Sufficiency and Effectiveness Review of the HM Protocol, Chapter A3

Draft, 21 May 2023

Results of modelling and mapping of critical loads of lead, cadmium

and mercury and critical concentrations of mercury in precipitation

and their exceedances in Europe

Gudrun Schütze

Jean-Paul Hettelingh

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1 Introduction

The development of the critical load approach for heavy metals within the framework of the Working Group on Effects was inspired by Article 6 (g) of the Protocol, which encourages work on an effects-based approach for the purpose of future optimised emission control strategies. The critical loads approach was deemed to be an appropriate way to link depositions of metals with effects on human health and the environment. The reason included the fact that optimised control strategies for acidifying and eutrophying air pollution had been successfully applied in support of abatement policies in Europe.

The scientific basis of the critical loads approach for Europe was developed over a period of 11 years with stepwise improvements of the methodologies according to the state of knowledge. It started with first exercises in the framework of the ESQUAD Project (Van den Hout (ed.) 1994) followed by meetings organised in the framework of the Convention’s International Cooperative Programme on Modelling and Mapping Critical Levels and Loads and Air Pollution Effects, Risks and Trends (ICP Modelling and Mapping). These meetings included heavy metal sessions of workshops of the Coordination Center for Effects (CCE) of ICP Modelling and Mapping from 1995 until 2005 and a series of scientific workshops and expert meetings (Gregor et al. 1998, Gregor et al. 1999, Curlík et al. 2000, Schütze et al. 2003). The first manuals for modelling and mapping of critical loads of heavy metals were published at the end of the last century (De Vries and Bakker 1996/1998, De Vries et al. 1996/1998). A first European wide dissemination and voluntary application by the network of National Focal Centres (this network currently covers European countries only) of the preliminary methodology was requested by the Working Group on Effects to the Coordination Center for Effects at its 20th session. Results are described in a collaborative report (Hettelingh et al., 2002) of the CCE and the EMEP Meteorological Synthesizing Centre – East. At its twenty-third session the Working Group on Effects then requested the Coordination Center for Effects to issue a call for data to its National Focal Centres on critical loads of cadmium, lead and mercury and Hg. The result of this call for data was reported to the 21st meeting of the Task Force ICP Modelling and Mapping, to the 24th session of the Working Group on Effects (EB.AIR/WG.1/2005/10/Add.1), and documented in a second collaborative report of the Coordination Center for Effects and the EMEP Meteorological Synthesizing Centre East (Slootweg et al, 2005).

EMEP Meteorological Synthesizing Centre East modelled depositions of cadmium (Cd), lead (Pb) and mercury (Hg) for 1990 and 2000. By comparing maps from the Meteorological Synthesizing Centre East of depositions with CCE maps of critical loads it has been possible to identify geographic locations in Europe where critical loads are exceeded.

In Slootweg et al. (2005) the response by the National Focal Centres, of European maps on critical loads of cadmium, lead, and mercury as well as of preliminary exceedance maps is described while updates received after the 21st meeting of the Task Force on ICP Modelling and Mapping are summarized in Hettelingh et al (2005). An evaluation to which extent the critical loads approach provides a satisfactory scientific basis for application in air pollution policy will be provided in a chapter of the Chairman’s Report to the Working Group on Strategies and Review (EB. AIR.WG.5/2006/XXX).

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2 Main principles of critical loads of heavy metals calculations

The methodology to calculate critical loads of Pb, Cd, and Hg is described in detail in the Convention’s Manual on Modelling and Mapping Critical Levels and Loads and Air Pollution Effects, Risks and Trends (Modelling and Mapping Manual 2004) and a related background document (De Vries et al. 2005). The description here is limited to the main principles and key variables and is excerpted from the Manual.

The critical load of a metal is the highest total metal input rate (g ha-1 a-1) below which harmful effects on human health and ecosystems will not occur in an infinite time perspective, according to present knowledge. While critical loads explore the sensitivity of ecosystems against metal inputs, the risk of effects can only be described by the exceedances, i.e. by comparison of critical loads with the actual inputs.

The method to calculate critical loads of heavy metals is based on a steady state balance of ecosystem inputs and outputs of heavy metals. The underlying assumption of steady state for the fluxes as well as chemical equilibrium in an undetermined future is consistent with concepts of sustainability. Implications of this on the interpretation of critical loads and their exceedances are explained later in this chapter.

2.1 Effects, indicators and critical limits

Receptor dependent critical loads of Pb, Cd, and Hg as well as the critical concentration of Hg in precipitation can be calculated when a representative effect indicator is identified. The metal concentration in an environmental compartment is such an indicator. The critical limit is the maximum of this indicator concentration that will not cause harmful effects in the long-term. Approaches have been designed to assess critical limits addressing either ecotoxicological effects or human health effects. Not all effects have the same relevance for every metal. Table 1 lists effects and their indicators that have been addressed in the 2004 call for data of the CCE.

Table 1. Overview of effects indicators used in the calculation of critical thresholds.

Effect_no Effects (indicators) Ecosystems Metals

1 Human health effects (ground water quality in view of use for drinking water supply)

Terrestrial ecosystems Pb, Cd, Hg

2 Human health effects (food quality) Terrestrial ecosystems

(arable land only)

Cd

3 Ecotoxicological effects Terrestrial ecosystems Pb, Cd, Hg

4 Ecotoxicological effects Fresh water ecosystems

Pb, Cd

5 Human health effects (food quality) Fresh water ecosystems

Hg

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Oral uptake is the main pathway for human health effects of environmental Cd, Pb as well as Hg. Critical loads for terrestrial ecosystems addressing human health effects can be calculated, either in view of not violating food quality criteria in crops or in view of ground water protection regarding its potential use as drinking water. The derivation of critical limits used for critical loads to protect human health requires complex analyses including pathways of metal intake independent from the environmental situation (Cd in cigarette smoke, Hg in dental amalgam etc.). In agreement with the Joint Task Force on Health Aspects of Air Pollution it was decided to use internationally accepted critical limits. Therefore, critical limits to protect ground water quality (effect 1) were set to the recommendations for maximum metal concentration in drinking water of WHO (2004) Cd: 3 mg m-3; Pb: 10 mg m-3, Hg: 1 mg m-3.

An appropriate indicator for critical load calculations addressing human health effects via food intake (effect 2) is the Cd concentration in wheat. The EU regulation (EG) No.466/2001 uses a limit for Cd of 0.2 mg kg-1 fresh weight in wheat grains. This limit is, however, not based on effects (it was derived with the principle “As Low As Reasonably Achievable – ALARA). In the critical loads calculation a conservative effects-based food quality criterion for wheat of 0.1 mg/kg fresh weight was used (see De Vries et al. 2005, Appendix 4). Using this criterion the pathway of Cd to wheat leads to the lowest critical Cd content in soils therefore protecting also against effects on human health via other food and fodder crops (including also the quality of animal products, De Vries et al. 2003).

The quantification of the risk of human health effects of Hg through fish consumption (effect 5) is not related to a critical load (g ha-1 a-1) but to a critical concentration of Hg in precipitation (ng L-1). This critical Hg concentration in precipitation can be linked with a simple model to the Hg concentration in fish assuming a steady-state situation in the catchment. A limit of 0.3 mg kg-1 fresh weight on total Hg in fish is consistent with recommendations by the USEPA (2001) and the WHO/FAO (2003) and therefore used in the calculation. This limit is frequently exceeded in Nordic surface waters already now.

Among terrestrial ecosystems, critical loads of Cd and Pb are to be calculated from the viewpoint of ecotoxicology for areas covered by non-agricultural land (forests, semi-natural vegetation) or agricultural land (arable land and grassland). Soil toxicity data collated and accepted under the terms of current EU risk assessment procedures (Draft risk assessment report Cd (EC 2003), Voluntary risk assessment for Pb, draft report, status 2006, provided by the Lead Development Association (LDA International) were used as basis for the derivation of critical limit functions for free metal ion concentrations (Pb, Cd) in soil solution. The data covered chronic effects on plants, soil-dwelling invertebrates and microbial processes on the population-level. While the ecotoxicological database was harmonised with EU risk assessment, the critical limits derivation was done in a different way: The free metal ion approach was considered most appropriate for the assessment of the influence of the bioavailability of metals on the effects on related organism groups. The bioavailability of metals does, however, not only depend on the free metal ion concentration but also on the concentration of other cations, particularly H+. This was taken into account in deriving critical limits as a function of the pH in soil drainage water. The method of derivation of the critical limit function is described in detail in Lofts et al. (2004). Critical limits taking into account secondary poisoning were explored in the background paper for the Manual (DeVries et. al. 2005). They were however considered very uncertain and

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therefore not included in the Modelling and Mapping Manual (2004) and current critical loads calculations for Europe.

Organic forest (top)soils are considered as the best understood critical receptor with respect to atmospheric Hg pollution, based on knowledge on effects on microbial processes and invertebrates. The suggested critical limit for Hg is that the concentration in the humus layer of forest soils after normalization with respect to the organic matter content should not exceed 0.5 mg (kg org)-1 (Meili et al. 2003). The strong association of Hg with organic matter leaves virtually no free ions. Therefore the exposure of biota to Hg is controlled by the competition between biotic and other organic ligands. Furthermore the contamination of all types of organic matter is determined by the supply of organic matter relative to the supply of Hg at a given site (Meili 1991, 1997). As a result, the critical limit for Hg in soils is set for the organically bound Hg rather than for the free ion concentration, also in solution.

In order to describe ecotoxicological effects of Cd in surface waters the 5-percentile cut-off value of chronic toxicity from the Draft EU risk assessment report (EC 2003) was used as critical limit. Deviating from the draft EU risk assessment no additional assessment factor was applied. For Pb the critical limit is based on Crommentuijn et al. (1997). It represents the highest value of a range for critical limits (to be used in dependence on water chemistry) suggested by a workshop of ICP Waters on heavy metals, 2002, in Lillehammer, Norway (Skjelkvale and Ulstein, 2002) These critical limits of Cd, Pb are provided as total dissolved concentrations. The free metal ion approach could not be used to derive critical limits of Pb and Cd due to limitations in the effects database for aquatic systems.

2.2 Calculation of tolerable metal fluxes in terrestrial ecosystems

Critical loads of metals for terrestrial ecosystems are focussed on the top soil. The soil depth to be considered in the quantification of metal fluxes depends on the receptor type and effect that is addressed (see Table 1).

The internal metal cycling within an ecosystem is ignored, since its influence on critical load results is relatively small, at least for Pb, Cd, Hg (De Vries et al. 2005). Because weathering of Pb, Cd, Hg causes only a minor input flux to topsoils, while uncertainties of such calculations are high, this flux was also neglected. Re-emission (volatilization) of deposited Hg is ignored, because this flux is already treated as part of the atmospheric net deposition in the modelling by EMEP Meteorological Synthesizing Centre East (Ryaboshapko et al. 1999, Ilyin et al. 2001). In consequence the critical load of a Pb, Cd, Hg equals the sum of tolerable outputs from the considered system in terms of net metal uptake and metal leaching.

The metal net uptake (g ha-1 a-1) is calculated from the annual yield (removal or increment) of biomass (kg ha-1 a-1) times a metal concentration (g kg-1) in harvestable parts of plants. Information on annual yields can be obtained from agricultural statistics and forest growth tables. It can also be modelled by relating yields to site characteristics as soil quality, climate and land use. The site specific share of different crops on agricultural land has to be considered. There is hardly any close relationship between metal contents in soil or soil solution and metal concentrations in harvestable parts of plants (an exception is Cd in wheat grains). Therefore, in general the metal uptake is calculated independently from metal concentrations in soils using medians of measured metal concentrations from relatively unpolluted areas. These median

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values in general neither exceed limits for food and fodder nor phytotoxic limits. The related uptakes are therefore considered to be tolerable. However, a good relationship of Cd in wheat grains to Cd in the soil solution exists (R2 = 0.62 according to Roemkens et al. 2004). This relationship was used in the calculation of critical loads of Cd to protect food quality. In these calculations, which were only performed for arable land, it was assumed that hundred percent of the crop rotation is wheat.

The tolerable leaching flux of a metal (g ha-1 a-1) is calculated from the flux of drainage water leaching from the considered soil layer (m3 ha-1 a-1) times a critical total concentration of the metal in the soil drainage water (g m-3). The flux of drainage water can be calculated from data on climate, land cover and soil type of the site. Only the critical limits for drinking water protection are provided as total concentrations and can thus directly be used as critical concentration in soil drainage water. Transfer functions have to be used to relate i) the critical limit for Cd in wheat, or ii) the critical limit of Hg in humus layers of forests to a respective critical total concentration in soil drainage water. With respect to the ecotoxicological based critical limits the chemical speciation model WHAM (Tipping 1994, 1998) was applied to calculate critical total concentrations of Cd and Pb in soil drainage water from site specific (pH dependent) critical concentrations of free metal ions. The main principles and the choice of input parameters of this specific WHAM model application (pH value, organic matter content of the soil, the concentrations of dissolved organic carbon and suspended particulate matter in soil solution, partial pressure of CO2) are explained in the Modelling and Mapping Manual (2004).

2.3 Calculation of tolerable metal fluxes in aquatic ecosystemsAs with terrestrial ecosystems, the critical load of Cd and Pb for freshwaters is the acceptable total load of heavy metals inputs corresponding to the sum of tolerable outputs by harvest within, and outflow from a catchment. The uptake into harvestable parts of plants is calculated in analogy to terrestrial ecosystems, while harvest in this context means in general the harvest of wood in forested catchments. In the calculation of the critical outflow of metals the lateral water flux off the catchment is multiplied with the critical total concentration in the surface water. The latter is derived from the critical limit considering the binding of metals to suspended particulate matter (SPM) and the concentration of SPM in the surface water. In calculations for lakes also net retention of metals can be considered. For a more detailed description of methods see Modelling and Mapping Manual (2004).

3 Results of critical load of Cd, Pb, Hg and exceedances for Europe

3.1 Critical loads

Altogether, eighteen countries submitted critical loads of heavy metals to the CCE. Critical loads of cadmium, lead and mercury were computed by 17, 17 and 10 countries, respectively. However not all countries addressed all effects, as is shown in Table 2.

The twenty-first Task Force meeting on modelling and mapping recommended to separate the protection against adverse health effects (effects 1, 2) from the protection against adverse ecosystem effects (effects 3 and 4). The result is shown in Figure 1 that gives the maps of critical loads of cadmium (top), lead (middle) and mercury (bottom)

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that will protect 95% of the receptor area within one 50 x 50 km EMEP grid cell against adverse effects on human health (left) and on ecosystems (right). Within these two categories of critical load maps the minimum critical load was used, if for a site critical loads were calculated for more than one effect. White areas indicate parts of Europe for which the respective critical loads have not been submitted by NFCs.

Table 2. Overview of the country response on the call for critical loads of cadmium, lead and mercury and the five effects (for meaning of effects numbers see table 1)Source: CCE (2005)

Country Effect number (see Table 1)Cd Pb Hg

1 2 3 4 1 3 4 1 3 5Austria AT x x x x x x xBelarus BY x xBelgium BE x x x x x x x x xBulgaria BG x xCyprus CY x x x x x xCzech Republic CZ x x xFinland FI xFrance FR x xGermany DE x x x x x x xItaly IT x xNetherlands NL x x x x xPoland PL x x xRussia RU x x x xSlovakia SK x x xSweden SE x x x x xSwitzerland CH x x x x xUkraine UA x xUnited Kingdom GB x xTotal 18 10 5 14 1 10 14 1 5 7 3

Figure 1 shows that ecosystem effects prevail as endpoint for critical loads that have been submitted. The maps are dominated by critical loads for terrestrial ecosystems (for a detailed description of the European critical loads database see Slootweg et al. 2005). The critical loads values may vary across political borders in dependence of the effects and receptor ecosystems chosen by the countries, but also due to differences in the national databases for input parameters. Such cross border variability can also be caused by changes in environmental conditions, if political borders follow natural borders (e.g. mountains, rivers).

Critical loads for human health and ecosystem protection are mapped using value ranges that are the same for each metal. This enables metal specific comparison between maps on the left with those on the right (Figure 1) of areas where both endpoints are mapped. It shows that effects on human health may occur at lower depositions of cadmium than ecosystem effects. For Pb and Hg in most - but not all areas for which both types of critical loads are mapped – ecosystem health related critical loads are more sensitive than those to protect human health.

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Figure 1. Maps of critical loads of cadmium (top), lead (middle) and mercury (bottom) that will protect 95% of the ecosystems against adverse effects on human health (left) and on ecosystems (right). Red shadings correspond to the lowest range of critical load values (source: CCE, 2005).

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The 5-percentile of all submitted European critical loads data of Cd related to human health effects is 1.9 g ha-1 a-1 or 2.4 g ha-1 a-1 for drinking water protection or food protection, respectively, and high values are in excess of 10 g ha -1 a-1. Critical loads to protect terrestrial ecosystem functions vary between < 1 and > 10 g ha -1 a-1 over Europe with a 5-percentile of 2.1 g ha-1 a-1.

The 5-percentile of all submitted critical loads data of Pb is 14 g ha -1 a-1 (human health effects) and 5 g ha-1 a-1 (ecosystem functioning). High values for both critical load types exceed 50 g ha-1 a-1.

Most critical loads related to human health effects of Hg deposition (effect 1) are higher than 1 g ha-1 a-1. The 5-percentile is 0.59 g ha-1 a-1. Critical loads for forest ecosystem function protection range between 0.1 g ha-1 a-1 and 0.4 g ha-1 a-1 for most areas.

For critical concentrations of Hg in precipitation (effect 5) only Belgium (Walloon), Sweden and Finland have submitted data that were reported in CCE (2005). The most sensitive area is in southern Sweden. There, the critical Hg concentration in precipi-tation was 0.5 to 1.0 ng L-1 for most EMEP grids, but even lower for some (Figure 2).

Figure 2. The 5th percentile EMEP50 grid values of the critical rainfall concentration for effect 5 (Source: CCE, 2005)

Representative effect-based information on the risk of air pollution impacts (exceedances) on a pan-European scale requires a broad transboundary coverage of critical loads. A European background database containing relevant European forest soil information was used in the past for countries that never submitted critical loads data. In the case of acidification and eutrophication critical load maps, based on national contributions and background data have been approved by the Working Group on Effects since 1994. These were then used in support of the assessment of areas at risk during negotiations of the 1994 Oslo Protocol and the 1999 Gothenburg Protocol. This historic practice regarding the use of the European background database has also been the guiding principle to compile a European background data base of critical loads of heavy metals. Figure 3 illustrates critical loads maps of Cd and Pb based on the European background database.

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Figure 3. The 5th percentile map of critical loads (effect 3) for forest soils of cadmium (top left), lead (top right) and mercury illustrating the possible use of the CCE European background database (source: CCE, 2005).

According to critical loads maps derived from the European background database areas that are sensitive for cadmium (top left) are generally located in the south of Europe. Areas that are most sensitive to lead (top right) are in the north, while the risk of mercury (bottom left) turns out to be widespread with particularly sensitive areas in the north of Europe. The Task Force recommended using the European background database for the assessment of exceedances in countries that did not submit data and do not object to this. Since the EU background database includes only information on forest ecosystems, maps of critical loads of Cd, Pb and Hg based on this can be used to complete European maps of critical loads protecting ecosystem health (effect 3).

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3.2 Exceedance of critical loads

The term exceedance in this section refers to the “average accumulated exceedance” (AAE). The AAE is the area-weighted average of the deposition value – computed by EMEP for each EMEP 50x50 km2 grid cell - minus critical loads that have been assessed in that grid cell. Thus, not only the exceedance of the most sensitive ecosystem is computed. An AAE may be computed for all ecosystem categories within a grid cell, but also for one single ecosystem category (such as a forest) in a grid cell for which data points are submitted by a National Focal Center. If the AAE computed for an EMEP grid cell is positive, the ecosystem area in that grid cell is said to be at risk.

Table 3 gives an overview of the percentage of national receptor areas that are at risk of human health effects (effects 1 and 2) in countries that submitted critical loads of cadmium, lead and/or mercury. Table 4 gives analogous information on the percentage of national ecosystems at risk of ecotoxicological effects (effects 3 and 4).

Tables 3 and 4 show that the risks of effects of Pb are more widespread than those of Cd. The data also reveal that the area of excess deposition of Pb in 2000 is strongly reduced in comparison to 1990. This holds to a lesser extent also for Cd and Hg.

In Europe 33.8 % of the ecosystem area in 1990 is subjected to excess deposition of Pb for human health effects, which is reduced to 8.3% in 2000 (see Table 3). The risk for ecosystem effects of Pb in 1990 and 2000 is 65.7% and 28.7% respectively (see Table 4).

For Cd 2.7% of the European receptor area is at risk of human health effects in 1990. Areas at risk of human health effects turn out to cover 0.8% using 2000 deposition data. In 1990 about 0.1% of total ecosystem area (2,668,646 km2) for which ecotoxicological based critical loads of Cd are submitted are at risk. In 2000 this percentage converges to 0%.

14.2 % of the European receptor area for which human health relevant Hg critical load data were submitted is at risk in 1990. The area at risk turn out to cover 3,9% using 2000 deposition data. In 1990 about 77.4% of total ecosystem area for which Hg-critical loads data relevant for ecosystem functionality are submitted are at risk. In 2000 this percentage covers 51%.

The critical mercury concentration in precipitation was exceeded in 1990 in almost all grid cells for which they have been computed. This risk hardly diminished by 2000 as can be seen from the average accumulated concentrations that are mapped in Figure 4. The average accumulated concentration is computed in a way similar to the derivation of the Average Accumulated Exceedance (AAE) described hereabove.

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Table3. Percentages of national ecosystem areas that are at long-term risk of human health effects in countries that submitted critical loads of cadmium, lead and/or mercury (Source: CCE, 2005)

Country

Cadmium (Cd) Lead (Pb) Mercury (Hg)

Eco area

(km2)

1990at

risk (%)

2000at

risk (%)

Eco area(km2)

1990at

risk (%)

2000at

risk (%)

Eco area

(km2)

1990at

risk (%)

2000at

risk (%)

AT 61,371 0.0 0.0 61,371 24.0 0.0 61,371 0.0 0.0BE 5,228 0.0 0.0 5,228 62.3 18.2 5,228 22.7 6.1BG 48,330 42.0 14.8 48,330 99.9 77.2 - - -CH 2,200 0.0 0.0 2,218 72.0 2.3 - - -CY 7,973 1.3 0.8 7,973 74.1 70.4 7,973 4.2 4.1CZ 25,136 1.1 0.5 25,136 93.1 19.9 25,136 7.4 1.9DE 290,003 1.4 0.1 290,003 79.0 7.4 290,003 17.9 4.8NL 19,471 0.1 0.0 19,471 89.2 0.1 - - -RU 425,425 0.0 0.0 650,575 3.3 2.5 - - -SE 22,050 0.0 0.0 - - - - - -UA 18,002 0.0 0.0 18,002 91.6 41.4 - - -

EU25 431,232 1.1 0.1 409,182 71.8 8.1 389,711 14.2 3.9Europe 925,190 2.7 0.8 1,128,308 33.8 8.3 389,711 14.2 3.9

Table 4. Percentages of national ecosystem areas that are at long-term risk of ecosystem effects in countries that submitted critical loads of cadmium, lead and/or mercury (Source: CCE, 2005)

Country

Cadmium (Cd) Lead (Pb) Mercury (Hg)

Eco area(km2)

1990at

risk (%)

2000at

risk (%)

Eco area(km2)

1990at

risk (%)

2000at

risk (%)

Eco area

(km2)

1990at

risk (%)

2000at

risk (%)

AT 61,371 0.0 0.0 61,371 48.7 11.1 32,601 39.2 11.7BE 5,237 0.0 0.0 5,237 63.0 12.8 5,228 100.0 83.5BY 121,128 9.1 0.1 121,128 100.0 10.2 - - -CH 9,411 0.0 0.0 9,393 99.0 24.1 11,611 80.2 44.4CY 7,973 0.0 0.0 7,973 80.9 78.4 - - -DE 290,003 0.1 0.0 290,003 83.8 9.0 99,866 97.0 59.8FR 170,638 0.1 0.0 170,638 93.7 9.8 - - -GB 50,075 0.5 0.0 50,075 25.9 6.0 - - -IT 278,128 0.0 0.0 278,128 0.3 0.0 - - -NL 22,314 0.0 0.0 22,314 98.4 21.5 - - -PL 88,383 0.5 0.0 88,383 73.5 14.7 88,383 100.0 99.9RU 1,393,300 1.1 0.2 1,194,125 70.8 51.0 - - -SE 151,432 0.0 0.0 151,432 60.5 1.9 152,074 56.0 22.9SK 19,253 2.6 1.1 19,253 52.3 22.6 19,253 99.0 65.3

EU25 1,144,807 0.1 0.0 1,144,807 56.3 7.4 397,405 77.4 51.2Europe 2,668,646 1.0 0.1 2,469,453 65.7 28.7 409,016 77.4 51.0

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Figure 4. The average accumulated concentration of Hg in 2000. White areas indicate ‘no data’.

Finally, CCE conducted a preliminary analysis of the risks from inputs of Cd and Pb from use of mineral and organic fertilizers in agricultural areas. These inputs have to be added to atmospheric deposition, if critical loads exceedance on agricultural land is addressed. Fertilisation alone causes critical loads of Cd to be exceeded in a limited number of grid cells, whereas for Pb exceedances are higher and occur more broadly.

3.3 Analysing long-term effects of exceedances

Critical loads are based on a steady-state concept, they are the constant depositions an ecosystem can tolerate in the long run, i.e. after it has equilibrated with these depositions.

However, many ecosystems are not in equilibrium with present or projected depositions, since there are processes (‘buffer mechanisms’) at work, which delay the reaching of an equilibrium (steady state) for years, decades or even centuries (damage delay time). By definition, critical loads do not provide any information on these time scales. Therefore, in its seventeenth session in December 1999, the Executive Body of the Convention ‘... underlined the importance of ... dynamic modelling of recovery’ (ECE/EB.AIR/68 p. 14, para. 51. b) to enable the assessment of time delays of recovery in regions where critical loads cease being exceeded and time delays of damage in regions where critical loads continue to be exceeded.

With critical loads, i.e. in the steady-state situation, only two cases can be distin-guished when comparing them to deposition: (1) the deposition is below critical load(s), i.e. does not exceed critical loads, and (2) the deposition is greater than critical load(s), i.e. there is critical load exceedance. In the first case there is no (apparent) problem, i.e. no reduction in deposition is deemed necessary. In the second case there is, by definition, an increased risk of damage to the ecosystem. Thus a

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critical load serves as a warning as long as there is exceedance, since it states that deposition should be reduced. However, it is often assumed that reducing deposition to (or below) critical loads immediately removes the risk of ‘harmful effects’, i.e. the chemical criterion that links the critical load to the (biological) effect(s), immediately attains a non-critical (‘safe’) value and that there is immediate biological recovery as well. But the reaction of soils, especially their solid phase, to changes in deposition is delayed by (finite) buffers. These buffer mechanisms can delay the attainment of a critical chemical parameter.

Actually, the current metal concentrations in environmental media can be higher or lower than the critical limits. They are increasing or decreasing in dependence on historical and present pollution as well as the metal buffer capacity of the site. Reasonable situations and future developments assuming future constant metal inputs and constant site characteristics are illustrated in Figure 5. The cases illustrated there lean on possible developments of a (soil) chemical and biological variable in response to a ‘typical’ temporal deposition pattern summarized in Posch et al. (2003) and used in the Modelling and Mapping Manual (2004, chapter 6).

Figure 5. Expected development of metal concentrations in environmental media in comparison to critical limits for different environmental situations

If current concentrations exceed critical limits, current outputs from the system are higher than the critical load, if equilibrium partitioning of the metal in the ecosystem is assumed. The more the concentration approaches the critical limit, the more the output approaches the critical load until a steady state at critical load is reached. In the case of current concentrations lower than critical limit, the opposite development would appear. Independent from the current situation, the steady state concentration will adjust at or below critical limit and therefore be safe in the long-term, provided the actual metal inputs to the system remain exactly at or below critical load.

Therefore, critical loads exceedance should be interpreted in any case as long-term risk for the considered system and effect. Non exceedance of critical load at present can be connected with a current risk, if the present concentration is above critical limit. Current exceedance does, however, not mean immediate damage if the current concentration is lower than critical limit. The damage delay time depends on the degree of pollution and the binding capacity of the site. It can partly take long time periods (decades up to hundreds of years). In the opposite case it would also take such

Environmental situations:1 The present concentration is below critical

limit, inputs are above critical load;2 The present concentration is above critical

limit, inputs are above critical load; 3 The present concentration is above critical

limit, inputs are exactly at critical load; 4 The present concentration is at critical

limit, inputs are and remain stable exactly at critical load (the theoretical critical load situation);

5 The present concentration is below critical limit, inputs are exactly at critical load

6 The present concentration is below critical limit, inputs are below critical load;

7 The present concentration is above critical limit, inputs are below critical load.

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long times to recover from heavy metal accumulation. Therefore the precautionary principle requires to avoid harmful metal accumulation in ecosystems.

Today’s exceedance of critical limits can be caused by historical anthropogenic pollution, but also by geogenic sources. The possibilities to distinguish anthropogenic pollution of soils from naturally high metal contents in a regional scale are rather limited. Since countries select the receptor ecosystems for which they submit critical loads, they can exclude areas with known natural pollution. Such areas can, however, also remain in the critical loads database, if it is felt that additional anthropogenic pollution should be limited for them.

Non exceedance of the critical load will in any case lead to a safe and sustainable situation in the future. Therefore critical loads and their exceedances are considered as an appropriate measure to evaluate sustainability of metal input scenarios.

3.4 Uncertainties

In this section, the emphasis of the review of the uncertainty in the assessment of the risk of heavy metal deposition is on modelled atmospheric deposition and critical loads maps for the computation of exceedances.

To calculate critical loads, mass fluxes (biomass, leaching water) are multiplied with measured or modelled concentrations that do not cause harmful effects in ecosystems. The mass fluxes of water and biomass are derived by NFCs on the base of official databases such as soil maps, meteorological data, land use maps and agricultural and forest statistics. These fluxes are also basic information in the critical loads calculations for acidity and nutrient nitrogen and have been proven to be robust in critical loads assessments in support of European air pollution abatement policies both under the Convention as well as under the European Commission. Similar robustness can be expected for metal concentrations in biomass taken from measurements in background areas and therefore for the whole uptake term in the critical loads calculation.

A sensitivity analysis of critical loads and exceedance calculations (ecotoxicological effects) according to methodologies described in the Modelling and Mapping Manual (2004) was conducted for a single example forest plot within a comprehensive research study on critical loads and dynamic modelling of heavy metals in the United Kingdom (Ashmore et al. 2004). Critical loads results and exceedances were highly sensitive to the parameters of the critical limit function. Other sensitive input parameters were runoff, dissolved organic carbon and hydrogen ion concentration. The results of the uncertainty analysis for a fixed critical limit showed that coefficients of variation for critical loads were between 40% and 60% for different metals. This confirms the results of an earlier analysis of uncertainty for the Netherlands (de Vries and Bakker, 1998).

The uncertainty of exceedance calculations at the current status of knowledge depends considerably more on national emission data that are required for the assessments of atmospheric deposition than on the uncertainty in critical loads.

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Critical loads:

Uncertainties of ecotoxicologically critical loads of Pb and Cd are mainly related to the elements in the derivation of critical limits in soil solution that determine critical leaching. These include the selection of species and appropriate endpoints in laboratory experiments to represent effects in complex ecosystems, limitations in laboratory methods (e.g. spiking effects), the duration of the experiments, and extrapolation of results obtained in laboratory to the field situation. The use of the free metal ion approach, while accounting best for the bioavailability of metals in soils, is also prone to uncertainty. This is due to the need of transformation from metal added in the ecotoxicological experiments to free metal ion concentration and - in a next step - to total concentration in the soil drainage water. This type of uncertainty could only be avoided if the ecotoxicological effects could be directly measured for these categories of dissolved metal concentrations (Ashmore et al. 2004). This is, however, not the state of knowledge, and such data are not expected to become available to a sufficient extent within the next years.

In the critical loads approach no “assessment factors” are used to address uncertainties of critical limits. Such factors could lower the critical limit values as recommended in the OECD methodology, or enhance them, if it is probable that laboratory tests overestimate effects compared to the field situation. The scientific basis of such factors is, however, rather weak.

The critical loads approach for heavy metals, including methods to derive critical limits, is the result of careful review of internationally accepted methods. Its scientific basis is sound and represents the state of the art. Remaining uncertainties can be reduced as more or better information becomes available.

Atmospheric deposition:

EMEP Meteorological Synthesizing Centre East had emphasized that official national emission reports used to model total net depositions of Cd, Pb, Hg were highly uncertain (Slootweg et al. 2005). Therefore the robustness of deposition data used to calculate exceedances of critical loads in Slootweg et al. (2005) and CCE (2005) could not yet well be established. EMEP Meteorological Synthesizing Centre East has remarked that official national emissions are too low to explain the magnitude of measured depositions. This implies that exceedances that are computed on the basis of official emissions may under estimate the actual risk of heavy metal deposition.

4 Summary and conclusions

Critical loads for cadmium, lead and mercury were successfully computed and mapped by 18 Parties of the LRTAP Convention. Critical loads for cadmium, lead and mercury were computed by 17, 17 and 10 countries, respectively. The methodology that was recommended in the call for data was carefully reviewed and documented in the Modelling and Mapping Manual (2004). The methodology enabled the assessment of ecosystem specific critical loads to protect human as well as environmental health.

These critical loads were compared to preliminary computations of ecosystem specific depositions computed by EMEP Meteorological Synthesizing Centre East of the three heavy metals in 1990 and 2000. The robustness of deposition results can not yet well be established due to the uncertainty of reported emissions. Bearing these

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uncertainties in mind, it is shown that atmospheric deposition of cadmium does not cause widespread risk in 2000, that the risk of lead deposition decreases since 1990 but is still widespread in 2000 and that the risk caused by mercury remains important without much change throughout the years in most of the countries that provided data on mercury. For Pb and Hg in most countries reductions of depositions sufficient to protect human health would be insufficient to protect ecosystems.

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