SIXTH FRAMEWORK PROGRAMME
Project contract no. 502527
ESPREME
Estimation of willingness-to-pay to reduce risks of exposure to heavy metals and
cost-benefit analysis for reducing heavy metals occurrence in Europe
Specific Targeted Research Project
Research priority 1.6. Assessment of environmental technologies for support of policy decisions, in
particular concerning effective but low-cost technologies in the context of fulfilling environmental legislation
Workpackage 06 – D05b
Exposure-response-functions for HM impacts on ecosystems and crops
March 2007
Due date of delivery after extension: February 2007 (final version)
Actual submission date: 6st February 2007
Start date of project: 1st of January 2004
Duration: 36 months
(extended to March 2007)
Lead authors for this deliverable: Mohammed Belhaj, IVL
Project co-funded by the European Commission within the Sixth Framework Programme (2002-2006)
PU
PP
RE
CO
Dissemination Level
Public
Restricted to other programme participants (including the Commission Services)
Restricted to a group specified by the consortium (including the Commission Services)
Confidential, only for members of the consortium (including the Commission Services)
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Table of Contents
1.
INTRODUCTION ........................................................................................................................... 2
2.
EVALUATION METHODS .......................................................................................................... 2
2.1.
2.2.
2.3.
3.
HEALTH EFFECTS ....................................................................................................................... 3
EVALUATION OF HEALTH EFFECTS ............................................................................................. 6
ECOSYSTEM EFFECTS ................................................................................................................. 9
CONCLUSION.............................................................................................................................. 11
List of Tables
Table 2.1 Damage costs of air emissions, Euro per kg of emission ......................................................... 6
Table 2.2 Human toxicity due to heavy metals emissions (Euro/kg) ....................................................... 7
Table 2.3 Environmental taxes and fees existing in Sweden in 1998 *. ................................................... 7
Table 2.4 Options for a charge on cadmium in fertilizers ........................................................................ 8
Table 2.5 One-step weighting factors for human toxicity* ...................................................................... 9
Table 2.6 One-step weighting factors for aquatic ecotoxicity based on AEPs ....................................... 10
Table 2.7 One-step weighting factors for aquatic ecotoxicity for metals released to soil and air. ......... 10
Table 2.8 One-step weighting factors for terrestrial ecotoxicity for metals released to soil and air. ..... 11
List of Figures
Figure 2.1 The composition ot the total economic value .......................................................................... 3
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1. Introduction
Air pollution is a question of higher importance in Europe and upcoming legislation and revisions of
Convention (CLRTAP) are expected to enhance the European air pollution. Established in 1998 the
joint Task force on Health Aspects of Long-Range transboundary Air pollution (TF Health) by the
Executive Body and the World Health Organization/Regional Office for Europe analyzed the impacts
of particulate matter, and ozone, heavy metals and persistent organics. Based on the WHO review of
new evidence of health effects of common air pollutants, published in 2003, the TF Health reviewed
the available information on the sources, chemicals properties and spatial distribution of pollution from
cadmium (Cd), lead (Pb) and mercury (Hg) and evaluated the potential health effects in Europe. In the
case of mercury for instance, estimating the risk of dietary exposure is at present difficult to quantify
due to the very complex dynamics of Hg in soil and water, and the factors that influence the
biomagnification of methylmercury concentration in predatory fish.1
In order to estimate the willingness of EU countries to pay for preventive measures related to six heavy
metals, namely – arsenic, cadmium, chrome, lead, mercury and nickel, the ESPREME programme
aims at analysing costs and benefits of reducing the effects of HM on health and the ecosystem. Since
other work packages of the programme are concerned with the cost side, the objective of this study is a
literature survey related to willingness to pay to reduce the emissions. However, since studies assessing
willingness to pay to reduce the impacts of heavy metals are scarce, some surrogate methods are
presented and discussed in the case of health toxicity. In the case of ecosystems some methods will be
briefly discussed and the Ecotax method will be presented and the estimated values based on this
method are presented for ecological toxicity.
This study is structured as follows. Section 2 is a discussion of different evaluation methods and the
results based on some methods are presented. Section 3 concludes.
2. Evaluation methods
In order for physical measures of impacts to be commonly measurable, they must be valued in
monetary units. The monetary valuation of different effects is not a straightforward procedure since
many of the effects have no market value. In general we often talk about the total value of something.
As shown in Figure 2.1, this total value is often composed of both use and non-use values. The use
value is the value derived from actual use of a good or service. The non-use values, also referred to as
“passive use” values, are values that are not associated with actual use, or even the option to use a good
or service.
The use value includes direct, non-direct and option values. The direct use value is the value attributed
to direct utilisation of ecosystem services. Non-direct-use values or "functional" values relate to the
ecological functions performed for example by forests, such as the protection of soils and the regulation
of watersheds. Option value is the value that people place on having the option to enjoy something in
1
http://www.unece.org/env/wge/inf.doc.1.2004%20Substantive%20Report.draft.pdf
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the future, although they may not currently use it. On the other hand, the non-use values include both
bequest and existence value. Bequest value is the value that people place on knowing that future
generations will have the option to enjoy something. Existence value is the value that people place on
simply knowing that something exists, even if they will never see it or use it. In order to assess these
values, environmental economics uses several methods.
These methods may be based on stated preferences involving studies including questionnaires asking
respondents for their willingness to pay such as in the case of contingent valuation and stated
preference methods. Other methods are based on revealed preferences that are often based on
consumers´ or producers' behaviour or actions such as the hedonic price method that is used to estimate
the value of environmental effects on properties such as the effect of noise or air pollution on house
prices.
Figure 2.1 The composition ot the total economic value
2.1.
Health effects
In the case of health effects other methods than stated or revealed preference ones can be used to
estimate the impact of externalities. These methods may be HALY, DALY or QALY. The HALY is a
Health Adjusted Life Year i.e. a generic term that includes the two most popular measures, the QALY
or Quality Adjusted Life Year and the DALY or Disability Adjusted Life Year. The QALY is simpler.
A value of quality of life is assigned from 0 (dead) to 1 (perfect health). The DALY is different in that
the reference states are 0 for perfect health and 1 for dead, and it is estimated for particular diseases,
instead of as a health state.2
more details related to these methods see ENGRI
http://www.mse.cornell.edu/courses/engri119/Class_Notes/haly_daly_qaly.html
2
For
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DALY ( x)  D(Cx x )e  r ( x  a )
Where x is age, C is a constant with the value 0.16243, r is the discount rate, and β equals 0.04.
Except DALY other evaluations of health effects estimates are based on the value of statistical life
(VSL) or the value of lost years (VOLY) where the relation between the two is as follows: The
willingness to pay for Δs (the change in the risk to die) leads to the value of statistical life such as:
VSL  WTPi / sN
i
where N is the population at risk. Within the ExternE for instance mortality is evaluated according to
the reduction of life expectancy but there has been a lack of studies that determine the value of a life
year (VOLY). Therefore ExternE derives VOLY from the so called VSL for which there are a very
large number of studies, by assuming that VSL is the present value of a discounted series of annual
VOLY values such as:
VSL VOLY  x Pi / (1  r )i  x 
T
ix
where xPi is the conditional probability to live until year i for a person at age x. T is the maximum
expected life length and r is the discount rate.
The VSL method has been used for instance in the impact pathway approach that has been developed
in the EC funded ExternE project series. Based on dose response functions this approach estimates the
environmental damage costs by using a damage function and associates the environmental change to
the implied impacts on health and the environment. To estimate the damage cost on health for instance,
different methods have been used to calculate the value of a statistical life such as VSL and the value of
a year of lost life (YOLL).
However, using DALY to estimate the damage costs may lead to underestimations (Pearce et al
(2004)). In order to avoid this underestimation that is based on the equalisation of expenditure spent on
avoiding and treating the causes of DALY, a next best approach is developed and includes the
willingness to pay (WTP) of individuals to avoid health states in question. As the estimate of VSL is
more easily defended from a theoretical perspective, Pearce et al (2004) follow a procedure adopted in
a World Bank study of pollution control and assign a WTP value to a DALY based on an ‘anchor’
estimate of VSL and an implied value of VOLY.
As in the case of health effects, direct measures and monetarisation of ecosystem damage may be
difficult or impossible to find. Furthermore, studies based on willingness to pay are scarce. Therefore other
methods are used to study the effects on ecosystem methods such as:
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-The standard price method
The standard price approach estimates the revealed preferences of policy makers. It calculates the
benefits of emission reduction – as perceived by policy makers - based on the abatement costs to reach
a well-defined emission reduction target. These costs are a proxy for the benefits that policy makers
attribute to these reductions, as we assume that policy makers act as rational decision makers who
carefully balance (their perception of) abatement costs of emission reductions with (their perception of)
the benefits of these emissions. However, since the standard price approach is based on the current
preferences of policy makers, as reflected in air quality policies, it cannot be used for cost-benefit
analysis or policy advices related to these emission reduction policies (Vermoote (2003)).
-The Clean-up cost approach
The clean-up cost approach approximates the estimation of damage costs using costs linked to cleaning
up or reducing a particular pollution. The idea is that once the damage resulting from pollution is done,
then the costs of rehabilitation to achieve the pre-damage situation will appear as a measure of damage
done. From an economic point of view, this deduction is at the very least debatable. Strictly speaking,
the costs of cleaning up are not a substitute for damage costs (EC (2000)). As in the case of standard
price approach, the clean-up costs cannot be used in a cost benefit analysis and they need to be used
with caution if they are applied as measures of damage.
-The Ecotax method
The Ecotax method also called the valuation weighting method is a monetarisation method based on a
combination of environmental taxation with Life Cycle Analysis (LCA). This method which can
estimate both health effects and environmental effects relies on two basic assumptions where the first
one is that the members of parliament represent the will of the people, and the second is that the
environmental tax system represents the priorities of the parliament. This method is in line with the
assumption in Leksell (1987) and Carlsson et al (2003) and EC (2003) where the method is based on
the fact that since a tax level is set optimally, based on governmental objectives, then this number
reflects the social value per unit of emission reduction.
This Ecotax method proceeds in the following way to estimate damage costs of certain emission. In
cases where the tax or fee used is not on the reference substance of the characterization method (as is
the case for many substances), the calculations are made according to the principle that a contribution
to an impact category can be considered equally harmful independently of what caused it. The value of
one extraction or emission may be translated into another extraction or emission contributing to the
same impact category, by means of characterisation factors (Johansson (1999)). For example, an
emission of 1 kg of methane is, according to IPCC (1995), equivalent to 56 kg of carbon-dioxide over a
20 year time frame. Emissions of carbon-dioxide has the value 0.041 Euro per kg, and thus, the
emission of methane receives the value (56 kg/kg * 0.041/kg) = 2.30 Euro/kg. The performance of such
calculations results in weighting factors that include both the characterisation and the valuation substeps of impact assessment.
Hence, the functional form for the estimations is as follows:
X x  ( Aa / Bb ) *Yx
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(Xx) represents the value of substance x, which is the one-step weighting factor searched for, (Aa) is the
valuation weighting factor for substance a and (Ba) is the characterization factor for substance a and
(Yx) is the characterization factor for substance x, given by the characterization method.
2.2.
Evaluation of health effects
Health effects from heavy metal have been estimated using different methods. In a study on the
economic valuation of environmental externalities from landfill and incineration of waste (EC (2000))
damage costs due to heavy metals were collected. These costs are shown in table 2.1 where the
differences between the estimated values depend on the effects included and the methods used in the
estimation.
Table 2.1 Damage costs of air emissions, Euro per kg of emission
Study 1
Study 2
Study 3
Arsenic
Cadmium
150
999
1015735
18.3
81.4
125370
Chrome
(VI)
123
819
200642
Lead
Mercury
Nickel
16.8
34627
25909
2.53
101549
Source: Adapted from EC (2000).
Notes:
Study 1: Ralf et al (1998): Study on Health Risks of Air Pollution from Incinerators. NOx: NO2 (via xNO3) + NO2 (via
O3)
Study 2: EC (1996): Economic Evaluation of the Draft Incineration Directive. VOC: total TOC is indicated. German site
used as base case.
Study 3: ECON (1995): Miljøkostnader knyttet til ulike typer avfall. Conversions performed with exchange
In the table, studies 1 and 2 are based on ExternE for conventional pollutants.
Study 1 only includes health effects whereas study 2 includes other quantifiable effects as well (EC
(2000). As shown the variation of the costs is large. However, it appears that studies 1 and 2, which are
based on epidemiological studies, are in the same order of magnitude. In contrast the results of study 3
based on health indices are much higher. It could be argued that studies 1 and 2 do not look at all health
effects from these substances, because not all health effects could be quantified. However, the
theoretical validity of the valuation approach used in study 3 is questionable, thus bringing into
question the robustness of the cost estimates (ibid).
In a study by Spadaro and Rabl (1999) carcinogenic toxicity of arsenic, cadmium, chrome and nickel
was evaluated based on the impact pathway methodology integrating real impacts of air emissions in
Life Cycle Analysis.3 Only the inhalation dose was taken into account.
3
A life cycle analysis and assessment (also known as life cycle analysis, life cycle inventory, ecobalance, cradle-to-graveanalysis, well-to-wheel analysis, material flow analysis and dust-to-dust energy cost) is the assessment of the environmental
impact of a given product or service throughout its lifespan.
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Table 2.2 Human toxicity due to heavy metals emissions (Euro/kg)
Spadaro & Rabl (1999)
Arsenic
Cadmium
Chromium (VI)
Nickel
(*) EC (2000)
171
20.9
140
2.87
Intervals of values from
literature*
150 – 999
18.3 - 81.4
123 – 819
2.53 - 16.8
As shown in table 2.2 Sparado et al results that are assumed to be consistent (EC (2003)) lie in the low
range of values collected from other studies where overestimation may depend on the method used, the
data available or both.
- Ecotax method
In order to use the Ecotax method the different environmental taxes and fees existing in the Swedish
tax system shown in table 2.3 are linked to the appropriate impact category. Since the same substances
that are toxic to many other organisms in the ecosystems are also toxic to humans the substances
treated in this method that are related to ecotoxicological effects are also applicable on human
toxicological effects.
Except the fact that the levels of taxes and fees are assumed to be set optimally, the choice of the
Swedish values depends on their high level compared to the levels of taxes and fees in other EU
countries. The low level of taxes and fees, if they exist, in other EU countries would imply that the
taxes are lower than the damage costs they give rise to.
Table 2.3 Environmental taxes and fees existing in Sweden in 1998 *.
Impact category
Ecotoxicological
and human health
Taxes and fees
effects 2.2 Euro/kg pesticide
3333.3 Euro/kg cadmium
1.1 – 11.1 Euro/kg benzene
20-38.9 Euro/kg lead
Source: adapted from Johansson (1999) where non-toxicological effects is not accounted for. (*) based on 1 Euro= 9 SEK,
However since the level of these taxes and fees has not been changed the values are not adjusted to 2007.
In Europe, environmental taxes and fees in the agricultural sector that are connected to different
impacts are not available or they do not exist in many countries. EC (2000) studied the introduction of
charges on cadmium in Europe in order to estimate the value of the charge. Since there is a wide
variation within the EU in the sensitivity of soils to cadmium accumulation, as well as in the cadmium
levels in agricultural soils, the optimal tax rate would therefore differ by region from an environmental
point of view. In theory farmers in sensitive areas would pay more for high cadmium content fertilizer
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use than farmers in other areas. However, a ‘farm-to-farm’ differentiation of the tax rate seems to be
unfeasible. Differentiation by Member State is an alternative, although differences in soil sensitivity
and background concentrations of cadmium within one Member State can be considerable. However, 3
options of a charge on cadmium in fertilizers are calculated based on remediation costs. These charges
which are shown in table 2.4 are proposed to internalize the costs of the impacts.
Table 2.4 Options for a charge on cadmium in fertilizers
Uniform approach
Minimum
approach
Divergent approach
Degree of
harmonizati
on
Applied in all Member States at
a uniform charge rate
Applied in all
Member States
at charge rates
equal
to, or exceeding
an EU wide
minimum
Applied in some Member
States
only,
with
differences in charge rates
Moderate
version
Charge rate EUR 0.25 g/ Cd,
initially applying
to fertilizer with more than 60
mg Cd per kg P2O5. Threshold
lowered to 40 mg Cd per kg
P2O5 after 2 years and to 20 mg
Cd per kg P2O5 after another 2
years
Minimum
charge rate EUR
0.25 per gramme
Cd; thresholds
and phasing as
in
Uniform
approach
Charge rates on average
EUR 0.25 g/Cd; thresholds
as
in
Uniform
and
Minimum Rate
approach; phasing differing
between Member States
Stringent
version
Charge rate EUR 1.00 per
gramme Cd; no threshold
or phasing
Minimum
Charge
rate
EUR
1.00 g/ Cd; no
threshold
or
phasing
Charge rates on average
EUR 1.00 g/Cd;
no threshold or phasing
When comparing the Cd taxes in Sweden i.e. 3.34 Euro per gramme of Cd with the proposed maximum
charges in Europe i.e. 1 Euro per gramme of Cd , the Swedish tax is more than three times the
maximum charge proposed for Europe. However, since the charge on cadmium is based on
remediation’s costs which is based on equivalent assumptions as the clean-up cost approach, discussed
above, the values may be misleading.4 Therefore, the charges used in this study for the calculation of
damage costs based on the Swedish taxes and fees would be optimal.
-results of ecotax method
Based on the Ecotax method as well as the Swedish taxes and fees the one-step weighting factors for
human toxicity for emissions to outdoor air, emission to water and emission to soil are from Johansson
Since short-run own-price elasticity on the demand for chemical fertilisers ranges from –0.2 to –0.3 (Bäckmann, 1999),
the low charge would not lead to any significant reductions of the fertilizers.
4
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(1999)). The weighting factors are based on human toxicity potentials (HTPs) from Jollier and Crettaz
(1997).
Table 2.5 One-step weighting factors for human toxicity*
Human toxicity for Human
emissions to outdoor air toxicity for
emissions to
water
Based on Based
on Based
on
11.11
38.89
2.22 Euro/kg
Euro/kg
Euro/kg
copper
benzene
lead
Arsenic
Cadmium
Chromium
Lead**
Mercury
Nickel
7.5e+07
1.6e+08
3.1e+07
1.9e+07
3.8e+08
3.1e+06
1.4e+03
2.9e+03
5.6e+02
3.5e+02
7.0e+03
5.6e+01
1.4e+03
2.9e+03
5.6e+02
7.8e+02
7.1e+03
5.6e+01
Human
toxicity
emissions to soil
Based on
3333.3
Euro/kg
cadmium
Nonedible
crops
1.4e+04
3.0e+04
6.0e+03
1.2e+04
7.4e+04
6.0e+02
for
Based
on
3333.3 Euro/kg
cadmium
Edible crops
1.5e+04
3.0e+04
5.9e+03
3.7e+03
7.4e+04
5.9e+02
(*) based on 1 Euro= 9 SEK. (**) only if the inhalation route of exposure is taken into account
As shown in the table values related to human toxicity for emissions to outdoor air in general and
mercury in particular are the highest implying that the damage of this substance on human health is
considerable. However, it is difficult to compare the values shown in this table to the values shown for
the human health in the studies discussed above based mainly on the substances studied and the
methods used.
2.3.
Ecosystem effects
The toxicological effects are difficult to assess. The toxicological effects depend on many things, such
as the concentration of the substance that the organism is exposed to, and the route of exposure. There
may also be different degrees of sensitivity towards the same substance among different organisms.
Another thing determining the damage imposed by a chemical is its fate in the environment, i.e. how
long it stays there before it is degraded, what the degradation products are and how it is transferred
between different environmental compartments such as air, soil and water. Analysing the effects of
chemical substances releases requires a lot of data, which is often far from easily obtained. Out of the
approximately ten million known substances, about 20 000 - 60 000 are in commercial use in the
industrialised countries (Thunberg 1992). Even for the most commonly used chemicals many features
determining their toxicological properties remain unknown (Johansson (1999)).
Hence, two problems exist regarding environmental impact assessment (EC (2000)):
-Data are not sufficient to calculate the damage caused by a given impact to ecosystems;
-There is no generally accepted way to quantify quantifiable damage caused to ecosystems.
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- Evaluation of ecosystem effects
Based on the Ecotax method as well as the Swedish taxes and fees the following one step weighting
factors are estimated for:
- Aquatic ecotoxicity based on aquatic ecotoxicity potentials (AEPs) for emissions to water from Jolliet
and Crettaz (1997);
- Aquatic ecotoxicity for metals released to soil and air;
- Terrestrial ecotoxicity for metals released to soil and air.
In the case of aquatic ecotoxicity table 2.6 shows the calculated values based on AEPs for emissions to
water and the tax on the active substance in pesticides. As can be seen mercury have the largest impact
(as in the case of human health) followed by cadmium. The impact of the other heavy metals is
comparatively marginal.
Table 2.6 One-step weighting factors for aquatic ecotoxicity based on AEPs
Characterization factor
Substance
AEP emission to water
One step weighting factor
Euro/kg
Based on Euro 2.22 /kg * copper
Arsenic
Cadmium
Chromium
Lead
Mercury
Nickel
0.52
520
2.6
5.2
1300
0.79
0.2
222.2
1.1
2.2
555.5
33.3
* based on 1 Euro= 9 SEK
When it comes to aquatic ecotoxicity for metals released to soil and air the damage cost of mercury is
the largest being 8333 Euro/kg for emissions to soil and 6000 Euro/kg for emissions to the air. In the
case of cadmium its damage is lower than that for mercury but it is much larger than the damage of the
other substances shown in the table.
Table 2.7 One-step weighting factors for aquatic ecotoxicity for metals released to soil and air.
Emission to soil
Characterizatio
n factor
Substance
AEP
Arsenic
Cadmium
0.24
240
Emission to air
One
step Characterizat One step weighting factor
weighting
ion factor
factor
Euro/kg
Euro/kg
Based on AEP
Based on Based
on
Euro
180 Euro Euro/kg
3333.33/kg)
20/kg)
38.88/kg lead
cadmium
lead
98 octane
93 octane
3.3
0.08
1.2
2.3
3333.3
79
1222.2
2444.4
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Chromium 1.2
16.6
0.4
6.1
12.2
Lead
3.9
54.4
1.3
20
38.9
Mercury
600
8333.3
200
3111.1
6000
Nickel
0.4
5
0.12
1.9
3.7
When it comes to terrestrial ecotoxicity for metals released to soil and air the results show again that
mercury and cadmium have the largest impacts on the ecosystem.
Table 2.8 One-step weighting factors for terrestrial ecotoxicity for metals released to soil and air.
Emission to air
Characterization One step
factor
factor
Substance
Emission to soil
Characterization One
step
factor
weighting
factor
Euro /kg
AEP
Based on
30000
SEK/kg
(Euro
3333.33/kg)
cadmium
Arsenic
Cadmium
Chromium
Lead
Mercury
Nickel
2.3
9.6
0.3
0.4
18
1.1
0.8
3.1
0.08
0.13
5.9
0.35
800
3333.3
90
144.4
6333.3
377.8
AEP
Euro /kg
Based on
180
SEK/kg
(Euro
20/kg)
lead
93
octane
111.1
477.8
12.2
20
911.1
53.3
weighting
Based on
350
SEK/kg
(Euro
38.88)
lead
98 octane
222.2
944.4
24.4
38.9
1777.7
104.4
3. Conclusion
Based on the lack of WTP studies to assess the benefits of reducing the effects on heavy metals on
human health and the environment, several other methods can be used to estimate the damage cost of
the metals. In the case of health effects a combination of the impact pathway approach and DALYs
would be the second best choice.
When it comes to effects on the ecosystem, the Ecotax method combining taxes and fees with LCA
may also constitute a second best choice. However, one should keep in mind that the results presented
concerning the impacts relative to different substances depend on the taxes and fees being used as well
as the toxicity potentials. When it come to taxes and fees, the Swedish values that are used are higher
than both the average in EU as well as the suggested levels in EC (2000) related to the case of
cadmium. Therefore, the estimated values for the ecosystem might give an insight on the ranges of the
damage cost implied by the studied heavy metals especially if the values are meant to be used in a cost
benefit analysis.
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Further, based on the results of environmental effects the damage of mercury and cadmium are highest
compared to the other studied heavy metals. Compared for instance to the cost of decadmiation these
damage values are very high suggesting higher charges on cadmium.
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DELIVERABLE D05B
Appendix
Critical Surface-Time 95
Jolliet and Crettaz (1997) suggest a method for the characterisation of ecotoxicological effects within
the Critical Surface-Time 95 methodology for life cycle impact assessment. The ecotoxicity potentials
in this approach are based on the assumption that two emissions are equivalent if they generate their
respective no effect concentration (NEC) during one year in the entire ecosystem considered. The
effect is assumed to be linear both with concentration and polluted volume. Thus, a concentration of a
pollutant of 0.5 mg/m3 in a volume of 1000 m3 is considered to have the same effect as a concentration
of 1 mg/m3 in a volume of 2000 m3. Fate and exposure are taken into account by the degradation,
dilution, inter-media transfer and food chain/bioconcentration routes. The mean residence time of the
substance in the medium per height of dilution is also considered. For the aquatic ecotoxicity potentials
(AEPs) the reference substance is zinc emitted to water. When determining the NEC, extrapolation
factors of 100 is used if three LC50 are available (for algae, crustacean and fish) and of 1000 if less than
three LC50 are available. Only surface fresh water is considered. The reference substance for the
terrestrial ecotoxicity potentials (TEPs) is zinc emitted to soil. The NECs are based on the LC 50 of earth
worms with an extrapolation factor of 1000. Where no data is available for earth worms, terrestrial
ecotoxicity is extrapolated on the basis of the LC50 for Tubifex worms in water. AEPs are presented for
emissions of eight metals to water, soil and air and for emissions of oil, phenol, phosphate and BOD to
water.
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