www.elsevier.com/locate/chemosphere

Chemosphere 65 (2006) 1966–1975

Mercury conversion processes in Amazon soils evaluatedby thermodesorption analysis

Claudia M. do Valle a, Genilson P. Santana b, Claudia C. Windmoller c,*

a Gerencia de Quımica, DE, Centro Federal de Educacao Tecnologica do Amazonas, Centro, 69020-120 Manaus, AM, Brazilb Departamento de Quımica, ICE, Universidade Federal do Amazonas, Coroado, 69077-000 Manaus, AM, Brazil

c Departamento de Quımica, 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 2006Available 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 collectednear Manaus, AM, Brazil. The samples were incubated by adding Hg(0) and HgCl2 to dry soil. Solid phase Hg speciation analysis wascarried out using a Hg thermodesorption technique with the aim of distinguishing elemental Hg(0) from Hg(II) binding forms. In the firstcase, 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 oxi-dation 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 lowestvalues corresponded to the samples with the lowest clay contents. The kinetics of conversion of Hg(0) as well as HgCl2 were roughlyfitted to the two first order reactions, a fast one and a slow one. It was not possible to evaluate differences between sampling sitesand 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) thanthe one obtained by the addition of Hg(0) (t1/2 = 148 d). Previous studies have shown the predominance of organically bound Hg in thesesamples. Thus, the kinetic difference between Hg oxidation and reduction in combination with the efficient retention processes by OMmay explain the high background values found in Amazon soils.� 2006 Elsevier Ltd. All rights reserved.

Keywords: Mercury; Speciation; Amazon soils; Redox process; Thermodesorption

1. Introduction

Depending on the redox conditions, Hg occurs in threedifferent 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 param-eters in determining speciation of Hg in soil solution andthe chemical transformations that might occur (Lin andPehkonen, 1999; Ravichandran, 2004). In soils that are richin OM, the Hg transformations frequently occur in the top

0045-6535/$ - see front matter � 2006 Elsevier Ltd. All rights reserved.

doi:10.1016/j.chemosphere.2006.07.001

* Corresponding author. Tel.: +55 31 3499 5752; fax: +55 31 3499 5700.E-mail address: claudia@zeus.qui.ufmg.br (C.C. Windmoller).

horizon. It is continuously remobilized and volatilized asorganomercury and Hg(0), respectively (Rocha et al.,2000; Beldowski and Pempkowiak, 2003). Mierle andIngram (1991) and Barrow and Cox (1992) proposed thatOM is important for the transport of Hg from upland soilsand wetlands to water and for its retention in soils.

The presence of sulfide is also important, because itoften forms the very insoluble HgS salt. In addition tochemical reactions, transformations may also be mediatedby microbial activity such as methylation (Steinnes,1997). The knowledge of Hg speciation and reactions(redox, complexation, and others) is important in orderto explain the retention and the mobility of this elementin soil.

In the conversion of elemental Hg to methylmercury, theoxidation of Hg(0) is an essential step. This oxidation is

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975 1967

considered as occurring in natural aquatic environmentsand is strongly affected by the concentration of Cl�

(Magalhaes and Tubino, 1995; Yamamoto, 1996; Lin andPehkonen, 1999). When analyzing soil samples in contactwith Hg vapor, Wang et al. (2003) concluded that Hg(0)transformations take place as soon as this element is depos-ited on the soil. Approximately 25% of the deposited Hgremained in the Hg(0) form, and 75% turned into the so-called active forms: soluble HCl, and organically boundand residual Hg. The oxidation of Hg(0) in contaminatedsoils and the effects of soil composition have has also beenreported (Windmoller et al., 1996; Thoming et al., 2000;Renneberg and Dudas, 2001).

Photoreduction is an important property of Hg (Costaand Liss, 2000), as the emission of volatile Hg speciesfrom natural soils is believed to be a significant contribu-tor to the atmospheric burden of Hg. Gustin et al. (2002)studied light-enhanced emission of Hg by substratesspiked with pure synthetic Hg species and by naturallyand anthropogenically Hg enriched substrates and demon-strated that the influence of light energy on Hg emissionby naturally enriched substrates is larger than that of soiltemperature.

The release of Hg associated to gold mining and defor-estation are two of the most important environmentalissues in the Amazon basin (Artaxo et al., 2000). Gold min-ing in the Amazon has been responsible for the release ofabout 2.000–3.000 tons of Hg over the last 20 years. Annu-ally, about 200 tons of Hg is released into the atmosphere(Lacerda, 1997). Roulet et al. (1998) calculated that morethan 97% of the Hg accumulated on the soil surface startedbefore gold mining (last 30 years). In addition, biomassburning in tropical forests also seems to have contributedsignificantly to the release of Hg to the atmosphere. Dis-cussing Hg emissions in relation to biomass burning inthe Amazon basin, Veiga et al. (1994) estimated a largeemission of Hg to the atmosphere, at levels of about90 tons per year. Therefore, Amazon soils seem to be alarge regional reservoir of Hg (Fadini and Jardim, 2001).Thus, a better understanding of the species and possiblereactions (redox, complexation, and others) of Hg in thesesoils would help to explain the high Hg levels in soils inregions where there is neither gold mining nor Hg contain-ing minerals. Further research is necessary to evaluate thekinetics of Hg transformations in various types of soilsand sediments. However, these studies depend on adequatetechniques to determine metal speciation.

The number of methods to determine Hg speciationhave increased over the years and so have their sophistica-tion, ranging from the visual identification of Hg phasesto sequential chemical extractions, to X-ray absorptionspectroscopy analysis (Sladek et al., 2002; Kim et al.,2003; Sladek and Gustin, 2003), and to Hg themor-desorption atomic absorption spectroscopy (TDAAS)(Windmoller et al., 1996; Biester and Scholz, 1997). Thesequential chemical extraction is based on the use of vari-ous solutions to extract Hg from different soil fractions.

However, instead of species-specific information, thismethod provides information on the differentiation of Hgcompounds according to behavioral classes, such as watersoluble, acetic acid solution, and other basic and acid solu-tions (Biester and Scholz, 1997; Bloom et al., 2003). X-rayabsorption spectroscopy analysis is a nondestructivemethod that uses high energy synchrotron-sourced X-rayradiation to identify specific species based on scatteringpatterns. This method is most useful in the identificationof specific species as long as they are included in a modeldatabase. However, the identification of Hg(0) is difficult(Sladek and Gustin, 2003) and these methods have so highlimits of detection (>100 lg g�1) that their applications arerestricted (Kim et al., 2003). TDAAS identifies Hg speciesby incremental heating and comparison of thermal releasepatterns to a compound database. More recently, it hasbeen used in investigations of Hg compounds in soils andsediments (Biester and Zimmer, 1998; Biester et al., 2000;Biester et al., 2002; Higueras et al., 2003). According to Sla-dek et al. (2002), TDAAS does not look at specific speciesas EXAFS does, but more at species bound to certainforms and it provide information on elemental Hg and dif-ferent Hg(II) binding forms.

The objective of this work was to study the redox behav-ior of Hg in soil samples collected in areas not impacted bygold mining near Manaus, AM, Brazil. The soil sampleswere prepared by incubation of Hg(0) and HgCl2 by drydilution and the TDAAS technique was used to monitorthe Hg(0) and Hg(II) binding forms periodically until thestabilization of the Hg thermodesorption curves. Somephysical and chemical characteristics of the samples arealso discussed.

2. Soil sampling

Soil samples were collected in three different areas ofManaus; BR 174 Highway (Spodosol, called km 10), Adol-pho Ducke Reserve (Oxisol, RD3 and Ultisol, RD4), andIndustrial District (Ultisol, DI) located in the east of theAmazon State, Brazil. The RD and ID samples were col-lected at depths 0/20, 20/40, 40/60, and 60/80 cm and thekm 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 thetop. The sampling sites were described in detail as coordi-nates, horizons, and Munsell color by do Valle et al. (2005).However, for a better understanding of this work, we pro-vided the following data: BR174 is an area located at km 10of the highway linking Manaus to Boa Vista; AdolphoDucke Reserve is an area away from any anthropogenicactivity, and the Industrial District is an area impactedby industrial activities. None of them has suffered anyimpact from gold mining activities. The samples werehomogenized and freeze-stored in disposable polyethylenebags. A fraction of each sample was preserved in a refrig-erator, whereas the remaining part was air-dried to con-stant weight, manually desegregated, sieved through a2 mm mesh sieve, and homogenized.

100 200 300 400 500

0.00.20.40.60.81.01.20.00.20.40.60.81.01.21.40.0

0.2

0.4

0.6

0.8

0.00.20.40.60.81.01.2

Hg(0)

Hg

rel

ease

Temperature °C

Hg2Cl2

HgCl2

HgO

Fig. 1. Thermograms of standard samples of mercury compounds.

1968 C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

3. Methods and materials

The soil fraction and texture, pH (water and KCl), totalHg (<2 mm and <0.053 mm fractions), sampling locations,and statistical analysis were previously described by doValle et al. (2005).

The nitrogen (N) and carbon (C) contents of the soilswere measured using a CHN Perkin Elmer model 2400 ana-lyzer. The mineralogical composition of the soils from theseareas had already been studied in detail (Bravard and Righi,1988; Righi et al., 1990). No carbonate is expected to bepresent, thus the C determined represents organic C. TheOM content was obtained by multiplying the organic C con-centration by a factor of 1.72, given that the OM C contentis generally around 58% (Radojevic and Bashkin, 1999).

Hg thermodesorption curves were determined by meansof an in-house apparatus consisting of an electronicallycontrolled heating unit and an Hg detection unit usedaccording to do Valle et al. (2005). For Hg detection, aquartz tube for purging thermally released Hg was placedin the optical system of an atomic absorption spectrometer(GBC 932-AA). Hg was detected at 253.7 nm. Analysiswas carried out at a heating rate of 33 �C min�1 undernitrogen gas flow of 200 ml min�1. Interferences, mainlyfrom pyrolytic OM products, were compensated by contin-uous deuterium background correction. Sample weight was�50–3000 mg, depending on the total Hg content. Thermo-grams of Hg standard samples were obtained in triplicateby adding Hg(0), Hg2Cl2, HgCl2 and HgO to thermallypretreated powder silica (500 �C for 2 h) up to the concen-tration of 30 mg kg�1 of total Hg by dry dilution. The pre-served fractions of original samples were analyzed induplicate and in triplicate as necessary.

Soil sample (fraction <2 mm) incubation was carried outby adding Hg(0) and HgCl2 up to 30 mg kg�1 of total Hg bydry dilution. This procedure was chosen in order to simulatethe best possible environmental conditions. Consequently,the addition of any substances to the soil samples wasavoided. They were manually macerated and immediatelyanalyzed by TDAAS in the same conditions as the standardsamples. Thermograms were obtained at intervals of hours,days, and weeks during one year under the same instrumen-tal condition until the Hg pattern became stable. The sam-ples were kept at room temperature during Hg monitoringto simulate natural conditions. In the beginning of the sam-ple preparation, it was not possible to make replicates of theanalyses due to the rapid modifications of Hg interactionswith the matrix. The interval between two analyses was�2 h. As soon as the thermograms showed that thesechanges gradually slowed down, the samples were analyzedin duplicate and in triplicate as necessary. The relative stan-dard deviations of the peak areas obtained were better than20%. The weights used to obtain original soil sample ther-mograms in this work were in the order of grams (>2 g),and for the analysis of the spiked samples it was 200 mgso that the signal related to natural Hg was only a smallfraction of total signal in incubated samples.

4. Results and discussion

The thermodesorption results of the Hg standard sam-ples are shown in Fig. 1. Hg(0) was released in the rangebetween room temperature and 200 �C with a maximumat 150 �C, Hg(I) compounds were released with a maxi-mum around 200 �C, and Hg(II) compounds were releasedat higher temperatures. Other studies (Biester and Scholz,1997; Biester et al., 2002; Sladek et al., 2002) have alsoshown that the thermodesorption temperatures of differentHg standards are different, indicating that the techniqueprovides data on speciation. Temperature ranges betweenstudies varied slightly due to differences in system opera-tional conditions, mainly heating rate and gas flow. How-ever, all studies agree that Hg(0) is released at a lowtemperature, followed by the Hg(II) binding forms. TheHg(I) release pattern was shown in only one earlier study(Windmoller et al., 1996), and the temperature range agreeswith the one obtained here. Thus, we have considered thethermograms peaks of the analyzed samples up to 150 �Cas Hg(0), and as Hg(I) when the maximum occurs ataround 220 �C, and as Hg(II) at higher temperatures.

The representative thermograms of the 15 original sam-ples are presented in Fig. 2. The Hg peaks that can be seenonly between 300 and 400 �C are Hg(II) binding forms.Sometimes the peaks were asymmetric or were split in twopeaks (sample RD3-60/80 for example), indicating that Hghad more than one interaction with the soil matrix. Adetailed discussion of speciation, concentration, and theinfluence of physical–chemical parameters on the behaviorof natural Hg (Table 1) in these samples was made by do

0.0

0.1

0.2

0.3

100 200 300 400 5000.0

0.2

0.4

0.6

0.8

0.0

0.1

0.2

0.3

Km 10-3

km 10-2Hg

rel

ease

Hg

rel

ease

km 10-1

100 200 300 400 500

0.00

0.05

0.10

0.150.00

0.05

0.10

0.150.00

0.05

0.10

0.150.00

0.05

0.10

0.15

DI-60/80

DI-40/60

DI-20/40

DI-0/20

100 200 300 400 500

0.0

0.1

0.2

0.3

0.4

0.0

0.1

0.2

0.3

0.4

0.0

0.1

0.2

0.3

0.4

0.0

0.1

0.2

0.3

0.4

RD3-60/80

RD3-40/60

RD3-20/40

RD3-0/20

100 200 300 400 500

0.0

0.1

0.2

0.30.0

0.1

0.2

0.30.0

0.1

0.2

0.30.0

0.1

0.2

0.3

RD4-60/80

RD4-40/60

RD4-20/40

RD4-0/20

Temperature °C Temperature °C

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

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975 1969

Valle et al. (2005), but one of the most important conclu-sions was that thermodesorption analysis revealed thepresence of only Hg(II). The comparison of the thermode-sorption range of Hg(II) with the result of a previous work(Biester et al., 2002) and a Hg(II) sample incubated in com-mercial humic acid indicated that it is predominantly organ-ically bound.

4.1. Incubation of soil samples with Hg(0)

The representative thermograms of the Hg(0) incubatedsamples from BR 174 Highway, Adolpho Ducke Reserveand Industrial District are shown in Fig. 3. It is possible

to see that the area of the Hg(0) peak (�150 �C) graduallydecreases and a peak appears at higher temperatures(>300 �C) and gradually increases. This occurred for allsamples, 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 toits volatization. The presence of three peaks can beobserved at times (Fig. 3, RD3-40/60, 30 d), which mayindicate that the added Hg(0) seems to partially oxidizeto Hg(I), and later to Hg(II). The similarity between thetemperature of the peak between Hg(0) and Hg(II) andthe Hg(I) standard, as shown in Fig. 1, reinforces the ideaof the presence of this oxidation state.

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

Samples Soil fraction (%)a (g kg�1)b Hgtotal (ng g�1)a,b (%)

Clay Silt Sand C/N OM <2 mm <0.053 mm Hg(0) oxid. Hg(II) red.

km 10–1 2.8 0.2 97.0 ndc 0.16 ± 0.01 <25 46 ± 4 40 19km 10–2 3.6 0.1 96.3 ndc 2.39 ± 0.03 764 ± 32 2320 ± 20 68 10km 10–3 2.7 0.3 97.0 ndc 1.29 ± 0.05 83 ± 2 3420 ± 130 57 20

RD3 0–20 72.9 4.9 22.2 18.16 28.1 ± 0.8 128 ± 14 244 ± 12 55 56RD3 20–40 64.2 4.5 31.3 11.18 23.1 ± 0.1 123 ± 12 231 ± 3 48 44RD3 40–60 71.3 2.0 26.7 16.80 17.4 ± 0.7 122 ± 17 256 ± 8 44 53RD3 60–80 67.4 5.8 26.8 13.31 16.0 ± 0.7 123 ± 20 271 ± 16 42 51RD4 0–20 31.0 3.7 65.3 10.88 24.4 ± 0.2 124 ± 11 313 ± 3 49 41RD4 20–40 35.6 2.0 62.4 9.45 17.9 ± 0.6 120 ± 1 367 ± 4 41 49RD4 40–60 35.3 5.8 58.9 11.90 16.4 ± 0.6 154 ± 9 475 ± 9 37 50RD4 60–80 39.8 4.9 55.3 13.32 11.5 ± 0.5 145 ± 7 349 ± 32 29 69

DI 0–20 27.8 5.7 66.5 23.85 24.7 ± 0.4 114 ± 12 363 ± 18 57 62DI 20–40 18.7 3.8 77.5 22.13 15.3 ± 0.3 108 ± 8 494 ± 52 44 55DI 40–60 15.7 3.6 80.7 31.27 16.2 ± 0.4 99 ± 10 459 ± 17 31 69DI 60–80 13.1 3.3 83.6 37.03 19.2 ± 0.5 92 ± 4 358 ± 11 28 48

RSD (%) 5 8 5

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

1970 C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

Season (dry and rainy) variation in the Amazon allowsOM to go through diverse transformations, includingredox processes (Rocha et al., 2000). It is known thatthe humic substances (HS) present in the OM can exertan antagonistic competition for Hg(II) ions (Rochaet al., 2000). While functional thiophenolic groups bindstrongly to Hg(II) (Xia et al., 1999), semiquinone groupsact as Hg(II) reducers (Allard and Arsenie, 1991). Struykand Sposito (2001) studied the redox properties of stan-dard humic acids (HA) and stated that it is possible thatinorganic constituents in HA, such as iron content, con-tribute to its redox reactions, either in conjunction withsemiquinone radicals or through another pathway. Theyalso say that others report a rapid reduction of Fe(III)by HA, suggesting that the iron strongly complexed withinthe HA structure (as Fe(III)) could be a potential source ofoxidation capacity. Based on the results obtained, theyforwarded a hypothesis for abiotic electron transfer reac-tions of HA involving Fe(III) as a mediator. Data onHg reduction by HS obtained from Rio Negro, in theAmazon region, the same investigated in this study, sug-gest a larger complexation capacity by strong adsorptionsites comparatively to reducing functional sites (Rochaet al., 2000).

Sample km 10–1 was the only one that presented twopeaks at the end of the monitoring process. The low-temperature peak is thought to be due to Hg(I), but itmight also be due to Hg(II) adsorption site in some miner-alogical phase as these Hg species appeared in a tempera-ture interval of 200–300 �C (Fig. 3). According to Biesteret al. (2002), Hg species adsorbed on the mineral surfaceare released in the range of 150–280 �C with a maximumat 200 �C, and above 300 �C Hg will be bound to OM.

The oxidation percentages of the added Hg(0) were calcu-lated through the Hg(II) peak areas of the stabilized samples(Fig. 4(a)). The value of the initial Hg(0) area was consideredas 100%, and the peak areas above 250 �C, which corre-spond to Hg(II), were calculated. One can see in Table 1 thatthe oxidation in the soil profile decreases with depth in thesame way as the OM contents does within the soil profile.The km 10–2 sample, which has the highest OM content ofall three samples from km 10, also displayed the highest oxi-dation percent (68%). Similarly, it can be observed throughprofiles RD3, RD4, and DI, that the layer the closest to thesurface (0/20 cm) and whose OM content is the highest alsoshowed the highest oxidation percents. The Pearson correla-tion of OM and % oxidation, excluding the km 10 sampleswhose properties were the most distinct, was significant forthe remaining 12 samples (0.78, p = 0.05). This indicatesthat there is a correlation between OM content and thepercent Hg oxidation. This could be explained by the OMcomplexation capacity, stabilizing the Hg(II) formed, favor-ing the oxidation of Hg(0) to Hg(II).

In order to obtain information on the kinetics of theadded Hg(0) conversion, graphs of % conversion as func-tion of days were made (Fig. 4(a)), except for km 10 sam-ples, because in this case too few thermograms wereobtained during monitoring. The graphs in Fig. 4(a) showa two-step oxidation process. Both were roughly fitted tothe first order reactions. The curves corresponding to thelayers at each sampling site did not show difference, anda single fit was made to obtain half-life values for eachset of four depths, with two exceptions (Fig. 4(b)). Thehalf-life values were calculated by the expression t1/2 =ln2/K, where K is the angular coefficient of the linear fit.As the thermograms showed a peak that may be attributed

100 200 300 400 500

0.00.4

0.81.2

1.6

0.0

0.40.8

1.21.6

0.0

0.0

0.4

0.4

0.8

0.8

1.2

1.2

km 10-1+Hg(0)

100 d

90 d

Hg

rel

ease

Hg

rel

ease

21 d

1 d

100 200 300 400 500

0.0

0.4

0.8

1.2

0.0

0.4

0.8

1.2

0.2

0.4

0.6

0.0

68 d

0.0

0.3

0.6

0.9

45 d

14 d

1 d

km 10-3+Hg(0)

100 200 300 400 500

0.0

0.1

0.2

0.3

0.0

0.1

0.2

0.3

0.4

0.0

0.2

0.4

0.6

0.8

0.0

0.10.2

0.30.4

137 d

71 d

30 d

1 d

RD3-40/60+Hg(0)

100 200 300 400 500

0.0

0.1

0.2

0.3

0.0

0.1

0.2

0.3

0.00.10.20.30.40.5

0.00.20.4

0.60.8

152 d

112 d

64 d

DI-40/60+Hg(0)

1 d

Temperature °C Temperature °C

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

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975 1971

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 inthe second step, which is much slower (t1/2 from 133 to178 d) Hg(I) oxidizes to Hg(II).

The samples from profiles RD3 and RD4, which arefrom the same sampling site, have the same half-life valuesin the second step (t1/2 = 133 d), while samples from theIndustrial District displayed a little slower kinetics (t1/2 =178 d). Although we cannot confirm this difference to besignificant, based on Table 1, we can notice that the C/Nrate for the DI samples is generally higher than that forthe RD samples. This can indicate that the level of degra-dation of the OM of the samples is lower, and that this dif-ference in half-life times may be due to the differences in

OM nature. However, more detailed studies must be car-ried out to confirm this.

The monitoring of Hg species conversion in soils usingthermodesorption has already been shown in a previousstudy (Windmoller et al., 1996), in which Hg(0) and Hg(II)were also added and dry-diluted in two types of soil col-lected at sites with records of Hg contamination. This studyhas shown that Hg(0) oxidation also occurred in the twocases, however, the final oxidation percents were 100%and 54%, leading to the conclusion that it depends on thechemistry of the soil in question. The soil with the highestoxidation percent came from an area containing chloride/soda production residues and the second from an areaimpacted by gold mining.

0 20 40 60 80 100 120 1400.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5H

g(II)

pea

k ar

ea in

crea

se

Profile RD3+Hg(0)

Step I(t1/2= 6 d)

Step II(t1/2= 133 d)

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

50 100 150 200 250 3000.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

0

60/80Profile RD3+Hg(II)

Step I(t1/2= 11 d)

Hg(II) peak area decrease

Layers 00/20

Step II(t1/2= 301 d)

Step I(t1/2= 4 d)

20/40 40/60

0 20 40 60 80 100 120 1400.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

Hg(

II) p

eak

area

incr

ease

Step I(t1/2= 11 d)

Step II(t1/2= 133 d)

Profile RD4+Hg(0)

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

0 50 100 150 200 250 3000.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

Step I(t1/2= 5 d)

Step II(t1/2= 385 d)

Step I(t1/2= 14 d)

Profile RD4+Hg(II)

Hg(II) peak area decrease

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

0 20 40 60 80 100 120 1400.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

4.0

20/40 40/60 60/80

Step II(t1/2= 178 d)

Time, days

Hg(

II) p

eak

area

incr

ease

Profile DI+Hg(0)

Step I(t1/2= 4 d)

Layers 00/20

0 50 100 150 200 250 3000.0

0.5

1.0

1.5

2.0

2.5

3.0

3.5

4.0

4.5

Step II(t1/2= 408 d)

Time, days

Hg(II) peak area decrease

Step I(t1/2= 6 d)

Profile DI+Hg(II)

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

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).

1972 C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

Simulating experiments to understand the accumulationand transport of Hg in soils in relation to atmosphericdeposit, Wang et al. (2003) concluded that Hg(0) transfor-mations take place as soon as this element is deposited onthe soil. Analysis of soil samples exposed to air containingHg(0) for two months indicated that approximately 25% ofthe deposited Hg remained in the Hg(0) form, and 75%turned into the so-called active forms: soluble HCl, organ-ically bound, and residual Hg. In another study, Bloomet al. (2003) added various Hg species to sediments and dis-cussed their conversion to methylmercury. They showedthat even the less soluble and more inert (Hg(0) and

HgS) species slowly convert to the species present in thesediment and undergo methylation, although to a lesserextend than the more soluble ones. In another study, Boen-ing (2000) affirms that Hg(0) may be oxidized to Hg(II),particularly in the presence of OM. In turn, Hg(II) maybe reduced to Hg(0) when reducing conditions are appro-priate.

Therefore, these results show how reactive these soilscan be when in contact with added Hg. The case of fre-quent atmospheric deposition of Hg(0) and its conversionto species that remain fixed in the soil longer may explainthe high Hg background values found for some soils. The

C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975 1973

study of do Valle et al. (2005) using the same samplesshowed the presence of only Hg(II) and the predominanceof organically bound Hg.

4.2. Incubation of soil samples with HgCl2

Fig. 5 shows some thermograms of the sequenceobtained by Hg(II) incubation. The first thermograms ofall the incubated samples showed a broad peak with a max-imum value from 280 to 300 �C. After 40 days, a decreasein this peak was observed and another peak appeared atlower temperatures (�180 �C). The latter disappearedgradually for RD3-40/60 and DI-40/60 samples. Theseresults indicate that part of the Hg(II) added might havebeen converted to Hg(I) in the case of the km 10–1 sample,

0.0

0.2

0.4

0.6

0.8

0.00.20.40.6

0.8

100 200 300 400 500

0.0

0.2

0.4

0.6

0.8

0.0

0.2

0.4

0.6

0.8

136 d

48 d

Hg

rel

ease

Hg

rel

ease

172 d

1 d

km 10-1+Hg(II)

0.00.10.20.30.40.5

0.0

0.2

0.4

0.6

100 200 300 400 500

0.0

0.2

0.4

0.6

0.00.2

0.40.6

0.8

82 d

280 d

138 d

RD3-40/60+Hg(II)

1 d

Temperature °C

Fig. 5. Thermograms of samp

and to Hg(0) in the other samples. In the last two samples(RD3-40/60 and DI-40/60), the reduced Hg peak disap-peared, suggesting that it volatilized from the sample.The occurrence of three peaks was also observed, suggest-ing three Hg oxidation states (samples km 10–3 and RD3).Only the km 10–1 and km 10–3 samples presented Hgpeaks at lower temperatures at the end of the monitoring.

However, it can be observed in the thermograms of sam-ple km 10–1 that the first peak appears at a temperaturesomewhat higher than that of km 10–3 (temperatures of190 and 150 �C, respectively). In the case of sample km10–1, this can be attributed either to the presence ofHg(I) or to the distribution of Hg(II) at different adsorp-tion sites of the mineralogical phases, as observed uponthe addition of Hg(0). In the case of sample km 10–3, the

100 200 300 400 5000.0

0.2

0.4

0.6

0.0

0.2

0.4

0.6

0.0

0.20.40.6

0.8

0.00.20.4

0.60.8

172 d

136 d

48 d

1 d

km 10-3+Hg(II)

100 200 300 400 500

0.0

0.2

0.4

0.6

0.0

0.2

0.4

0.6

0.0

0.2

0.4

0.6

0.8

0.00.30.60.91.21.5

DI-40/60+Hg(II)

260 d

116 d

47 d

1 d

Temperature °C

les incubated with HgCl2.

1974 C.M. do Valle et al. / Chemosphere 65 (2006) 1966–1975

peak is probably due to the presence of Hg(0) in combina-tion with Hg(I) as the temperature is lower.

The final conversion percent of added Hg(II) was notthe same in all samples. Samples (km 10) showed a lowerpercent of Hg(II) (average 20%) than the others (Table 1)did. These samples differed from the others for their muchlower clay and OM contents. The difference between thepercents observed for all other samples was not consideredsignificant.

In contrast to the Hg(0) conversion process, no Pearsoncorrelation was observed between the OM content and thepercent of Hg(II) conversion (Table 1). Although km 10–2and RD4-0/20 soil samples, which had the highest OMcontent within the profile, revealed the lowest conversionpercent, Rocha et al. (2000), who studied the efficiency ofreduction of HS from Rio Negro, also observed a decreasein efficiency as OM concentration increased. However,Matthiessen (1998) obtained different results for syntheticHS. This contradicting result was attributed to the differentnatures of the HS studied.

The graphs in Fig. 4(b) show the percent of Hg(II) con-version as a function of time. The data were obtainedthrough the thermogram peak area values. The peakobtained in the first thermogram was considered 100%,and the reduction of this area was calculated in terms ofpercent. These graphs show the conversion process in twosteps, both of which were roughly fitted to the first orderreactions and are attributed to the reduction of Hg(II) toHg(I) and to Hg(0). For samples RD3 and RD4, a slowerkinetics was observed in the first step (t1/2 = 11 and 14 d)for the upper layers (0/20 cm), which are exactly the onesthat have high OM content, and t1/2 = 4 and 5 days forthe other layers. However, these differences are too smallto be conclusive. In order to reach better conclusions, othertests using more samples are necessary. The second step ofthe reaction, slower than the first one, was similar for thelayers of all profiles. Only the half-life for the IndustrialDistrict samples (t1/2 = 408 d) was slightly longer.

Rocha et al. (2000) studied the HS reduction processesin Rio Negro, whose matrix composition is equivalent tothat of the samples of this work, and obtained reductioncurves similar to those of Fig. 4(b). Their curves were alsofitted to the two first-order reactions, attributed to the firststep, in which from 60% to 70% Hg is reduced, and a sec-ond 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 inthe order of hours, much shorter than those obtained inthis work, which are in the order of days. It is clear thatin a solid dry matrix the interactions are less efficient andtherefore slower, and also that other matrix componentsmay contribute to slow down the reduction process.

Although the differences in dynamics between the sam-ples studied remain limited due to the small number ofsamples studied, the results show that the half-lives in thefirst order reaction, for the addition of both Hg(0) andHg(II), were comparable. In the case of the second reac-

tions, however, the incubation with Hg(II) is much higherthan the one with Hg(0):

Hgð0Þ !4–11 d

HgðIÞ !133–178 d

HgðIIÞ ð1Þ

HgðIIÞ !11–14 d

HgðIÞ !301–408 d

Hgð0Þ ð2Þ

Therefore, the results showed that both processes mayoccur in the soils studied and, in general, the dynamics ofone of the processes (either oxidation or reduction) mayprevail. In case of prevailing oxidation, a greater retentionof metal may occur over time, since it is the reduced speciesthat may return to the atmosphere through volatilization.Miretzky et al. (2005) investigated the sorption potentialof Hg(II) in Amazon soils and showed the importance ofOM, rather than of clay content, in the sorption of Hg(II)in the Amazon soil. In addition, their study suggests thatthis process is not reversible. According to Rocha et al.(2003), the degree of Hg(II) reduction is significantly influ-enced by the ratio of phenolic/carboxylic groups and thesulfur bound of HS, revealing a strong competitionbetween Hg(II) complexation and reduction. In otherwords, the differences in the dynamics of the oxidationand reduction processes, along with the efficiency of theHg retention processes, may explain the high backgroundvalues found in Amazon soil.

5. Conclusions

Monitoring soil samples collected near Manaus and dryincubated with Hg(0) and Hg(II) through thermodesorp-tion and comparison with standard samples showed achange in the speciation, indicating their respective oxida-tion and reduction.

A correlation between the OM content and the percentof oxidation upon the addition of Hg(0) was observedand also a decrease in oxidation in relation to the soil pro-file depth, following the natural behavior of the OM withinthis profile.

In order to compare the conversion rates between differ-ent kinds of soils, studies with more samples would be nec-essary. However, the results obtained so far have shownthat 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 intermediateform of Hg(I). The first step is faster (transition to Hg(I))than the second.

The results show the importance of the reactivity of thesoil/sediment in contact with added Hg. If gaseous elemen-tal Hg is dry deposited to soils, it has the potential to beconverted to reactive Hg and sequestered in the soil whileif reactive gaseous Hg is dry deposited to soils, it has thepotential to be reduced and re-emitted through volatiliza-tion. The chemistry of one of the processes (either oxida-tion or reduction) may prevail. In case of prevailingoxidation, a larger metal retention may occur after sometime, since it is the reduced species that returns to theatmosphere through volatilization and not the oxidized

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

one. In other words, the differences in the dynamics ofthe oxidation and reduction processes, together with theefficiency of the Hg retention processes may explain thehigh background values found for Amazon soils.

Acknowledgements

The authors would like to thank CNPq and FAPEMIGfor the financial support.

References

Allard, B., Arsenie, I., 1991. Abiotic reduction of mercury by humicsubstances in aquatic system – An important process for the mercurycycle. Water Air Soil Pollut. 56, 457–464.

Artaxo, P., Campos, R.C., Fernandes, E.T., Martins, J.V., Xiao, Z.,Lindqvist, O., Fernandez-Jimenez, M.T., Maenhaut, W., 2000. Largescale mercury and trace element measurements in the Amazon basin.Atmos. Environ. 34, 4085–4096.

Barrow, N.J., Cox, V.C., 1992. The effects of pH chloride concentrationon mercury sorption. I. By goethite. J. Soil Sci. 43, 295–304.

Beldowski, J., Pempkowiak, J., 2003. Horizontal and vertical variabilitiesof mercury concentration and speciation in sediments of the GdanskBasin, Southern Baltic Sea. Chemosphere 52, 645–654.

Biester, H., Scholz, C., 1997. Determination of mercury binding forms incontaminated soils: Mercury pyrolysis versus sequential extractions.Environ. Sci. Technol. 31, 233–239.

Biester, H., Zimmer, H., 1998. Solubility changes of mercury bindingforms in contaminated soils after immobilization treatment. Environ.Sci. Technol. 32 (18), 2755–2772.

Biester, H., Gosar, M., Covelli, S., 2000. Mercury speciation in sedimentsaffected by dumped mining residues in the drainage area of the Idrijamercury mine, Slovenia. Environ. Sci. Technol. 34, 3330–3336.

Biester, H., Muller, G., Scholer, H.F., 2002. Binding and mobility of insoils contaminated by emissions from chlor-alkali plants. Sci. TotalEnviron. 284, 191–203.

Bloom, N.S., Preus, E., Katon, J., Hiltner, M., 2003. Seletiveextractions to assess the biogeochemically relevant fractionation ofinorganic mercury in sediments and soils. Anal. Chim. Acta. 479,233–248.

Boening, D.W., 2000. Ecological effects, transport, and fate of mercury: Ageneral review. Chemosphere 40, 1335–1351.

Bravard, S., Righi, D., 1988. Characterization of clays in an Oxisol–Spodosol toposequence in Amazonia (Brazil). Clay Miner. 23, 279–289.

Costa, M., Liss, P., 2000. Photoreduction and evolution of mercury fromseawater. Sci. Total Environ. 261, 125–135.

do Valle, C.M., Santana, G.P., Augusti, R., Egreja Filho, F.B.,Windmoller, C.C., 2005. Speciation and quantification of mercury inOxisol, Ultisol and Spodosol from Amazon (Manaus, Brazil).Chemosphere 58, 779–792.

Fadini, P.S., Jardim, W.F., 2001. Is the Negro River Basin (Amazon)impacted by naturally occurring mercury. Sci. Total Environ. 275, 71–82.

Gustin, M.S., Biester, H., Kim, C.S., 2002. Investigation of the light-enhanced emission of mercury from naturally enriched substrates.Atmos. Environ. 36, 3241–3254.

Higueras, P., Oyarzun, R., Biester, H., Lillo, J., Lorenzo, S., 2003. A firstinsight into mercury distribution and speciation in soils from theAlmaden mining district, Spain. J. Geochem. Explor. 80, 95–104.

Kim, C.S., Bloom, N.S., Rytuba, J.J., Brown Jr., G.E., 2003. Mercuryspeciation by X-ray absorption sine structure spectroscopy andsequential chemical extractions: A comparison of speciation methods.Environ. Sci. Technol. 37, 5102–5108.

Lacerda, L.D., 1997. Global mercury emissions from gold and silverminining. Water Air Soil Pollut. 97, 209–221.

Lin, C.-J., Pehkonen, S.O., 1999. Aqueous phase reactions of mercurywith free radicals and chlorine: Implications for atmospheric mercurychemistry. Chemosphere 38, 1253–1263.

Magalhaes, M.E.A., Tubino, M., 1995. A possible path for mercury inbiological systems: The oxidation of metalic mercury by molecularoxygen in aqueous solutions. Sci. Total Environ. 170, 229–239.

Matthiessen, A., 1998. Reduction of divalent mercury by humic sub-stances kinetic and quantitative aspects. Sci. Total Environ. 213, 177–183.

Mierle, G., Ingram, R., 1991. The role of humic substances in themobilization of mercury from watersheds. Water Air Soil Pollut. 56,349–357.

Miretzky, P., Bisinoti, M.C., Jardim, W.F., 2005. Sorption of mercury(II)in Amazon soils from column studies. Chemosphere 60, 1583–1589.

Radojevic, M., Bashkin, V.N., 1999. Practical Environmental Analysis.The Royal Society of Chemistry, Cambridge, pp. 325–330.

Ravichandran, M., 2004. Interactions between mercury and dissolvedorganic matter – A review. Chemosphere 55, 319–331.

Renneberg, A.J., Dudas, M.J., 2001. Transformations of elementalmercury to inorganic and organic forms in mercury and hydrocarbonco-contaminates soils. Chemosphere 45, 1103–1109.

Righi, D., Bravard, S., Chauvel, A., Ranger, J., Robert, M., 1990. In situstudy of soil processes in an Oxisol–Spodosol sequence of Amazonia(Brazil). J. Soil Sci. 150 (1), 438–445.

Rocha, J.C., Sargentini Jr., E., Zara, L.F., Rosa, A.H., Santos, A., Burba,P., 2000. Reduction of mercury(II) by tropical river humic substances(Rio Negro) – A possible process of the mercury cycle in Brazil.Talanta 53, 551–559.

Rocha, J.C., Sargentini Jr., E., Zara, L.F., Rosa, A.H., Santos, A., Burba,P., 2003. Reduction of mercury(II) by tropical river humic substances(Rio Negro) – Part II. Influence of structural features (molecular size,aromaticity, phenolic groups, organically bound sulfur). Talanta 61,699–707.

Roulet, M., Lucotte, M., Canuel, R., Rheault, I., Tran, S., De FreitosGogh, Y.G., Farella, N., Souza do Valle, R., Souza Passos, C.J., DeJesus Da Silva, E., Mergler, D., Amorim, M., 1998. Distribution andpartition of total mercury in waters of the Tapajos River basin,Brazilian Amazon. Sci. Total Environ. 213, 203–211.

Sladek, C., Gustin, M.S.C., 2003. Evaluation of sequential and selectiveextraction methods for determination of mercury speciation andmobility in mine waste. Appl. Geochem. 18, 567–576.

Sladek, C., Gustin, M.S., Biester, H., Kim, C., 2002. Application of threemethods for determining mercury speciation in mine waste. Geochem.Explor. Environ. A. 2, 369–375.

Steinnes, E., 1997. Mercury. In: Alloway, B.J. (Ed.), Heavy Metal in Soils,second ed. Blackie Academic and Professional, London, pp. 245–259.

Struyk, Z., Sposito, G., 2001. Redox properties of standard humic acids.Geoderma 102, 329–346.

Thoming, J., Kliem, B.K., Ottosen, L.M., 2000. Electrochemicallyenhanced oxidation reactions in sandy soil polluted with mercury.Sci. Total Environ. 261, 137–147.

Veiga, M.M., Meech, J.A., Onate, N., 1994. Mercury pollution fromdeforestation. Nature 368, 816–817.

Wang, D., Shi, X., Wei, S., 2003. Accumulation and transformation ofatmospheric mercury in soil. Sci. Total Environ. 304, 209–214.

Windmoller, C.C., Wilken, R-D., Jardim, W.F., 1996. Mercury speciationin contaminated soils by thermal release analysis. Water Air SoilPollut. 89, 399–416.

Xia, K., Skyllberg, U.L., Bleam, W.F., Bloom, P.R., Nater, E.A., Helmke,P.A., 1999. X-ray absorption spectroscopic evidence for the complex-ation of Hg(II) by reduced sulfur in soil humic substances. Environ.Sci. Technol. 33, 257–261.

Yamamoto, M., 1996. Stimulation of elemental mercury oxidation in thepresence of chloride ion in aquatic environments. Chemosphere 32,1217–1224.