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Global Mercury Assessment

CHAPTER  6     Sources and cycling of mercury to the global environment

6.1         Overview

420.       The releases of mercury to the biosphere can be grouped in four categories:

• Natural sources - releases due to natural mobilisation of naturally occurring mercury from the Earth's crust, such as volcanic activity and weathering of rocks;

• Current anthropogenic (associated with human activity) releases from the mobilisation of mercury impurities in raw materials such as fossil fuels – particularly coal, and to a lesser extent gas and oil – and other extracted, treated and recycled minerals;

• Current anthropogenic releases resulting from mercury used intentionally in products and processes, due to releases during manufacturing, leaks, disposal or incineration of spent products or other releases;

• Re-mobilisation of historic anthropogenic mercury releases previously deposited in soils, sediments, water bodies, landfills and waste/tailings piles.

421.       Figure 6.1 shows these release categories with main types of possible control mechanisms.

Figure 6.1    Categorisation of sources of mercury releases to the environment and main control
                   options.

422.       The recipients of mercury releases to the environment include the atmosphere, water environments (aquatic) and soil environments (terrestrial). There are continuing interactions – fluxes of mercury – between these compartments.  These are described in section 6.4 on pathways of mercury to – and in – the environment. The speciation – the chemical form – of the released mercury varies depending on the source types and other factors as described in chapter 2. This also influences the impacts on human health and environment as different mercury species have different toxicity.

423.       Given the understanding of the global mercury cycle, current releases add to the global pool of mercury in the biosphere – mercury that is continuously mobilised, deposited on land and water surfaces, and re-mobilised. Being an element, mercury is persistent – it cannot be broken down to less toxic substances in the environment. The only long-term sinks for removal of mercury from the biosphere are deep-sea sediments and, to a certain extent, controlled landfills, in cases, where the mercury is physio-chemically immobilised and remains undisturbed by anthropogenic or natural activity (climatic and geological). This also implies that even as the anthropogenic releases of mercury are gradually eliminated, decreases in some mercury concentrations – and related environmental improvements – will occur only slowly, most likely over several decades or longer.  However, improvements may occur more quickly in specific locations or regions that are largely impacted by local or regional sources.

Local releases – global effects

424.       The origins of atmospheric mercury deposition (flow of mercury from air to land and oceans) are local and regional as well as hemispherical or global. Several large studies have supported the conclusion that, in addition to local sources (such as chlor-alkali production, coal combustion and waste incineration facilities), the general background concentration of mercury in the global atmosphere contributes significantly to the mercury burden at most locations. Similarly, virtually any local source contributes to the background concentration – the global mercury pool in the biosphere - much of which represents anthropogenic releases accumulated over the decades (see for example US EPA, 1997; Munthe et al., 2001). Also, the ocean currents are media for long-range mercury transport, and the oceans are important dynamic sinks of mercury in the global cycle.

425.       The majority of atmospheric anthropogenic emissions are released as gaseous elemental mercury. This is capable of being transported over very long distances with the air masses. The remaining part of air emissions are in the form of gaseous divalent compounds (such as HgCl₂) or bound to particles present in the emission gas. These species have a shorter atmospheric lifetime than elemental vapour and will deposit via wet or dry processes within roughly 100 to 1000 kilometres. However, significant conversion between mercury species may occur during atmospheric transport, which will affect the transport distance.

426.       The atmospheric residence time[1] of elemental mercury is in the range of months to roughly one year. This makes transport on a hemispherical scale possible and emissions in any continent can thus contribute to the deposition in other continents. For example, according to the modelling of the intercontinental mercury transport performed by EMEP/MSC-E (Travnikov and Ryaboshapko, 2002) up to 50 percent of anthropogenic mercury deposited to North America is from external sources. Similarly, contributions of external sources to anthropogenic mercury depositions to Europe and Asia were estimated to be about 20 percent and 15 percent, respectively. 

427.       Furthermore, as mentioned, mercury is also capable of re-emissions from water and soil surfaces. This process greatly enhances the overall residence time of mercury in the environment. Recent findings by Lindberg et al. (2001) indicate re-emission rates of approximately 20 percent over a two-year period, based on stable mercury isotope measurements in north-western Ontario, Canada.

Anthropogenic sources of mercury releases

428.       A large portion of the mercury present in the atmosphere today is the result of many years of releases due to anthropogenic activities. The natural component of the total atmospheric burden is difficult to estimate, although a recent study (Munthe et al., 2001) has suggested that anthropogenic activities have increased the overall levels of mercury in the atmosphere by roughly a factor of 3.

429.       While there are some natural emissions of mercury from the earth’s crust, anthropogenic sources are the major contributors to releases of mercury to the atmosphere, water and soil.

430.       There are significant uncertainties in the available release inventories, not only by source, but also by country.  Nonetheless, the best available estimates of mercury emissions to air from various significant sources are shown in table 6.1 below.

431.       The emissions from stationary combustion of fossil fuels (especially coal) and incineration of waste materials accounts for approximately 70 percent of the total quantified atmospheric emissions from significant anthropogenic sources. As combustion of fossil fuels is increasing in order to meet the growing energy demands of both developing and developed nations, mercury emissions can be expected to increase accordingly in the absence of the deployment of control technologies or the use of alternative energy sources. Control technologies have been developed for coal combustion plants and waste incinerators with the primary intention of addressing acidifying substances (especially SO₂ and NO_X), and particulate matter (PM). Such existing technologies may provide some level of mercury control, but when viewed at the global level, currently these controls result in only a small reduction of mercury from these sources. Many control technologies are significantly less effective at reducing emissions of elemental mercury compared to other forms. Optimised technologies for mercury control are being developed and demonstrated, but are not yet commercially deployed.

432.       Available global estimates of atmospheric emissions from waste incineration, as well as other releases originating from intentional uses of mercury in processes and products, are deemed underestimated and to some degree incomplete. However, recorded virgin mercury production has been decreasing from about 6000 to about 2000 metric tons per year during the last two decades, and consequently, related releases from mining and usage of mercury may also be declining.

433.       Anthropogenic emissions from a number of major sources have decreased during the last decade in North America and Europe due to reduction efforts. Also, total anthropogenic emissions to air have been declining in some developed countries in the last decade.  For example, Canadian emissions were reduced from about 33 metric tons to 6 metric tons between 1990 and 2000 (Canadian comments, comm-20-gov; Canadian submission, sub42gov).

Table 6.1     Estimates of global atmospheric releases of mercury from a number of significant
                  anthropogenic sources in 1995 (metric tons/year). Releases to other media are not 
                  accounted for here. *1

Notes

 1   Releases to aquatic and terrestrial environments, as well as atmospheric releases from a number of other sources, 
are not included in the table, as no recent global estimates are available. See chapter 6 for description of issue.

2    Considered underestimated by authors of the inventory, see notes to table 6.10.

3    Represents total of the sources mentioned in this table, not all known sources. Sums are rounded and may therefore
not sum up precisely.

4    Estimated emissions from artisanal gold mining refer to late 1980's/early 1990's situation. A newer reference 
(MMSD, 2002) indicates that mercury consumption for artisanal gold mining - and thereby most likely also mercury 
releases - may be even higher than presented here.

5    Production of non-ferrous metals releasing mercury, including mercury, zinc, gold, lead, copper, nickel.

Natural sources of mercury release

434.       Natural sources include volcanoes, evaporation from soil and water surfaces, degradation of minerals and forest fires.  The natural mercury emissions are beyond our control, and must be considered part of our local and global living environment. It is necessary to keep this source in mind, however, as it does contribute to the environmental levels.  In some areas of the world, the mercury concentrations in the Earth's crust are naturally elevated, and contribute to elevated local and regional mercury concentrations in those areas.

435.       Today’s emissions of mercury from soil and water surfaces are composed of both natural sources and re-emission of previous deposition of mercury from both anthropogenic and natural sources. This makes it very difficult to determine the actual natural mercury emissions. For global estimates of natural emissions, see section 6.3.6.

436.       Published estimates of natural versus anthropogenic mercury emissions show significant variation, although more recent efforts have emphasized the importance of human contributions (see for example Fitzgerald et al. (1998), Jackson (1997) and Lamborg et al. (2002)). Attempts to directly measure natural emissions are ongoing (see for example Coolbaugh et al., 2002).  Nonetheless, available information indicates that natural sources account for less than 50 percent of the total releases.

437.       On average around the globe, there are indications that anthropogenic emissions of mercury have resulted in deposition rates today that are 1.5 to 3 times higher than those during pre-industrial times. In and around industrial areas the deposition rates have increased by 2 to 10 times during the last 200 years (Lindquist et al., 1984; Bergan et al., 1999; see also section 6.4 on pathways).

Contributions from intentional uses versus impurities in high volume materials

438.       Regarding anthropogenic releases, the relative importance of intentional uses versus mobilisation of mercury impurities varies between countries and regions, particularly depending on:

• State of substitution of intentional uses (products and processes);

• Reliance on fossil fuels for energy production, particularly coal, and the presence of controls for other pollutants, which also reduce mercury emissions;

• Extent of mining and mineral extraction industry;

• Waste disposal pattern – incineration/landfilling;

• State of implementation of release control technologies in power production, waste incineration and various industrial processes.

439.       For a number of countries described in the submission from the Nordic Council of Ministers (sub84gov), estimated contributions of intentional uses vary between 10 and 80 percent of the total domestic emissions to air, depending on the influence of the factors listed above in each country. Rough estimates by main anthropogenic source types in each of these countries are shown in section 6.3.2.

440.       As an illustration, figure 6.2 shows the overall turnover of mercury in the Danish society in 1992/93. Denmark is a quite small country with relatively accurate monitoring of the flows of products and waste in the economy and the environment. Therefore, it has been possible to perform rather detailed balances, so-called substance flow assessments for mercury, which provide useful information on the contributions from different sectors to the total mercury burden in the society and the environment. As shown in the figure, the majority of the input – more than two thirds – originated from intentional uses (chlor-alkali production and products), and the contributions from intentional uses to releases to air in 1992/93[2] could roughly be estimated at 50-80 percent of the total releases to air from Denmark (submission from the Nordic Council of Ministers, sub84-gov). It should be noted that primary mineral extraction and processing is not as large a sector in Denmark, as in many other countries.

Figure 6.2    The turnover of mercury in the Danish society in 1992/93, kilograms mercury/year (based
                   on Maag et al., 1996). Please note that inputs and outputs do not balance because
                   outputs reflect higher inputs from previous years. Net change in stocks was negative.

441.       Examples of national distributions of anthropogenic mercury releases from different individual source types are given in section 6.3.4.  In countries where mercury mining or intentional use of mercury for small-scale gold mining is taking place, these sources can be significant (see for example Colombian submission, sub14gov).

442.       Parts of the descriptive text in this chapter were based on the submission from the Nordic Council of Ministers (sub84gov) and to a lesser extent Pirrone et al. (2001).

6.2         Natural sources of mercury

443.       Natural sources include volcanoes, evaporation from soil and water surfaces, degradation of minerals and forest fires. Mercury in small, but varying concentrations can be found virtually in all geological media. Elemental and some forms of oxidized mercury are permanently coming to the atmosphere due to their volatility. High temperature in the Earth mantle results in high mercury mobility and mercury continuously diffuses to the surface. In the zones of deep geological fractures these processes go on more intensively. Here are located so-called mercury geochemical belts where mercury concentrations in the upper layer appreciably exceed their average values. In some parts of mercury belts the intensive accumulation of mercury resulted in the formation of (extractable) deposits (Jonasson and Boyle, 1971; Bailey et al., 1973). Regions with high concentrations in surface rocks are characterized by high mercury emissions to the atmosphere.

444.       Today’s releases of mercury from soil and water surfaces are, however, not only natural, but are significantly influenced by previous deposition of mercury from anthropogenic sources. This makes it extremely difficult to determine the actual natural mercury emissions. For example, total estimates of re-emission from soil and water surfaces in Europe exist, but they include mercury originating from both natural and anthropogenic sources (Pirrone et al., 2001).

445.       A considerable portion of the mercury emissions from forest fires may also be re-emitted anthropogenic mercury (USA; comm-24-gov).

446.       A number of attempts have been made to estimate the regional and global natural emissions of mercury. It is, however, difficult to do so with any precision and research is still done in this field at several institutions (AMAP, 2000).

447.       Understanding of the global mercury cycle, shown schematically in figure 6.3, has improved significantly with continuing study of source releases, mercury fluxes to the earth's surface, and the magnitude of mercury reservoirs that have accumulated in soils, watersheds and ocean waters.  Although considerable uncertainty still exists, it has become increasingly evident that anthropogenic emissions of mercury to the air rival or exceed natural inputs. Recent estimates place the annual amounts of mercury released into the air by human activities at between 50 and 75 percent of the total yearly input to the atmosphere from all sources (US EPA, 1997).

448.       Mason et al. (1994) estimated the global natural emissions at about 1650 metric tons/year. In an update performed by Lamborg et al. (2002) it was estimated at 1400 metric tons/year (as illustrated in figure 6.3). MSC-E and EMEP (comm-4-ngo) quote Bergan and Rohde (2001) for an estimated global natural emission of about 2400 metric tons, of which 1320 was emitted from land and 1100 was emitted from oceans.

Figure 6.3     Comparison of estimated pre-industrial and current mercury budgets and fluxes. All
                    fluxes (arrows) and pools (in frames) in metric tons (adapted from Lamborg et al. (2002);
                    the original authors note that the cycle is seen as unsteady.)

6.3         Anthropogenic sources of mercury

6.3.1      Mobilisation of mercury impurities in materials

449.       Mercury is naturally present in coal and other fossil fuels, as well as in minerals like lime for cement production and soils (such as agricultural soils subject to acidification management) and metal ores including for example zinc-, copper- and gold ore. Coal-fired power production is today deemed the single largest global source of atmospheric mercury emissions (Pacyna and Pacyna, 2000). This is due to the increasing global power consumption, and also to the fact that emissions from intentional use of mercury are gradually diminishing in many of the industrialised countries.

450.       As an example, China reports the following regarding the emissions of mercury from coal combustion in the country:

“According to information from research, the average mercury content of coal is 0.038 – 0.32 mg/kg. The total amount of mercury emissions from coal combustion was about 296-302.9 metric tons annually in the middle of the 1990’s, including 213.8 metric tons in the atmosphere and 89.07 metric tons in ash and cinder. The average content of organic mercury in coals collected from 15 provinces and cities was 0.037 mg/kg, occupying 18.1 percent of the mercury. The average contents of organic mercury in fly ashes of burning coal 0.045 mg/kg, occupying 28.1 percent of total mercury in ash. From 1978 through 1995, mercury emission had been increasing at average 4.8 percent per year.” (Comments from China, comm-19-gov).

451.       When relating this information to estimates of global mercury emissions from major quantified sources (approx. 2100 metric tons/year in 1990), as shown in table 6.11, it is clear that the importance of emissions from coal combustion is significant.

452.       Mercury impurities in primary and recycled materials constitute major contributions to the total global mercury burden. Measures to reduce these releases are described in chapter 8.

453.       Processing of secondary raw materials, like iron and steel for example, can also be a significant source of mercury releases, and emission control technologies are often necessary. In this case the origin of the mercury may be both natural impurities and a result of intentional use of mercury in products/components in iron/steel scrap (switches, air-bag activators etc.)

454.       Many industrialised countries have legislative structures for emission control covering mercury in place today. The national submissions to UNEP for this assessment indicate that the situation might be different for a number of countries with other conditions, see chapter 9 and the separate Appendix to this report – Overview of existing and future national actions, including legislation, relevant to mercury.

6.3.2      Releases from intentional use of mercury in products and processes

455.       As described in chapter 7, mercury is used in many products and industrial processes. Despite decreasing consumption in many industrialised countries during the last two decades, the intentional use of mercury in products and processes is still deemed a significant source of mercury to the environment. The recorded global primary production of virgin mercury is still large compared to current estimates of global atmospheric mercury emissions.

456.       When assessing the releases of mercury to the environment, it is generally difficult to quantify diffuse releases from the life cycle of mercury-containing products. These sources have not always been included fully in regional or global inventories for mercury releases to the environment. Some national studies do however give a certain insight in the contributions from this source category (see below).

457.       The releases of mercury from waste treatment and storage can be very difficult to assess, but national balances (“substance flow assessments”) can cast light on some of the aspects needed. Such substance flow assessments have been performed with varying detail in for example the Netherlands, USA and Denmark.

458.       Also, some research performed in the US indicates that releases from products via normal use, spills, breakage, scrap metal processing and disposal are significant sources that may be under-estimated in some release inventories (USA; comm-24-gov).

459.       Much of the mercury brought into use with products and for consumer purposes will be incinerated or end up in landfills with collected waste. In many parts of the world it may be lost, dumped or incinerated diffusely and informally directly in the environment. Under present circumstances, therefore, a significant part of the total consumption of mercury is expected to end up in the environment rather directly and quickly. How much this amounts to on a global level has not been seriously estimated. As indicated in section 6.4 on pathways, some of the mercury used, collected and treated under more controlled conditions may also be spread to the environment over a longer period of time.

460.       Examples of quantified contributions from different intentional uses to national mercury releases are given in section 6.3.4 below.

Relative importance to air emissions of intentional uses versus impurities

461.       When considering the abatement measures to choose, it is relevant to examine the relative importance of intentional mercury uses versus mercury impurities in materials. This poses some difficulties, as not all flows are registered in detail, and particularly as the origin of mercury in the waste flow cannot always be allocated with precision. Nevertheless, it is possible to form rough impressions of this relationship in individual countries, notably where substance flow assessments (SFAs) on mercury have been performed. In table 6.2, rough estimates of the distribution of air emissions on main source categories are presented on the basis of the data presented in this section. In the table, note that releases to water and soil are not included, but may be of significance. Landfilled amounts are not included, except for some recorded air emissions from landfills. Additionally, inventory methodology may vary among the countries and not all emissions are necessarily recorded in all cases.

462.       The contribution from intentional mercury uses in a number of products in the European region was also assessed by Munthe and Kindbom (1997). They found that in the mid-1990's three dominating groups of intentional mercury uses in products[3] contributed about 18 percent of the total mercury emissions to air in this region. Additional contributions from dental amalgam use were not included in the assessment.

463.       See also the discussion of relative importance of main source categories, and factors influencing this distribution, in section 6.1.

Table 6.2    Rough estimates of relative importance of main source categories to recorded
                 anthropogenic emissions to air – examples (submission of the Nordic Council of Ministers,
                 sub84gov).

Note:       Data sources mentioned in table 6.4 in section 6.3.4.

464.       Table 6.3 from US EPA (1997) shows estimates/projections for US discards of mercury used intentionally in products. For comparison, the reported consumption of mercury with intentional uses in the USA was estimated at 711 metric tons in 1990 and 372 metric tons in 1996 (Sznopek and Goonan, 2000; corresponding to 784 and 410 short tons respectively). It should be noted that mercury contents in waste products reflect earlier, higher mercury consumption rates (see figure 9.2 in section 9.2.4 UNITED STATES).

465.       According to new information from the USA (comm-24-gov), table 6.3 may not account for the use of mercury switches in cars, or may undercount estimated discards in the USA.  The State of Maine's Department of Environmental Protection estimates an average of roughly 2/3 switch per car on the road today, at 0.8 grams of mercury per switch.  While it is true that these switches do not typically wind up in MSW disposal facilities per se, they are largely discards from households and are unaccounted for in typical product discard inventories such as table 6.3.

Table 6.3     Estimated discards of mercury used intentionally in products in municipal solid waste in
                   the USA as estimated/projected in 1992^a, unless noted (US EPA, 1997).

^Notes:

^a       US EPA, 1992 (except for fluorescent lamps estimates).

^b       Discards before recovery.

^d       New information from the USA (in comm-24-gov, 2002).

*       Since 1992 several States have restricted the mercury content of alkaline batteries and/or banned the sale of mercuric oxide batteries. Federal legislation to restrict mercury use in batteries went into effect in May 1996.  The battery industry has eliminated mercury as an intentional additive in alkaline batteries, except in button cells. Although no current estimate of mercury releases from batteries was available for these years, according to the National Electrical Manufacturers Association (NEMA), the entire USA battery industry used only approximately 6.6 tons of mercury in 1994 (NEMA, 1996).

1.3.3             Mobilisation of mercury due to changes in land use

466.       Under some conditions anthropogenic changes in land use may result in substantial mobilisation of mercury already present in the environment (originating from natural and/or anthropogenic sources). For example, in some environments, anthropogenic modifications including farmlands, recent clear-cuttings and water reservoirs (hydroelectric, aquaculture, irrigation) may considerably enhance the release of mercury to aquatic systems and the bio-accumulation of mercury in organisms. There is a growing body of evidence that the soils of forested watersheds contain considerable stores of both methylmercury and inorganic divalent forms. Both in North America and in Northern Europe, evidence is gradually accumulating, which points to the effect of terrain disturbance as a factor in the mobilisation and transport of both the inorganic and methylmercury stored in watersheds, and apparently also in the production of methylmercury. Investigations in connection with hydro-electric reservoirs revealed the importance of understanding transport phenomena involving flooded soils. Watershed-scale hydrology is emerging as an increasingly important explanatory variable (Canadian comments; comm-20-gov).

1.3.4             Examples of national mercury releases distributed on source types

467.       As mentioned above, the relative contributions to releases of mercury from different source types varies between countries depending on a number of factors. In order to illustrate possible loads from individual source types, examples of distributions of releases to air, water, soil and landfills are given for a number of countries in tables 6.4-6.7 below (aggregated in the submission from the Nordic Council of Ministers, sub84gov, except for Mexican data added here).

468.       Attention must be paid to the often significant differences between countries, which again are related to differences with respect to the sources that exist, different equipment or standards for cleaning operations like flue gas cleaning, as well as different year and methodology of investigation. Though a deeper investigation and description of the background and quality of the presented data would be useful, this has not been possible for this report, principally due to limitations in time and resources. This also implies that direct comparison between countries is not relevant.

469.       The countries mentioned in the following three tables were chosen as illustrative examples only. A number of other countries also regularly collect and publish release data and a number of data sets were submitted by other countries.  These can be found in national submissions and comments available on the Global Mercury Assessment webpage www.chem.unep.ch/mercury.

470.       An important source of mercury releases that is not covered in the tables given in this section, but which is occurring in an increasing number of countries, is the use of mercury for gold and silver extraction. Based on available - and probably incomplete - information, it has been estimated that the present annual global mercury input to the environment from gold mining alone may be upwards of 500 metric tons, two-thirds of it emitted to the atmosphere, with the other third going to soils and waters (Lacerda, 1997; MMSD, 2002). Similarly, in those countries where primary mercury is still mined, this mining may also represent a significant source of mercury releases.

471.       It should be noted that relatively few data are available regarding the releases of mercury on the total life cycle of oil and natural gas (from extraction to combustion or disposal). Both are consumed in large quantities globally. Additional research on this important question is ongoing in the USA and scheduled for reporting in 2003 (USA, comm-24-gov). 

472.       Thailand has described their efforts in the management of mercury releases to the aquatic environment from oil and gas extraction (Thai submission, sub53gov).

Emissions to the atmosphere

473.       Besides the sources mentioned above, important sources for atmospheric emissions include certain industrial activities, waste treatment and disposal, as well as combustion and fossil fuel energy generating processes in general.

474.       Combustion of waste is a major source of mercury releases to the environment. It should be kept in mind, that the source of this mercury is the mercury contents in the products constituting the waste, both in the form of intentionally used mercury and unintentional presence of mercury (either as a natural impurity or as an anthropogenic trace pollutant in the raw materials used).

475.       Concerning cremation, it may be noted that crematories are normally not equipped with flue gas cleaning facilities for removal of mercury. The emissions from cremation are primarily due to the use of mercury in amalgam for dental purposes.

Table 6.4     Mercury emissions to air - examples *1

Notes:

1       From US EPA (1997); OSPAR (2000); Maag et al. (1996); Norwegian Pollution Control Authority (2001); Finnish Environment Institute (1999); Mukherjee et al. (2000); Mexican information submission, and KEMI (1998). The presented distribution of sources, as intentional/unintentional, was made in the submission from the Nordic Council of Ministers (sub84gov), except for Mexican numbers. 

2       Covers incineration of municipal waste, medical waste, hazardous waste and sewage sludge.

3       Assumptions USA ~ 264 million capita; UK ~ 59 million capita; Denmark ~ 5.3 million capita; Norway ~ 4.4 million capita; Sweden ~ 8.5 million capita; Finland ~ 5.2 million capita; Mexico ~ 99 million capita.

4       In the reference (Mukherjee et al, 2000), emissions from manufacturing are aggregated and include both mercury from unintentional mobilisation and intentional uses. There are indications, however, that the first mentioned is the dominating source category from manufacturing. Total emissions from manufacturing are therefore mentioned under “Mobilised Hg impurities – Manufacturing”.

5       The relatively high figure for waste incineration in Denmark in 1992-93 was caused by the widespread use of incineration for treatment for municipal solid waste in the country. In 1993, around 78 percent of all municipal solid waste in Denmark was directed to incineration, and at that time only 86 percent of the incineration capacity was equipped with cleaning facilities for acidic flue gas cleaning (Maag et al, 1996).

6       Categorised in Mexican submission as follows: thermo-electric plants (0.13 metric tons/y), carbo-electric plants (0.79), industrial commercial boilers (0.095) and residential wood combustion (1.2 metric tons/y).

7       The USA (in comm-24-gov) provided updated information on national emissions from certain source categories and totals (1996 emissions in metric tons): Chlor-alkali: 9, lamps breakage: 1, dental preparation: 0.7, waste incineration: 74, landfills: 0.2, Cement: 4, pulp and paper: 2, coal boilers: 55 (uncertain), oil and natural gas: 1, and total quantified emissions for 1996: 170. For 1999 the total quantified emissions were estimated at 125 metric tons. 

8       According to OSPAR (2001b, as cited by Greenpeace), the chlor-alkali industry in the United Kingdom reported mercury releases of 1.4 metric tons in 1999.

Releases to water

476.       Concerning releases to water a dominant source in western countries will typically be outlets from municipal sewage treatment plants, as municipal wastewater may contain mercury originating from e.g. dentist clinics, miscellaneous measurement and monitoring equipment as well as laboratories (originating from intentional uses). In some countries, direct discharges of mercury-containing wastewater may be relatively larger. Also, several other sources of mercury releases to aquatic environments are not listed in the table below, and the quantities may not reflect the situation in countries with less developed controls. This may particularly be the case if a country has large industrial sectors applying mercury, such as chlor-alkali production with mercury-cell technology.

477.       The Norwegian data indicate that offshore oil activities may be a significant source of releases to the marine environment. A similar release may also likely take place into Danish waters (and possibly other locations), but has so far not been quantified.

Table  6.5    Mercury releases to water – examples *1

Notes:

1       From Maag et al. (1996), Norwegian Pollution Control Authority (2001) and KEMI (1998). Presented distribution of sources as intentional or unintentional was made in the submission from the Nordic Council of Ministers (sub84gov).

2       Assumptions: Denmark ~ 5.3 million capita; Norway ~ 4.4 million capita; Sweden ~ 8.5 million capita. 

Releases to soil

478.       Concerning releases to soil in the examples from the Nordic countries, the dominant sources seem to be:

• Cemeteries, which is primarily due to the use of mercury as amalgam for dental purposes; and

• Land application of sewage sludge from municipal wastewater treatment (originating from intentional uses as described).

Table 6.6    Mercury releases to soil - examples *1

Notes:        From Maag et al. (1996) and Norwegian Pollution Control Authority (2001).  Distribution of sources
as intentional or unintentional is the responsibility of the authors of this report solely. 
Assumptions: Denmark ~ 5.3 million capita; Norway ~ 4.4 mil capita.

6.3.5      Landfilling

479.       Besides the direct releases to the environment indicated above significant quantities of mercury are directed to landfills either as manufacturing waste, end-of-pipe technology waste or contained in products disposed of as municipal solid waste or hazardous waste.

480.       Information on mercury quantities landfilled in a number of countries is presented in table 6.7. The noted differences between countries may be explained by:

• Different activities, e.g. whether mining and metal extraction takes place in the country or not;

• Different environmental policies, e.g. mercury is extracted from zinc extraction residues in Finland, whereas in Norway such residues are directed to landfill;

• The sources included may differ between inventories.

Table 6.7     Mercury directed to landfills or collected as hazardous waste (aggregated in the 
                  submission from the Nordic Council of Ministers, sub84gov) *1.

Notes:

1       Original references: Sznopek & Goonan 2000, OSPAR 2000, Endre et al., 1999.

2       Assumptions USA ~ 264 million capita; UK ~ 59 million capita; Denmark ~ 5.3 million capita;
Finland ~ 4.8 million capita; Norway ~ 4.4 million capita; Sweden ~ 8.5 million capita.

3       Reference: Norwegian Pollution Control Authority (2001).

6.3.6      Global and regional release estimates

481.       The total global anthropogenic and natural releases of mercury are not known with high precision. Several attempts have been made, however, at quantifying these totals. Table 6.8 shows global totals as estimated by different authors. As can be seen, the numbers are relatively uncertain. This is reasonable given the complexity of the quantification. As mentioned in section 6.3.7, generally it has not been possible to include all significant contributions to global releases in the estimates. An impression of which source types may be included and which may often be omitted can be seen in table 6.9.

Table  6.8     Estimates of total releases of mercury to the global environment (table presented by 
                   OECD, 1994, with estimates by Mason et al. (1994), Pirrone et al. (1996) and Lamborg
                   et al. (2002) added here).

Notes:  1    Anthropogenic releases and totals: Numbers include an estimated re-emission (net increase of evasion from
oceans) of 1400 metric tons/year originating from previous anthropogenic releases (new anthropogenic releases are thus estimated at 4150 metric tons/year in this study). 

2    Anthropogenic releases and totals: Numbers include an estimated re-emission (net increase of evasion from oceans) of 400 metric tons/year originating from previous anthropogenic releases (new anthropogenic releases are thus estimated at 2600 metric tons/year in this study). 

Table 6.9     Estimates of Worldwide Releases of Mercury to the Atmosphere, Soil and Water in 1983
                  with quantified and omitted contributions stated (metric tons per year, from Nriagu and
                  Pacyna (1988) and Nriagu (1989) as presented by OECD (1994), presentation of
                  summation slightly corrected and question marks added here).

Notes: +? means that real totals may be larger, as inputs given as “no estimate” are not included in presented totals.

Notes in OECD (1994): 

*       Insignificant contributions to the atmosphere from: oil combustion, zinc-cadmium production, secondary non-ferrous production, steel and iron manufacturing, cement production, and mobile sources (eds. comment here: These input may actually be of interest, as described elsewhere in this chapter). 

**      Landfills included.

482.       A number of more recent release inventories have been performed. Generally, they only include major sources of atmospheric releases – mainly from mercury impurities in high volume materials and to a lesser extent from the lifecycles of mercury in intentional uses. The totals from these studies are presented in table 6.10 along with the totals from table 6.9.

Table 6.10    Newer estimates of atmospheric releases from some major anthropogenic sources, as
                   compared to totals from table 6.9  (metric tons/year).

Notes:   +? means that real totals may be larger, as inputs given as “no estimate” are not included in presented totals.

1       Includes also 172 metric tons of mercury releases from chlor-alkali production and other smaller sources (Pirrone et al., 2001).

2       Not including releases from gold extraction (has been estimated by Lacerda (1997) at up to 460 metric tons/year at about 1990, of which most was released to the atmosphere). Also not including releases from chlor-alkali production and "other sources". The authors of the inventory state that releases from waste incineration are most likely underestimated due to lack of national data on wastes (Pirrone et al., 2001).

3       The uncertainty on the total is significant – the authors mention that an estimation accuracy of less than 50 percent can be assigned for mercury in Europe (Pirrone et al., 2001). Most likely, the inaccuracy is higher for large parts of the world.

483.       In table 6.11, the results of the global atmospheric emission inventory for 1995 are presented by source types included in the quantification, and geographical continents (as presented in Pirrone et al., 2001). Here, the highest contribution within each source type has been put in bold text.

484.       Pirrone et al. (2001) comment the trends in the geographic distribution of emissions as follows:

“There have been major changes in emissions in 1995 compared to 1990, with respect to the location of major emission regions contributing the most to the global emission survey of the element. Whereas the mercury emissions in Europe and North America have decreased quite substantially during the period from 1990 through 1995, emissions in Asia, particularly in China and India, have increased significantly. The Asian sources contributed about 30 percent to the total emissions of mercury in 1990, compared to 56 percent in 1995. An increase of more than 250 metric tons was estimated for China between the years 1990 and 1995. The increase of mercury emissions in China from 1990 through 1995 is clearly related to the increase of coal combustion in the country. The mercury emission increase due to the increased combustion of coal has overcome a slight reduction of emission of air pollutants in the country due to the installation of high efficient emission control devices starting in the mid-1990's.

Decrease of mercury emissions in Western Europe, the United States, and Canada can be explained by further installation of emission control equipment, particularly various flue gas desulfurisation (FGD) technologies, as already mentioned. Relatively low temperatures found in wet scrubber systems allow many of the more volatile trace elements to condense from the vapour phase and thus to be removed from the flue gases.

Decrease of mercury emissions from combustion sources, as well as other industrial sources in Central and Eastern Europe from 1990 to 1995 was also caused by a general decrease of industrial activities and resulting decrease of the consumption of raw materials.”

Table 6.11    Estimates of global atmospheric emissions of mercury from a number of major
                   anthropogenic sources in 1995 (metric tons/year;Pirrone et al.,1996; 2001).

Notes:

1       Estimates of maximum values, which are regarded as close to the best estimate value by the authors of the inventory.  Totals represent total of the sources mentioned in this table, not all known sources.  

2       The total emission estimate for 1990 also includes 171.7 metric tons from chlor-alkali production and other “less significant” sources.

3       Considered underestimated by authors of the inventory, see notes to table 6.9 above.

4       Does not include gold extraction, chlor-alkali production and “other sources”, see notes to table 6.9 above.

485.       For more information about emission control technologies and efficiencies, see section 8.3.

486.       Geographical distribution of 1990 emissions to the atmosphere from major sources are visualised in figure 6.4 below. The designations in the figure are associated with uncertainties and not all source types are included. The figure gives, however, a good presentation of the global character of the mercury pollution problem.

6.3.7      Quantification of mercury releases

487.       It should be kept in mind that almost any attempt to quantify anthropogenic releases to the environment would – in principle – tend to underestimate the total releases as compared to the true releases occurring. The reason is that whereas the contributions from actually quantified individual source types can – in principle – be both overestimated and underestimated, contributions from all relevant source types are very rarely quantified. Often, the collection of the data required for quantification of releases from less homogenous source types will demand larger resources.

Figure 6.4    Spatial distribution of global emissions of mercury to air within a 1*x1* grid. Source of
                  data: J. Pacyna pers. comm., Canadian Global Emissions Interpretation Centre (CGEIC),
                  as presented by AMAP (1998). Original figure presented courtesy of AMAP, Norway.

488.       Types of releases, which are often not included in aggregated inventories or are included with higher uncertainties, are:

• Releases to water, directly from industry and from public wastewater systems;

• Diffuse releases from uncollected waste containing products and materials with mercury content;

• Diffuse releases from informally incinerated waste;

• Diffuse releases from informal, unprotected waste dumps;

• Evaporation of mercury from controlled landfills and informal waste dumps;

• Releases (to air water and soil) from smaller industrial point sources;

• Releases from small-scale/artisanal gold extraction activities;

• Releases from end-of-pipe technology derived wastes.

489.       A mass balance approach like national substance flow analyses (“SFAs” as for example described by Hansen and Lassen, 2000), where release estimates are evaluated versus the inputs to the economy, may be among the methods that give a more complete description of releases and paths of substances like mercury in society, provided a set of basic data are available or can be produced.

490.       National and regional inventories of atmospheric emissions exist for European countries, USA, Canada and possibly a few other countries (see overview of European countries in Pirrone et al., 2001, chapter 2). A number of governmental and intergovernmental submissions to UNEP present data on estimated national and regional atmospheric emissions. The data may possibly add to the understanding of the anthropogenic release patterns as part of a more detailed study. This has, however, not been possible within the time period and the resources available at this stage of UNEP’s mercury assessment process.

6.4         Pathways of mercury to – and in – the environment

491.       The aim of this section is to give an overview of the ways mercury mobilised by humans is released to the environment and how it is distributed, re-distributed and stored in and between environmental compartments. The further pathways leading to adverse effects on humans and the environment are described in chapters 4 and 5 respectively.

6.4.1      Mercury is persistent in the environment

492.       A fact that is basic for the understanding of mercury’s pathways in society and the environment is that it is an element and therefore cannot be broken down or degraded to harmless substances. As described, mercury may change between different states and species in its cycle, but its simplest form is elemental mercury, which itself is harmful to humans and the environment. This means that once mercury has been brought into circulation in the biosphere by human activity it does not “disappear” again in time spans comparable to human lifetime.

6.4.2      Fate of mercury introduced into society through intentional uses

493.       In spite of strongly reduced use of mercury in some regions of the world, global mercury consumption for intentional uses in products and processes is still significant. Releases to the environment from these uses appear generally to be underestimated, most likely due to the complexity and higher research resources needed for such efforts.

494.       Much of the mercury brought into use through products and for consumer purposes will end up in landfills with collected waste, or be lost, dumped or incinerated diffusely and informally directly in the environment. A significant part is expected to end up in the environment rather directly and quickly. How much this amounts to on a global level is difficult to estimate, though it may be possible to form rough estimates given sufficient resources. As indicated below, also the mercury used, collected and treated under more controlled conditions, may partly be spread to the environment over a longer range of time.

495.       In some parts of the world waste collection is informal, inefficient or non-existing. In such cases, mercury in waste will be spread diffusely in the environment, incinerated under informal conditions, or disposed of in informal dumps with no protection against local soil and groundwater contamination.

496.       In parts of the world with more regulated waste collection and disposal, landfills are often controlled and equipped with membranes for collection of water passing through the stored waste (“leachate”) and facilities for the cleaning of the same leachate. After operation time, the waste is often covered with soil and vegetation. In these cases most of the mercury will normally be retained in the stored waste for decades or centuries, as the amounts leaching out with water and evaporating to the atmosphere is generally believed to be minimal in the initial phases of the landfills existence (for vapourisation, see below). On the other hand, leaching and evaporation of mercury will continue for decades, maybe centuries, and will require continued treatment of the leaching water. If this treatment is done in the general wastewater treatment system (normal procedure) most of the leaching mercury will end up in sewage sludge, which is sometimes spread as a fertiliser on farmland, and thus will add to the mercury release to the environment. Or the mercury content will prevent this utilisation of the sludge and therefore it will be incinerated, deposited or treated in some other manner. In the long term (centuries, millenniums), the fate of mercury in normal surface landfills cannot be considered well defined. Can we expect the cleaning of the leachate to continue for centuries? Will former landfills situated near urban areas become attractive for construction and housing activities and thus be exposed by excavation activities (a quite common situation already)? When will, ultimately, geological or climatic processes disturb the sealing of the landfill and potentially spread the deposited mercury over large areas (for Northern conditions, for example, this is likely to happen (ice age), the question is when)?

497.       Some countries rely on controlled waste incineration, which reduce the waste volume and make use of the energy bound in the waste materials. Because of its low boiling point, most of the mercury is thermally released during the combustion, and will be emitted directly to the atmosphere, unless the exhaust gas is filtered effectively.  In some industrialised countries filter facilities on waste incinerators have been improved during the last decade or two, and this is also reflected in decreased emissions of mercury (AMAP, 2000). Generally, only about 35-85 percent of the mercury is retained by filtering (Pirrone et al., 2001), and parts of the mercury will still be emitted directly to the environment. However, carbon injection followed by filtration can increase the retention rate significantly.  Mercury retention close to 100 percent is not normal (see section 8.3). The mercury eliminated from the exhaust gas is retained in incineration residues and, for some types of filtering technology, in solid residues from wastewater treatment (from scrubbing process). These residues are stored in landfills with the implications described above, or – depending of their content of pollutants – used for special construction purposes (under roads or similar). In some cases such solid residues are stored in special deposits for hazardous waste, which are additionally secured with top membranes eliminating or reducing evaporation and leachate production from the waste.

498.       As mentioned, many countries additionally make an effort to separate products with high mercury contents from the general waste stream. It has, however, proved difficult to reach high collection rates, particularly when the separation is to be done by the consumers. A high degree of information and motivation is necessary for successful separation by consumers, and the simplest possible separation system meeting the requirements, should be preferred. Irrespective of collection set-up, separate collection and treatment implies significant extra costs for the society.

499.       Also, mercury is known to evaporate from landfills. For example, Canada has reported that in Ontario atmospheric mercury concentrations over three landfills were measured at 360-4,470 ng/m³ compared to ambient mercury concentrations of 1.5-2.0 ng/m³ across Canada (Pilgrim, 1998).  On the other hand, a more recent survey by Environment Canada indicated a mercury concentration of approximately 10 ng/m³ in landfill gas (Canadian submission, sub42gov). Meanwhile, recent studies (Lindbergh et al., 2001, among others quoted by the USA, Comm-24-gov) indicate that mercury emissions from landfills may be higher than previously estimated. It is clear that as long as a fuller understanding of the significance of this mercury release pathway on a global scale has not been reached, it should be a focus of ongoing research.

500.       Lindbergh et al. (2001) has also found methylmercury being emitted from investigated municipal waste landfills. Based on the knowledge of chemical transformation processes in landfills, transformation of mercury to the more toxic methylmercury (or dimethylmercury) could likely be a general phenomenon in municipal waste landfills. This pathway bypasses the bio-transformation in the aquatic environments and - of particular concern if landfill emissions are significant – adds directly to the methylmercury load and related impacts on humans and the environment.

6.4.3      Fate of mobilised mercury impurities in high volume materials

501.       Significant parts of the mercury mobilised by humans through the use of materials with low natural contents of mercury impurities are released diffusely into the environment with no ways of retaining the mercury. A significant example is the use of coal and other fossil fuels in households and many industrial boilers.

502.       In the increasing industrial activity of the world, large volumes of materials (coal, metal and ore, lime, plastics, some high volume chemicals, etc.) are used, which contain small traces of mercury impurities. As mentioned in other sections of this chapter, these sources constitute major parts of both national and global releases of mercury to the environment.

503.       In some parts of the world, major point sources of releases of mercury from impurities in high volume materials are equipped with emission reduction technology, which reduces the direct outlets of mercury, as well as other pollutants, to the environment. As mentioned in section 6.3.6, the use of such measures have been increasing in Europe and North America during the last decade or two, and implementation has also begun in latter years in other regions, such as East Asia.

504.       Descriptions of the capacity of such measures to retain mercury from direct releases to the environment are given in chapter 8. Mercury retained in solid residues from emission reduction technology is generally stored in landfills implying intermediate protection and potential long-term releases as described for intentional mercury uses above. Liquid (water) residues from certain emission reduction technologies are generally treated in waste water cleaning plants integrated in the facility, or in the public waste water treatment system, with the implications mentioned for landfill effluents above.

6.4.4      The global mercury cycle

505.       It is important to understand that the origins of atmospheric mercury deposition (flow of mercury from air to land and oceans) are local as well as hemispherical or global. Several large studies have supported the understanding that besides local sources, like industry, coal combustion and waste incineration, also the general background concentrations in the global/hemispherical air contributes significantly to the mercury burden at any location (see for example US EPA, 1997; Munthe et al., 2001; Pirrone et al., 2001). Similarly, virtually any local source contributes to the background levels – the global mercury pool in the biosphere. Also the ocean currents are media for long-range mercury transport, and the oceans are important dynamic sinks of mercury in the global cycle.

Movements of mercury in and between environmental compartments

506.       As mentioned, mercury is a natural element that cannot be created or destroyed and the same amount has existed on the planet since the earth was formed.  A significant amount of research indicates that natural and human (anthropogenic) activities can redistribute this element in the atmospheric, soil and water ecosystems through a complex combination of transport and transformations. Figure 6.5 below illustrates the main interactions between the environmental compartments.

 Figure 6.5    Dynamic interactions of mercury distribution between the environmental compartments
                   (based on Lamborg et al., 2002, as adapted from Mason et al., 1994).

Air

507.       Mercury is emitted to the atmosphere from a variety of point and diffuse sources and is dispersed and transported in the air, deposited to the earth and stored in or redistributed between water, soil and atmospheric compartments. Therefore, mercury cycling and mercury partitioning between different environmental compartments are complex phenomena that depend on numerous environmental parameters. Wet deposition was, until recently, assumed to represent the primary mechanism for transfer of mercury and its compounds from the atmosphere to aquatic and terrestrial receptors.  However, studies by US EPA, the Florida Department of Environmental Protection and US Department of Energy have all shown that dry deposition of divalent gaseous mercury species can be equal or greater than wet deposition, even in moist climatic areas such as the Florida Everglades and the Great Lakes Region with relatively high annual precipitation (Rea et al., 2000; 2001; Vette et al., 2002; Landis et al., 2002).  The chemical and physical form of mercury in air affects the mechanisms by which it is transferred to the earth surface and ultimately influences the total depositional flux. An increase in ambient air concentrations of mercury will result in an increase of direct human exposure and an increase of mercury flux entering terrestrial and aquatic ecosystems leading to elevated concentrations of methylmercury in freshwater and marine biota. Extensive research conducted on mercury deposition in Boreal forests systems has shown that the main source of mercury and methylmercury to the forest floor is litterfall, i.e. needles, branches (Iverfeldt, 1991; Munthe et al., 1995). This mercury and methylmercury mainly originates from the atmosphere (not via root uptake) and adsorbs on plants surfaces via dry deposition.

508.       Monitoring networks for wet deposition of mercury have been established in North America and in Europe for the purposes of providing an indication of the magnitude of depositional flux and to provide data for evaluation and testing of atmospheric mercury simulation models. Figure 6.6 shows the monitoring stations established for North America within the Mercury Deposition Network as of 2001 with observations of accumulated wet deposition of mercury and average mercury concentration in precipitation for that year. Figure 6.7 shows similar results from the wet deposition network established in Sweden. The Swedish stations are a part of the EMEP monitoring activity under the UNECE LRTAP Convention. Similar networks for monitoring dry deposition of mercury are needed to provide a complete measure of depositional flux and to provide data for simulation model testing and evaluation.

Figure 6.6     Monitoring stations established for North America within the Mercury Deposition Network 
                   as of 2001 with observations of accumulated wet deposition of mercury and average
                   mercury concentration in precipitation for that year. Figures are from the MDN web page:
                   http://nadp.sws.uiuc.edu/mdn/.

Figure 6.7   Monitoring stations and results (wet deposition measured in years noted) from the wet
                 deposition network established in Sweden. The Swedish stations are part of the EMEP
                 monitoring activity under the UNECE LRTAP Convention. Figure provided by John Munthe,
                  IVL, Sweden.

509.       A recent evaluation of mercury levels in 1 kg pike in Swedish lakes has revealed a decrease when comparing concentrations found between the periods 1981-1987, and those measured in 1988-1995. The decrease may be attributed to documented decreases in atmospheric deposition during this period (Johansson et al., 2001). Similar effects appear to be occurring in the Florida Everglades in the USA following the implementation of mercury emission controls on waste incinerators in the Miami area, but these results are preliminary and have not been published in the peer-reviewed scientific literature.

Water

510.       Once in aquatic ecosystems, mercury can exist in dissolved and/or particulate forms and can undergo chemical/microbial transformation to methylmercury as described in section 2.3.  Contaminated sediments at the bottom of surface waters can serve as an important mercury reservoir, with sediment-bound mercury recycling back into the aquatic ecosystem for decades or longer.

Soil

511.       Mercury has a long retention time in soil and as a result, the mercury accumulated in soil may continue to be released to surface waters and other media for long periods of time, possibly hundreds of years.

Environmental long-range transport

512.       Mercury pollution is transported over long distances by air as well as water movements. In particular, air transport is believed to be important for mercury, as mercury in the form of vapourised elemental mercury may be transported quickly over long distances and thus air transport may be responsible for the distribution of mercury to the most remote parts of the Earth. For example, the AMAP-assessment (AMAP, 1998) points at mining and metallurgical sources in the Northern part of Russia, besides industrial regions in Europe and North America, as the dominating sources of other heavy metals in the air in the High Arctic during winter time. Contrary to other heavy metals, the large volume of atmospheric mercury emissions is emitted as the element in a gaseous state. Mercury vapour is capable of being transported over long distances with the air masses. Newer evidence suggests that the background levels of mercury in the atmosphere (from anthropogenic and natural sources) contribute significantly to the mercury burden in remote areas like the Arctic. The remaining part of mercury air emissions are in the states of ionic or gaseous mercury compounds/ions, which are deposited by both dry and wet atmospheric processes close to the source, mainly within a radius of a few hundred kilometres.

513.       A group of scientists, including several of the world’s top specialists on atmospheric mercury research, conclude the following in a recent review of the environmental impact of mercury in Europe (Pirrone et al., 2001):

“Long-range transport of mercury in Europe was first observed in the late seventies in Sweden (Brosset, 1982). Since then long-term monitoring activities carried out in Scandinavia have shown a clear gradient in wet deposition of mercury with elevated fluxes in the south-western part of the region, i.e. closer to the main emission sources in Central Europe (Iverfeldt, 1991; Munthe et al., 2001a). Similar patterns have been shown in North America. The Scandinavian studies have also revealed a significant decrease in wet deposition after a reduction of mercury emissions around 1990 (Iverfeldt et al., 1995; Munthe et al., 2001a).

514.       Recent research projects conducted within the Environment and Climate Research Programme have revealed that the anthropogenic influence on atmospheric mercury levels in Europe are still considerable despite reductions in emissions during the last decade (Pirrone et al., 2000; Munthe et al., 2001). These research projects have also clearly shown the influence of the hemispherical/global cycling of mercury. The authors conclude that, despite the significant decreases in mercury emissions during the last decade, the atmospheric deposition is still significantly increased in comparison to pre-industrial times. They state that, according to their judgement, further reductions are needed to protect sensitive ecosystems and to prevent and decrease levels of methylmercury in freshwater fish in Scandinavia and elsewhere.  A significant influence from background contributions was also noted. The authors assume that a large portion of the mercury present in the global atmosphere today is a result of decades of emissions from anthropogenic activities. They state that the natural component of the total atmospheric burden is difficult to estimate, but is probably on the order of 20 to 40 percent and that anthropogenic activities have thus increased the levels of mercury in background air by roughly a factor
of 3.

515.       A similar understanding is expressed by the US EPA (1997) in their “Mercury report to Congress”.

“The polar sunrise mercury depletion incidence”

516.       A special phenomenon has been shown to influence the deposition of mercury in the Polar regions. It has been termed “the polar sunrise mercury depletion incidence” or “the mercury sunrise”, as a highly elevated deposition of mercury is taking place during the first few months of the Polar sunrise (best studied in the Arctic). It appears that the solar activity and present ice crystals influence the atmospheric transformation of elemental gaseous mercury to divalent mercury, which is more rapidly deposited. The mercury depletion happens at the same time as the surface-level ozone depletion (a separate phenomenon from the better known ozone depletion in the stratosphere).

517.       This polar phenomenon poses a special challenge to atmospheric mercury transport modellers, because they need to understand the mechanism of the phenomenon to predict mercury exchange and deposition in and around the Polar regions.

518.       The net atmospheric input to Polar ecosystems resulting from this phenomena is not known in detail. Re-emissions of mercury occur from the snow surface and during snowmelt, but the depletion events may still result in significant input to the aquatic environment. In case this phenomenon shows up to be resulting in higher yearly mercury deposition rates in the Polar regions than in other regions of the world, this could mean that the Polar regions serve as “mercury cold traps” collecting an un-proportionally high part of the global mercury emissions. This would fit well with the observed high mercury concentrations in the Arctic aquatic environment.

519.       Mercury depletion has now been observed in Alert, Canada (Schroeder et al., 1998; Lu et al., 2001), in Barrow, Alaska, USA (Lindberg et al., 2002b), Svalbard (Berg et al., 2002), in Greenland (Skov, 2002) as well as in the Antarctic (Ebinghaus et al., 2002), and can thus be described as a generally occurring polar phenomena which may influence the total input to Polar ecosystems.

520.       Suggestions for other references for further reading about the polar mercury depletion incidence are Schroeder et al. (1998) and Lu et al. (2001).

Accumulated anthropogenic mercury burden

521.       Mercury from natural sources is present in the environment, but the anthropogenic contribution to the environmental mercury burden is evident. On average around the globe, there are indications that anthropogenic emissions of mercury have since pre-industrial times resulted in 50-300 percent increases in deposition rates, and in and around industrial areas the deposition rates have increased by a factor of 2-10 during the last 200 years (Bergan et al., 1999; Lindquist et al., 1984; as cited in von Rein and Hylander, 2000). Such information can be derived from mercury concentration profiles in lake and ocean sediments and peat bogs, and from geographical trends in soil mercury concentrations, among others.

522.       Profiles of mercury concentrations in different depths from the sediment surface give a picture of the changes in the mercury burden over time. Several natural conditions, such as local currents, oxygen concentrations and biological activity, influence the immobilisation and re-mobilisation of mercury bound in sediments. Therefore, the locality from which mercury profiles are taken for this purpose should be selected with care, and the result should be interpreted cautiously, particularly for the most recent upper layers, which may still be affected by re-mobilisation (HELCOM, 2001).

523.       However, in a very recent paper Schuster et al. (2002) used a glacial ice-core record to study atmospheric mercury deposition during the last 270 years.  Among other observations, they concluded that the anthropogenic contribution during the last 100 years rose to 70 percent of the total.  On the other hand, declines in atmospheric mercury deposition were apparent in both the ice-core record and in sediment-core records over the last ten years (Schuster et al., 2002, as cited by the World Chlorine Council: Comm-4-ngo).  While keeping in mind the caveat above, this may suggest that the major anthropogenic influence on atmospheric mercury deposition during the industrial era is now beginning to decline.

524.       As an example of indications of the accumulated mercury burden over time in different geographical regions, mercury concentrations in marine sediments from the Arctic, Skagerrak in the greater North Sea area (OSPAR waters of North Europe), and the Baltic (HELCOM waters), are presented in the figures 6.8-6.11 (the selection of illustrations is somewhat arbitrary – many examples exist in the literature). It is notable that most profiles show the same trend of increased mercury concentrations in industrial times.

525.       For the Arctic (figure 6.8), it is concluded in the AMAP assessment (1998) that several data sets indicate widespread accumulation of mercury in surficial Arctic sediments. The enrichments occur particularly in the upper 2-10 cm of the sediments, even at the North Pole. The report states that this phenomenon could indicate global scale input to the marine environment in recent times, but that more investigations are required, before definite conclusions could be drawn about the source of the observed enrichment.

Figure 6.8    Examples of mercury concentrations in sediment profiles from the Arctic marine area
                  (AMAP, 1998).[4] Original figure presented courtesy of AMAP, Norway.

526.       The profiles from Skagerrak (figure 6.9) and the Baltic area (figures 6.10 and 6.11) have been dated. Here, mercury concentrations have risen during the last century. For the Baltic profiles there is an indication that the mercury burden has decreased during the last few decades. This appears reasonable, as control of regional releases has been strengthened considerably in Scandinavia in this period (Submission from the Nordic Council of Ministers, sub84gov). The overall pattern in the European waters is that the mercury concentrations in marine sediments are highest close to shores and river outlets with high anthropogenic activity and industrial sources (like pulp and paper industry and chlor-alkali industry) (OSPAR, 2000; HELCOM, 2001).

Figure 6.9    Examples of mercury in sediment profiles from Skagerrak south of Norway in the OSPAR
                  Convention marine area (Oddvar and Thorsnes, 1997). Blue line: Profile from station 2
                  nearest the Oslo Fjord and the Swedish coast. Original figures presented courtesy of
                  Geological Survey of Norway (NGO).

527.       A number of researchers have questioned under what conditions sediment profiles in remote areas without local pollution can be taken as evidence for an elevated global or hemispherical atmospheric background mercury concentration, as otherwise commonly agreed among scientist in this field. The question is whether mercury is mobile enough to change physical position in the upper sediment layers during early geo-chemical changes, so-called “diagenesis”.

528.       The question has been the subject of some discussion in recent literature. For example, Fitzgerald et al. (1998) has “examined the weaknesses in interpretation and the choice of information that has been used to argue against atmospheric mercury contamination” and reviews several sets of data from other investigations, which cannot – from their judgement – be explained by diagenesis. Among other arguments, Fitzgerald et al. (1998) claim the older ice core studies from Greenland, used for argumentation by the critics, were not using the ultra-clean sampling techniques known today, and produced results contaminated by the sampling equipment. Newer studies from the Greenland ice cap support the conclusions of a general increase of atmospheric background levels due to anthropogenic emissions. Fitzgerald et al. (1998) conclude that despite uncertainties in current understanding, there is a broad and geologically consistent data base indicating that, over large regions of the globe, anthropogenic mercury emissions have increased relative to natural sources since the onset of the industrial period.

529.       Investigations of mercury concentrations in lake sediment profiles performed in the 1980’s in Sweden clearly show an increase in mercury concentrations in surface sediments (Johansson, 1985). The increase is large in sediment profiles taken in the southern part of the country whereas lake sediments from lakes in the north show very little increase. This clearly indicates the influence of long-range transport from source areas on the European continent. In more recent lake cores from Southwestern Sweden, mercury concentrations in surficial sediments decrease corresponding to a reduced atmospheric input in the 1990’s (Munthe et al., 1995).

Figure 6.10    Mercury levels in surface sediment in the Baltic Marine Area (mg//kg dw; salt-corrected
                    values). The yellow and orange areas indicate the deep basins. Orientation: Denmark is
                    situated in the lower left corner of the figure - Russia in the higher right corner. Figure
                    from HELCOM (2001), original presented courtesy of HELCOM, Finland.

Figure 6.11    Vertical distribution of mercury (mg/kg dry-weight basis) in sediment from Lübeck Bay,
                    Gdansk Bay, Gulf of Finland (GF-2), Bothnian Bay (F-2) and Bothnian Sea (EB-1) in
                    1993. The age of the sediment is indicated. Figure from HELCOM (2001), original
                    presented courtesy of HELCOM, Finland.

1.4.5             Atmospheric transport models for mercury

530.       For a couple of decades efforts have been invested in developing models able to describe the often complex picture of atmospheric cycling of mercury in different regions of the world. Today models exist for parts of the Northern Hemisphere allowing scientist to describe air transport of substances like mercury and predict mercury deposition rates as related to geographical position, as well as monitoring the consequences of changes in emission patterns.  Simulation modeling plays a critical role in developing a better understanding of atmospheric mercury cycling when combined with basic observational study.  When model results are compared to observations, the incidents of poor agreement are used to isolate important scientific uncertainties which can be addressed by further basic research.  The models are then updated to reflect any new atmospheric source or process information obtained and tested against observations once again.  This iterative cycle of modelling and basic research continues until the desired model accuracy with respect to observation is demonstrated.  At this time, there remain serious discrepancies between model simulations and observed atmospheric mercury concentrations and deposition fluxes, and model inter-comparison studies have shown differing results from various models when simulating identical circumstances (see Ryaboshapko et al., 2001).  This suggests that our scientific understanding of atmospheric mercury remains flawed, incomplete, or a combination of both.

531.       Atmospheric models of mercury transport have been developed for the last decade both on regional and global/hemispheric scales. Regional models cover North America (Bullock et al., 1997; Pai et al., 1997; Seigneur et al., 2001; Bullock and Brehme, 2002) and Europe, including European part of Russia (Petersen et al., 2001; Ilyin et al., 2001). Global or hemispheric models could be divided into box-type ones describing general cycling of mercury in the environment by means of large reservoirs (Mason et al., 1994; Lamborg et al., 2002) and grid models calculating long-range mercury transport and deposition over the globe (Bergan et al., 1999; Seigneur et al., 2001; Travnikov and Ryaboshapko, 2002). Comparison and evaluation of different mercury transport models are performed within the mentioned models intercomparison campaign (Ryaboshapko et al., 2001). Creation of global models has also been attempted, based on mass balance (Mason et al., 1994) or meteorological transport (Bergan et al., 1999 and Shia et al., 1999) approaches.

532.       An impression of the development and state of the art of mercury transport modelling can be obtained in the following documents:

1996:  Global and regional mercury cycles: Sources, fluxes and mass balances (Baeyens et al., 1996).

1999:  Proceedings from the WMO/EMEP/UNEP workshop on modelling of atmospheric transport and deposition of persistent organic pollutants and heavy metals (WMO/EMEP/UNEP, 2000).

2000:  Current methods and research strategies for modelling atmospheric mercury (Bullock, 2000); which gives a description of current methods and research strategies for modelling atmospheric mercury transport, transformation and deposition in North America and Europe.

2001:  Summary and findings from the AMAP-NMR-MEPOP International Workshop on Mercury and POPs, held in Roskilde, Denmark, 10-12 September 2001 (Annex 3 of the submission from the Nordic Council of Ministers, sub84gov).

2001:  EU Ambient Air Pollution by Mercury (Hg) – Position Paper (Pirrone et al., 2001); which gives a good description of atmospheric mercury transport and deposition modelling as well as the most recent results from Northern Europe and the Mediterranean area.

2002:  Comparison of mercury chemistry models (Ryaboshapko et al., 2002); which describes a comparison of the model treatments for mercury in cloud/fog water in various long-range transport models under development in North America and Europe.

533.       Quite a number of documents on mercury transport modelling and its results have been produced in connection with the EMEP programme and other activities relating to the LRTAP Convention performed under the auspices of UN ECE, see the submissions from UN ECE (attachments to sub9igo).

534.       Annual operational calculations of mercury transboundary transport and depositions within the European region are performed by Meteorological Synthesizing Centre East of EMEP (MSC-E). Mercury concentration levels in the ambient air and deposition fields for each Party of the LRTAP Convention are assessed along with mutual mercury transport between countries. Besides, recently developed within the joint project of EMEP and AMAP, a hemispheric model allows for the evaluation of mercury contamination in the Northern Hemisphere. Results of the hemispheric mercury transport modelling are presented in Figure 6.12 (Travnikov and Ryaboshapko, 2002). The modelling results show that gaseous mercury is more or less uniformly distributed in the Northern Hemisphere (note differences in scale), while deposition fluxes vary significantly (up to 2 orders of magnitude) from industrialized to remote regions. It is possible to distinguish the three areas most contaminated by mercury: Southeastern Asia, Europe and the eastern part of North America.

535.       Some national submissions to UNEP give information on ambient air concentrations, which might add further to the understanding of the atmospheric transport of mercury.

   a)  Mean annual concentration of total                   b) Total annual mercury deposition
        gaseous mercury

Figure 6.12    Mean annual concentration of total gaseous mercury (a) and total annual mercury
                    deposition (b) in the Northern Hemisphere – note differences in scale. From Travnikov and
                    Ryaboshapko (2002); submitted by MSC-E of EMEP (comm-4-igo).

6.4.6             Watershed cycling models for mercury

536.       As with modelling the atmospheric cycling of mercury, modelling efforts to address the watershed cycling of mercury have received increasing attention in the past two decades. The modelling is also complex, due to the numerous species and transformation processes possible, and the difficulty in quantifying each. One modelling effort that has been developed and applied in North America is the Mercury Cycling Model. The model, initially developed for lakes, considers mercury inputs to and losses from the water body, reaction processes (e.g. methylation and demethylation, reduction of dissolved reactive mercury to elemental mercury, etc.), fluxes between compartments (e.g. particle settling to sediments, sediment resuspension), and other components. The model, which has undergone various modifications, has been applied in several locations, including a temperate lake in northern Wisconsin and the Florida Everglades, as part of US EPA’s pilot mercury Total Maximum Daily Load (TMDL) project. For information on the model see Hudson et al. (1994).