Mercury conversion processes in Amazon soils evaluated by thermodesorption analysis

Chemosphere 65 (2006) 1966–1975 www.elsevier.com/locate/chemosphere

Mercury conversion processes in Amazon soils evaluated by thermodesorption analysis Cla´udia M. do Valle a, Genilson P. Santana b, Cla´udia C. Windmo¨ller

c,*

a

c

Gereˆncia de Quı´mica, DE, Centro Federal de Educac¸a˜o Tecnolo´gica do Amazonas, Centro, 69020-120 Manaus, AM, Brazil b Departamento de Quı´mica, ICE, Universidade Federal do Amazonas, Coroado, 69077-000 Manaus, AM, Brazil Departamento de Quı´mica, ICEx, Universidade Federal de Minas Gerais, Campus Pampulha, 31270-901 Belo Horizonte, MG, Brazil Received 30 March 2005; received in revised form 30 June 2006; accepted 3 July 2006 Available online 14 August 2006

Abstract This paper reports on the speciation study and the Hg redox behavior in Amazon soils not influenced by gold mining and collected near Manaus, AM, Brazil. The samples were incubated by adding Hg(0) and HgCl2 to dry soil. Solid phase Hg speciation analysis was carried out using a Hg thermodesorption technique with the aim of distinguishing elemental Hg(0) from Hg(II) binding forms. In the first case, we observed the conversion of Hg(0) to Hg(II) binding forms in the range of 28–68% and a correlation between the percent of oxidation and OM content. Samples incubated with Hg(II) showed the formation of Hg(I) and/or Hg(0) in the range of 19–69%. The lowest values corresponded to the samples with the lowest clay contents. The kinetics of conversion of Hg(0) as well as HgCl2 were roughly fitted to the two first order reactions, a fast one and a slow one. It was not possible to evaluate differences between sampling sites and types of soils, but the mean half-life of the first order reaction obtained by the addition of Hg(II) was slower (t1/2 = 365 d) than the one obtained by the addition of Hg(0) (t1/2 = 148 d). Previous studies have shown the predominance of organically bound Hg in these samples. Thus, the kinetic difference between Hg oxidation and reduction in combination with the efficient retention processes by OM may explain the high background values found in Amazon soils. Ó 2006 Elsevier Ltd. All rights reserved. Keywords: Mercury; Speciation; Amazon soils; Redox process; Thermodesorption

1. Introduction Depending on the redox conditions, Hg occurs in three different valence states, Hg(0), Hg(I), and Hg(II), Hg(0) and Hg(II) being normally found in soils (Boening, 2000). In addition to the redox potential, pH, Cl ions, and organic matter (OM) concentration are the key parameters in determining speciation of Hg in soil solution and the chemical transformations that might occur (Lin and Pehkonen, 1999; Ravichandran, 2004). In soils that are rich in OM, the Hg transformations frequently occur in the top

*

Corresponding author. Tel.: +55 31 3499 5752; fax: +55 31 3499 5700. E-mail address: [email protected] (C.C. Windmo¨ller).

0045-6535/$ - see front matter Ó 2006 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2006.07.001

horizon. It is continuously remobilized and volatilized as organomercury and Hg(0), respectively (Rocha et al., 2000; Beldowski and Pempkowiak, 2003). Mierle and Ingram (1991) and Barrow and Cox (1992) proposed that OM is important for the transport of Hg from upland soils and wetlands to water and for its retention in soils. The presence of sulfide is also important, because it often forms the very insoluble HgS salt. In addition to chemical reactions, transformations may also be mediated by microbial activity such as methylation (Steinnes, 1997). The knowledge of Hg speciation and reactions (redox, complexation, and others) is important in order to explain the retention and the mobility of this element in soil. In the conversion of elemental Hg to methylmercury, the oxidation of Hg(0) is an essential step. This oxidation is

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

considered as occurring in natural aquatic environments and is strongly affected by the concentration of Cl (Magalha˜es and Tubino, 1995; Yamamoto, 1996; Lin and Pehkonen, 1999). When analyzing soil samples in contact with Hg vapor, Wang et al. (2003) concluded that Hg(0) transformations take place as soon as this element is deposited on the soil. Approximately 25% of the deposited Hg remained in the Hg(0) form, and 75% turned into the socalled active forms: soluble HCl, and organically bound and residual Hg. The oxidation of Hg(0) in contaminated soils and the effects of soil composition have has also been reported (Windmo¨ller et al., 1996; Tho¨ming et al., 2000; Renneberg and Dudas, 2001). Photoreduction is an important property of Hg (Costa and Liss, 2000), as the emission of volatile Hg species from natural soils is believed to be a significant contributor to the atmospheric burden of Hg. Gustin et al. (2002) studied light-enhanced emission of Hg by substrates spiked with pure synthetic Hg species and by naturally and anthropogenically Hg enriched substrates and demonstrated that the influence of light energy on Hg emission by naturally enriched substrates is larger than that of soil temperature. The release of Hg associated to gold mining and deforestation are two of the most important environmental issues in the Amazon basin (Artaxo et al., 2000). Gold mining in the Amazon has been responsible for the release of about 2.000–3.000 tons of Hg over the last 20 years. Annually, about 200 tons of Hg is released into the atmosphere (Lacerda, 1997). Roulet et al. (1998) calculated that more than 97% of the Hg accumulated on the soil surface started before gold mining (last 30 years). In addition, biomass burning in tropical forests also seems to have contributed significantly to the release of Hg to the atmosphere. Discussing Hg emissions in relation to biomass burning in the Amazon basin, Veiga et al. (1994) estimated a large emission of Hg to the atmosphere, at levels of about 90 tons per year. Therefore, Amazon soils seem to be a large regional reservoir of Hg (Fadini and Jardim, 2001). Thus, a better understanding of the species and possible reactions (redox, complexation, and others) of Hg in these soils would help to explain the high Hg levels in soils in regions where there is neither gold mining nor Hg containing minerals. Further research is necessary to evaluate the kinetics of Hg transformations in various types of soils and sediments. However, these studies depend on adequate techniques to determine metal speciation. The number of methods to determine Hg speciation have increased over the years and so have their sophistication, ranging from the visual identification of Hg phases to sequential chemical extractions, to X-ray absorption spectroscopy analysis (Sladek et al., 2002; Kim et al., 2003; Sladek and Gustin, 2003), and to Hg themordesorption atomic absorption spectroscopy (TDAAS) (Windmo¨ller et al., 1996; Biester and Scholz, 1997). The sequential chemical extraction is based on the use of various solutions to extract Hg from different soil fractions.

1967

However, instead of species-specific information, this method provides information on the differentiation of Hg compounds according to behavioral classes, such as water soluble, acetic acid solution, and other basic and acid solutions (Biester and Scholz, 1997; Bloom et al., 2003). X-ray absorption spectroscopy analysis is a nondestructive method that uses high energy synchrotron-sourced X-ray radiation to identify specific species based on scattering patterns. This method is most useful in the identification of specific species as long as they are included in a model database. However, the identification of Hg(0) is difficult (Sladek and Gustin, 2003) and these methods have so high limits of detection (>100 lg g1) that their applications are restricted (Kim et al., 2003). TDAAS identifies Hg species by incremental heating and comparison of thermal release patterns to a compound database. More recently, it has been used in investigations of Hg compounds in soils and sediments (Biester and Zimmer, 1998; Biester et al., 2000; Biester et al., 2002; Higueras et al., 2003). According to Sladek et al. (2002), TDAAS does not look at specific species as EXAFS does, but more at species bound to certain forms and it provide information on elemental Hg and different Hg(II) binding forms. The objective of this work was to study the redox behavior of Hg in soil samples collected in areas not impacted by gold mining near Manaus, AM, Brazil. The soil samples were prepared by incubation of Hg(0) and HgCl2 by dry dilution and the TDAAS technique was used to monitor the Hg(0) and Hg(II) binding forms periodically until the stabilization of the Hg thermodesorption curves. Some physical and chemical characteristics of the samples are also discussed. 2. Soil sampling Soil samples were collected in three different areas of Manaus; BR 174 Highway (Spodosol, called km 10), Adolpho Ducke Reserve (Oxisol, RD3 and Ultisol, RD4), and Industrial District (Ultisol, DI) located in the east of the Amazon State, Brazil. The RD and ID samples were collected at depths 0/20, 20/40, 40/60, and 60/80 cm and the km 10 samples on a slope surface of 10/40 cm (km 10–1), 40/60 cm (km 10–2) and 60/80 cm (km 10–3) from the top. The sampling sites were described in detail as coordinates, horizons, and Munsell color by do Valle et al. (2005). However, for a better understanding of this work, we provided the following data: BR174 is an area located at km 10 of the highway linking Manaus to Boa Vista; Adolpho Ducke Reserve is an area away from any anthropogenic activity, and the Industrial District is an area impacted by industrial activities. None of them has suffered any impact from gold mining activities. The samples were homogenized and freeze-stored in disposable polyethylene bags. A fraction of each sample was preserved in a refrigerator, whereas the remaining part was air-dried to constant weight, manually desegregated, sieved through a 2 mm mesh sieve, and homogenized.

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C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

3. Methods and materials

4. Results and discussion

The soil fraction and texture, pH (water and KCl), total Hg (<2 mm and <0.053 mm fractions), sampling locations, and statistical analysis were previously described by do Valle et al. (2005). The nitrogen (N) and carbon (C) contents of the soils were measured using a CHN Perkin Elmer model 2400 analyzer. The mineralogical composition of the soils from these areas had already been studied in detail (Bravard and Righi, 1988; Righi et al., 1990). No carbonate is expected to be present, thus the C determined represents organic C. The OM content was obtained by multiplying the organic C concentration by a factor of 1.72, given that the OM C content is generally around 58% (Radojevic´ and Bashkin, 1999). Hg thermodesorption curves were determined by means of an in-house apparatus consisting of an electronically controlled heating unit and an Hg detection unit used according to do Valle et al. (2005). For Hg detection, a quartz tube for purging thermally released Hg was placed in the optical system of an atomic absorption spectrometer (GBC 932-AA). Hg was detected at 253.7 nm. Analysis was carried out at a heating rate of 33 °C min1 under nitrogen gas flow of 200 ml min1. Interferences, mainly from pyrolytic OM products, were compensated by continuous deuterium background correction. Sample weight was 50–3000 mg, depending on the total Hg content. Thermograms of Hg standard samples were obtained in triplicate by adding Hg(0), Hg2Cl2, HgCl2 and HgO to thermally pretreated powder silica (500 °C for 2 h) up to the concentration of 30 mg kg1 of total Hg by dry dilution. The preserved fractions of original samples were analyzed in duplicate and in triplicate as necessary. Soil sample (fraction <2 mm) incubation was carried out by adding Hg(0) and HgCl2 up to 30 mg kg1 of total Hg by dry dilution. This procedure was chosen in order to simulate the best possible environmental conditions. Consequently, the addition of any substances to the soil samples was avoided. They were manually macerated and immediately analyzed by TDAAS in the same conditions as the standard samples. Thermograms were obtained at intervals of hours, days, and weeks during one year under the same instrumental condition until the Hg pattern became stable. The samples were kept at room temperature during Hg monitoring to simulate natural conditions. In the beginning of the sample preparation, it was not possible to make replicates of the analyses due to the rapid modifications of Hg interactions with the matrix. The interval between two analyses was 2 h. As soon as the thermograms showed that these changes gradually slowed down, the samples were analyzed in duplicate and in triplicate as necessary. The relative standard deviations of the peak areas obtained were better than 20%. The weights used to obtain original soil sample thermograms in this work were in the order of grams (>2 g), and for the analysis of the spiked samples it was 200 mg so that the signal related to natural Hg was only a small fraction of total signal in incubated samples.

The thermodesorption results of the Hg standard samples are shown in Fig. 1. Hg(0) was released in the range between room temperature and 200 °C with a maximum at 150 °C, Hg(I) compounds were released with a maximum around 200 °C, and Hg(II) compounds were released at higher temperatures. Other studies (Biester and Scholz, 1997; Biester et al., 2002; Sladek et al., 2002) have also shown that the thermodesorption temperatures of different Hg standards are different, indicating that the technique provides data on speciation. Temperature ranges between studies varied slightly due to differences in system operational conditions, mainly heating rate and gas flow. However, all studies agree that Hg(0) is released at a low temperature, followed by the Hg(II) binding forms. The Hg(I) release pattern was shown in only one earlier study (Windmo¨ller et al., 1996), and the temperature range agrees with the one obtained here. Thus, we have considered the thermograms peaks of the analyzed samples up to 150 °C as Hg(0), and as Hg(I) when the maximum occurs at around 220 °C, and as Hg(II) at higher temperatures. The representative thermograms of the 15 original samples are presented in Fig. 2. The Hg peaks that can be seen only between 300 and 400 °C are Hg(II) binding forms. Sometimes the peaks were asymmetric or were split in two peaks (sample RD3-60/80 for example), indicating that Hg had more than one interaction with the soil matrix. A detailed discussion of speciation, concentration, and the influence of physical–chemical parameters on the behavior of natural Hg (Table 1) in these samples was made by do

1.2 1.0 0.8 0.6 0.4 0.2 0.0

HgO

0.8 0.6

Hg release

0.4

HgCl2

0.2 0.0 1.4 1.2 1.0 0.8 0.6 0.4 0.2 0.0 1.2 1.0 0.8 0.6 0.4 0.2 0.0

Hg2Cl 2

Hg(0)

100

200

300

400

500

Temperature °C Fig. 1. Thermograms of standard samples of mercury compounds.

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

1969

0.15 0.3

0.10 0.2 0.1

0.05

km 10-1

0.0

0.10

Hg release

0.3

0.05

0.10

km 10-2

0.0

0.05

0.8

0.00 0.15

0.6

DI-40/60

0.10

0.4 0.2

DI-20/40

0.00 0.15

0.2 0.1

DI-0/20

0.00 0.15

0.05

Km 10-3

DI-60/80

0.00

0.0 100

200

300

400

100

500

200

300

400

500

0.3

0.4 0.3

0.2

0.2 0.1

0.1

RD3-0/20

0.0 0.4

RD4-0/20

0.0 0.3

0.3

0.2

Hg release

0.2 0.1

0.1

RD4-20/40

RD3-20/40

0.0 0.4

0.0 0.3

0.3

0.2

0.2 0.1

0.1

RD3-40/60

RD4-40/60

0.0 0.4

0.0 0.3

0.3

0.2

0.2 0.1

0.1

RD3-60/80

RD4-60/80

0.0

0.0 100

200

300

400

500

Temperature °C

100

200

300

400

500

Temperature °C

Fig. 2. Depth profile thermograms of original soil samples from the three sampling areas.

Valle et al. (2005), but one of the most important conclusions was that thermodesorption analysis revealed the presence of only Hg(II). The comparison of the thermodesorption range of Hg(II) with the result of a previous work (Biester et al., 2002) and a Hg(II) sample incubated in commercial humic acid indicated that it is predominantly organically bound. 4.1. Incubation of soil samples with Hg(0) The representative thermograms of the Hg(0) incubated samples from BR 174 Highway, Adolpho Ducke Reserve and Industrial District are shown in Fig. 3. It is possible

to see that the area of the Hg(0) peak (150 °C) gradually decreases and a peak appears at higher temperatures (>300 °C) and gradually increases. This occurred for all samples, clearly indicating a conversion of Hg(0) to Hg(II). Although no total Hg determination has been carried out, the decrease in the Hg(0) peak is probably partially due to its volatization. The presence of three peaks can be observed at times (Fig. 3, RD3-40/60, 30 d), which may indicate that the added Hg(0) seems to partially oxidize to Hg(I), and later to Hg(II). The similarity between the temperature of the peak between Hg(0) and Hg(II) and the Hg(I) standard, as shown in Fig. 1, reinforces the idea of the presence of this oxidation state.

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Table 1 Sample physical–chemical characterization and Hg natural concentration; percent (%) oxidation and reduction of incubated samples Samples

km 10–1 km 10–2 km 10–3 RD3 RD3 RD3 RD3 RD4 RD4 RD4 RD4 DI DI DI DI

0–20 20–40 40–60 60–80 0–20 20–40 40–60 60–80

0–20 20–40 40–60 60–80

(g kg1)b

Soil fraction (%)a

b c

(%)

Clay

Silt

Sand

C/N

OM

<2 mm

<0.053 mm

Hg(0) oxid.

Hg(II) red.

2.8 3.6 2.7

0.2 0.1 0.3

97.0 96.3 97.0

ndc ndc ndc

0.16 ± 0.01 2.39 ± 0.03 1.29 ± 0.05

<25 764 ± 32 83 ± 2

46 ± 4 2320 ± 20 3420 ± 130

40 68 57

19 10 20

72.9 64.2 71.3 67.4 31.0 35.6 35.3 39.8

4.9 4.5 2.0 5.8 3.7 2.0 5.8 4.9

22.2 31.3 26.7 26.8 65.3 62.4 58.9 55.3

18.16 11.18 16.80 13.31 10.88 9.45 11.90 13.32

28.1 ± 0.8 23.1 ± 0.1 17.4 ± 0.7 16.0 ± 0.7 24.4 ± 0.2 17.9 ± 0.6 16.4 ± 0.6 11.5 ± 0.5

128 ± 14 123 ± 12 122 ± 17 123 ± 20 124 ± 11 120 ± 1 154 ± 9 145 ± 7

244 ± 12 231 ± 3 256 ± 8 271 ± 16 313 ± 3 367 ± 4 475 ± 9 349 ± 32

55 48 44 42 49 41 37 29

56 44 53 51 41 49 50 69

27.8 18.7 15.7 13.1

5.7 3.8 3.6 3.3

66.5 77.5 80.7 83.6

23.85 22.13 31.27 37.03

24.7 ± 0.4 15.3 ± 0.3 16.2 ± 0.4 19.2 ± 0.5

114 ± 12 108 ± 8 99 ± 10 92 ± 4

363 ± 18 494 ± 52 459 ± 17 358 ± 11

57 44 31 28

62 55 69 48

RSD (%) a

Hgtotal (ng g1)a,b

5

8

5

do Valle et al. (2005). Average concentration ± standard deviation (three replicates). nd – Nitrogen not detected.

Season (dry and rainy) variation in the Amazon allows OM to go through diverse transformations, including redox processes (Rocha et al., 2000). It is known that the humic substances (HS) present in the OM can exert an antagonistic competition for Hg(II) ions (Rocha et al., 2000). While functional thiophenolic groups bind strongly to Hg(II) (Xia et al., 1999), semiquinone groups act as Hg(II) reducers (Allard and Arsenie, 1991). Struyk and Sposito (2001) studied the redox properties of standard humic acids (HA) and stated that it is possible that inorganic constituents in HA, such as iron content, contribute to its redox reactions, either in conjunction with semiquinone radicals or through another pathway. They also say that others report a rapid reduction of Fe(III) by HA, suggesting that the iron strongly complexed within the HA structure (as Fe(III)) could be a potential source of oxidation capacity. Based on the results obtained, they forwarded a hypothesis for abiotic electron transfer reactions of HA involving Fe(III) as a mediator. Data on Hg reduction by HS obtained from Rio Negro, in the Amazon region, the same investigated in this study, suggest a larger complexation capacity by strong adsorption sites comparatively to reducing functional sites (Rocha et al., 2000). Sample km 10–1 was the only one that presented two peaks at the end of the monitoring process. The lowtemperature peak is thought to be due to Hg(I), but it might also be due to Hg(II) adsorption site in some mineralogical phase as these Hg species appeared in a temperature interval of 200–300 °C (Fig. 3). According to Biester et al. (2002), Hg species adsorbed on the mineral surface are released in the range of 150–280 °C with a maximum at 200 °C, and above 300 °C Hg will be bound to OM.

The oxidation percentages of the added Hg(0) were calculated through the Hg(II) peak areas of the stabilized samples (Fig. 4(a)). The value of the initial Hg(0) area was considered as 100%, and the peak areas above 250 °C, which correspond to Hg(II), were calculated. One can see in Table 1 that the oxidation in the soil profile decreases with depth in the same way as the OM contents does within the soil profile. The km 10–2 sample, which has the highest OM content of all three samples from km 10, also displayed the highest oxidation percent (68%). Similarly, it can be observed through profiles RD3, RD4, and DI, that the layer the closest to the surface (0/20 cm) and whose OM content is the highest also showed the highest oxidation percents. The Pearson correlation of OM and % oxidation, excluding the km 10 samples whose properties were the most distinct, was significant for the remaining 12 samples (0.78, p = 0.05). This indicates that there is a correlation between OM content and the percent Hg oxidation. This could be explained by the OM complexation capacity, stabilizing the Hg(II) formed, favoring the oxidation of Hg(0) to Hg(II). In order to obtain information on the kinetics of the added Hg(0) conversion, graphs of % conversion as function of days were made (Fig. 4(a)), except for km 10 samples, because in this case too few thermograms were obtained during monitoring. The graphs in Fig. 4(a) show a two-step oxidation process. Both were roughly fitted to the first order reactions. The curves corresponding to the layers at each sampling site did not show difference, and a single fit was made to obtain half-life values for each set of four depths, with two exceptions (Fig. 4(b)). The half-life values were calculated by the expression t1/2 = ln 2/K, where K is the angular coefficient of the linear fit. As the thermograms showed a peak that may be attributed

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

Hg release

1.6 1.2 0.8 0.4 0.0

1.2

km 10-1+Hg(0)

km 10-3+Hg(0)

0.8

1d

0.4

1d

0.0 1.2

1.6 1.2 0.8 0.4 0.0

0.8

21 d

0.4

14 d

0.0

1.2

0.9

0.8

0.6 0.3

0.4

45 d

90 d

0.0

0.0

1.2

0.6

0.8

0.4

0.4

100 d

0.0

0.2

68 d

0.0

100

200

300

400

100

500

200

0.8 0.6

RD3-40/60+Hg(0)

0.4 0.3 0.2 0.1 0.8

0.0

0.5 0.4 0.3 0.2 0.1 0.0

0.4

0.3

0.6 0.4 0.2

300

400

30 d

0.3

500

DI-40/60+Hg(0)

0.4 0.2 0.0

1d

0.0

Hg release

1971

1d

64 d

0.2

0.2 0.1

71 d

0.1 0.0

0.0

0.3

0.3

0.2

112 d

0.2

0.1

137 d

0.0

0.1

152 d

0.0

100

200

300

400

500

Temperature °C

100

200

300

400

500

Temperature °C

Fig. 3. Thermograms of samples incubated with Hg(0).

to Hg(I), it is suggested that in the first step, which is faster (t1/2 from 4 to 11 d), Hg(0) oxidizes to Hg(I), and that in the second step, which is much slower (t1/2 from 133 to 178 d) Hg(I) oxidizes to Hg(II). The samples from profiles RD3 and RD4, which are from the same sampling site, have the same half-life values in the second step (t1/2 = 133 d), while samples from the Industrial District displayed a little slower kinetics (t1/2 = 178 d). Although we cannot confirm this difference to be significant, based on Table 1, we can notice that the C/N rate for the DI samples is generally higher than that for the RD samples. This can indicate that the level of degradation of the OM of the samples is lower, and that this difference in half-life times may be due to the differences in

OM nature. However, more detailed studies must be carried out to confirm this. The monitoring of Hg species conversion in soils using thermodesorption has already been shown in a previous study (Windmo¨ller et al., 1996), in which Hg(0) and Hg(II) were also added and dry-diluted in two types of soil collected at sites with records of Hg contamination. This study has shown that Hg(0) oxidation also occurred in the two cases, however, the final oxidation percents were 100% and 54%, leading to the conclusion that it depends on the chemistry of the soil in question. The soil with the highest oxidation percent came from an area containing chloride/ soda production residues and the second from an area impacted by gold mining.

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C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975 4.5

Step I (t1/2= 6 d)

4.0

Step I (t1/2= 4 d)

4.0

3.5

Step II (t1/2= 301 d)

Step II (t1/2= 133 d)

3.0 2.5 2.0

2.0

1.5

Step I (t1/2= 11 d)

Layers 00/20 20/40 40/60 60/80

1.0

Profile RD3+Hg(0)

Profile RD3+Hg(II)

Layers 00/20 20/40 40/60 60/80

0.0

1.5 1.0 0.5 0.0

0

20

40

60

80

100

120

140

0

50

100

150

200

250

300

4.5

4.5

Step I (t1/2= 5 d)

4.0

Step I (t1/2= 11 d)

3.5 3.0

4.0

Step II (t1/2= 133 d)

2.5

3.5

Step II (t1/2= 385 d)

3.0 2.5 2.0

2.0 1.5 1.0 0.5

Profile RD4+Hg(0)

Step I (t1/2= 14 d)

Layers 00/20 20/40 40/60 60/80

Profile RD4+Hg(II)

Layers 00/20 20/40 40/60 60/80

20

40

60

80

100

1.0 0.5 0.0

0.0 0

1.5

Hg(II) peak area decrease

Hg(II) peak area increase

3.0 2.5

0.5

120

140

0

50

100

150

200

250

300

4.5

4.5

Step I (t1/2= 4 d)

4.0

4.0

Step II (t1/2= 408 d)

3.5 3.0

3.5 3.0

Step II (t1/2= 178 d)

2.5

2.5

Step I (t1/2= 6 d)

2.0

Layers 00/20 20/40 40/60 60/80

1.5 1.0 0.5

Profile DI+Hg(0)

2.0

Profile DI+Hg(II)

Layers 00/20 20/40 40/60 60/80

1.5 1.0 0.5

Hg(II) peak area decrease

Hg(II) peak area increase

3.5

Hg(II) peak area decrease

Hg(II) peak area increase

4.5

0.0

0.0 0

20

40

60

80

100

120

140

Time, days

0

50

100

150

200

250

300

Time, days

Fig. 4. Hg(II) peak area increase, by addition of Hg(0), as a function of time (a); Hg(II) peak area decrease, by addition of HgCl2, as a function of time (b).

Simulating experiments to understand the accumulation and transport of Hg in soils in relation to atmospheric deposit, Wang et al. (2003) concluded that Hg(0) transformations take place as soon as this element is deposited on the soil. Analysis of soil samples exposed to air containing Hg(0) for two months indicated that approximately 25% of the deposited Hg remained in the Hg(0) form, and 75% turned into the so-called active forms: soluble HCl, organically bound, and residual Hg. In another study, Bloom et al. (2003) added various Hg species to sediments and discussed their conversion to methylmercury. They showed that even the less soluble and more inert (Hg(0) and

HgS) species slowly convert to the species present in the sediment and undergo methylation, although to a lesser extend than the more soluble ones. In another study, Boening (2000) affirms that Hg(0) may be oxidized to Hg(II), particularly in the presence of OM. In turn, Hg(II) may be reduced to Hg(0) when reducing conditions are appropriate. Therefore, these results show how reactive these soils can be when in contact with added Hg. The case of frequent atmospheric deposition of Hg(0) and its conversion to species that remain fixed in the soil longer may explain the high Hg background values found for some soils. The

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

study of do Valle et al. (2005) using the same samples showed the presence of only Hg(II) and the predominance of organically bound Hg. 4.2. Incubation of soil samples with HgCl2 Fig. 5 shows some thermograms of the sequence obtained by Hg(II) incubation. The first thermograms of all the incubated samples showed a broad peak with a maximum value from 280 to 300 °C. After 40 days, a decrease in this peak was observed and another peak appeared at lower temperatures (180 °C). The latter disappeared gradually for RD3-40/60 and DI-40/60 samples. These results indicate that part of the Hg(II) added might have been converted to Hg(I) in the case of the km 10–1 sample,

0.8 0.6 0.4

1d

0.2 0.0

Hg release

and to Hg(0) in the other samples. In the last two samples (RD3-40/60 and DI-40/60), the reduced Hg peak disappeared, suggesting that it volatilized from the sample. The occurrence of three peaks was also observed, suggesting three Hg oxidation states (samples km 10–3 and RD3). Only the km 10–1 and km 10–3 samples presented Hg peaks at lower temperatures at the end of the monitoring. However, it can be observed in the thermograms of sample km 10–1 that the first peak appears at a temperature somewhat higher than that of km 10–3 (temperatures of 190 and 150 °C, respectively). In the case of sample km 10–1, this can be attributed either to the presence of Hg(I) or to the distribution of Hg(II) at different adsorption sites of the mineralogical phases, as observed upon the addition of Hg(0). In the case of sample km 10–3, the

0.8 0.6

km 10-1+Hg(II)

km 10-3+Hg(II)

0.4 0.2 0.0

0.8

0.8

0.6 0.4 0.2 0.0

0.6 0.4 0.2

48 d

1973

1d

48 d

0.0

0.8

0.6

0.6

0.4

0.4

136 d

0.2 0.0

136 d

0.2 0.0

0.8

0.6

0.6 0.4

0.4

0.2

172 d

0.0

0.0 100

Hg release

0.8 0.6 0.4 0.2 0.0

172 d

0.2

200

300

400

500

RD3-40/60+Hg(II) 1d

0.5 0.4 0.3 0.2 0.1 0.0

100

1.5 1.2 0.9 0.6 0.3 0.0

200

300

400

500

DI-40/60+Hg(II) 1d

0.8 0.6 0.4

82 d

47 d

0.2 0.0

0.6

0.6

0.4

0.4

0.2

138 d

0.0

0.2

116 d

0.0

0.6 0.6

0.4

0.4

0.2

280 d

0.0

260 d

0.2 0.0

100

200

300

400

Temperature °C

500

100

200

300

400

Temperature °C

Fig. 5. Thermograms of samples incubated with HgCl2.

500

1974

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

peak is probably due to the presence of Hg(0) in combination with Hg(I) as the temperature is lower. The final conversion percent of added Hg(II) was not the same in all samples. Samples (km 10) showed a lower percent of Hg(II) (average 20%) than the others (Table 1) did. These samples differed from the others for their much lower clay and OM contents. The difference between the percents observed for all other samples was not considered significant. In contrast to the Hg(0) conversion process, no Pearson correlation was observed between the OM content and the percent of Hg(II) conversion (Table 1). Although km 10–2 and RD4-0/20 soil samples, which had the highest OM content within the profile, revealed the lowest conversion percent, Rocha et al. (2000), who studied the efficiency of reduction of HS from Rio Negro, also observed a decrease in efficiency as OM concentration increased. However, Matthiessen (1998) obtained different results for synthetic HS. This contradicting result was attributed to the different natures of the HS studied. The graphs in Fig. 4(b) show the percent of Hg(II) conversion as a function of time. The data were obtained through the thermogram peak area values. The peak obtained in the first thermogram was considered 100%, and the reduction of this area was calculated in terms of percent. These graphs show the conversion process in two steps, both of which were roughly fitted to the first order reactions and are attributed to the reduction of Hg(II) to Hg(I) and to Hg(0). For samples RD3 and RD4, a slower kinetics was observed in the first step (t1/2 = 11 and 14 d) for the upper layers (0/20 cm), which are exactly the ones that have high OM content, and t1/2 = 4 and 5 days for the other layers. However, these differences are too small to be conclusive. In order to reach better conclusions, other tests using more samples are necessary. The second step of the reaction, slower than the first one, was similar for the layers of all profiles. Only the half-life for the Industrial District samples (t1/2 = 408 d) was slightly longer. Rocha et al. (2000) studied the HS reduction processes in Rio Negro, whose matrix composition is equivalent to that of the samples of this work, and obtained reduction curves similar to those of Fig. 4(b). Their curves were also fitted to the two first-order reactions, attributed to the first step, in which from 60% to 70% Hg is reduced, and a second step, in which the remaining Hg is slowly volatilized. They worked with aqueous solutions and controlled pH, and the reaction constants and half-life obtained are in the order of hours, much shorter than those obtained in this work, which are in the order of days. It is clear that in a solid dry matrix the interactions are less efficient and therefore slower, and also that other matrix components may contribute to slow down the reduction process. Although the differences in dynamics between the samples studied remain limited due to the small number of samples studied, the results show that the half-lives in the first order reaction, for the addition of both Hg(0) and Hg(II), were comparable. In the case of the second reac-

tions, however, the incubation with Hg(II) is much higher than the one with Hg(0): Hgð0Þ ! HgðIÞ 4–11 d

HgðIIÞ ! HgðIÞ 11–14 d

!

133–178 d

!

HgðIIÞ

301–408 d

Hgð0Þ

ð1Þ ð2Þ

Therefore, the results showed that both processes may occur in the soils studied and, in general, the dynamics of one of the processes (either oxidation or reduction) may prevail. In case of prevailing oxidation, a greater retention of metal may occur over time, since it is the reduced species that may return to the atmosphere through volatilization. Miretzky et al. (2005) investigated the sorption potential of Hg(II) in Amazon soils and showed the importance of OM, rather than of clay content, in the sorption of Hg(II) in the Amazon soil. In addition, their study suggests that this process is not reversible. According to Rocha et al. (2003), the degree of Hg(II) reduction is significantly influenced by the ratio of phenolic/carboxylic groups and the sulfur bound of HS, revealing a strong competition between Hg(II) complexation and reduction. In other words, the differences in the dynamics of the oxidation and reduction processes, along with the efficiency of the Hg retention processes, may explain the high background values found in Amazon soil. 5. Conclusions Monitoring soil samples collected near Manaus and dry incubated with Hg(0) and Hg(II) through thermodesorption and comparison with standard samples showed a change in the speciation, indicating their respective oxidation and reduction. A correlation between the OM content and the percent of oxidation upon the addition of Hg(0) was observed and also a decrease in oxidation in relation to the soil profile depth, following the natural behavior of the OM within this profile. In order to compare the conversion rates between different kinds of soils, studies with more samples would be necessary. However, the results obtained so far have shown that the kinetics of these processes displayed two steps, which may possibly be extended to other types of soils. In both incubation cases, it seems to occur an intermediate form of Hg(I). The first step is faster (transition to Hg(I)) than the second. The results show the importance of the reactivity of the soil/sediment in contact with added Hg. If gaseous elemental Hg is dry deposited to soils, it has the potential to be converted to reactive Hg and sequestered in the soil while if reactive gaseous Hg is dry deposited to soils, it has the potential to be reduced and re-emitted through volatilization. The chemistry of one of the processes (either oxidation or reduction) may prevail. In case of prevailing oxidation, a larger metal retention may occur after some time, since it is the reduced species that returns to the atmosphere through volatilization and not the oxidized

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