Environ Monit Assess (2009) 157:125–135 DOI 10.1007/s10661-008-0522-0

Mercury fractionation in stream sediments from the Quadrilátero Ferrífero gold mining region, Minas Gerais State, Brazil Eduardo V. V. Varejão · Carlos R. Bellato · Maurício P. F. Fontes

Received: 26 December 2007 / Accepted: 11 September 2008 / Published online: 27 September 2008 © Springer Science + Business Media B.V. 2008

Abstract The Iron Quadrangle (IQ) region, located in the state of Minas Gerais, has been the most important gold producing area in Brazil since the end of seventeenth century. The use of mercury for gold amalgamation in small scale mines has been responsible for large release of Hg to aquatic and terrestrial environments during 300 years of mining. The present work sought to evaluate the fractionation of Hg in stream sediments is the southern region of the IQ by utilizing sequential extraction. Since mobility and availability of Hg are related to its distribution among sediment partitions, fractionation methods provide detailed information on the ecotoxicological impact and risks associated to the presence of Hg in sediments. The total Hg concentration varied from 179.3 to 690.1 μg kg−1 and Hg0 accounted

E. V. V. Varejão · C. R. Bellato (B) Departamento de Química, Universidade Federal de Viçosa, Av. PH Holfs, s/n, 36571-000, Viçosa, Minas Gerais, Brazil e-mail: [email protected] M. P. F. Fontes Departamento de Solos, Universidade Federal de Viçosa, Av. PH Holfs, s/n, 36571-000, Viçosa, MG, Brazil

for the majority at all sample sites, ranging from 42% to 56% of the total. Keywords Mercury · Sediment · Fractionation

Introduction Small-scale gold mining, known as garimpos in Portuguese, is a rudimentary artisan activity which typically involves the use of Hg0 to improve gold recovery by amalgamation in the final stage of the ore dressing process. The obtained amalgam is heated and large quantities of mercury are released to the environment. This practice has been in uses since the time of the Romans (Lacerda and Salomons 1998) and continues in developing countries. Two types of pollution occur: the discarded material is returned to the waterways near the mining sites or when the amalgamation is roasted in open air or in a retort, Hg vapor escapes to the atmosphere (Hacon et al. 1995; Klingerman et al. 2001). Elevated concentrations of Hg in river beds are frequently associated with mining activities (Nriagu 1994; Lacerda 1997; Slowey et al. 2005). Aquatic ecosystems are particularly sensitive to Hg contamination because the biological and physical-chemical conditions can facilitate transformation of different Hg species. A speciation analysis is a critical factor

126

which must be considered for the determination of mobility and bioavailability (Ullrich et al. 2001; Biester et al. 2002; Kim et al. 2004). The ecological and toxicological effects are highly dependent on the speciation of Hg (Clarkson 2002). In aquatic environments, mercury can be encountered in the elemental (Hg0 ), inorganic (Hg+ and Hg++ ) or organic forms (Lindqvist and Rodhe 1985; Loux 1998; Barbosa et al. 2003). Aqueous organic mercury can be divided into two categories: covalent-bounded organomercurials such as monomethylmercury (MMHg, CH3 Hg+ ) and dimethylmercury (DMHg, CH3 HgCH3 ), and compounds of mercury and organic material such as humic substances (Gill and Bruland 1990; Leermakers et al. 2005). It is well known that methylmercury is the most common organomercurial compound, and more toxic that inorganic mercury, affecting enzymes, the nervous system and immunity (Watras and Bloom 1992). According to Nriagu (1994), the cumulative release of mercury to the environment due to the mining of precious metals in the Americas from 1580 to 1900 was estimated to be 257,000 tons. In Brazil, the contamination of aquatic ecosystems related to the use of Hg for gold amalgamation in mining activities is principally found documented in the Amazon region (Martinelli et al. 1988; Filho and Maddock 1997; Malm 1998; Guimarães et al. 1999; Mascarenhas et al. 2004). Although various studies indicate the relation between mining activities and the release of Hg as the source of environmental mercury contamination, in some regions it is controversial as to whether the Hg is release naturally or by anthropogenic means, requiring greater debate on the specific subject (Wasserman et al. 2003). The Iron Quadrangle (IQ), located in the state of Minas Gerais in Southeast Brazil, is a large and well known mineral deposit. Not only are extensive iron deposits found, but Au mineralization is also encountered (Dorr 1969; Ladeira 1988; Vieira 1988, 1991; Ribeiro-Rodrigues 1998; Vial 1988; Oliveira 1998). Since the end of the seventeenth century, the region has been the most important gold producing area. In the southern portion of the IQ, the municipalities of Ouro Preto and Mariana are known for their ancient

Environ Monit Assess (2009) 157:125–135

auriferous regions in Brazil, where both aquatic and terrestrial environments have been contaminated by Hg liberation used for the amalgamation of gold during 300 years of exploration. Sediments contaminated with tailings constitute basins of bioavailable trace elements and act as sources of inorganic mercury, which can be methylated in aquatic environments (Gleyzes et al. 2002; Rytuba 2003). The mobility and bioavailability of trace elements depends not only on their total concentration but also on their different forms of association with distinctive sediment fractions (Filgueiras et al. 2004). Understanding of the Hg geochemical cycle in sediments is essential for the comprehension of modes of transfer mercury to overlaying water and the biota (He et al. 2007). The total concentration is used as criteria for the determination of the ecotoxicological effects of the contamination of sediments assumes that all trace element forms had a uniform impact on the environment; this approach provides limited information on mobility and bioavailability (Tessier et al. 1979). Speciation analyses provide more detailed information on the amounts of elements of interest associated with different partitions (fractions) of sediments (Davidson et al. 1998; Žemberyová et al. 2006; Raposo et al. 2006). The speciation of Hg can be obtained by application of a sequential extraction method used to subdivide the Hg content in sediment into different somewhat soluble species groups (Rubio and Rauret 1996). Due to the fact that different Hg species present varying degrees of solubility, the species present in sediments and their respective proportions in a region contaminated by Hg can greatly influence the release and transport of mercury in the medium (Kim et al. 2004). The objective of the present work was to study the fractionation of Hg in contaminated stream sediments collected in the region between Ouro Preto and Mariana, in the southern portion of the Iron Quadrangle. Up until now, information about the distribution of Hg in sediments in this area remains scarce. The acquisition of fractionation data from the present study aimed to contribute to the understanding of the distribution and mobility of Hg in the region and to infer about the potential environmental impacts.

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127

Material and methods

abandoned cinnabar (HgS) mine was found located near the Botafogo Creek, a tributary of the Tripuí Creek. The Köppen climate classification defines that region as a Maritime Temperate (Cwb). Yearly rainfall is typical of a tropical regime, presenting an average of 1,723 mm year−1 . The wet season (from October to March) represents 89.6% of the annual precipitation, while the dry season (from April to September) represents 10.4% of the total precipitation. The average annual temperature is 18.5◦ C, where January is the hottest month with a maximum average annual temperature of 21.2◦ C, and July the coldest with a minimum average annual temperature of 15.5◦ C (Coelho 1994). Superficial water and stream sediment samples were collected at four sites along the Tripuí creek and Carmo river (from S1 to S4) in March of 2006 (Fig. 1). In all of the sampling sites along the rivers, the sediments observed throughout the river channels contain vestiges from the extensive mining activities developed in the region.

Description of the sample sites The Carmo River Sub-Basin, located in the southeast part of the Iron Quadrangle, is part of the Rio Doce Basin, which comprises a dredging area of 83.400 km2 and includes 222 municipalities and 461 districts. The Carmo River, formed in the Ouro Preto municipality by junction of the Tripuí and Funil creeks, has been affected by ancient gold mining activities (e.g., Passagem de Mariana, Chico Rei, Scliar, Quartzito and Lajes). The Passagem gold mine, located in the Village of Passagem de Mariana, 7 km east of Ouro Preto, produced at least 60 tons of gold between the seventeenth century and 1954. Along the Tripuí creek and Carmo River between the municipalities of Ouro Preto and Mariana, diverse abandoned gold mines are encountered along with the active exploitation of gemstones and metal ores; artisanal small-scale gold mining can still be found. An

419° 45! S

20° 00! S N

Iron Quadrangle 0

State of Minas Gerais

Brazil

44° 00! W

42° 22! W

43°37!30"

Carm o er riv

4

43°22!30"

20°15!00"

43°52!30"

Ouro Preto

15 km

3

C

ek

ek

river mo ar

Mariana 2

1

Sampling site

20°30!00"

cre

re

uí

lc ni Fu

Tri p

SCALE

Tripuí Ecologic Station N

Bauxite mining Bauxite melt Minicipalities border

0

Fig. 1 Study area

2

4 km

0

5

10

15

20km

128

Reagents, solutions and flasks Sampling flasks and glassware was rigorously washed with distilled water, immersed in nitric acid 10% (v/v), and rinsed with distilled/deionized water. All reagents used were of analytical grade (Vetec and Merk) and distilled/deionized water was used to prepare all solutions. Sampling, pretreatment and storage of the samples At each sampling site three to five sub-samples (0–20 cm deep) were collected inside a 1 m2 area using plastic shovels. Samples were transferred to plastic bags and transported to the laboratory in coolers where they were naturally dried, sifted through a to 2 mm sieve, carefully ground and homogenized. No preservatives were added. Physical–chemical parameters of the water quality The “in situ” measurements of physical-chemical parameters related to the water quality were done at all sample sites. A digital pH meter (WTW, model 340i) was used to determine acidity or alkalinity. Redox potential, conductivity, temperature, total dissolved solids and salinity were measured using a portable device (SCHOT, model handylab LF1) and the dissolved oxygen concentrations were determined using an oxygen sensitive membrane electrode (HANNA, model HI-9142). Sediment analyses The pH, organic matter, cation exchange capacity (CEC) and granulometry of the samples were determined in accordance with EMBRAPA (EMBRAPA 1997). Part of the sieved sediment sample (<2 mm) was submitted to chemical, physical and mineralogical characterization. Difratograms were obtained using powder X-ray diffraction, performed with a Rigaku D-Max difratometer equipped with a cobalt tube (Co-Kα radiation, λ = 1.790269), graphite curved crystal monochromator and operated at 40 kV and 30 mA. A scan was performed using step-by-step method, on the interval

Environ Monit Assess (2009) 157:125–135

between 4◦ and 50◦ 2θ, with an increment of 0.05◦ and time of 1 s per step. Another part of the samples was sieved to <0.63 mm and submitted to Hg fractionation, using the sequential extraction method in six steps as proposed by Lechler et al. (1997). Digestion of the sediments was done using a microwave oven (MILESTONE, EthosPlus model). Determination of Hg concentrations was executed by cold vapor atomic spectrometry, using a VARIAN SpectrAA 200 spectrometer coupled to a VARIAN VGA 77 vapor generator. The complete procedure for sequential extraction is reported below: Step 1: 0.25 g of sample was transferred to a Teflon® digestion tube where 9 mL of HNO3 , 5 mL of HF and 2 mL of HCl were added. This mixture was submitted to acid digestion in a microwave oven, starting at ambient temperature and increasing 180◦ C at 5 min intervals until reaching 180◦ C where the temperature was maintained for 10 min with a power of 1,000 W. The extracts were diluted with 3 mol L−1 HCl and the total Hg content was determined. Step 2: Two 10.0 g portions of each sample were collected in two 250 mL beakers and placed in a block digestor for 48 h at 180◦ C to liberate elemental Hg. After heating, the sample in one of the two beakers was submitted to acid digestion as reported in step 1. The difference between the Hg detected in this second step and total Hg (step 1) corresponds to elemental Hg. The sample in the second beaker was used for the detection of other Hg species in the following steps. Step 3: 50 mL of 0.5 mol L−1 MgCl2 was added to the beaker containing the sample that was previously heated and the mixture was agitated at room temperature for 2 h and left standing for 60 min. Forty milliliters of the supernatant was decanted and centrifuged and used for the determination of exchangeable Hg. The residue was washed with 50 mL of water, left standing for 60 min or until

Environ Monit Assess (2009) 157:125–135

the supernatant became clear, and the water was then discarded. The residue was submitted to step 4. Step 4: Utilizing the residue from step 3, 50 ml of 0.5 mol L−1 HCl was added and this mixture was agitated for 2 h at room temperature and left to decant for 60 min. The following proceedings were performed as reported for step 3, and the supernatant was analyzed for the determination of strongly-bound Hg. Step 5: 25 mL of 0.2 mol L−1 NaOH was added to the residue from step 4 and agitated for 2 h at room temperature. Twentyfive milliliters of 4% (v/v) glacial acetic acid was added, followed by extraction just as in steps 3 and 4. From the extracts obtained in step 5, organic Hg was determined. Step 6: Subtraction of the elemental, exchangeable, strongly bound and organic Hg contents from the total Hg level, permitted for the calculation of residual Hg (encapsulated in Hg silica or sulfides).

Results and discussion Results of the “in situ” physical–chemical parameters for water quality are presented in Table 1. Total Hg percentages and the values obtained by mercury fractionation can be found in Table 2. The total Hg content showed a sharp decrease downstream, ranging from 690.2 to 179.7 μg kg−1 , but the percentage of the different Hg fractions remained similar among the studied sites (Fig. 2).

Table 1 Physical– chemical parameters related to water quality

a Values

established by Brazilian Environmental Standards to human supply destined waters (CONAMA 2005)

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Elemental mercury was the predominant species, representing 42% to 56 % of the total. Similar results were obtained in studies conducted in other gold mining regions of Brazil (Pestana et al. 2000). Eh, pH and organic matter content proved to be important parameters, exerting control on the speciation of Hg and its distribution in aquatic environments. The field values of pH and Eh varied from 6.9 to 7.6 and from 184 to 281 mV, respectively. Within these intervals, the Hg0 specie was expected to be predominate (Brookins 1988), as encountered in the present work and also observed in previous studies (CETEM 1992; Silva et al. 1993; Pestana et al. 2000). According to Lacerda and Salomons (1991), Hg found at processing sites is likely to occur in the metallic form. The presence of significant amounts of mercury and the prevalence of the elemental form are evidence of the release of Hg from gold mining in the riverbeds of the region. Environmental and human health risks related to the presence of Hg in sediments is of great concern. Since Hg0 has a relatively low solubility in water, it persists in the medium (Lindqvist and Rodhe 1985; Wang et al. 2004) and remains available for chemical and biological transformations (Biester et al. 2002). Methylation is an important step in which Hg enters the aquatic food chain and the ingestion of fish contaminated by Hg, the greatest bioaccumulator of MMHg in aquatic systems, representing an important source of human intoxication in certain regions (Hempel et al. 1995; Gray et al. 2000; Hess 2002; Barbosa et al. 2003; Krantz and Dorevitch 2004). In addition to the predominant elemental Hg, significant amounts of exchangeable, strongly

Water physicochemical parameters pH Redox potential (Eh), mV Temperature, ◦ C Electrical conductivity, μS cm−1 Total dissolved solids (TDS), mg L−1 Salinity Dissolved oxygen (DO), mg O2 L−1

BESa

Sampling sites S1

S2

S3

S4

7.6 281 15.6 1484 1414 0.5 14.9

7.5 210 18.1 512 486 0 9.3

7.3 233 20.8 263 250 0 8.4

6.9 184 20.4 180 171 0 11.3

6.0–9.0

< 500 < 0.5 >5

130 Table 2 Mercury fractionation in sediment samples

a Mean of three repetitions ± standard deviation

Environ Monit Assess (2009) 157:125–135 Concentration (μg kg−1 )a

Hg fractions

Sampling sites Elemental Exchangeable Strongly-bounded Organic Residual Total

S1

S2

S3

S4

327.8 ± 14.5 114.5 ± 9.1 159.9 ± 5.7 5.84 ± 0.8 81.9 ± 6.2 690.2 ± 56.5

198.4 ± 19.9 54.6 ± 2.6 45.5 ± 1.9 12.7 ± 1.6 82.4 ± 5.3 392.6 ± 32.6

133.6 ± 12.7 42.7 ± 3.6 30.1 ± 1.3 6.6 ± 0.5 31.8 ± 1.9 243.9 ± 12.1

79.9 ± 7.1 38.2 ± 2.6 29.8 ± 1.4 17.7 ± 1,5 19.4 ± 1.6 179.7 ± 19.1

bound and organic Hg were found in all sample sites. These fractions correspond to the Hg associated to diverse sediment components such as Fe and Mn oxyhydroxides, a portion of organically bonded Hg which is exchanged through protonation of organic sites, and liberated from surface mineral sites where it is found strongly adsorbed (Pestana et al. 2000). Since elemental mercury is sensitive to redox variations (Stein et al. 1996), Hg0 present in sediments may suffer oxidation, leading to the formation of Hg2+ (Windmöller et al. 1996; Wang et al. 2003; Valle et al. 2006), which presents great susceptibility to be absorbed by metallic oxihydroxides and organic superficies (Ribeiro et al. 1999; Rytuba 2003; Sanei and Goodarzi 2006).

Fig. 2 Mercury distribution (in percentage) among the different sediment fractions. S1 to S4 Sampling sites

The physical–chemical characteristics of the sediments are shown in Table 3. Despite being predominantly negatively charged as confirmed by the negative pH values (difference between pHH2O and pHKCl ), the evaluated sediment samples did not show a relatively high cationic exchangeable capacity (CEC) which might be explained by the absolute prevalence of the sand fraction in all samples. However, presence of the silt fraction containing secondary minerals such as kaolinite, gibbsite, goethite and hematite (Fig. 3) explains the absorption of Hg, present as exchangeable and strongly bound species. Despite the importance of organic material for the distribution of Hg in sediments to be known (Gibbs 1973; Orem et al. 1986; Chin and Gschwend 1991),

100 90 80 70

Percentage

60 50 40 30 20 10 0

1

4

3

2

Sediment Samples Elemental

Exchangeable

Strongly-bonded

Oxidizable

Residual

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131

Table 3 Physicochemical characterization of the sediments collected along the Tripuí Creek and Carmo River  pH

pH S1 S2 S3 S4

H2 O

KCl

6.58 6.04 5.77 5.62

7.04 6.97 6.33 6.47

a Cation

Organic matter

Total CECa

Effective CECa

Granulometry (%)

(mg kg−1 ) −0.46 −0.93 −0.56 −0.85

3.10 9.83 7.93 5.86

0.80 0.65 0.76 0.88

0.71 0.50 0.47 0.39

Coarse sand

Fine sand

Silt

Clay

50 80 65 53

41 12 24 44

4 4 8 2

5 4 3 1

exchange capacity

organic Hg corresponds to a small percentage of the total justified by the low levels of organic matter present in the samples as well as low clay mineral content (Table 3). The low percentages of clay and organic material in sediments are due to the intense leaching process which occurs during gold mining. Although monomethylmercury (MMHg) is considered the most toxic species of Hg (Barbosa et al. 2003; Krantz and Dorevitch 2004), it has been demonstrated that this organic form of mercury corresponds to a insignificant fraction of the total Hg present in sediment samples, (Bonzongo et al. 1996; Leermakers et al. 2005) and therefore was not considered during the speciation process proposed by Lechler et al. (1997).

Fig. 3 X-ray diffraction spectra of the sediment samples. S1 to S4 Sampling sites

Residual Hg corresponds to a fraction of mercury encapsulated in silicate minerals or Hg sulfide (Lechler et al. 1997). The presence of Hg sulfide in sediments is generally related to natural occurrences or related to “in situ” production where sulfate-reducing bacteria increase their activity sediments (Lechler et al. 1997). An abandoned cinnabar (HgS) mine is located on the bank of the Botafogo creek, a tributary of the Tripuí creek. Since cinnabar is a fragile mineral and easily broken into small particles (Fernández-Martinéz et al. 2005), its transport and deposit downstream may be responsible for the presence of Hg sulfide present in the sediments analyzed, detected as residual fraction.

Quartz + Mica

Quartz Goethite

500

Goeth. + Hemat.

Gibbsite Kaolinite

Kaolinite Mica

Hematite Feldspar

Goethite

Mica

400

Intensity

S1

300 S2

200 S3

100

0

10000

20000

30000 2θ

40000

50000

132

Small size particles, which present a larger surface area per volume, are capable of retaining large quantities of trace elements by adsorption and precipitation processes on the particle surfaces. For this reason, mercury and other heavy metals tend to be concentrated in fine particles present in sediments such as clay minerals and organic material (Förstner and Wittman 1979; Salomons and Förstner 1984; Horowitz and Elrick 1987; Wang and Chen 2000; Boszke et al. 2004; Chen et al. 2007). All analyzed samples showed very low levels of clay and organic material, confirming that the collected sediments had a low adsorption capacity. In its elemental form, mercury released from process sites showed relatively low solubility and mobility, tending to be concentrated in the vicinity of its release site signifying that Hg transport in aquatic mediums is controlled by sediment suspensions (Lacerda and Salomons 1991; Filho and Maddock 1997; Pestana et al. 2000; Straaten 2000). In the present work, in spite of the considerable quantities of exchangeable and strongly bound Hg detected in the samples, it can be assumed that the low adsorption capacity of the evaluated sediments allowed that a large portion of Hg2+ formed by oxidation of Hg0 may be released. The capacity of Hg2+ to complex with suspended particulate matter (Cranston and Buckley 1972) increases the mobility of mercury in the studied waterways.

Conclusions The total Hg contents and its fractionation in sediments collected from the Tripuí creek and the Carmo River in the southern part of the Iron Quadrangle were evaluated. The region constitutes the oldest Brazilian gold mining area, where Hg was historically used for the amalgamation of gold in small-scale mining activities. Sediment samples collected showed total Hg levels varying between 179.7 and 690.2 μg kg−1 . Although smallscale mining activities have diminished in recent decades due to the scarceness of easily extractable minerals and the more rigorous inspection and control of activity by local authorities, elemental

Environ Monit Assess (2009) 157:125–135

mercury was detected as the predominant form (42% to 56%). The residual fraction present in sediments can be attributed to Hg sulfides descendant from a cinnabar mine located upstream and transported with particulate material to be deposited downstream. Low percentages of clay and silt confer a poor adsorption capacity of the studied sediments, which is confirmed by the low cationic exchangeable capacity (CEC) values observed. However, the incidence of secondary minerals can be responsible for the presence of Hg2+ species. On the other hand, since significant quantities of Hg2+ have been detected as exchangeable and strongly bound Hg, the relatively low absorption capacity of the analyzed sediments permits for the assumption that a large portion of Hg2+ produced from the oxidation of Hg0 does not remain in the sediments. This fact, in association with the high complexation capacity of Hg2+ with suspended particulate material, may be responsible for the increase in mobility of Hg in waterways of the studied region. Acknowledgements We thank the Fundação de Amparo à Pesquisa do Estado de Minas Gerais (FAPEMIG) and the Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq), Brazil, for financial support.

References Barbosa, A. C., Souza, J., Dórea, J. G., Jardim, W. F., & Fadini, P. S. (2003). Mercury biomagnification in a tropical black water, Rio Negro, Brazil. Archives of Environmental Contamination and Toxicology, 45, 235–246. doi:10.1007/s00244-003-0207-1. Biester, H., Müller, G., & Schöler, H. F. (2002). Binding and mobility of mercury in soils contaminated by emissions from chlor-alkaki plants. The Science of the Total Environment, 284, 191–203. doi:10.1016/ S0048-9697(01)00885-3. Bonzongo, J. C., Heim, K. J., Warwick, J. J., & Lyons, W. B. (1996). Mercury levels in surface waters of the Carson River–Lahontan Reservoir system, Nevada: Influence of historic mining activities. Environmental Pollution, 92, 193–201. doi:10.1016/0269-7491(95)00102-6. Boszke, L., Kowalski, A., & Siepak, J. (2004). Grain size partitioning of mercury in sediments of the Middle Odra River (Germany/Poland). Water, Air, and Soil Pollution, 159, 125–138. doi:10.1023/B: WATE.0000049171.22781.bd.

Environ Monit Assess (2009) 157:125–135 Brookins, D. G. (1988). Eh –pH diagrams for geochemistry (p. 176). New York: Springer. CETEM (1992). In L. H. Farid (Ed.), Preliminary diagnosis of the environmental impacts caused by gold prospecting in Alta Floresta (MT) (p. 185). Rio de Janeiro: Centro de Tecnologia Mineral. Chen, C.-W., Kao, C.-M., Chen, C.-F., & Dong, C.-D. (2007). Distribution and accumulation of heavy metals in the sediments of Kaohsiung Harbor, Taiwan. Chemosphere, 66, 1431–1440. doi:10.1016/j. chemosphere.2006.09.030. Chin, Y., & Gschwend, P. M. (1991). The abundance, distribution, and configuration of porewater organic colloids in recent sediments. Geochimica et Cosmochimica Acta, 55, 1309–1317. doi:10.1016/00167037(91)90309-S. Clarkson, T. W. (2002). The three modern faces of mercury. Environmental Health Perspectives, 110, 11–23. Coelho, I. G. D. (1994). Clima. In P. R. J. Alvim (Ed.), Desenvolvimento Ambiental de Ouro Preto– Microbacia do Ribeirão do Funil. MG. Belo Horizonte: CETEC/IGA-SAE/PR. CONAMA–Conselho Nacional de Meio Ambiente (2005). Resolução no. 357, Ministério do Desenvolvimento Urbano e Meio Ambiente, Brasília, Brazil. Cranston, R. E., & Buckley, D. E. (1972). Mercury pathways in river and estuary. Environmental Science & Technology, 6, 274–278. doi:10.1021/es60062a007. Davidson, C. M., Duncan, A. L., Littlejohn, D., Ure, A. M., & Garden, L. M. (1998). A critical evaluation of the three-stage BCR sequential extraction procedure to asses the potential mobility and toxicity of heavy metals in industrially-contaminated land. Analytica Chimica Acta, 363, 45–55. doi:10.1016/ S0003-2670(98)00057-9. Dorr, J. V. N., II (1969). Physiographic, stratigraphic and structural development of the QF, Brazil. Prof. Paper 641-A (pp. 109). Washington: DNPM/USGS. EMBRAPA (1997). Manual de métodos de análise do solo (2nd edn.). Rio de Janeiro: Centro Nacional de Pesquisa de Solos. Fernández-Martinéz, R., Loredo, J., Ordóñez, A., & Rucandio, M. I. (2005). Distribution and mobility of mercury in soils from an old mining área in Mieres, Asturias (Spain). The Science of the Total Environment, 346, 200–212. doi:10.1016/j.scitotenv.2004. 12.010. Filgueiras, A. V., Lavilla, I., & Bendicho, C. (2004). Chemical sequential extraction for metal partitioning in environmental solid samples. Journal of Environmental Monitoring, 4, 823–857. doi:10.1039/ b207574c. Filho, S. R., & Maddock, J. E. L. (1997). Mercury pollution in two gold mining areas of the Brazilian Amazon. Journal of Geochemical Exploration, 58, 231–240. doi:10.1016/S0375-6742(97)00006-X. Förstner, U., & Wittman, G. T. (1979). Metal pollution in the aquatic environment (p. 486). Berlin: Springer.

133 Gibbs, R. J. (1973). Mechanisms of trace metal transport in rivers. Science, 181, 71–73. doi:10.1126/science.180. 4081.71. Gill, G. A., & Bruland, K. W. (1990). Mercury speciation in surface freshwater systems in California and other areas. Environmental Science & Technology, 24, 1392– 1400. doi:10.1021/es00079a014. Gleyzes, C., Tellier, S., & Astruc, M. (2002). Fractionation studies of trace elements in contaminated soils and sediments: A review of sequential extraction procedures. Trends in Analytical Chemistry, 21, 451–467. doi:10.1016/S0165-9936(02)00603-9. Gray, J. E., Theodorakos, P. M., Bailey, E. A., & Turner, R. R. (2000). Distribution, speciation and transport of mercury in stream-sediment, stream-water, and fish collected near abandoned mercury mines in southwestern Alaska, USA. The Science of the Total Environment, 260, 21–33. doi:10.1016/S00489697(00)00539-8. Guimarães, J. R. D., Forstier, A. H., Forti, M. C., Melfi, J. A., Kehrig, H., Mauro, J. B. N., et al. (1999). Mercury from human and environmental samples from two lakes in Amapa, Brazilian Amazon. Ambio, 28, 196–301. Hacon, S., Artaxo, P., Gerab, F., Yamasoe, M. A., Campos, R. C., Conti, L. F., et al. (1995). Atmospheric mercury and trace elements in the region of Alta Floresta in the Amazon Basin. Water, Air, and Soil Pollution, 80, 273–283. doi:10.1007/BF01189677. He, T., Lu, J., Yang, F., & Feng, X. (2007). Horizontal and vertical variability of mercury species in pore water and sediments in small lakes in Ontario. The Science of the Total Environment, 386, 53–64. doi:10.1016/j.scitotenv.2007.07.022. Hempel, M., Chau, Y. K., Dutka, B. J., McInnis, R., Kwan, K. K., & Liu, D. (1995). Toxicity of organomercury compounds: Bioassay results as a basis for risk assessment. Analyst (London), 120, 721–724. doi:10.1039/an9952000721. Hess, E. V. (2002). Environmental chemicals and autoimmune disease: Cause and effect. Toxicology, 181–182, 65–70. doi:10.1016/S0300-483X(02)00256-1. Horowitz, A. J., & Elrick, K. A. (1987). The relation of stream sediment surface area, grain size and composition to trace element chemistry. Applied Geochemistry, 2, 437–451. doi:10.1016/0883-2927(87)90027-8. Kim, C. S., Rytuba, J. J., & Brown, G. E., Jr. (2004). Geological and anthropogenic factors influencing mercury speciation in mine wastes: An EXAFS spectroscopy study. Applied Geochemistry, 19, 379–393. doi:10.1016/S0883-2927(03)00147-1. Klingerman, D. C., La Rovere, E. L., & Costa, M. A. (2001). Management challenges on small-scale gold mining activities in Brazil. Environmental Research, 87, 181–198. doi:10.1006/enrs.2001.4301. Krantz, A., & Dorevitch, S. (2004). Metal exposure and common chronic diseases: A guide for the clinician. Disease-a-Month, 50, 215–262. doi:10.1016/j. disamonth.2004.04.001.

134 Lacerda, L. D. (1997). Global mercury emissions from gold and silver mining. Water, Air, and Soil Pollution, 97, 209–221. Lacerda, L. D., & Salomons, W. (1991). Mercury in the Amazon: A chemical time bomb? Dutch Ministry of housing, physical planning and environment. Chemical time bomb project (p. 46). Amsterdam, The Netherlands. Lacerda, L. D., & Salomons, W. (1998). Mercury form gold and silver mining: A chemical time bomb? (p. 146). Berlin: Springer. Ladeira, E. A. (1988). Metalogenia dos depósitos de ouro do Quadrilátero Ferrífero, Minas Gerais, Brasil. In C. Schobbenhaus & C. E. S. Coelho (Eds.), Principais Depósitos Minerais do Brasil (pp. 301–375). Brasília: DNPM/CVRD, 3. Lechler, P. J., Miller, J. R., Hsu, L.-C., & Desilets, M. O. (1997). Mercury mobility at the Carson River superfund site, west-central Nevada, USA: Interpretation of mercury speciation data in mill tailings, soils, and sediments. Journal of Geochemical Exploration, 58, 259–267. doi:10.1016/S0375-6742(96)00071-4. Leermakers, M., Baeyens, W., Quevauviller, P., & Horvat, M. (2005). Mercury in environmental samples: Speciation, artifacts and validation. Trends in Analytical Chemistry, 24, 383–393. doi:10.1016/j.trac.2004.01.001. Lindqvist, O., & Rodhe, H. (1985). Atmospheric mercury: A review. Tellus, 37B, 136–159. Loux, N. T. (1998). An assessment of mercury-speciesdependent binding with natural organic carbon. Chemical Speciation and Bioavailability, 10, 127–136. doi:10.3184/095422998782775754. Malm, O. (1998). Gold mining as a source of mercury exposure in the Brazilian Amazon. Environmental Research, 77, 73–78. doi:10.1006/enrs.1998.3828. Martinelli, L. A., Ferreira, L. R., Forsberg, B. R., & Victoria, R. L. (1988). Mercury contamination in the Amazon: A gold rush consequence. Ambio, 17, 252–254. Mascarenhas, A. F. S., Brabo, E. S., Silva, A. P., Fayal, K. F., Jesus, I. M., & Santos, E. C. O. (2004). Avaliação da concentração de mercúrio em sedimentos e material particulado no rio Acre, Estado do Acre, Brasil. Acta Amazonica, 34, 61–68. doi:10.1590/ S0044-59672004000100008. Nriagu, J. (1994). Mercury pollution from the past mining of gold and silver in America. The Science of the Total Environment, 149, 167–181. doi:10.1016/00489697(94)90177-5. Oliveira, F. R. (1998). Contribuição ao estudo da geologia estrutural e da gênese do depósito aurífero de Passagem de Mariana-MG. MSc Thesis, Unicamp, São Paulo. Orem, W. H., Hatcher, P. G., Spiker, E. C., Szeverenyi, N. M., & Maciel, G. E. (1986). Dissolved organic matter in anoxic waters from Mangrove Lake, Bermuda. Geochimica et Cosmochimica Acta, 69, 4881–4894. Pestana, M. H. D., Lechler, P., Formoso, M. L. L., & Miller, J. (2000). Mercury in sediments from gold and copper exploration áreas in the Camaquã River Basin, Southern Brazil. Journal of South American

Environ Monit Assess (2009) 157:125–135 Earth Sciences, 13, 537–547. doi:10.1016/S0895-9811 (00)00039-0. Raposo, J. C., Zuloaga, O., Sanz, J., Villanueva, U., Crea, P., Extrebaria, N., et al. (2006). Analytical and thermodinamical approach to understand the mobility retention of arsenic species from the river to estuary. Marine Chemistry, 99, 42–51. doi:10.1016/j.marchem. 2005.02.004. Ribeiro, M. G., Jr., Silva Filho, E. V., Souza, M., & Lacerda, L. D. (1999). Mercury burden in soils from Central Amazon. In Book of abstract of the 5th international conference mercury as a global pollutant. Rio de Janeiro, Brazil. Ribeiro-Rodrigues, L. C. (1998). Gold mineralization in Archean banded iron-formation of the QF, Minas Gerais, Brazil. The Cuiaba Mine. PhD Thesis, RTWH Aachen. Rubio, R., & Rauret, G. (1996). Validation of the methods for heavy metal speciation in soils and sediments. Journal of Radioanalytical and Nuclear Chemistry, 208, 529–540. doi:10.1007/BF02040070. Rytuba, J. J. (2003). Mercury from mineral deposits and potential environmental impact. Environmental Geology, 43, 326–338. Salomons, W., & Förstner, U. (1984). Metals in hydrocycle (p. 349). Berlin: Springer. Sanei, H., & Goodarzi, F. (2006). Relationship between organic matter and mercury in recent lake sediment: The physical-geochemical aspects. Applied Geochemistry, 21, 1900–1912. doi:10.1016/j.apgeochem.2006. 08.015. Silva, A. P., Ferreira, N. L. S., Pádua, H. B., Veiga, M. M., Silva, G. D., Castro e Silva, E. O., et al. (1993). Mobilidade do mercúrio no Pantanal de Poconé. Ambiente, 7, 52–56. Slowey, A. J., Rytuba, J. J., & Brown, G. E., Jr. (2005). Speciation of mercury and mode of transport from placer gold mine tailings. Environmental Science & Technology, 39, 1547–1554. doi:10.1021/es049113z. Stein, E. D., Cohen, Y., & Winer, A. M. (1996). Environmental distribution and transformation of mercury compounds. Critical Reviews in Environmental Science and Technology, 26, 1–43. Straaten, P. V. (2000). Mercury contamination associated with small-scale gold mining in Tanzania and Zimbabwe. The Science of the Total Environment, 259, 105–113. doi:10.1016/S0048-9697(00)00553-2. Tessier, A., Campbell, P. G. C., & Bisson, M. (1979). Sequential extraction procedure for the speciation of particulate trace metals. Analytical Chemistry, 51, 844–851. doi:10.1021/ac50043a017. Ullrich, S. M., Tanton, T. W., & Abdrashitova, S. A. (2001). Mercury in the aquatic environment: A review of the factors affecting methylation. Critical Reviews in Environmental Science and Technology, 31, 241–293. doi:10.1080/20016491089226. Valle, C. M., Santana, G. P., & Windmöller, C. C. (2006). Mercury conversion process in Amazon soils evaluated by thermodesorption analysis. Chemosphere, 65, 1966–1975. doi:10.1016/j.chemosphere.2006.07.001.

Environ Monit Assess (2009) 157:125–135 Vial, D. S. (1988). Mina de ouro de Cuiabá, QF, Minas Gerais. In C. Schobbenhaus & C. E. S. Coelho (Eds.), Principais depósitos minerais do Brasil (Vol. 3, pp. 413–419). Brasilia: DNPM/CVRD. Vieira, F. W. R. (1988). Processos epigenéticos de formação dos depósitos auríferos e zonas de alteração hidrotermal do Grupo Nova Lima, Quadrilátero Ferrífero, Minas Gerais. In Proc 35th congr Brasileiro de geologia, 6–13 November 1988 (Vol. 1, pp. 76–87) Belém. Sociedade Brasileira de Geologia, Belém. Vieira, F. W. R. (1991). Textures and processes of hydrothermal alteration and mineralization in the Nova Lima Group, Minas Gerais, Brasil. In E. A. Ladeira (Ed.), The economics, geology, geochemistry and genesis of gold deposits. Proc symp Brazil gold’91, 13–17 May 1991 (pp. 319–325). Belo Horizonte. Balkema, Rotterdam. Wang, F., & Chen, J. (2000). Relation of sediment characteristics to trace metal concentrations: A statistical study. Water Research, 34, 694–698. doi:10.1016/S00431354(99)00184-0. Wang, Q., Kim, D., Dionysiou, D. D., Sorial, G. A., & Timberlake, D. (2004). Sources and remediation of mercury contamination in aquatic systems—a litera-

135 ture review. Environmental Pollution, 131, 323–336. doi:10.1016/j.envpol.2004.01.010. Wang, D., Shi, X., & Wei, S. (2003). Accumulation and transformation of atmospheric mercury in soil. The Science of the Total Environment, 304, 209–214. doi:10.1016/S0048-9697(02)00569-7. Wasserman, J. C., Hacon, S., & Wasserman, M. A. (2003). Biogeochemistry of mercury in the Amazonian environment. Ambio, 32, 336–342. doi:10.1639/00447447(2003)032[0336:BOMITA]2.0.CO;2. Watras, C. J., & Bloom, N. S. (1992). Mercury and methylmercury in individual zooplankton: Implications for bioaccumulation. Limnology and Oceanography, 37, 1313–1318. Windmöller, C. C., Wilken, R.-D., & Jardim, W. F. (1996). Mercury speciation in contaminated soils by thermal release analysis. Water, Air, and Soil Pollution, 89, 399–416. doi:10.1007/BF00171644. Žemberyová, M., Barteková, J., & Hagarová, I. (2006). The utilization of modified BCR tree-step sequential extraction procedure for the fractionation of Cd, Cr, Cu, Ni, Pb and Zn in soil reference materials of different origins. Talanta, 70, 973–978. doi:10.1016/ j.talanta.2006.05.057.