PPP.

Acknowledgements

Financial support from the Swedish National Energy Administration is

gratefully acknowledged.

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