RT = 5076 Joules
Abstract
Mercury pollution in the Amazon region represents today one of the most serious environmental issues faced by mankind. Quantities from 70 to 170 tonnes of mercury are discharged into the Amazonian environment annually from gold infomal mining operations. Vegetation fire is pointed as another major source of Hg emission. Mercury is likely emitted in a reactive form form vegetation and the extent of the biota contamination is wider than the Hg emitted by mining activities. Stability of metallic mercury in aquatic environments is investigated. In contact with organic-rich sediments, mercury-organic complex formation is favourable as indicated by thermodynamic and laboratory studies. Although methylation process from dissolved complexes is an unclear issue, mercury mobility as complexes might increase bioavailability.
Introduction
The high levels of mercury in fish in the Amazon region has been reported by many authors (Martinelli et al., 1988; Pfeiffer et al., 1989;Fernandes et al.,1990; Malm et al., 1990; Pfeiffer et al., 1991; CETEM, 1991a; CETEM, 1991b, Veiga and Meech, 1992; CETEM, 1993; Barbosa et al., 1994). Mining activities has been pointed to be the main source of Hg pollution. Fish is the main diet of most Amazonian communities and people living distant from mining activities have shown high levels of mercury in blood (GEDEBAM, 1992).
Mercury emission from vegetation is another source that received little attention from researchers. Nriagu (1989) pointed that biologically-mediated volatilization can be another important source of Hg emissions. Non-methane hydrocarbon compounds, such as isoprene and terpenes from plants may form strong complexes with Hg and other trace metals and thus play a part in the transfer of metals to the atmosphere. Particulate organic carbon, which is the dominant component of atmospheric aerosols is another Hg bearing phase.
Natural forest fires were attributed as an additional source of mercury emission in Manitoba, Canada (Williamson et al, 1986). Recently, forest fires were pointed by Veiga et al. (1994) as an additional and major source of Hg emission in the Amazon.
Although Hg is not allowed to be used in "garimpos", in fact amalgamation is the main process used. Cleary (1990) reported only one "garimpeiro" who did not use Hg and he was regarded as eccentric by his peers. More than 90% of the gold present in gravity concentrates can be trapped in amalgam according to field observations at some operations. Price is not an impediment for reducing use. Even at 5 times the international price, Hg is still a cheap reagent for extracting gold, with a cost equivalent to 0.012 g of gold per tonne processed. (Veiga & Fernandes, 1990). The mining and amalgamation methods used in "garimpos" are variable which together with the fate of contaminated tailings and Au-Hg separation procedures define the extent of Hg losses.
This paper reports the main sources of mercury emission in the Amazon and comments the Hg stability in aquatic systems.
Emissions from Mining Activities
Brazil is not a mercury producer and imports around 340 tonnes annually. From 1972 to 1984, Mexico was the main Hg supplier to Brazil. Since 1984 this picture has changed and non Hg-producing countries (the Netherlands, Germany and England) are responsible for almost 80% of the Hg entering Brazil. Mercury imports are allowed only for registered industrial uses, however the declared uses (electronic industries, chlorine plants, paints, dental, etc.) are declining. The updated Brazilian laws (Norm 434 - Aug.9/89 and Norm 14 - Jan.15/90) intend to exert more control on Hg imports. In 1989 this represented about 22% of the total 340 t of mercury. The remainder was imported for re-sale to industries, but it is estimated that over 170 t were illegally diverted to "garimpos" - informal mining activities (Ferreira & Appel, 1991).
Pfeiffer & Lacerda (1988) reported that Hg losses due to dredge mining range from 30 to 45% of the Hg introduced in the process. When retorts are not used, the losses include 45% released into rivers and 55% into atmosphere.
Farid et al (1991) evaluated a type of "garimpo" which used a grinding operation (hammer crusher) and gravity concentration (sluice or centrifuge). Figure 1 shows a simplified flowsheet of the operations involved in this type of "garimpo", while an Hg-balance is provided in Figure 2. These operations are conducted on a lode ore and its weathered part. Erosion of the quartz vein hosted by ferruginous and carbonaceous phyllites spread out gold into the weathered layer. The gold grade is poorer than in quartz veins but easier to mine. Large production can be achieved, such as 3 million tonnes of run of mine/year, but gold recovery in the gravity circuit is usually lower than 50% due to poor liberation.
Concentrates are usually amalgamated in barrels or pans and the mineral portion is separated from amalgam by panning. This operation takes place either in waterboxes or in pools excavated in the ground. The method used to remove the excess mercury from amalgams is filtration using a piece of fabric to squeeze by hand. The amalgam obtained, usually with 60% gold content is retorted or simply burnt in pans. The bullion still contains 5% residual mercury which is released during melting operations in gold shops. Mercury entering the atmosphere can represent as much as 50% of that introduced into the amalgamation process when retorts are not used (Fig. 2).
Fig. 1 - Flowsheet of a typical "garimpo" in Poconé, MT, Brazil
Fig. 2 - Balance of mercury in the amalgamation steps (adapted from Farid et al, 1991)
As a rough estimate, if we assume losses of 40% of 170 tonnes Hg used in "garimpos" in 1989 (Ferreira & Appel, op.cit.), 68 tonnes/year are calculated as losses due to poorly conducted amalgamation practice. This is similar to the range of 50 to 70 tonnes Hg/year reported by Pfeiffer and Lacerda (1988). The ration Hgconsumed : Auproduced is sometimes used to calculate Hg losses. This seems to be an inaccurate approach since gold output from these mining activities is not well established and is difficult to estimate, ranging from 34 (official production by DNPM, 1989) to 220 (Fernandes & Portela, 1991) tonnes/year. In addition, stockpiling is not taken into account by this ratio and this may actually be a preferred practice by "garimpeiros" because of the "illegal" nature of this commodity.
The Hg:Au ratio provides a picture of mercury consumption on a large scale. For instance in the Amazon region this ratio ranges from 0.6 to 1.3. A distribution of mercury losses can be done as follows (CETEM, 1989) :
Emissions from Vegetation
Forest fires may be expected to mobilize Hg contained in biomass and redistribute it into the atmosphere either as vapour or attached to particulates.
Fish from reservoirs in northern Manitoba showed high Hg levels. No man-made source of Hg could be precisely identified. The high Hg background of organic soils associated with impoundments stimulates biomethylation and subsequent incorporation of Me-Hg in the aquatic biota. Natural forest fires were also attributed as an additional source of mercury emission. The amount released annually to the atmosphere from natural fires in the boreal forest region of Manitoba was calculated at 20 g Hg/ha representing only 0.02% of the provincial annual emissions from natural sources which creates high short-term emissions in the form of a pulse. The evaluation assumed 0.4 ppm as the Hg concentration in timber, but about 0.08 ppm was considered lost during fires (Williamson et al, 1986).
Worldwide, wild forest fires are estimated as responsible for releasing 20 tonnes of Hg to the atmosphere, which comprise less than 1% of natural emissions (Nriagu, 1989). However, intentional wood combustion represents 60 to 300 tonnes of Hg (about 5% of all man-made emissions in 1983) (Nriagu & Pacyna, 1988). This estimate was based on a range of 0.1 to 0.5 ppm Hg in wood. Today with the very high rate of deforestation by fires in the Amazon, Hg emissions derived from wood combustion must be a more important source. Between 35,000 (1987) and 50,000 (1988) km2 of the Amazon are estimated to be deforested per year becoming mostly cattle pasture, although other expansion reasons are also apparent (Fearnside, 1989; Terborgh, 1992).
Natural Hg levels in plants range from 0.001 to 0.1 ppm (dry weight). In forest ecosystems, this increases to 0.01 to 0.3 ppm (Pendias & Pendias, 1992), while crops grown in soils containing less than 0.04 ppm Hg vary from 0.004 to 0.09 ppm (Gracey & Stewart, 1974). Little is known about Hg distribution in the Amazon flora. Samples of macrophyte (Potenderia lanceolata) from uncontaminated areas around Brazilian "garimpos" show Hg levels of 0.10 ppm (Lacerda, 1991). Other aquatic macrophytes, Victoria amazonica and Eichornia crassipes collected at polluted sites in Madeira River "garimpos" show levels up to 1 ppm Hg (Martinelli et al., 1988).
A wide range of temperatures can be encountered in vegetation fires; between 650 and 1100°C are reported (Raison et al, 1985). At 200 - 300°C destructive distillation of up to 85% of organic substances occurs (Benscoter & Neuenschwander, 1989). Losses of trace metals during fires occurs through bonding with particulate matter (M. Feller, Dept. Forestry, UBC, personal comm.). Most mercury compounds are volatile at temperatures between 25 - 450°C, in which organic mercurials usually have lower boiling points than inorganic compounds. Mercury emission from fossil fuels is partly or fully in vapour form. In a coal fired power plant, it was observed that 92% of Hg was released in vapour form, 7% in particulate form and 1% was retained in the ash. When coal is burnt, loss of 90 - 99% of Hg is reported (Mukherjee, 1991; Kaiser & Tolg, 1980). In combustion plants 50% of mercury emitted is elemental and 50% in divalent form. This latter can be in gaseous form or bound to particles (Hall et al, 1991). When chloride is present in a combustion process, such as in waste incinerators, HgCl2 is reported to be the predominant form emitted (Pacyna & Münch, 1991). This compound is stable at high temperatures, but HgO decomposes above 500 °C. In a combustion process HgO can be formed from gaseous mercury at temperatures between 300 and 500 °C, during the cooling process (Hall et al, 1990). HgO formation was observed by Hall et al. (1991) only in the presence of a catalyst (activated carbon). Obviously soot or ash particles could assist in this process.
Kaufman et al. (1992) estimated that the combustion efficiency average 91% and 97% in tropical deforestation and cerrado fires respectively. These authors analyzed about 1,000 ppm Hg in smoke particles smaller than 2.5 mm but the source of this metal was not identified. The use of Hg by "garimpos" was inferred by the authors as a possible source.
Considering that most mercury is present in wood in organic form, it seems reasonable to assume that about 90% of Hg is lost from above ground biomass (M. Feller - Dept. Forestry, UBC, personal comm.). It is also reasonable that, similar to other metals, the remaining portion of Hg becomes weakly bound to the ash after burning and is easily leached by runoff water.
Table 1 - Estimated Hg emissions from deforestation
We have calculated the amount of Hg emitted by deforestation from estimates of biomass distribution in the Amazon (Jordan, 1989). We assumed that the majority of the Hg compounds will be released from above ground biomass even without complete combustion, while only a minor amount is volatilised from organics in the soil surface. Our evaluation of the efficiency of Hg release is shown in the Table 1. Using an estimate of Hg levels in plants and organic matter of 0.05 and 0.1 ppm respectively, about 15 g/ha (1.5 kg/km2) of Hg are being emitted. Considering that 50,000 km2 were burnt in 1988, 75 tonnes of Hg likely were emitted to the atmosphere that year.
The total area of forest lost until 1991 is estimated at 404,000 km2 (Abril, 1993). So over 610 tonnes of Hg have been emitted into the atmosphere from this source. Applying a range of 0.1 to 0.5 ppm Hg in wood as claimed by Nriagu & Pacyna (1988), the Hg emitted from above ground wood alone would range from 950 to 4730 tonnes up to 1991.
Following deforestation of an area, the land is generally used for pasture. On a cycle between 2 to 7 years, the pastures are reburned in order to control pests and to restore adequate nutrient levels in the soil. Cerrado vegetation (mainly grasses and bush) takes up mercury from soil and deposition by rain. Today, it is estimated that only 10% of the biomass burnt is from to deforestation (A. Seltzer, Brazil's National Inst. Space Research, personal comm.) and 90% is from cerrado burning. This mercury is also released upon burning and although the amount of biomass is much lower than the forest, the extent of burning is considerably higher. The above ground biomass of cerrado estimated by Setzer & Pereira (1991) is 43 tonnes/ha. Using the same relationship between above ground and below ground biomass in the Table 1, we calculated the cerrado biomass and Hg released by cerrado burning (Table 2).
Mercury is also released from soil organics and humus material following a forest fire because of organic decay. Release occurs in two ways - by volatilization and by surface water leaching. Rate of removal is considerably slower than from the initial release by burning. The rate of organic decay depends on many factors - climate, type of original soil, geology of the underlying rock, land use, weather, etc. The exact model is thus extremely uncertain and site-dependent.
Table 2 - Estimated Hg emissions from cerrado burning
Setzer & Pereira (1991) estimated that 200,000 km2 were burnt in the Amazon (40% deforestation, 60% cerrado) in 1987. Using these numbers, we calculated the Hg released by deforestation (200,000 x 0.4 x 1.5 kg Hg/km2 = 120 tonnes Hg) and Hg emitted by cerrado burning (200,000 x 0.6 x 0.225 kg Hg/km2 = 27 tonnes Hg). This shows that Hg released by fires is substantially higher than Hg emitted by "garimpos" (about 70 to 100 tonnes/y).
Mercury Reactivity in Natural Aquatic Systems
The form in which mercury is released into the environment determines its reactivity and transformation rate. Pacyna & Münch (1991) showed that more than 50% of Hg released from coal-fired boilers is in a water soluble form. No information was found about the Hg form in wood combustion gases but we assume that part of this mercury is in the elemental form and oxidized mercuric species may be formed during cooling. This represents an imminent danger since these forms of Hg can be easily transformed into methylated compounds in the waterways.
From mining activities, metallic mercury vapour, Hg° likely constitutes by far the largest component of the total gaseous Hg concentration in the atmosphere with perhaps some minor amount of divalent Hg (II). Owing to its high vapour pressure (0.246 Pa or 1.85 x 10-3 Torr at 25 °C) mercury in ambient air is predominantly in the gaseous phase rather than associated with particulate matter as with other transition metals (e.g. Cd, Zn, Cu, Ni, Pb). The process in which Hgo vapour oxidizes in the atmosphere is not well understood. Oxidation is certainly accelerated in clouds in the presence of ozone (O3) and chlorine (Cl2) but reduction of Hg(II) to Hgo is also a feasible process (Schroeder, 1991).
Lindqvist et al. (1991) pointed that mercury can remain in the atmosphere for 6 to 24 months in a dry climate. However the reactions that occur in clouds result in a shorter residence time as mercury returns to the earth through rainfall .
The extent of metallic mercury dispersion due to amalgam burning in pans is not quantified. Neither air analysis nor soil samples up to 500 m from gold shops show significant mercury concentration (CETEM, 1991a, 1991b, 1993) (Fig. 3). According to , Marins et al. (1991)the majority of Hg is deposited near the emission source (i.e. with 1 km).
Fig. 3 - Hg distribution in soils around gold shops of Alta Floresta (Source: CETEM, 1991b)
Precipitated mercury or that which is dumped with amalgamation tailing enters the aquatic system predominantly in metallic form. How this mercury is transformed in soluble compounds depends mainly on sediment (soil) composition and physico-chemical characteristics.
If we assume that metallic mercury is in equilibrium in a simple
aquatic system, the predominant mercury species in solution would
be undissociated mercury, Hgo (aq.). Data from Balej (1985):, RT = 5076 Joules
Hg0(1)=Hg0(aq) ............(1)
K=10-6.5............(2)
Mercury has to be oxidized to become more soluble, i.e. to form Hg(II) species or complexes which are far more reactive. Mechanisms of methylmercury formation are faster when Hg (II) compounds exist (Bisogni & Lawrence, 1975; Imura et al., 1971). The dominance of each Hg (II) species is controlled basically by pH, Eh, chloride concentration, sulphide concentration and the presence of other soluble substances, such as organic matter (Björnberg at al., 1988).
The stability of mercury compounds can be studied using Eh-pH diagrams. The Eh-pH diagram presented in Fig. 4 was built with the CSIRO Thermodata program. The system was simplified by assuming concentrations of [Hg]= 2ppb (or 10-8 M), [Cl]= 3.5ppm and [S]= 3.2ppm (or 10-4 M for each). Hgo (aq) with a nearly constant solubility of 63 ppb (eq. 2), is predominant but other species also exist in aerated waters. Surface waters (aerated) and terrestrial soils can exhibit Eh > 0.4 V favouring stability of other Hg species that are more soluble and reactive than Hgo (aq.), such as Hg(OH)2 and HgCl2. (Gavis & Ferguson, 1972; Schuster, 1991). In an oxidizing condition, if Hg(II) is present, HgCl2 (aq) or Hg(OH)2 should be the predominant inorganic species in solution depending on chloride concentration.
Fig. 4 - Eh (redox potential) versus pH for the main inorganic Hg species.
It seems that the information obtained from Eh-pH diagrams with respect to natural systems, must be used carefully. The theoretical values are applied to a system in equilibrium. In natural waters, it is common to find non-equilibrium conditions, as transformation rates to more stable compounds can be very slow (Baeyens et al., 1979). The most toxic form of mercury, methylmercury is an example. It is thermodynamically less stable than inorganic species. So equilibrium of inorganic and organic mercury species at high organic concentrations is not possible (Hem, 1970; Stumm & Morgan, 1981). In addition methylmercury is mainly produced in sediments and afterwards released into the water column to be accumulated rapidly by biota.
The thermodynamic analysis based on Eh-pH diagrams (Fig. 4) suggests that metallic mercury dumped in an aquatic environment with a sediment redox potential (Eh) below 0.4 V should be stable. However, the presence of soluble organic acids in the sediments changes this conclusion.
When dissolved organic matter (say fulvic acid) is present at concentrations higher than 1 mg/l (ppm), the complex formed (Hg-FA) is more stable and predominant than any of the inorganic species (Duinker, 1980; Xu and Allard, 1991) (Fig. 5). The presence of fulvic acids (FA), is an important parameter that enhances solubility of organic matter and associated mercury. The more FA present in the aquatic system, the more the metal becomes water-soluble as a complex. When the ratio FA:Hg > 2, formation of water soluble complexes is favoured. Solubility of such complexes increases with pH and they are more stable than inorganic Hg complexes, preventing Hg compounds from precipitating. Schnitzer and Kerndorff (1981) have shown that over a large range of pH (4 to 9), when more than 20 ppm of FA is added to solution, Hg becomes very soluble. The authors pointed out that Hg interacts with fulvic acid in partly hydrolyzed forms.
It is known that Hg-organic complexes are formed from reaction of mercuric compound solutions (Lövgren & Sjöberg, 1989; Ramamoorthy & Rust, 1976). However, no information is found in the literature to predict the complex formation when metallic mercury is brought into contact with organic-rich solutions, as occurs when Hg is condensed from vapour or amalgamation tailing is dumped into Amazon creeks that bear sediments with high organic levels. Since natural organic acids have extremely variable chemical composition, thermodynamic data on metallic-complexes are difficult to estimate.
Fig. 5 - Relative predominance of the complex Hg-fulvic acid in relation to Hg(OH)2 (Source: Xu and Allard, 1991)
Lövgren and Sjöberg (op.cit.) evaluated experimentally the formation of complexes from reacting a mixture of organic compounds found in bog waters with HgCl2. The ligand is represented as a diprotic acid H2L. Two complexes are formed, HgL and Hg(H-1 L)- according to the reactions:
.............
...............(3)
.. 
......(4)
Using the the solubility constants that the authors calculated for HgCl2o (aq) and H2L equilibrium, we calculated, the potential for the Hgo(aq)/complex equilibria. The Nernst Equations become:
............................................ (5)
.................................... (6)
The equations 5 and 6 are plotted in Figure 6. The upper line represents equilibrium between complexes and Hgo (aq) in common Amazonian darkwaters, in which the dissolved organic concentration is around 10-4 M (Walker, 1990). In this situation, the redox potential of acidic waters (pH 4 to 5.5) must be above 0.4 V to favour Hg-organic complex formation. The higher the pH, the lower the redox potential necessary to form Hg-complexes.
Fig. 6 - Equilibrium boundaries of Hgo(aq) and Hg-organic complexes.
Organic-rich solutions, such as interstitial waters, when in contact with metallic mercury, can form soluble complexes at lower Eh levels than those observed in the Eh-pH diagram for inorganic species (Fig. 4). As we do not know the chemical composition of the organic acid which provides the ligand that complexes with mercury, we assumed that the molecular weight is 1000 g and there is 100 g/l (or 0.1 M) of this ligand in the contaminated sediment. This is a reasonable assumption for interstitial waters of organic sediments according to Dr. L. Lowe (Dept. Soil Science, UBC, personal comm.). This condition is represented in Fig. 6 as the second full line.
The dotted lines in the Fig. 6 represents a situation in which the concentration of the Hg complex in the interstitial water is 1000 times lower than the Hgo(aq) concentration. If we consider a Hgo (aq) concentration of 63 ppb, then in this situation, Hg-complex concentration would be 0.063 ppb. This level is close to background for natural waters (Fitzgerald, 1979). It can be reasonably assumed that under these condition, there is no likelihood of Hg bioaccumulation or danger since no significant complex concentration exists in solution.
Fresh water, even with low concentrations of ligand in contact with metallic mercury, at equilibrium with Hgo (aq) at the metal surface, might favour the stability of Hg complexes at low levels of redox potential. The Eh-pH conditions measured in bottom sediments of Poconé from June 1990 (dry season) until January 1991 showed that Eh ranges from 0.07 to 0.38 V and pH from 5.8 to 7.6.The highest measurements were obtained in the rainy season (Silva et al. 1991). Fig. 6 shows that Hg complex formation is possible in most investigated environments.
In order to investigate the possibility of Hg-complex formation, we contacted 5 ml of metallic mercury in 5 different vials with 30 ml of 0.1 M tannic acid solution. The contact surface of metallic mercury with tannic solution was 7 cm2. The solutions were agitated manually once per week to crudely simulate the natural conditions to which dumped Hg might be subjected in contact with organic-rich interstitial waters. Mercury concentration in solution and Eh were analyzed at the beginning of the test, 7, 21 , 44 and 100 days of contact. At the indicated date, the solutions were siphoned, centrifuged and and Hg analysis conducted by flameless atomic absorption spectrometry. The pH changed slightly from 3.2 to 2.8 after 100 days.
As shown in Fig. 7, the Hg concentration increase was accompanied by solution Eh. The Eh measured is actually the result of a mixed potential of complex formation and organic oxidation. As oxygen is likely the main electron donor in the complex formation reaction, Hg oxidation is controlled by oxygen diffusion in water. So, we can conclude that the worst situation for Hg complexation with organics is when "hot spots" exist in shallow creeks with considerable dissolved oxygen available. For deep sediments, the available oxygen is likely to be extremely low and non-replenished.
Fig. 7 - Metallic Hg in contact with tannic acid solution - Hg in solution and Eh variation
How these organic complexes transform into methylmercury is unclear. Bioaccumulation of these complexes likely is feasible through biotic and abiotic processes (Mannio et al., 1986; Verta et al., 1986)since fulvic acids are methyl group donors.
Humic substances, such as fulvic acids can also adsorb Me-Hg produced in the sediments (Hintelmann et al., 1994). This mechanism facilitates Me-Hg transport, and increases residence time of the toxic substance in water (Watras et al., 1994). The toxicity of Me-Hg bound to fulvic acid (a soluble complex) is considered lower than free Me-Hg but is still very toxic (H. Hintelmann, Trent Univ., Peterborough, Ontario - personal communication).
Conclusions
Mercury emission in Amazon derived from informal mining activities represents a significant source of local pollution. About 70 tonnes of Hg are annually emitted in which poor amalgam distillation practice account for significant part of this total.
Stability of metallic mercury mercury condensed from vapour or dumped into watercourses is not high in the presence of organic rich sediments. Formation of Hg-organic complexes is the main process to mobilize Hg from sediments. Transformation of soluble organic complexes into methylmercury (predominant form in fish) by biotic or abiotic mechanisms is still unclear but definitely such complexes increase bioavailability of the pollutant.
Vegetation fires likely release more Hg compounds into the aquatic environment in a much more reactive form than mining activities. This emission source is likely responsible for regional or even continental Hg contamination. Based on this work, monitoring programs are now being established to measure the extent of this source of emission in the Amazon region as well as to control and remedy Hg bioaccumulation in darkwater systems.
References
Abril 1993. Almanaque Abril. Ed. Abril, São Paulo, Brazil, - in Portuguese.
Cleary, D., 1990. Anatomy of the Amazon Gold Rush. Macmillan Ed., Oxford, 250 p.
Stumm,W. and Morgan, J.J., 1981. Aquatic Chemistry. 2ed. John Wiley & Sons, New York, 780p.
Terborgh, J. 1992. Diversity and the Tropical Rain Forest. Sci. American Library, New York.
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