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Research Paper |
1 Area de Edafoloxía e Química Agrícola, Facultade de Ciencias de Ourense, Universidade de Vigo, 32004 Ourense, 2 Departamento de Edafoloxía e Química Agrícola, Facultade de Farmacia, Universidade de Santiago de Compostela, 15706 Santiago de Compostela, 3 Departamento de Química Física, Facultade de Ciencias de Ourense, Universidade de Vigo, 32004 Ourense, Spain
* E-mail: mastevez{at}uvigo.es
(Received 10 February 2003; revised 3 October 2003)
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ABSTRACT |
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 TOP ABSTRACT MATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSACKNOWLEDGMENTSREFERENCES |
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KEYWORDS: mercury, humic acid, kaolin, adsorption, desorption
In addition to being natural constituents, heavy metals enter the soil via beneficial agricultural additives such as lime, fertilizer, manure, herbicides, fungicides, sewage sludge and compost. Other sources such as mine wastes, fly ash and atmospheric deposits can increase the heavy metal concentration in soils (Kabata Pendias & Pendias, 2001; Arias et al., 2001a).
Adsorption and desorption are important processes that affect the mobility and fate of heavy metals in soils. The knowledge of both processes will provide useful information concerning the capacity of soil colloids to immobilize heavy metals and to prevent the pollution of surface and subsurface waters, as well as the inclusion of these heavy metals in the trophic chain through their absorption by plants or micro fauna.
The adsorption capacity of heavy metals by different adsorbents (e.g. sepiolite, kaolinite, illite, palygorskite and bentonite) has been extensively studied under different experimental conditions (Viraraghavan & Kapoor, 1994; Garcia Sanchez et al., 1999; Arias et al., 2001b). Because Hg is highly toxic, even at low concentrations, the release of Hg and its compounds in the environment represents a serious environmental problem, although anthropogenic emissions have been reduced in the last decade (Allard & Arsenie, 1991). The most significant anthropogenic activities giving rise to Hg emission are Cu and Zn smelting, the burning of fossil fuels (mainly coal), the Hg cell chlor-alkali process for production of Cl2 and caustic soda and consumption-related discharges, including waste incineration (Steinnes, 1995).
Mercury shows a peculiar physical-chemical behaviour compared to other heavy metals. Hg compounds dissociate at lower pH values; in very dilute solutions the value of pKa1 is 3.4 and of pKa2 is 2.77 (Baes & Mesmer, 1976). While other heavy metals exhibit an increase in adsorption when pH increases, Hg(II) does not show this behaviour (Barrow & Cox, 1992). Another factor must also be considered: Hg(II) shows a strong tendency to form complexes with chloride, with the resulting chloride-Hg species showing little to no affinity to surfaces (Baes & Mesmer, 1976; Yin et al., 1996; Sarkar et al., 1999).
Organic matter is one of the most important soil components controlling Hg adsorption, mainly in acidic soils (Anderson, 1979; Yin et al., 1996). Humic acids constitute a large fraction of humic substances in soil (Schnitzer, 1986), which in turn account for 70–80% of all soil organic matter (Schnitzer, 1978). They possess a high molecular weight and exhibit low mobility in acidic soil, which endows them with a high capacity to immobilize Hg dumped into the soil.
Although a sizeable portion of organic matter in soil is bound to the soil clays forming complexes of varying stabilities and properties (Greenland, 1965), there is a lack of experimental data for the adsorption of Hg(II) by organic compounds associated with mineral surfaces. Furthermore, there is a scarcity of available information about metal desorption from this type of system, especially relevant with a view to determining the potential of soil as a purifying system and to anticipating the potential phytotoxic effects of metals on specific crops.
In the current study, the interaction of Hg(II) with humic acids previously adsorbed on kaolin was studied using concentrations of Hg(II) between 5 x 10–7 and 2.5 x 10–5 M. Also, we studied a case where previously adsorbed Hg(II) can be desorbed at different pH values with a view to estimating the efficiency with which this mineral association can control Hg(II) levels in the soil solution. The potential of Cu(II), a frequent co-pollutant in Hg-polluted environments which shows high affinity to natural organic matter, to displace adsorbed Hg(II) in the different samples has also been studied as a function of HA concentration and Hg(II) surface coverage.
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MATERIALS AND METHODS |
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TOPABSTRACTMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSACKNOWLEDGMENTSREFERENCES |
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Humic acid extraction was carried out as follows: soil (Umbric Leptosol under coniferous forest) was shaken with 0.1 M NaOH under a N2 atmosphere for 24 h. The alkaline supernatant was separated from the residue by centrifuging, adjusted to pH 2 with 2 M HCl, and allowed to stand for 24 h at room temperature (Schnitzer, 1982). The precipitated HA was separated from the supernatant by centrifuging, shaken with an aqueous mixture of 0.5% HCl and 0.5% HF (48 h) and then dialysed against distilled water to remove free acid and salt. The purified HA was isolated by freeze-drying the dialysed solution.
Humic acid (HA) composition was 51.8% C, 4% N and 4% H; the distribution of functional groups was 599 cmol kg–1 of total acidity, 330 cmol kg–1 of COOH groups together with a total of 291 cmol kg–1 of OH groups.
Kaolin-HA complexes have been described previously (Arias et al., 1996, 2002). Mixtures of HA and kaolin were prepared as follows: 75 g of kaolin were stirred in 0, 7.5, 15, 30 and 60 cm3 of the HA stock solution (66.7 g dm–3). The pH was adjusted to 5 with 5 M NaOH, and the volume was made up to 160 cm3 with distilled water. These solutions were left to stand at room temperature for 24 h, whereupon they were dialysed against distilled water, and then centrifuged (1000 g, for 20 min) to bring down the suspended material, which was isolated and dried in the open air.
HA was indirectly estimated by multiplying the organic C concentration obtained by the wet combustion method described by Guitián & Carballas (1976) by a factor of 1.93, which is the ratio of organic matter to carbon in HA, as determined by elemental analysis. The final concentration of HA in the kaolin-HA complexes ranged from 0 to 26.9 mg g–1 (0, 6.6, 12.2, 19.5 and 26.9 mg g–1 of kaolin).
Adsorption experiments
A batch technique was used to perform all adsorption experiments: 20 ml of 0.01 M NaNO3 solutions with concentrations of Hg(II) [as Hg(NO3 )2 ] ranged between 5 x 10–7 and 2.5 x 10–5 M at pH 4 were added to 100 mg of sample (kaolin or kaolin-HA) in 50 ml Pyrex centrifuge tubes. The centrifuge tubes were capped with ParafilmTM and the suspensions were shaken for 1 h on a rotatory shaker at 200 rpm at room temperature (2032°C). Preliminary studies had shown that Hg concentrations in solution remained constant after 1 h of shaking. Control tubes without sorbent were prepared to check precipitation or adsorption to the centrifuge tubes. After equilibration time, the samples were centrifuged (10 min at 4000 rpm) and filtered using 10 ml polypropylene syringes connected to a filter holder containing a 0.22 µm polysulfone membrane (Gelman Supor Acrodisc 25), and the first 5 ml of filtrate were discarded. The Hg concentrations in the filtrate were analysed by atomic absorption spectrophotometry using the cold-vapor technique (CVAAS, detection limit 0.4 µg l–1). The quantity of Hg adsorbed by the samples was calculated as the difference between the initial concentration of Hg and the concentration remaining in the solution at equilibrium. The dissolved organic carbon (DOC) and pH of equilibrium were also determined using a TOC analyser and a combined glass electrode, respectively. All assays were carried out at least in triplicate.
Desorption experiments
For the study of desorption, samples previously reacted with three Hg(II) concentrations were used (2.5 x 10–6, 5 x 10–6 and 2.5 x 10–5 M). At the end of the adsorption period (1 h), the same volume of 0.01 M NaNO3 (two samples) or 0.01 M NaNO3 plus 0.33 x 10–3 M Cu(NO3)2 (two samples) replaced three quarters of the supernatant of each sample. The pH value of these solutions was adjusted to 4 as for the adsorption experiments. The samples were resuspended and treated as in the adsorption experiments. The desorption steps were repeated four times.
Effect of pH on adsorption-desorption
For the study of the adsorption of Hg(II) as a function of pH, the pH values of the suspensions prepared as described above were adjusted in the range from 2.5 to 6.5 using 0.1 M HNO3 or 0.1 M NaOH. The final pH values were 2.630.0, 4.130.1, 5.630.1 and 6.530.2, respectively (n = 15). The Hg(II) initial concentration was 2.5 x 10–5 M [as Hg(NO3)2]. The ionic strength was 0.01 M NaNO3.After the equilibrium period (1 h), the suspensions were centrifuged (10 min at 4000 rpm) and filtered, then Hg was analysed in the filtrate by CVAAS. Calculations performed using the MINTEQA2 computer speciation program (Allison et al., 1991) showed that no Hg precipitation occurs within the experimental pH range.
In order to characterize the Hg-organic species present in the systems studied we used the computer ionic speciation model WHAM (Windermere Humic Aqueous Model) (Tipping, 1994). WHAM is a combination of several submodels including models for inorganic solution speciation and a humic metal ion binding model (Model V). Model V (Tipping & Hurley, 1992) is a discrete site electrostatic model which takes into account both specific and non-localized (diffuse-layer) binding.
The desorption experiments were carried out in the same pH range (~2.5–6.5) and for samples previously equilibrated with a constant concentration of Hg (2.5 x 10–5 M [as Hg(NO3)2]). The experimental procedure is similar to that described by Cavallaro & McBride (1984). The supernatant of the adsorption experiments was completely substituted by 20 ml of a free-Hg solution (0.01 M NaNO3). The samples were shaken for 1 h. The unadsorbed Hg from the adsorption step was evaluated by mass differences.
Electrokinetic measurements
The electrokinetic measurements were conducted using a microelectrophoretic apparatus (Malvern Zetasizer 3000 HS) and expressed as 
-potential values. Suspensions of samples were prepared (100 mg l–1) in 0.01 M NaNO3 and in 0.01 M NaNO3 containing 5 µM Hg(II) (as Hg(NO3)2). Suspension pH was adjusted to the pH range 2–7 with 0.1 M HNO3 or 0.1 M NaOH. The suspensions were shaken for 1 h, then an aliquot of the suspensions (10 ml) was pipetted, diluted and injected into the electrophoresis cell. For each system, at least 15 measurements were recorded. All the suspensions were prepared with filtered (<0.22 µm) MilliQplus water and all measurements were made at 2032°C.
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RESULTS AND DISCUSSION |
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TOPABSTRACTMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSACKNOWLEDGMENTSREFERENCES |
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). The final pH at equilibrium was equal to 4.030.2 (n = 80) in all the systems studied. Freundlich (equation 1) and Langmuir (equation 2) isotherms have been employed to describe the adsorption behaviour of Hg(II). These equations are written as:
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 | (1) |
 | (2) |
where X is the amount of solute retained per unit weight of the adsorbent (µmol g–1), C is the equilibrium concentration of the solute remaining in the solution (µM); Kf and 1/n are the Freundlich coefficients in equation 1; KL (l µmol–1) is a constant related to the energy of adsorption, and Xm (µmol g–1) is the maximum adsorption capacity of the sample.
Freundlich and Langmuir parameters for Hg(II) adsorption obtained from the fits of X vs. C data to equations 1 and 2 are shown in Table 1. As seen, the Freundlich isotherm yielded best fits to the experimental data, except for the kaolin-HA complex with CHA = 26.9 mg g–1. For this complex, the adsorption data were better described by the Langmuir equation, despite the theoretical limitations (Barrow, 1999). The Freundlich equation applies very well for solids with heterogeneous surface properties and generally for heterogeneous solid surfaces (Stumm & Morgan, 1981). This equation is an empirical equation, where Kf can be interpreted as the amount adsorbed when C = 1. In this sense, the value of Kf increases when the humic acid concentration in the kaolin-HA complex also increases. The Kf value increases from 0.36 (in the case of kaolin without HA) to 3.76 corresponding with the kaolin-HA complex which contains the largest amount of humic acid (26.9 mg g–1) (Table 1). The exponential parameter 1/n can be interpreted as the ‘adsorption intensity’ or the degree of dependence of the adsorption process on the solute concentration. The value of 1/n decreases as the amount of humic acid present in the adsorbent increases (Table 1). However, an increase of 1/n was observed for the highest concentration of humic acid (Table 1). This increase observed in the 1/n value could be due to the organic matter dissolved that increases the DOM concentration to 12.5 mgl–1 (Table 1). For this system, the Hg(II) adsorption was lowest except at the highest Hg(II) concentration, in which the adsorption of Hg(II) increases as the HA adsorbed on kaolin increases, as occurs in the systems containing low levels of HA. This effect can be related to two facts (Yin et al., 1996): (1) the decrease in the number of organic-based surface adsorption sites as a consequence of the solubilization of humic acid, and/or (2) the formation of a soluble Hg complex with the dissolved humic acid which diminishes Hg(II) adsorption on the solid particles.
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This is a well documented factor that affects Hg(II) adsorption behaviour. Yin et al. (1996, 1997) showed that organic matter is the most important soil component controlling Hg(II) retention and release. They observed that an increase in the DOM significantly decreased Hg(II) adsorption to soils, which was attributed to the complexation of Hg(II) by the DOM. This complexation effect was suggested as the cause of the S-type isotherms obtained for the Hg(II) adsorption by soils with large organic-matter contents (Barrow & Cox, 1992; Yin et al., 1997). Drexel et al. (2002) observed the same competition effect for Hg(II) between the sorbent (two Everglades peats) and the DOM released from the sorbent, that acts as a competitive ligand. They obtained conditional binding constants for the strong sites on the peat (Kpeat = 1021.8±0.1 and 1022.0±0.1 M–1) less than those determined for the DOM strong sites (KDOM= 1022.8±0.1 and 1023.2±0.1 M–1), resulting in Hg(II) binding by DOM rather than by peat at low Hg(II) concentrations, thus causing nonlinearities in the Hg(II) sorption isotherms onto peat.
In our experiments, lower Hg(II) adsorption was observed for the kaolin-HA complex with the highest HA concentration (26.9 mg g–1) than for the complexes with lower HA concentration, except for the highest Hg(II) concentration employed. Therefore, this behaviour seems to be related to the initial Hg(II) concentration. At low Hg(II) initial concentration (10–6 M) adsorption for the complexes with CHA=26.9 mg g –1 was 7.0±0.5% lower than Hg(II) adsorption for the complexes with CHA=19.5 mg g–1. When Hg(II) initial concentration increased from 10–6 Mto 10–5 M, only a reduction of 2.2±1.1% in Hg(II) adsorption was observed for the same complexes. In contrast, at the highest Hg(II) initial concentration, this behaviour was not observed, i.e. as the HA concentration of the kaolin increased, the amount of Hg(II) in solution decreased (Fig. 1
).
Haitzer et al. (2002) showed that the Hg-to-DOM concentration ratio strongly affects the binding of Hg(II) to dissolved organic matter. The Hg/DOM ratio was proposed by these authors as an important parameter that affects the Hg-DOM binding constants, therefore affecting sorption reactions. They observed very strong Hg-DOM interactions at Hg/DOM ratios below ~0.001, while at Hg/DOM ratios above ~0.01, they observed weaker Hg-DOM interactions as denoted by the much lower conditional distribution coefficients determined for these ratios. This means that as Hg concentration increases, more inorganic Hg becomes available for adsorption onto the solids, thus increasing adsorption (Barrow & Cox, 1992; Yin et al., 1996, 1997). In our study, for a DOM concentration of 12.5 mg/l (corresponding to the systems with a HA concentration of 26.9 mg g–1), the ratio Hg/DOM increases from 0.008 to 0.4 as the Hg(II) concentration increases from 5 x 10–7 Mto 2.5 x 10–5 M. So, for high Hg(II) concentrations, the interaction of DOM with Hg(II) in these systems is weaker, reducing the competitive effect between the DOM (that acts as a ligand) and the sorbent (Kaolin-HA complexes), resulting in increasing Hg(II) adsorption by the Kaolin-HA complexes as the HA content increases. For the lower Hg(II) concentrations, the strong interaction between the Hg and the DOM released from the Kaolin-HA complex is capable of reversing this tendency, increasing the Hg concentration in solution, though the HA concentration in the sorbent is greater (Fig. 1).
Table 2 shows the degree of complexation of Hg(II) for the systems Kaolin-HA complex-Hg(II) calculated using WHAM. As the DOM concentration increases, the percentage of DOM-complexed Hg increases for all the Hg(II) concentration range. Also, calculations show that for a given DOM concentration, the percentage of complexed Hg(II) decreases as the Hg(II) concentration increases. So, for the systems with a DOM concentration of 12.5 mg l–1, only at the highest Hg(II) concentration studied, was the degree of DOM-complexed Hg <99% (Table 2). It is known that DOM-Hg complexes are less adsorbed by the surfaces (Schuster, 1991). For the highest Hg(II) concentration employed, the capacity of the DOM to complex Hg(II) approaches saturation. Thus for this high Hg(II) concentration, more free Hg(II) species are available for adsorption; as a consequence, Hg(II) adsorption on the kaolin-HA complexes increases as the HA concentration increases.
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Desorption experiments
Desorption experiments were carried out for samples previously reacted with three different initial Hg concentrations (2.5 x 10–6, 5 x 10–6 and 2.5 x 10–5 M). Adsorption-desorption hysteresis was evident in all systems studied. Experimental results are shown in Fig. 2a and b for kaolin and for the desorption with 0.01 MNaNO3 and 0.01 M NaNO3 plus 0.33 x 10–3 M Cu(NO3)2 respectively. The percentage of Hg desorbed was <1% in both cases. Similar results (data not shown) were found for the kaolin-HA complexes, the desorption percentage being <1%. This result proves a strong affinity of Hg(II) to both organic surfaces and kaolinite. This suggests that at low surface coverage, the Hg(II) was bound by high-energy sites, independent of the adsorbent surface. These high-energy sites could either be the sites that form very stable surface complexes with Hg(II), or the micro pores that trap Hg(II) and require high activation energy to release Hg (Schultz et al., 1987; Yin et al., 1996). The high-energy sites are therefore mainly from organic matter and/or silt and clay.
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). The experimental behaviour was the same as that obtained using 0.01 M NaNO3 as a release medium. The addition of high concentrations of Cu(II) does not increase the amount of Hg(II) desorbed, even bearing in mind that Cu(II) forms the most stable complexes with natural organic ligands of all the cations of the Irving series showing a strong affinity to organic matter (Schuster, 1991, McBride et al., 1998).
This result provides indirect evidence that the interaction of Hg(II) with kaolin and kaolin-HA complexes at pH 
4 is mainly not electrostatic, due to the low percentages of Hg(II) desorbed and the high hysteresis observed in the adsorption-desorption process. On the other hand, the distribution of inorganic species in solution calculated using the WHAM model shows that at pH 4, the predominant Hg(II) inorganic species is the soluble Hg(OH)2, bearing in mind the fact that Hg-DOM organic complexes dominate the aqueous Hg(II) speciation at low Hg concentrations (Table 2). The concentration of charged species (Hg2+ and Hg(OH)+) are negligible against Hg(OH)2 at experimental pH (pH 4) (Fig. 3), suggesting a specific interaction between the uncharged species of Hg(II) and the surfaces. Also, one has to consider that not only cations or hydroxo complexes but any metal complex may be specifically adsorbed onto solid surfaces (Schuster, 1991).
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-potential of the adsorbent surface (kaolin and humic-acid kaolin complex) has been measured in the presence and absence of 5 µM of Hg(II) at different pH values (Fig. 4). Electrokinetic measurements allow us to differentiate between the adsorption of particles via electrostatic interaction of the positively charged species with a negatively charged surface, or adsorption via a chemical reaction between Hg species and the surface functional groups (Yin et al., 1997). In this sense, differences between the surface functional groups for Hg(II) adsorption on kaolin and the humic acid must be emphasized. Kaolin is a multi-component oxide system, containing silanol (
SiOH) and aluminol (AlOH) surface functional groups. Sarkar et al. (2000) using the triple-layer model (TLM), suggested that the silanol group was responsible for retaining the bulk (>85%) of the adsorbed Hg(II) through the formation of the SiO–-HgOH+ outer-sphere and the SiOHg(OH)2 inner-sphere species. However Hg(II) binding to humic acid surfaces is dominated by interactions with organic thiol functional groups in combination with carboxyl and phenol functional groups (Xia et al., 1999; Hesterberg et al., 2001; Drexel et al., 2002).
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-potential in the presence and absence of Hg. This 
G can be obtained by equation 3 (James & Healy, 1972; Stumm & Morgan, 1981):
 | (3) |
where z is the valence of the adsorbing species, F is the Faraday constant, and 
is the potential at the plane of adsorption. The value is approximated as potential (Yin et al., 1997). The calculated Gcoul was –2.9 kJ mol –1 and 1.5 kJ mol–1at pH 2.5 for kaolin and K-HA complex with CHA= 26.9 mg g–1, respectively. The electrostatic binding energy calculated as a function of pH shows an increase of the electrostatic contribution to the sorption process as the pH decreases. As pH increases Gcoul decreases, since both the surface potential and the percentage of charged Hg2+ and Hg(OH)+ species decrease. This means that at experimental pH (pH 4), where uncharged species dominate, the electrostatic contribution is negligible.
Taking into account the electrokinetic measurements, together with the theoretical calculations, a chemical interaction between uncharged Hg species and our surface must be occurring. This conclusion is consistent with that of Tiffreau et al. (1995) and Sarkar et al. (1999, 2000).
Effect of pH in adsorption-desorption
The effect of pH on the adsorption-desorption process has been studied at an initial Hg(II) concentration of 2.5 x 10–5 M. The experiments were carried out at four different pH values (2.5, 4.5, 5.5 and 6.5) (Table 3). Kaolin showed an adsorption maximum at pH 4.5 (pHmax), then a slight decrease in adsorption was observed as the pH value increased. This same behaviour was also observed previously for the Hg(II) adsorption by kaolinite (Sarkar et al., 2000). Farrah & Pickering (1978) found that the adsorption of Hg(II) on kaolinite and illite was nearly constant as pH varied from 3.5 to 9.0. Similarly, Viraraghavan & Kapoor (1994) found that adsorption of Hg(II) in mont-morillonite from wastewater decreased as pH increased from 3.5 to 7.5. In the absence of humic acid, the aqueous speciation of Hg(II) does not change above pH 5, where Hg(OH)2 is the predominant inorganic species in solution (Fig. 3). However, we observed a decrease in the Hg(II) adsorption on kaolin at pH values above the pHmax. This decrease may be caused by the increasing OH– concentration in solution as the pH increases, that acts as a competitive ligand for the surface functional groups (Sarkar et al., 2000) and as a consequence of the loss of exchangeable ligand (H2O or OH–) from the surface (Thanabalasingam and Pickering, 1985; Sarkar et al., 2000). The gradual decrease in Hg(II) adsorption may also be due to the decrease in HgOH+ species as solution pH increases above pH 4.5 (Barrow & Cox, 1992).
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). For this system, the complexation of Hg by DOM also contributed to the decrease observed in the adsorption at pH >pHmax, since the DOM increases as pH increases (data not shown). Therefore, the degree of complexation also increases (Yin et al., 1996). In fact, WHAM calculations showed that the percentage of DOM-complexed Hg increases from 2% to 79%, as the pH increases from 4.5 to 6.5. On the other hand, the presence of HA increased the adsorption of Hg(II) over all the pH range studied. For the complexes with CHA = 12.2, CHA = 19.5 and CHA = 26.9 mg g–1, an adsorption percentage >90% was observed. For the complex with CHA = 26.9 mg g–1 and for pH values of 5.5 and 6.5, a relative decrease in Hg(II) adsorption was observed compared to the complexes with CHA = 19.5 mg g–1, due to a partial redissolution of humic acid at these pH values, that increases the DOM concentration at pH 6.5 from 13.6 mg l–1 to 17.5 mg l–1. Again, the DOM complexation is responsible for this relative decrease observed; what’s more, as the pH increases from 4.5 to 6.5 the degree of Hg complexation by DOM increases from 80% to 89% for the complex with CHA = 26.9 mg g –1 and from 29% to 78% for the complex with CHA = 19.5 mg g– 1.
The percentages of desorption were low for the kaolin-HA complexes (Table 3). Only in the absence of HA and pH = 2.5 was desorption observed. In this case the percentage of desorption was >50% (Table 3). This behaviour could be related to the predominant Hg species in solution at this pH value (Hg2+) (Fig. 3). In these conditions, the electrostatic contribution to the interaction between the surface and the Hg(II) is important, as we have shown above, while for the other pH values the predominant species is Hg(OH)2, and hence the interaction between adsorbent and adsorbate is non-electrostatic. For these pH values, the percentage of desorption observed was negligible (<1%) (Table 3). Sarkar et al. (2000) postulated that Hg(II) adsorption on kaolinite is characterized by the formation of both inner- and outer-sphere complexes. Their calculations, employing the TLM showed an increase of the electrostatic interactions XO–HgOH+ at low pH, a fact that is congruent with the results we obtained for Hg desorption.
The presence of HA dramatically reduced the percentage of Hg desorbed, even at pH = 2.5 (Table 3) and facilitates the formation of bonds between adsorbent and adsorbate with greater covalent character. This increase in the covalent interaction prevents Hg mobilization and hence we can avoid problems related to Hg toxicity in natural media.
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CONCLUSIONS |
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TOPABSTRACTMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSACKNOWLEDGMENTSREFERENCES |
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The adsorption isotherms showed that as HA concentration in the sorbent increases, the sorption of Hg(II) also increases, except for low Hg(II) concentrations and high concentrations of DOM. This reflects the importance of DOM in the Hg(II) sorption processes, and how the DOM released from the solid fraction to the soil solution may act as a competitive ligand, possibly increasing the mobility of Hg.
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ACKNOWLEDGMENTS |
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TOPABSTRACTMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSACKNOWLEDGMENTSREFERENCES |
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REFERENCES |
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TOPABSTRACTMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSACKNOWLEDGMENTSREFERENCES |
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) CHA = 0 mg g–1; (
) CHA = 6.6 mg g–1; (
) CHA = 12.2 mg g–1; (
) CHA = 19.5 mg g–1; (•) CHA = 26.9 mg g–1. Error bars represent standard deviations. The absence of error bars implies that the error is less than the size symbol of the symbol.
) 5 x 10–6 M; (•) 2.5 x 10–6 M.
), humic acid complex –CHA = 26.9 mg g–+ 5 µM Hg (