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Humic acid adsorption and surface charge effects on schwertmannite and goethite in acid sulphate waters
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Humic acid adsorption and surface charge effects on schwertmannite and goethite in acid sulphate waters Sirpa Kumpulainena,b,, Frank von der Kammerb, Thilo Hofmannb a
Department of Geology, University of Helsinki, P.O. Box 64, 00014 Helsinki, Finland Department of Environmental Geosciences, Althanstrasse 14, 1090 Vienna, Austria
b
art i cle info
ab st rac t
Article history:
In acid conditions, as in acid mine drainage waters, iron oxide particles are positively
Received 13 September 2007
charged, attracting negatively charged organic particles present in surrounding natural
Received in revised form
waters. Schwertmannite (Fe8O8(OH)6SO4) and goethite (a-FeOOH) are the most typical iron
17 December 2007
oxide minerals found in mine effluents. We studied schwertmannite formation in the
Accepted 18 December 2007
presence of humic acid. Further, surface charge and adsorption of humic acid on synthetic
Available online 27 December 2007
schwertmannite and goethite surfaces in pH 2–9 and in humic acid concentrations of
Keywords: Acid mine drainage Schwertmannite Humic acid Adsorption Surface charge Particles
0.1–100 mg/L C were examined. Schwertmannite did precipitate despite the presence of humic acid, although it contained more sulphate and had higher specific surface area than ordinary schwertmannite. Specific surface area weighted results showed that schwertmannite and goethite had similar humic acid adsorption capacities. Sulphate was released from schwertmannite surfaces with increasing pH, resulting in an increase in specific surface area. Presence of sulphate in solution decreased the surface charge of schwertmannite and goethite similarly, causing coagulation. In acid conditions (pH 2–3.5), according to the zeta potential, schwertmannite is expected to coagulate even in the presence of high concentrations of humic acid (p100 mg/L C). However, at high humic acid concentrations (10–100 mg/L C) with moderate acid conditions (pH43.5), both schwertmannite and goethite surfaces are strongly negatively charged (zeta potential o30 mV) thus posing a risk for colloid stabilization and colloidal transport. & 2007 Elsevier Ltd. All rights reserved.
1.
Introduction
Weathering of sulphides in tailings and waste rock piles produces acidic waters enriched with sulphate, iron and toxic elements (Nordstrom and Alpers, 1999). Schwertmannite (Fe8O8(OH)6(SO4)) and goethite (a-FeOOH) are the most typical iron(III) oxide minerals precipitated from acidic (pH 2.5–4.5), sulphate-rich (o3000 mg/L) waters (Bigham et al., 1996; Regenspurg et al., 2004; Kumpulainen et al., 2007). The formation of schwertmannite is thermodynamically favoured over ferrihydrite (Fe5HO8 4H2O) in the presence of sulphate in pH 2–7.5 (Majzlan et al., 2004). Schwertmannite is kinetically
unstable, and transforms to goethite over time (Bigham et al., 1996; Jo¨nsson et al., 2005). Adsorption of natural organic matter (NOM) suppresses the phase transformation from ferrihydrite to goethite (Cornell and Schwertmann, 1979), and is therefore expected to similarly retard the transformation of schwertmannite. Crystallization of iron oxides may be inhibited in the presence of organic ligands due to formation of complexes with dissolved iron (Cornell and Schwertmann, 1979). Sulphate adsorbs on goethite surfaces through the formation of both inner- and outer-sphere complexes. At pHo6, or high ionic strengths, formation of inner-sphere complexes is
Corresponding author. Department of Geology, University of Helsinki, P.O. Box 64, 00014 Helsinki, Finland. Fax: +358 9 191 50826.
E-mail address: [email protected] (S. Kumpulainen). 0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.12.015
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favoured (Peak et al., 1999; Rietra et al., 2001). Sulphate adsorption on schwertmannite is similar to that of goethite. One-third of sulphate in schwertmannite is specifically adsorbed to the mineral surface (Bigham et al., 1990), but cannot be discounted from mineral formula, because it has a stabilizing effect on schwertmannite particles (Bigham et al., 1996). Iron oxides are important adsorbents of heavy metals and other harmful elements. Adsorption capacity arises from large specific surface area and from development of surface charge. Specific surface area of iron oxides varies typically in mining environments 100–200 m2/g (Bigham et al., 1990; Gagliano et al., 2004). Iron oxide surfaces are protonated, and therefore positively charged in acid conditions, and dehydroxylated and negatively charged in alkaline conditions. The pH at which the surface has equal amount of positive and negative adsorption sites is called the point of zero charge (PZC) and is measured by potentiometric titration. If no specifically adsorbed ions are present, PZC equals the isoelectric point (IEP) measured by electrophoresis. IEP is the point where no movement of particles occurs in an electrical field and where zeta potential is zero (Stumm, 1992). Pure synthetic goethite has PZC at pH 9.0 (van Geen et al., 1994; Gaboriaud and Ehrhardt, 2003). Natural iron oxides have a much lower PZC than synthetic oxides because of impuor rities. Surface-adsorbed anionic elements such as CO2 3 can significantly reduce the surface charge, whereas SO2 4 adsorption of cationic substances such as Cu2+ or Zn+ increase the surface charge (Russell et al., 1975; van Geen et al., 1994; Ren and Packman, 2005). Not only adsorption tendency but also aggregation tendency has great impact on the transport of heavy metals. Iron oxides tend to coagulate in near-neutral pH or in high ionic strengths (Matijevic, 1980; Schudel et al., 1997), and therefore, generally considered to settle in acid mine waters. However, colloidal size particles have been identified from mine waters (Karlsson et al., 1988; Schemel et al., 2000; Za¨nker et al., 2002; Regenspurg et al., 2004). When waters from mine sites discharge into the surrounding environments, they come into contact with waters enriched with NOM. Median concentration of dissolved organic carbon in stream waters in Europe is 5 mg/L, but varies up to several tens of mg/L (Salminen et al., 2005). Adsorption of NOM on goethite and hematite has been studied by various authors (Tipping, 1981; Gu et al., 1994; Au et al., 1999), but investigations focusing on conditions typical for mine environments, such as Jo¨nsson et al. (2006), have been scarce. Dilution of acid mine drainage waters with humic acid (HA)-containing waters can cause either retention or mobilization of metals depending on the metal to organic matter ratio and pH (Suteerapataranon et al., 2006). Organic amendments added to tailings may induce aggregation of iron oxides cementing the tailings (Stjernman Forsberg and Ledin, 2003). On the other hand, some organic acids produced by plants, such as citrate, dissolve already formed cements, mobilizing the colloids (Slowey et al., 2005). Iron has been shown to be strongly associated with NOM in a wetland receiving acid mine drainage. In the presence of high concentrations of NOM (4100 mg/L), iron particles were stabilized in the colloidal size range (Peiffer et al., 1999).
NOM consists of various organic substances that are formed during decomposition of vegetation and animal tissues. Due to variability of chemical and physical properties, different fractions of NOM exhibit different adsorption behaviours. In general, carboxyl and hydroxyl functional groups are responsible for adsorption (Gu et al., 1995). Adsorption of NOM occurs through ligand exchange reactions and formation of surface complexes. Carboxylic groups adsorb especially at low pH, whereas adsorption of hydroxylic groups becomes more important at high pH (Filius et al., 2003). HAs are high molecular weight organic materials, which are precipitated generally in a pH range of 2–3 in contrast to fulvic acids (FAs), which are lighter, and also soluble at pHo2. Adsorption of organic matter on iron oxide surfaces lowers the surface charge, causing a charge reversal even at small loadings (0.1–0.2 mg/L), followed by aggregation and settling of particles. Addition of larger amounts of organic matter increases the negative surface charge and colloidal stability (Herrera Ramos and McBride, 1996; Kretzschmar and Sticher, 1997). Adsorption of NOM on iron oxide surface also affects the mobility of harmful elements. In the presence of NOM, adsorption of Cu and U and to a lesser extent also Co, has been shown to be enhanced (Murphy and Zachara, 1995; Ali and Dzombak, 1996; Zuyi et al., 2000; Jo¨nsson et al., 2006). The presence of sulphate can decrease adsorption of NOM due to competitive ion effect (Gu et al., 1994; Ali and Dzombak, 1996; Jo¨nsson, 2003). In high NOM concentrations, some metals, such as Cu, might also tend to become more mobile (Alcacio et al., 2001; Jo¨nsson, 2003). This has been explained by coating of iron hydroxide surfaces with organic material and complexing tendency of soluble organic acids towards Cu. Trace metals bind directly to the iron hydroxide surfaces in circum neutral pH, but in more acidic conditions they bind to the functional groups of organic matter adsorbed on iron hydroxides (Tessier et al., 1996). The objective of the present paper was to study the interactions of NOM and iron oxides typical for acid mine environments, their effect on surface charge of iron oxides, and further effect of NOM adsorption on colloidal stability. First, schwertmannite formation in the presence of HA was studied. Then, a series of adsorption experiments with schwertmannite and goethite were done to evaluate the adsorption of HA and its effects on surface charge in variable pH and HA concentrations.
2.
Material and methods
2.1.
Precipitate synthesis
Schwertmannite (ideally Fe8O8(OH)6(SO4)) was synthesized with a method described by Loan et al. (2004). Five hundred millilitres of preheated MilliQ water was mixed with 2.506 g of exsiccated Fe2(SO4)3 xH2O (purissimum p.a. grade) in a PE bottle. The suspension was stirred in a water bath at 85 1C for 1 h. The pH of the suspension was 1.8. The suspension was filtered through 0.45 mm pore size Whatman cellulose filter, the residue was washed twice with MilliQ water, and freeze dried. The method produced 0.5 g of freeze-dried precipitate.
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Similarly, schwertmannite was synthesized in the presence of HAs (0.1, 1, 10 and 100 mg/L C) adding HA before iron sulphate addition. Goethite (a-FeOOH) was synthesized according to the method of Schwertmann and Cornell (2000). The suspension was filtered through 0.1 mm pore size Whatman cellulose filter, the residue was washed twice with MilliQ water, and freeze dried. Synthetic schwertmannite had a composition of Fe8O8(OH)4.92(SO4)1.54 and a specific surface area of 130 m2/g. Synthetic goethite had a specific surface area of 72 m2/g. Both schwertmannite and goethite contained small amounts of carbon impurities, 0.19% (w/w) and 0.87% (w/w), respectively.
2.2.
Humic acid purification
Adsorption experiments were conducted using commercial HA (Fluka). The HA was first purified from clay particles (initial ash content 20%), trace metals, humin and FA fractions with the method described by Vermeer et al. (1998). Purified HA was freeze dried. Purified HA consisted of 55.3% (w/w) C, 3.87% (w/w) H, 1.15% (w/w) N, and 1.54% (w/w) S. Oxygen content was calculated: 38.2% (w/w) O. HA stock solution of 2 g/L was prepared and adjusted to pH 10 with NaOH in order to dissolve all humic material. Stock solution was allowed to equilibrate for 24 h. Solutions of 0.1, 1, 10 and 100 mg/L C were diluted from stock solution for the batch adsorption experiments. HA concentrations (0.1–100 mg/L C) were selected to correspond to values typically encountered in stream waters.
2.3.
Batch adsorption experiments
Adsorption of HA on schwertmannite and goethite surfaces was studied with batch adsorption experiments using iron oxide to solution ratio of 1 g/L. The ratio was the same as in schwertmannite synthesis in the presence of HA. Suspension pH was adjusted before HA addition to 2, 3, 4, 5, 6, 7, 8 and 9 by adding HCl or NaOH, shaking for 24 h and adjusting the pH once more, if required, before HA addition. HA solution with strength of 0.1, 1, 10 or 100 mg/L C and pH adjusted similarly to the schwertmannite suspensions was added and shaken for 24 h. After batch adsorption, all precipitates were washed twice, filtrated and freeze dried.
2.4.
Chemical analysis
Iron hydroxide Fe content was determined by direct current plasma optical emission spectrometry. Mean standard error was 71.1% (w/w). C, H, N and S content were determined by combustion (Vario Macro elemental analyser). All measurements were analysed in duplicate, and mean standard error was 70.01% (w/w) for C, 70.004% (w/w) for H, 70.01% (w/w) for N and 70.05% (w/w) for S.
2.5.
Mineralogical analysis
Mineralogy of synthesized precipitates was determined by X-ray powder diffraction. CuKa-radiation and a Philips X0 Pert
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diffractometer equipped with a vertical goniometry and a curved Cu diffracted beam monochromator were used. The patterns were scanned from 2 to 701 2y with 0.021 steps and counting rate 1 s/step.
2.6.
Particle size
Particle size distribution of synthetic oxides and selected precipitates after batch adsorption experiments was studied with laser diffraction using a Malvern Mastersizer Hydro 2000 mP by calculating volume-based size distribution according to the Mie scattering theory. Refractive index of schwertmannite and goethite was set to 2.17, and adsorption to 0.1. All measurements were performed in triplicate. Mean particle size for synthetic schwertmannite and goethite after resuspension and moderate mixing rate (1500 rpm) was 10 and 29 mm, respectively. Minimum particle size was determined from mixed and sonicated suspension. Minimum particle size for schwertmannite and goethite was 270 and 30 nm, respectively. The particle size distribution results were used to select suitable pore size filters, 0.45 mm for schwertmannite and 0.1 mm for goethite, in order to minimize the loss of particles in the filtration process and to keep the time of filtration reasonable.
2.7.
Surface charge
Surface charge of synthetic schwertmannites and goethites was evaluated by determining electrophoretic mobility of suspended particles with Malvern Zetasizer Nano ZS. The Henry equation and Smoluchowski approximation were applied to calculate zeta potential. All measurements were conducted at +23 1C and calculated by triplicate analysis. Colloidal suspension was considered electrostatically stable, when zeta potential of particles was 4+30 mV oro30 mV (Evans and Wennerstro¨m, 1999). IEP for pure oxides in different concentrations of NaCl (0.001, 0.01 and 0.1 M) and Na2SO4 (0.1 and 1 mM) was determined by titration. Before titration, suspension was allowed to equilibrate and settle for 20 h. Titration of 10 mL of precipitate suspension (1 g/L) with 0.1 M NaOH or 0.1 M HCl was carried out from pH 2 to 9, or from 9 to 2, depending on the zeta potential in the start of the titration, in order to avoid crossing of IEP in the beginning. Zeta potentials of single samples were determined after batch adsorption experiments with HA solutions. Suspensions were left to settle for at least 2 h. A small volume of suspension was collected underneath the surface for zeta potential measurements. Surface charge determinations were carried out in the presence of atmospheric oxygen and carbon dioxide, which correspond to conditions in shallow surface waters.
2.8.
Specific surface area
Specific surface area of synthesized precipitates and selected schwertmannites after batch adsorption experiments was determined using the Brunauer, Emmett and Teller multipoint method (Brunauer et al., 1938). Before measurements, freeze-dried samples were degassed in 110 1C for 30 min. Temperatures between 90 and 120 1C are necessary to remove
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hygroscopic water (Weidler, 1997) and should not induce mineralogical changes (Regenspurg et al., 2004). Determination of specific surface area was by Quantachrome 2200, measuring adsorption of N2 in cryogenic temperatures from 6 points in the range Po/P 0.05–0.3. Determinations were performed in duplicate, and mean standard error was 71.2 m2/g.
3.
Results and discussion
3.1. Schwertmannites synthesized in the presence of humic acid Series of schwertmannite produced in the presence of different concentrations of HA showed no significant effect on the degree of schwertmannite crystallinity. Precipitation of schwertmannite in the presence of HA was expected since negatively charged surface ligands on bacterial cells are known to act as passive nucleation sites for iron oxides in acid environments (Ferris et al., 1989). Hydrolysis of iron and precipitation of schwertmannite resulted in a low pH (pH 1.8), which further led to coprecipitation of HA together with schwertmannite. Fe/S ratios of forming schwertmannite decreased with increasing HA concentrations (Table 1). Decreasing molar Fe/S ratios (Table 1), increased sulphate contents (Table 1) and low electric conductivities of washing water after synthesis indicated that coprecipitated HA retarded outward diffusion of adsorbed sulphate from schwertmannite particles. Specific surface area of forming schwertmannite increased in the presence of 0.1–10 mg/L C HA (Table 1). However, in the presence of higher HA concentrations (100 mg/L C), specific surface area decreased probably due to the blocking effect of coprecipitating HA. Even so, specific surface area was higher than that of pure schwertmannite.
3.2.
Effects of sulphate and humic acid on surface charge
Zeta potential of pure synthetic schwertmannite in NaCl solutions always fell within 730 mV irrespective of solution pH, and it promoted flocculation and settling (Fig. 1a). Opposite to this, synthetic goethite suspension showed typical zeta values 4+30 mV at pH 2–6, promoting electrostatic stabilization of colloids (Fig. 1b). The IEP (IEPpH) in
0.1–0.001 M NaCl solutions for synthetic schwertmannite was 6.9–8.2, and for synthetic goethite 8.5. The IEPpH values for synthetic schwertmannite were in good agreement with the value (at pH 7.2) reported by Jo¨nsson et al. (2005) for natural schwertmannite. The IEPpH for goethite was slightly lower than earlier reports (van Geen et al., 1994; Gaboriaud and Ehrhardt, 2003) due to the presence of atmospheric carbon dioxide, which can adsorb on iron oxide surfaces thus lowering the surface charge. Common point of intersection (cpi), the point where titration curves intersect each other, was same as IEPpH for goethite, but schwertmannite showed a cpi at pH 5.4 and at zeta potential +10 mV. In potentiometric titrations, the different cpi from PZC indicates the presence of specifically adsorbing anions (Lyklema, 1984), such as sulphate. Schwertmannite is unstable with respect to goethite. After resuspension, specifically adsorbed sulphate starts to desorb from schwertmannite surfaces. Sulphate release continues until schwertmannite has transformed fully to goethite (Bigham et al., 1996). In the presence of reactive solute (Na2SO4), zeta potential decreased due to adsorption of sulphate anions on iron oxide surfaces (Fig. 1c–d). In a sulphate concentration of 0.1 mM, zeta potential of schwertmannite was similar to that in NaCl solutions (Fig. 1a). Small amounts of sulphate (0.1 mM) considerably decreased the zeta potential of goethite from +35 mV in 1 mM NaCl solution to +20 mV in 0.1 mM Na2SO4 solution at pH 2–5 (Fig. 1b, d). In sulphate concentrations of 1 mM, the IEP for schwertmannite decreased to 5.7 and for goethite to 6.8. Above pH 8, sulphate adsorption should not occur on goethite surfaces (Persson and Lo¨vgren, 1996), and schwertmannite becomes unstable (Majzlan et al., 2004). Schwertmannite and goethite surfaces behaved similarly in the presence of sulphate, and had zeta potentials within 730 mV, promoting flocculation and settling. In the presence of 0.1, 1, 10 and 100 mg/L C of HA, the IEPs for schwertmannite were 4.5, 3.1, 2.8 and 2.5, respectively (Fig. 1e). The IEPs for goethite in the presence of 0.1, 1 and 10 mg/L C of HA were 8.2, 6.9 and 2.2, respectively (Fig. 1f). In the presence of 100 mg/L C HA, zeta potential of goethite was always negative. The difference in surface charges between low (0.1 and 1 mg/L C) and high (10 and 100 mg/L C) HA concentrations for goethite (Fig. 1f) was thought to be related to surface
Table 1 – Chemical composition (Fe, SO4 and C) and specific surface area (SSA) of precipitates formed after schwertmannite synthesis in the presence of humic acid (HA)
Synthetic schwertmannite (SS) SS+0.1 mg/L C HA in synthesis SS+1 mg/L C HA in synthesis SS+10 mg/L C HA in synthesis SS+100 mg/L C HA in synthesis
Fe/S (mol/mol)
Fe % (w/w)
SO4 % (w/w)
C % (w/w)
SSA (m2/g)
5.19 4.95 4.83 4.70 4.62
46.1 45.3 45.6 44.4 40.0
15.2 15.7 16.2 16.2 14.9
0.19 0.11 0.20 0.90 5.51
130 232 233 261 152
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Fig. 1 – Zeta potentials of synthetic schwertmannite (left) and synthetic goethite (right) in different pH and concentrations of NaCl, Na2SO4 and humic acid (HA) solutions. Initial iron oxide concentration was 1 g/L. Zeta potentials for iron oxides in NaCl and Na2SO4 solutions were measured by titration. Iron oxides in humic acid solutions were measured by determining zeta potentials of single samples after manual pH adjustment, where pH represents the final pH after batch adsorption experiment.
coverage: in high HA concentrations (10–100 mg/L C) goethite particles were assumed to be fully covered with HA masking the electrochemical properties of initial iron oxide surfaces. This was examined by evaluating the extent of adsorption of HA on goethite surfaces (see Section 3.4). The difference in zeta potentials in the presence of low concentrations of HA between schwertmannite and goethite was, on the other hand, thought to be related to the continuous and pH-
dependent sulphate release from schwertmannite, which lowers the zeta potential (Fig. 1a, c, e).
3.3. Desorption of SO4 from schwertmannite and effect on surface area Sulphate content of schwertmannite after batch adsorption experiments (Fig. 2) was lower (p13.7% (w/w)) than that of
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Fig. 2 – Sulphate content of schwertmannite after humic acid batch adsorption experiments in various initial humic acid concentrations (0.1, 1, 10 and 100 mg/L C). pH is final pH after batch adsorption experiment.
original schwertmannite (15.2% (w/w)). Further, electric conductivities of schwertmannite suspensions after batch adsorption experiments amounted to 100–250 mS/cm higher than those of goethite suspensions. The sulphate content of schwertmannite varied according to different pH values (Fig. 2), decreasing with increasing pH. The amount of released sulphate after pH adjustment and batch adsorption experiments with HA increased with pH, and amounted to 0.15–1.35 mmol/g. These values are consistent with those reported by Jo¨nsson et al. (2005) for sulphate release from natural schwertmannite in water with variable pH. In high HA concentrations (100 mg/L C), the release of sulphate was slightly higher than in low HA concentrations (0.1–10 mg/L C), which cannot be explained only by changes in mass due to HA adsorption or coprecipitation. Thus, adsorption of HA at high HA concentrations (100 mg/L C) must be accompanied by enhanced desorption of sulphate from schwertmannite. Specific surface area of selected precipitates was studied after batch adsorption experiments. Specific surface area of schwertmannite increased with pH in 0.1 mg/L C HA suspension, from 108 m2/g at pH 2 to 290 m2/g at pH 8 (Fig. 3). The increase was linear with increasing pH until pH 7, which is consistent with Jo¨nsson et al. (2005). One-third of sulphate in schwertmannite is adsorbed on surfaces, corresponding with a maximum sulphate adsorption capacity of 2.6 mmol/m2 (Bigham et al., 1990). Desorption of sulphate from schwertmannite in this study was higher than maximum sulphate adsorption capacity in HA concentrations of 0.1–10 mg/L C above pH 7. In high HA concentrations (100 mg/L C), sulphate release from schwertmannite was higher than maximum sulphate adsorption capacity in the entire pH range (2–9). Thus, part of the sulphate must have originated from the schwertmannite structure at pH 47 in low HA concentrations (0.1–10 mg/L C), and at pH 2–9 in high HA concentrations (100 mg/L C). At pH 47.5, the schwertmannite phase becomes unstable and was expected to begin to dissolve.
Fig. 3 – Specific surface area of selected schwertmannite samples after batch adsorption experiments in 0.1 mg/L C humic acid solutions. Dashed line represents the specific surface area of pure schwertmannite before pH adjustment and humic acid addition.
3.4.
Adsorption and coprecipitation of humic acid
Pure HA that was used in the batch adsorption experiments precipitated at pH 1.5. In HA batch adsorption experiments with schwertmannite, HA coprecipitated and/or adsorbed completely at pH 2, resulting in a schwertmannite that contained 9.5% (w/w) C HA after reaction with 100 mg/L C HA solution (Fig. 4a). In pH 3–5, coprecipitation and/or adsorption of HA decreased gradually with increasing pH. Within the pH range from 6 to 9, the adsorption of HA on schwertmannite was equal. However, HA adsorption equilibrium was not achieved for schwertmannite in any of the tested pH. Failure to reach equilibrium resulted, on the one hand, from instability of the schwertmannite phase at high pH, and on the other hand, from coprecipitation of HA on schwertmannite surfaces at low pH. Moreover, the observed disequilibrium could have resulted from moderately low HA concentrations, corresponding to values encountered in natural waters. Maximum HA content in goethite after batch adsorption was lower (6.5% (w/w) C at pH 2) than that of schwertmannite (Fig. 4a, b). The HA adsorption curves bend towards typical adsorption equilibration patterns with increasing pH. Coprecipitation of HA in both schwertmannite and goethite suspensions at pH 2–5 was also observed during filtration after washing of the suspension, which considerably slowed down in the presence of high concentrations of HA. In HA concentrations of 0.1–1 mg/L C, all added HA adsorbed on schwertmannite and goethite surfaces, thus indicating incomplete surface coverage and supporting our surface charge observations (see Section 3.2; Fig. 1e–f). Further, in HA concentration of 10 mg/L C, all added HA adsorbed on schwertmannite surfaces, but not on goethite; adsorption of HA on goethite was complete only in a pH range of 2–4. Percentages of adsorbed HA depended on pH, and varied (in HA concentration of 100 mg/L C) between 35% and 100% for schwertmannite and 5% and 65% for goethite (Fig. 5). The specific surface area weighted results for synthetic schwertmannite and goethite were alike (Fig. 6). In HA
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Fig. 4 – Final C content of (a) schwertmannite and (b) goethite after equilibrium with 0.1, 1, 10 and 100 mg/L C humic acid solutions with various pH values.
Fig. 5 – Percentages of adsorbed C from total available C for schwertmannite (SS, open symbols) and goethite (SG, closed symbols) in high (100 mg/L C) humic acid concentration. pH is final pH after batch adsorption experiment.
concentration of 10 mg/L C, the surface area weighted HA adsorption on schwertmannite was slightly lower than on goethite. But unlike goethite surfaces, schwertmannite surfaces were not fully covered with HA after the 10 mg/L C HA batch adsorption experiment. In high HA concentration (100 mg/L C), especially with pH values above 6, the adsorption of HA was noticeably higher (0.1 mg/m2) on schwertmannite than on goethite (Fig. 6). However, taking into account the changes in specific surface area due to pH-dependent sulphate release (Section 3.3), the surface area weighted HA adsorption was very similar for schwertmannite and goethite. Sulphate release increases the reactive surface area of schwertmannite at pHo7.5 (Section 3.3; Jo¨nsson et al., 2005). In high HA concentrations (100 mg/L C), when surfaces were fully covered, adsorption minimum for schwertmannite was approximately at pH 7 (Fig. 6a), corresponding with the IEP for schwertmannite (Section 3.2). Schwertmannites synthesized in the presence of HA contained less HA and more sulphate than schwertmannites at pH 2 after batch adsorption experiments (Table 1; Fig. 2;
Fig. 4). Gu et al. (1994) and Jo¨nsson (2003) stated that adsorption of HA decreased in the presence of sulphate due to the competitive effects of SO2 4 with HA for surfaces sites. Likewise in our study, the competitive ion effect of sulphate explains the differences in adsorption of HA on schwertmannite between batch adsorption and synthesis experiments. In addition, HA adsorption on schwertmannite may have decreased in synthesis solutions due to the formation of organic ligand–iron complexes. Our HA adsorption results for schwertmannite and goethite are consistent with earlier studies, although they obtained slightly higher values (Gu et al., 1994; Jo¨nsson et al., 2006). However, the results are not directly comparable due to the different origins of NOM and different iron oxides used. For example, Jo¨nsson et al. (2006) examined natural schwertmannite, which inherently contained significant amounts of carbon. Further, they failed to take into account pHdependent changes in specific surface area. Commercial HA, which we used, is well characterized, but differs significantly from NOM occurring at natural environments by having a higher molecular weight, higher aromaticity and lower oxygen content than for example, the aquatic HAs and FAs of the International Humic Substances Society (O’Loughlin et al., 2000). Due to the higher molecular weight of the HA which we used, a slight increase in the adsorption was expected.
3.5.
Particle size effects
Adsorption of HA on iron oxides has been shown to increase with increasing ionic strength (Saito et al., 2004; Ille´s and Tomba´cz, 2004; Weng et al., 2006). Particle size of iron oxides increases simultaneously with ionic strength (Ille´s and Tomba´cz, 2006). Primary particles of schwertmannite are reported to be 2–4 nm thick and 60–90 nm in length, but aggregated densely to larger-sized, 200–500 nm, particles (Bigham et al., 1994). In this study, minimum particle size of initial schwertmannite and goethite in water, determined by laser diffraction, was 270 and 30 nm, respectively (Fig. 7). After batch adsorption experiment, goethite particles formed
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Fig. 6 – Specific surface area (initial SSA) weighted humic acid adsorption on (a) schwertmannite and (b) goethite surfaces in different initial humic acid concentrations and pH, whereby pH is final pH after batch adsorption experiment. Continuous lines with filled symbols represent data calculated using initial surface area, and dashed lines with empty symbols represent data calculated using surface area of schwertmannite determined after batch adsorption experiment in 0.1 mg/L C humic acid solution.
4.
Conclusions
This paper examined the adsorption of HA on iron oxide surfaces, the changes in surface charge, the effect of surface area and particle size to evaluate the colloidal stability of iron oxides in variable pH and in the presence of sulphate and HA. The following conclusions were drawn:
Fig. 7 – Volumetric particle size distribution of pure synthetic schwertmannite (SS) and goethite (SG) resuspended in water, stirred with 1500 rpm (thick line) and 1500 rpm using ultrasonic (thin line). Schwertmannite is represented by the continuous line and goethite by the dashed line.
non-transparent suspension in high (100 mg/L C) or low (0.1 mg/L C) HA concentration, which was expected, based on highly negative or positive zeta potentials (Fig. 1f). In contrast, schwertmannite suspensions were transparent after batch adsorption experiments, and settled despite the strongly negative zeta potentials in moderately acid pH (Fig. 1e). Due to continuous release of sulphate from schwertmannite (Section 3.3), sulphate is always present in small amounts in a suspension consisting of schwertmannite, which may lead to aggregation and settling. As with pure schwertmannites, schwertmannites synthesized in the presence of HA aggregated rapidly and settled. Aggregates had the size of several hundred nanometres to several tens of micrometres. Further investigations are been carried out to determine particle size effects of sulphate on schwertmannite and goethite colloids in acid mine drainage environments.
(1) The formation of schwertmannite can also occur in the presence of HA. Adsorbed or coprecipitated HA retards outward diffusion of sulphate from schwertmannite. (2) Sulphate similarly decreases schwertmannite and goethite surface charges, inducing coagulation. Release of sulphate from schwertmannite increases with increasing pH (up to pH 7). Sulphate release from schwertmannite surfaces leads to an increase in the surface area. (3) In reasonably high HA concentrations (10–100 mg/L C) with moderately acid conditions (pH 43.5), both schwertmannite and goethite surfaces are strongly negatively charged (zeta potential o30 mV), thus promoting colloid stabilization and colloidal transport. (4) The specific surface area weighted HA adsorption is similar for schwertmannite and goethite. The presence of sulphate on schwertmannite surfaces or in solution acts as a competitive ion, which decreases HA adsorption. However, due to its much larger surface area, schwertmannite is a more efficient adsorbent of HA than goethite. (5) Despite the highly negative surface charges of schwertmannite in moderately acid environments (pH 43.5), large particle size of schwertmannite and presence of sulphate on schwertmannite surfaces seem to diminish the risk of colloid stabilization and particle transport.
Acknowledgements This study was completed with the support of Ernst Mach stipend granted by the Austrian Federal Ministry of Education,
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Science and Culture (BMBWK) and stipend granted by the Kone Foundation. We thank two anonymous reviewers for their valuable comments and Dr. John Plant, Vienna, for checking the language. R E F E R E N C E S
Alcacio, T.E., Hesterberg, D., Chou, J.W., Martin, J.D., Beauchemin, S., Sayers, D.E., 2001. Molecular scale characteristics of Cu(II) bonding in goethite–humate complexes. Geochim. Cosmochim. Acta 65, 1355–1366. Ali, M.A., Dzombak, D.A., 1996. Interactions of copper, organic acids, and sulfate in goethite suspensions. Geochim. Cosmochim. Acta 60, 5045–5053. Au, K.-K., Penisson, A.C., Yang, S., O’Melia, C.R., 1999. Natural organic matter at oxide/water interfaces: complexation and conformation. Geochim. Cosmochim. Acta 63, 2903–2917. Bigham, J.M., Schwertmann, U., Carlson, L., Murad, E., 1990. A poorly crystallized oxyhydroxysulfate of iron formed by bacterial oxidation of Fe(II) in acid mine waters. Geochim. Cosmochim. Acta 54, 2743–2758. Bigham, J.M., Carlson, L., Murad, E., 1994. Schwertmannite, a new iron oxyhydroxysulphate from Pyha¨salmi, Finland, and other localities. Mineralogic. Mag. 58, 641–648. Bigham, J.M., Schwertmann, U., Traina, S.J., Winland, R.L., Wolf, M., 1996. Schwertmannite and the chemical modeling of iron in acid sulfate waters. Geochim. Cosmochim. Acta 60, 2111–2121. Brunauer, S., Emmett, P.H., Teller, E., 1938. Adsorption of gases in multimolecular layers. J. Am. Chem. Soc. 60, 309–319. Cornell, R.M., Schwertmann, U., 1979. Influence of organic anions on the crystallization of ferrihydrite. Clays Clay Miner. 27, 402–410. Evans, D.F., Wennerstro¨m, H., 1999. The Colloidal Domain: Where Physics, Chemistry, Biology, and Technology Meet, second ed. Wiley-VCH, New York, 632p. Ferris, F.G., Tazaki, K., Fyfe, W.S., 1989. Iron oxides in acid mine drainage environments and their association with bacteria. Chem. Geol. 74, 321–330. Filius, J.D., Meeussen, J.C.L., Lumsdon, D.G., Hiemstra, T., van Riemsdijk, W.H., 2003. Modelling the binding of fulvic acid by goethite: the speciation of adsorbed FA molecules. Geochim. Cosmochim. Acta 67, 1463–1474. Gaboriaud, F., Ehrhardt, J.-J., 2003. Effects of different crystal faces on the surface charge of colloidal goethite (a-FeOOH) particles: an experimental and modeling study. Geochim. Cosmochim. Acta 67, 967–983. Gagliano, W.B., Brill, M.R., Bigham, J.M., Jones, F.S., Traina, S.J., 2004. Chemistry and mineralogy of ochreous sediments in a constructed mine drainage wetland. Geochim. Cosmochim. Acta 68, 2119–2128. Gu, B., Schmitt, J., Chen, Z., Liang, L., McCarthy, J.F., 1994. Adsorption and desorption of natural organic matter on iron oxide: mechanisms and models. Environ. Sci. Technol. 28, 38–46. Gu, B., Schmitt, J., Chen, Z., Liang, L., McCarthy, J.F., 1995. Adsorption and desorption of different organic matter fractions on iron oxide. Geochim. Cosmochim. Acta 59, 219–229. Herrera Ramos, A.C., McBride, M.B., 1996. Goethite dispersibility in solutions of variable ionic strength and soluble organic matter content. Clays Clay Miner. 44, 286–296. Ille´s, E., Tomba´cz, E., 2004. The role of variable surface charge and surface complexation in the adsorption of humic acid on magnetite. Colloids Surf. A: Physicochem. Eng. Aspects 230, 99–109. Ille´s, E., Tomba´cz, E., 2006. The effect of humic acid adsorption on pH-dependent surface charging and aggregation of magnetite nanoparticles. J. Colloid Interface Sci. 295, 115–123.
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Jo¨nsson, J., 2003. Phase transformations and surface chemistry of secondary iron minerals formed from acid mine drainage. Doctoral Thesis, Department of Chemistry, Inorganic Chemistry, Umea˚ University, Sweden, 71p. ISBN 91-7305-552-2. Jo¨nsson, J., Persson, P., Sjo¨berg, S., Lo¨vgren, L., 2005. Schwertmannite precipitated from acid mine drainage: phase transformation, sulphate release and surface properties. Appl. Geochem. 20, 179–191. Jo¨nsson, J., Sjo¨berg, S., Lo¨vgren, L., 2006. Adsorption of Cu(II) to schwertmannite and goethite in presence of dissolved organic matter. Water Res. 40, 969–974. Karlsson, S., Allard, B., Ha˚kansson, K., 1988. Characterization of suspended solids in a stream receiving acid mine effluents, Bersbo, Sweden. Appl. Geochem. 3, 345–356. Kretzschmar, R., Sticher, H., 1997. Transport of humic-coated iron oxide colloids in a sandy soil: influence of Ca2+ and trace metals. Environ. Sci. Technol. 31, 3497–3504. Kumpulainen, S., Carlson, L., Ra¨isa¨nen, M.-L., 2007. Seasonal variations of ochreous precipitates in mine effluents in Finland. Appl. Geochem. 22, 760–777. Loan, M., Cowley, J.M., Hart, R., Parkinson, G.M., 2004. Evidence on the structure of synthetic schwertmannite. Am. Mineral. 89, 1735–1742. Lyklema, J., 1984. Points of zero charge in the presence of specific adsorption. J. Colloid Interface Sci. 99, 109–117. Majzlan, J., Navrotsky, A., Schwertmann, U., 2004. Thermodynamics of iron oxides: part III. Enthalpies of formation and stability of ferrihydrite (Fe(OH)3), schwertmannite (FeO(OH)3/4(SO4)1/8), and e-Fe2O3. Geochim. Cosmochim. Acta 68, 1049–1059. Matijevic, E., 1980. Colloid chemical aspects of corrosion of metals. Pure Appl. Chem. 52, 1179–1193. Murphy, E.M., Zachara, J.M., 1995. The role of sorbed humic substances on the distribution of organic and inorganic contaminants in groundwater. Geoderma 67, 103–124. Nordstrom, D.K., Alpers, C.N., 1999. Geochemistry of acid mine waters. In: Plumlee, G.S., Logsdon, J.J. (Eds.), The Environmental Geochemistry of Mineral Deposits, Part A. Processes, Techniques, and Health Issues, vol. 6A. Society of Economic Geologists, Reviews in Economic Geology, pp. 133–160. O’Loughlin, E.J., Traina, S.J., Chin, Y.-P., 2000. Association of organotin compounds with aquatic and terrestrial humic substances. Environ. Toxicol. Chem. 19, 2015–2021. Peak, D., Ford, R.G., Sparks, D.L., 1999. An in situ ATR-FTIR investigation of sulfate bonding mechanisms on goethite. J. Colloid Interface Sci. 218, 289–299. Peiffer, S., Walton-Day, K., Macalady, D.L., 1999. The interaction of natural organic matter with iron in a wetland (Tennessee Park, Colorado) receiving acid mine drainage. Aquat. Geochem. 5, 207–223. Persson, P., Lo¨vgren, L., 1996. Potentiometric and spectroscopic studies of sulfate complexation at the goethite–water interface. Geochim. Cosmochim. Acta 60, 2789–2799. Regenspurg, S., Brand, A., Peiffer, S., 2004. Formation and stability of schwertmannite in acidic mining lakes. Geochim. Cosmochim. Acta 68, 1185–1197. Ren, J., Packman, A.I., 2005. Coupled stream—subsurface exchange of colloidal hematite and dissolved zinc, copper and phosphate. Environ. Sci. Technol. 39, 6387–6394. Rietra, R.P.J.J., Hiemstra, T., van Riemsdijk, W.H., 2001. Comparison of selenate and sulfate adsorption on goethite. J. Colloid Interface Sci. 240, 384–390. Russell, J.D., Paterson, E., Fraser, A.R., Farmer, V.C., 1975. Adsorption of carbon dioxide on goethite (a-FeOOH) surfaces, and its implications for anion adsorption. J. Chem. Soc., Faraday Trans. I 71, 1623–1630. Saito, T., Koopal, L.K., van Riemsdijk, W.H., Nagasaki, S., Tanaka, S., 2004. Adsorption of humic acid on goethite: isotherms,
ARTICLE IN PRESS 2060
WAT E R R E S E A R C H
42 (2008) 2051– 2060
charge adjustments, and potential profiles. Langmuir 20, 689–700. Salminen, R., Batista, M.J., Bidovec, M., Demetriades, A., De Vivo, B., De Vos, W., Duris, M., Gilucis, A., Gregorauskiene, V., Halamic, J., Heitzmann, P., Lima, A., Jordan, G., Klaver, G., Klein, P., Lis, J., Locutura, J., Marsina, K., Mazreku, A., ˚ ., Ottesen, R.-T., Petersell, V., Plant, O’Connor, P.J., Olsson, S.A J.A., Reeder, S., Salpeteur, I., Sandstro¨m, H., Siewers, U., Steenfelt, A., Tarvainen, T., 2005. Geochemical Atlas of Europe. Part 1: Background Information, Methodology and Maps. Geological Survey of Finland, 525p. Schemel, L.E., Kimball, B.A., Bencala, K.E., 2000. Colloid formation and metal transport through two mixing zones affected by acid mine drainage near Silverton, Colorado. Appl. Geochem. 15, 1003–1018. Schudel, M., Behrens, S.H., Holthoff, H., Kretzschmar, R., Borkovec, M., 1997. Absolute aggregation rate constants of hematite particles in aqueous suspension: a comparison of two different surface morphologies. J. Colloid Interface Sci. 196, 241–253. Schwertmann, U., Cornell, R.M., 2000. Iron Oxides in the Laboratory, second ed. VCH, Weinheim, 188p. Slowey, A.J., Johnson, S.B., Rytuba, J.J., Brown Jr., G.E., 2005. Role of organic acids in promoting colloidal transport of mercury from mine tailings. Environ. Sci. Technol. 39, 7869–7874. Stjernman Forsberg, L., Ledin, S., 2003. Effects of iron precipitation and organic amendments on porosity and penetrability in sulphide mine tailings. Water Air Soil Pollut. 142, 395–408. Stumm, W., 1992. Chemistry of the Solid–Water Interface. Processes at the Mineral–Water and Particle–Water Interface in Natural Systems. Wiley, New York, 428p.
Suteerapataranon, S., Boyby, M., Geckeis, H., Fangha¨nel, T., Grudpan, K., 2006. Interaction of trace elements in acid mine drainage solution with humic acid. Water Res. 40, 2044–2054. Tessier, A., Fortin, D., Belzile, N., DeVitre, R.R., Leppard, G.G., 1996. Metal sorption to diagenetic iron and manganese oxyhydroxides and associated organic matter: narrowing the gap between field and laboratory measurements. Geochim. Cosmochim. Acta 60, 387–404. Tipping, E., 1981. Adsorption by goethite (a-FeOOH) of humic substances from three different lakes. Chem. Geol. 33, 81–89. Vermeer, A.W.P., van Riemsdijk, W.H., Koopal, L.K., 1998. Adsorption of humic acid to mineral particles. 1. Specific and electrostatic interactions. Langmuir 14, 2810–2819. van Geen, A., Robertson, A.P., Leckie, J.O., 1994. Complexation of carbonate species at the goethite surface: implications for adsorption of metal ions in natural waters. Geochim. Cosmochim. Acta 58, 2073–2086. Weidler, P.G., 1997. BET sample pretreatment of synthetic ferrihydrite and its influence on the determination of surface area and porosity. J. Porous Mater. 4, 165–169. Weng, L., van Riemsdijk, W.H., Koopal, L.K., Hiemstra, T., 2006. Adsorption of humic substances on goethite: comparison between humic acids and fulvic acids. Environ. Sci. Technol. 40, 7494–7500. Za¨nker, H., Moll, H., Richter, W., Brendler, V., Hennig, C., Reich, T., Kluge, A., Hu¨ttig, G., 2002. The colloid chemistry of acid rock drainage solution from an abandoned Zn–Pb–Ag mine. Appl. Geochem. 17, 633–648. Zuyi, T., Taiwei, C., Jinzhou, D., XiongXin, D., Yingjie, G., 2000. Effect of fulvic acids on sorption of U(VI), Zn, Yb, I and Se(IV) onto oxides of aluminium, iron and silicon. Appl. Geochem. 15, 133–139.
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Available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Humic acid adsorption and surface charge effects on schwertmannite and goethite in acid sulphate waters Sirpa Kumpulainena,b,, Frank von der Kammerb, Thilo Hofmannb a
Department of Geology, University of Helsinki, P.O. Box 64, 00014 Helsinki, Finland Department of Environmental Geosciences, Althanstrasse 14, 1090 Vienna, Austria
b
art i cle info
ab st rac t
Article history:
In acid conditions, as in acid mine drainage waters, iron oxide particles are positively
Received 13 September 2007
charged, attracting negatively charged organic particles present in surrounding natural
Received in revised form
waters. Schwertmannite (Fe8O8(OH)6SO4) and goethite (a-FeOOH) are the most typical iron
17 December 2007
oxide minerals found in mine effluents. We studied schwertmannite formation in the
Accepted 18 December 2007
presence of humic acid. Further, surface charge and adsorption of humic acid on synthetic
Available online 27 December 2007
schwertmannite and goethite surfaces in pH 2–9 and in humic acid concentrations of
Keywords: Acid mine drainage Schwertmannite Humic acid Adsorption Surface charge Particles
0.1–100 mg/L C were examined. Schwertmannite did precipitate despite the presence of humic acid, although it contained more sulphate and had higher specific surface area than ordinary schwertmannite. Specific surface area weighted results showed that schwertmannite and goethite had similar humic acid adsorption capacities. Sulphate was released from schwertmannite surfaces with increasing pH, resulting in an increase in specific surface area. Presence of sulphate in solution decreased the surface charge of schwertmannite and goethite similarly, causing coagulation. In acid conditions (pH 2–3.5), according to the zeta potential, schwertmannite is expected to coagulate even in the presence of high concentrations of humic acid (p100 mg/L C). However, at high humic acid concentrations (10–100 mg/L C) with moderate acid conditions (pH43.5), both schwertmannite and goethite surfaces are strongly negatively charged (zeta potential o30 mV) thus posing a risk for colloid stabilization and colloidal transport. & 2007 Elsevier Ltd. All rights reserved.
1.
Introduction
Weathering of sulphides in tailings and waste rock piles produces acidic waters enriched with sulphate, iron and toxic elements (Nordstrom and Alpers, 1999). Schwertmannite (Fe8O8(OH)6(SO4)) and goethite (a-FeOOH) are the most typical iron(III) oxide minerals precipitated from acidic (pH 2.5–4.5), sulphate-rich (o3000 mg/L) waters (Bigham et al., 1996; Regenspurg et al., 2004; Kumpulainen et al., 2007). The formation of schwertmannite is thermodynamically favoured over ferrihydrite (Fe5HO8 4H2O) in the presence of sulphate in pH 2–7.5 (Majzlan et al., 2004). Schwertmannite is kinetically
unstable, and transforms to goethite over time (Bigham et al., 1996; Jo¨nsson et al., 2005). Adsorption of natural organic matter (NOM) suppresses the phase transformation from ferrihydrite to goethite (Cornell and Schwertmann, 1979), and is therefore expected to similarly retard the transformation of schwertmannite. Crystallization of iron oxides may be inhibited in the presence of organic ligands due to formation of complexes with dissolved iron (Cornell and Schwertmann, 1979). Sulphate adsorbs on goethite surfaces through the formation of both inner- and outer-sphere complexes. At pHo6, or high ionic strengths, formation of inner-sphere complexes is
Corresponding author. Department of Geology, University of Helsinki, P.O. Box 64, 00014 Helsinki, Finland. Fax: +358 9 191 50826.
E-mail address: [email protected] (S. Kumpulainen). 0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.12.015
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favoured (Peak et al., 1999; Rietra et al., 2001). Sulphate adsorption on schwertmannite is similar to that of goethite. One-third of sulphate in schwertmannite is specifically adsorbed to the mineral surface (Bigham et al., 1990), but cannot be discounted from mineral formula, because it has a stabilizing effect on schwertmannite particles (Bigham et al., 1996). Iron oxides are important adsorbents of heavy metals and other harmful elements. Adsorption capacity arises from large specific surface area and from development of surface charge. Specific surface area of iron oxides varies typically in mining environments 100–200 m2/g (Bigham et al., 1990; Gagliano et al., 2004). Iron oxide surfaces are protonated, and therefore positively charged in acid conditions, and dehydroxylated and negatively charged in alkaline conditions. The pH at which the surface has equal amount of positive and negative adsorption sites is called the point of zero charge (PZC) and is measured by potentiometric titration. If no specifically adsorbed ions are present, PZC equals the isoelectric point (IEP) measured by electrophoresis. IEP is the point where no movement of particles occurs in an electrical field and where zeta potential is zero (Stumm, 1992). Pure synthetic goethite has PZC at pH 9.0 (van Geen et al., 1994; Gaboriaud and Ehrhardt, 2003). Natural iron oxides have a much lower PZC than synthetic oxides because of impuor rities. Surface-adsorbed anionic elements such as CO2 3 can significantly reduce the surface charge, whereas SO2 4 adsorption of cationic substances such as Cu2+ or Zn+ increase the surface charge (Russell et al., 1975; van Geen et al., 1994; Ren and Packman, 2005). Not only adsorption tendency but also aggregation tendency has great impact on the transport of heavy metals. Iron oxides tend to coagulate in near-neutral pH or in high ionic strengths (Matijevic, 1980; Schudel et al., 1997), and therefore, generally considered to settle in acid mine waters. However, colloidal size particles have been identified from mine waters (Karlsson et al., 1988; Schemel et al., 2000; Za¨nker et al., 2002; Regenspurg et al., 2004). When waters from mine sites discharge into the surrounding environments, they come into contact with waters enriched with NOM. Median concentration of dissolved organic carbon in stream waters in Europe is 5 mg/L, but varies up to several tens of mg/L (Salminen et al., 2005). Adsorption of NOM on goethite and hematite has been studied by various authors (Tipping, 1981; Gu et al., 1994; Au et al., 1999), but investigations focusing on conditions typical for mine environments, such as Jo¨nsson et al. (2006), have been scarce. Dilution of acid mine drainage waters with humic acid (HA)-containing waters can cause either retention or mobilization of metals depending on the metal to organic matter ratio and pH (Suteerapataranon et al., 2006). Organic amendments added to tailings may induce aggregation of iron oxides cementing the tailings (Stjernman Forsberg and Ledin, 2003). On the other hand, some organic acids produced by plants, such as citrate, dissolve already formed cements, mobilizing the colloids (Slowey et al., 2005). Iron has been shown to be strongly associated with NOM in a wetland receiving acid mine drainage. In the presence of high concentrations of NOM (4100 mg/L), iron particles were stabilized in the colloidal size range (Peiffer et al., 1999).
NOM consists of various organic substances that are formed during decomposition of vegetation and animal tissues. Due to variability of chemical and physical properties, different fractions of NOM exhibit different adsorption behaviours. In general, carboxyl and hydroxyl functional groups are responsible for adsorption (Gu et al., 1995). Adsorption of NOM occurs through ligand exchange reactions and formation of surface complexes. Carboxylic groups adsorb especially at low pH, whereas adsorption of hydroxylic groups becomes more important at high pH (Filius et al., 2003). HAs are high molecular weight organic materials, which are precipitated generally in a pH range of 2–3 in contrast to fulvic acids (FAs), which are lighter, and also soluble at pHo2. Adsorption of organic matter on iron oxide surfaces lowers the surface charge, causing a charge reversal even at small loadings (0.1–0.2 mg/L), followed by aggregation and settling of particles. Addition of larger amounts of organic matter increases the negative surface charge and colloidal stability (Herrera Ramos and McBride, 1996; Kretzschmar and Sticher, 1997). Adsorption of NOM on iron oxide surface also affects the mobility of harmful elements. In the presence of NOM, adsorption of Cu and U and to a lesser extent also Co, has been shown to be enhanced (Murphy and Zachara, 1995; Ali and Dzombak, 1996; Zuyi et al., 2000; Jo¨nsson et al., 2006). The presence of sulphate can decrease adsorption of NOM due to competitive ion effect (Gu et al., 1994; Ali and Dzombak, 1996; Jo¨nsson, 2003). In high NOM concentrations, some metals, such as Cu, might also tend to become more mobile (Alcacio et al., 2001; Jo¨nsson, 2003). This has been explained by coating of iron hydroxide surfaces with organic material and complexing tendency of soluble organic acids towards Cu. Trace metals bind directly to the iron hydroxide surfaces in circum neutral pH, but in more acidic conditions they bind to the functional groups of organic matter adsorbed on iron hydroxides (Tessier et al., 1996). The objective of the present paper was to study the interactions of NOM and iron oxides typical for acid mine environments, their effect on surface charge of iron oxides, and further effect of NOM adsorption on colloidal stability. First, schwertmannite formation in the presence of HA was studied. Then, a series of adsorption experiments with schwertmannite and goethite were done to evaluate the adsorption of HA and its effects on surface charge in variable pH and HA concentrations.
2.
Material and methods
2.1.
Precipitate synthesis
Schwertmannite (ideally Fe8O8(OH)6(SO4)) was synthesized with a method described by Loan et al. (2004). Five hundred millilitres of preheated MilliQ water was mixed with 2.506 g of exsiccated Fe2(SO4)3 xH2O (purissimum p.a. grade) in a PE bottle. The suspension was stirred in a water bath at 85 1C for 1 h. The pH of the suspension was 1.8. The suspension was filtered through 0.45 mm pore size Whatman cellulose filter, the residue was washed twice with MilliQ water, and freeze dried. The method produced 0.5 g of freeze-dried precipitate.
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Similarly, schwertmannite was synthesized in the presence of HAs (0.1, 1, 10 and 100 mg/L C) adding HA before iron sulphate addition. Goethite (a-FeOOH) was synthesized according to the method of Schwertmann and Cornell (2000). The suspension was filtered through 0.1 mm pore size Whatman cellulose filter, the residue was washed twice with MilliQ water, and freeze dried. Synthetic schwertmannite had a composition of Fe8O8(OH)4.92(SO4)1.54 and a specific surface area of 130 m2/g. Synthetic goethite had a specific surface area of 72 m2/g. Both schwertmannite and goethite contained small amounts of carbon impurities, 0.19% (w/w) and 0.87% (w/w), respectively.
2.2.
Humic acid purification
Adsorption experiments were conducted using commercial HA (Fluka). The HA was first purified from clay particles (initial ash content 20%), trace metals, humin and FA fractions with the method described by Vermeer et al. (1998). Purified HA was freeze dried. Purified HA consisted of 55.3% (w/w) C, 3.87% (w/w) H, 1.15% (w/w) N, and 1.54% (w/w) S. Oxygen content was calculated: 38.2% (w/w) O. HA stock solution of 2 g/L was prepared and adjusted to pH 10 with NaOH in order to dissolve all humic material. Stock solution was allowed to equilibrate for 24 h. Solutions of 0.1, 1, 10 and 100 mg/L C were diluted from stock solution for the batch adsorption experiments. HA concentrations (0.1–100 mg/L C) were selected to correspond to values typically encountered in stream waters.
2.3.
Batch adsorption experiments
Adsorption of HA on schwertmannite and goethite surfaces was studied with batch adsorption experiments using iron oxide to solution ratio of 1 g/L. The ratio was the same as in schwertmannite synthesis in the presence of HA. Suspension pH was adjusted before HA addition to 2, 3, 4, 5, 6, 7, 8 and 9 by adding HCl or NaOH, shaking for 24 h and adjusting the pH once more, if required, before HA addition. HA solution with strength of 0.1, 1, 10 or 100 mg/L C and pH adjusted similarly to the schwertmannite suspensions was added and shaken for 24 h. After batch adsorption, all precipitates were washed twice, filtrated and freeze dried.
2.4.
Chemical analysis
Iron hydroxide Fe content was determined by direct current plasma optical emission spectrometry. Mean standard error was 71.1% (w/w). C, H, N and S content were determined by combustion (Vario Macro elemental analyser). All measurements were analysed in duplicate, and mean standard error was 70.01% (w/w) for C, 70.004% (w/w) for H, 70.01% (w/w) for N and 70.05% (w/w) for S.
2.5.
Mineralogical analysis
Mineralogy of synthesized precipitates was determined by X-ray powder diffraction. CuKa-radiation and a Philips X0 Pert
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diffractometer equipped with a vertical goniometry and a curved Cu diffracted beam monochromator were used. The patterns were scanned from 2 to 701 2y with 0.021 steps and counting rate 1 s/step.
2.6.
Particle size
Particle size distribution of synthetic oxides and selected precipitates after batch adsorption experiments was studied with laser diffraction using a Malvern Mastersizer Hydro 2000 mP by calculating volume-based size distribution according to the Mie scattering theory. Refractive index of schwertmannite and goethite was set to 2.17, and adsorption to 0.1. All measurements were performed in triplicate. Mean particle size for synthetic schwertmannite and goethite after resuspension and moderate mixing rate (1500 rpm) was 10 and 29 mm, respectively. Minimum particle size was determined from mixed and sonicated suspension. Minimum particle size for schwertmannite and goethite was 270 and 30 nm, respectively. The particle size distribution results were used to select suitable pore size filters, 0.45 mm for schwertmannite and 0.1 mm for goethite, in order to minimize the loss of particles in the filtration process and to keep the time of filtration reasonable.
2.7.
Surface charge
Surface charge of synthetic schwertmannites and goethites was evaluated by determining electrophoretic mobility of suspended particles with Malvern Zetasizer Nano ZS. The Henry equation and Smoluchowski approximation were applied to calculate zeta potential. All measurements were conducted at +23 1C and calculated by triplicate analysis. Colloidal suspension was considered electrostatically stable, when zeta potential of particles was 4+30 mV oro30 mV (Evans and Wennerstro¨m, 1999). IEP for pure oxides in different concentrations of NaCl (0.001, 0.01 and 0.1 M) and Na2SO4 (0.1 and 1 mM) was determined by titration. Before titration, suspension was allowed to equilibrate and settle for 20 h. Titration of 10 mL of precipitate suspension (1 g/L) with 0.1 M NaOH or 0.1 M HCl was carried out from pH 2 to 9, or from 9 to 2, depending on the zeta potential in the start of the titration, in order to avoid crossing of IEP in the beginning. Zeta potentials of single samples were determined after batch adsorption experiments with HA solutions. Suspensions were left to settle for at least 2 h. A small volume of suspension was collected underneath the surface for zeta potential measurements. Surface charge determinations were carried out in the presence of atmospheric oxygen and carbon dioxide, which correspond to conditions in shallow surface waters.
2.8.
Specific surface area
Specific surface area of synthesized precipitates and selected schwertmannites after batch adsorption experiments was determined using the Brunauer, Emmett and Teller multipoint method (Brunauer et al., 1938). Before measurements, freeze-dried samples were degassed in 110 1C for 30 min. Temperatures between 90 and 120 1C are necessary to remove
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hygroscopic water (Weidler, 1997) and should not induce mineralogical changes (Regenspurg et al., 2004). Determination of specific surface area was by Quantachrome 2200, measuring adsorption of N2 in cryogenic temperatures from 6 points in the range Po/P 0.05–0.3. Determinations were performed in duplicate, and mean standard error was 71.2 m2/g.
3.
Results and discussion
3.1. Schwertmannites synthesized in the presence of humic acid Series of schwertmannite produced in the presence of different concentrations of HA showed no significant effect on the degree of schwertmannite crystallinity. Precipitation of schwertmannite in the presence of HA was expected since negatively charged surface ligands on bacterial cells are known to act as passive nucleation sites for iron oxides in acid environments (Ferris et al., 1989). Hydrolysis of iron and precipitation of schwertmannite resulted in a low pH (pH 1.8), which further led to coprecipitation of HA together with schwertmannite. Fe/S ratios of forming schwertmannite decreased with increasing HA concentrations (Table 1). Decreasing molar Fe/S ratios (Table 1), increased sulphate contents (Table 1) and low electric conductivities of washing water after synthesis indicated that coprecipitated HA retarded outward diffusion of adsorbed sulphate from schwertmannite particles. Specific surface area of forming schwertmannite increased in the presence of 0.1–10 mg/L C HA (Table 1). However, in the presence of higher HA concentrations (100 mg/L C), specific surface area decreased probably due to the blocking effect of coprecipitating HA. Even so, specific surface area was higher than that of pure schwertmannite.
3.2.
Effects of sulphate and humic acid on surface charge
Zeta potential of pure synthetic schwertmannite in NaCl solutions always fell within 730 mV irrespective of solution pH, and it promoted flocculation and settling (Fig. 1a). Opposite to this, synthetic goethite suspension showed typical zeta values 4+30 mV at pH 2–6, promoting electrostatic stabilization of colloids (Fig. 1b). The IEP (IEPpH) in
0.1–0.001 M NaCl solutions for synthetic schwertmannite was 6.9–8.2, and for synthetic goethite 8.5. The IEPpH values for synthetic schwertmannite were in good agreement with the value (at pH 7.2) reported by Jo¨nsson et al. (2005) for natural schwertmannite. The IEPpH for goethite was slightly lower than earlier reports (van Geen et al., 1994; Gaboriaud and Ehrhardt, 2003) due to the presence of atmospheric carbon dioxide, which can adsorb on iron oxide surfaces thus lowering the surface charge. Common point of intersection (cpi), the point where titration curves intersect each other, was same as IEPpH for goethite, but schwertmannite showed a cpi at pH 5.4 and at zeta potential +10 mV. In potentiometric titrations, the different cpi from PZC indicates the presence of specifically adsorbing anions (Lyklema, 1984), such as sulphate. Schwertmannite is unstable with respect to goethite. After resuspension, specifically adsorbed sulphate starts to desorb from schwertmannite surfaces. Sulphate release continues until schwertmannite has transformed fully to goethite (Bigham et al., 1996). In the presence of reactive solute (Na2SO4), zeta potential decreased due to adsorption of sulphate anions on iron oxide surfaces (Fig. 1c–d). In a sulphate concentration of 0.1 mM, zeta potential of schwertmannite was similar to that in NaCl solutions (Fig. 1a). Small amounts of sulphate (0.1 mM) considerably decreased the zeta potential of goethite from +35 mV in 1 mM NaCl solution to +20 mV in 0.1 mM Na2SO4 solution at pH 2–5 (Fig. 1b, d). In sulphate concentrations of 1 mM, the IEP for schwertmannite decreased to 5.7 and for goethite to 6.8. Above pH 8, sulphate adsorption should not occur on goethite surfaces (Persson and Lo¨vgren, 1996), and schwertmannite becomes unstable (Majzlan et al., 2004). Schwertmannite and goethite surfaces behaved similarly in the presence of sulphate, and had zeta potentials within 730 mV, promoting flocculation and settling. In the presence of 0.1, 1, 10 and 100 mg/L C of HA, the IEPs for schwertmannite were 4.5, 3.1, 2.8 and 2.5, respectively (Fig. 1e). The IEPs for goethite in the presence of 0.1, 1 and 10 mg/L C of HA were 8.2, 6.9 and 2.2, respectively (Fig. 1f). In the presence of 100 mg/L C HA, zeta potential of goethite was always negative. The difference in surface charges between low (0.1 and 1 mg/L C) and high (10 and 100 mg/L C) HA concentrations for goethite (Fig. 1f) was thought to be related to surface
Table 1 – Chemical composition (Fe, SO4 and C) and specific surface area (SSA) of precipitates formed after schwertmannite synthesis in the presence of humic acid (HA)
Synthetic schwertmannite (SS) SS+0.1 mg/L C HA in synthesis SS+1 mg/L C HA in synthesis SS+10 mg/L C HA in synthesis SS+100 mg/L C HA in synthesis
Fe/S (mol/mol)
Fe % (w/w)
SO4 % (w/w)
C % (w/w)
SSA (m2/g)
5.19 4.95 4.83 4.70 4.62
46.1 45.3 45.6 44.4 40.0
15.2 15.7 16.2 16.2 14.9
0.19 0.11 0.20 0.90 5.51
130 232 233 261 152
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Fig. 1 – Zeta potentials of synthetic schwertmannite (left) and synthetic goethite (right) in different pH and concentrations of NaCl, Na2SO4 and humic acid (HA) solutions. Initial iron oxide concentration was 1 g/L. Zeta potentials for iron oxides in NaCl and Na2SO4 solutions were measured by titration. Iron oxides in humic acid solutions were measured by determining zeta potentials of single samples after manual pH adjustment, where pH represents the final pH after batch adsorption experiment.
coverage: in high HA concentrations (10–100 mg/L C) goethite particles were assumed to be fully covered with HA masking the electrochemical properties of initial iron oxide surfaces. This was examined by evaluating the extent of adsorption of HA on goethite surfaces (see Section 3.4). The difference in zeta potentials in the presence of low concentrations of HA between schwertmannite and goethite was, on the other hand, thought to be related to the continuous and pH-
dependent sulphate release from schwertmannite, which lowers the zeta potential (Fig. 1a, c, e).
3.3. Desorption of SO4 from schwertmannite and effect on surface area Sulphate content of schwertmannite after batch adsorption experiments (Fig. 2) was lower (p13.7% (w/w)) than that of
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Fig. 2 – Sulphate content of schwertmannite after humic acid batch adsorption experiments in various initial humic acid concentrations (0.1, 1, 10 and 100 mg/L C). pH is final pH after batch adsorption experiment.
original schwertmannite (15.2% (w/w)). Further, electric conductivities of schwertmannite suspensions after batch adsorption experiments amounted to 100–250 mS/cm higher than those of goethite suspensions. The sulphate content of schwertmannite varied according to different pH values (Fig. 2), decreasing with increasing pH. The amount of released sulphate after pH adjustment and batch adsorption experiments with HA increased with pH, and amounted to 0.15–1.35 mmol/g. These values are consistent with those reported by Jo¨nsson et al. (2005) for sulphate release from natural schwertmannite in water with variable pH. In high HA concentrations (100 mg/L C), the release of sulphate was slightly higher than in low HA concentrations (0.1–10 mg/L C), which cannot be explained only by changes in mass due to HA adsorption or coprecipitation. Thus, adsorption of HA at high HA concentrations (100 mg/L C) must be accompanied by enhanced desorption of sulphate from schwertmannite. Specific surface area of selected precipitates was studied after batch adsorption experiments. Specific surface area of schwertmannite increased with pH in 0.1 mg/L C HA suspension, from 108 m2/g at pH 2 to 290 m2/g at pH 8 (Fig. 3). The increase was linear with increasing pH until pH 7, which is consistent with Jo¨nsson et al. (2005). One-third of sulphate in schwertmannite is adsorbed on surfaces, corresponding with a maximum sulphate adsorption capacity of 2.6 mmol/m2 (Bigham et al., 1990). Desorption of sulphate from schwertmannite in this study was higher than maximum sulphate adsorption capacity in HA concentrations of 0.1–10 mg/L C above pH 7. In high HA concentrations (100 mg/L C), sulphate release from schwertmannite was higher than maximum sulphate adsorption capacity in the entire pH range (2–9). Thus, part of the sulphate must have originated from the schwertmannite structure at pH 47 in low HA concentrations (0.1–10 mg/L C), and at pH 2–9 in high HA concentrations (100 mg/L C). At pH 47.5, the schwertmannite phase becomes unstable and was expected to begin to dissolve.
Fig. 3 – Specific surface area of selected schwertmannite samples after batch adsorption experiments in 0.1 mg/L C humic acid solutions. Dashed line represents the specific surface area of pure schwertmannite before pH adjustment and humic acid addition.
3.4.
Adsorption and coprecipitation of humic acid
Pure HA that was used in the batch adsorption experiments precipitated at pH 1.5. In HA batch adsorption experiments with schwertmannite, HA coprecipitated and/or adsorbed completely at pH 2, resulting in a schwertmannite that contained 9.5% (w/w) C HA after reaction with 100 mg/L C HA solution (Fig. 4a). In pH 3–5, coprecipitation and/or adsorption of HA decreased gradually with increasing pH. Within the pH range from 6 to 9, the adsorption of HA on schwertmannite was equal. However, HA adsorption equilibrium was not achieved for schwertmannite in any of the tested pH. Failure to reach equilibrium resulted, on the one hand, from instability of the schwertmannite phase at high pH, and on the other hand, from coprecipitation of HA on schwertmannite surfaces at low pH. Moreover, the observed disequilibrium could have resulted from moderately low HA concentrations, corresponding to values encountered in natural waters. Maximum HA content in goethite after batch adsorption was lower (6.5% (w/w) C at pH 2) than that of schwertmannite (Fig. 4a, b). The HA adsorption curves bend towards typical adsorption equilibration patterns with increasing pH. Coprecipitation of HA in both schwertmannite and goethite suspensions at pH 2–5 was also observed during filtration after washing of the suspension, which considerably slowed down in the presence of high concentrations of HA. In HA concentrations of 0.1–1 mg/L C, all added HA adsorbed on schwertmannite and goethite surfaces, thus indicating incomplete surface coverage and supporting our surface charge observations (see Section 3.2; Fig. 1e–f). Further, in HA concentration of 10 mg/L C, all added HA adsorbed on schwertmannite surfaces, but not on goethite; adsorption of HA on goethite was complete only in a pH range of 2–4. Percentages of adsorbed HA depended on pH, and varied (in HA concentration of 100 mg/L C) between 35% and 100% for schwertmannite and 5% and 65% for goethite (Fig. 5). The specific surface area weighted results for synthetic schwertmannite and goethite were alike (Fig. 6). In HA
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Fig. 4 – Final C content of (a) schwertmannite and (b) goethite after equilibrium with 0.1, 1, 10 and 100 mg/L C humic acid solutions with various pH values.
Fig. 5 – Percentages of adsorbed C from total available C for schwertmannite (SS, open symbols) and goethite (SG, closed symbols) in high (100 mg/L C) humic acid concentration. pH is final pH after batch adsorption experiment.
concentration of 10 mg/L C, the surface area weighted HA adsorption on schwertmannite was slightly lower than on goethite. But unlike goethite surfaces, schwertmannite surfaces were not fully covered with HA after the 10 mg/L C HA batch adsorption experiment. In high HA concentration (100 mg/L C), especially with pH values above 6, the adsorption of HA was noticeably higher (0.1 mg/m2) on schwertmannite than on goethite (Fig. 6). However, taking into account the changes in specific surface area due to pH-dependent sulphate release (Section 3.3), the surface area weighted HA adsorption was very similar for schwertmannite and goethite. Sulphate release increases the reactive surface area of schwertmannite at pHo7.5 (Section 3.3; Jo¨nsson et al., 2005). In high HA concentrations (100 mg/L C), when surfaces were fully covered, adsorption minimum for schwertmannite was approximately at pH 7 (Fig. 6a), corresponding with the IEP for schwertmannite (Section 3.2). Schwertmannites synthesized in the presence of HA contained less HA and more sulphate than schwertmannites at pH 2 after batch adsorption experiments (Table 1; Fig. 2;
Fig. 4). Gu et al. (1994) and Jo¨nsson (2003) stated that adsorption of HA decreased in the presence of sulphate due to the competitive effects of SO2 4 with HA for surfaces sites. Likewise in our study, the competitive ion effect of sulphate explains the differences in adsorption of HA on schwertmannite between batch adsorption and synthesis experiments. In addition, HA adsorption on schwertmannite may have decreased in synthesis solutions due to the formation of organic ligand–iron complexes. Our HA adsorption results for schwertmannite and goethite are consistent with earlier studies, although they obtained slightly higher values (Gu et al., 1994; Jo¨nsson et al., 2006). However, the results are not directly comparable due to the different origins of NOM and different iron oxides used. For example, Jo¨nsson et al. (2006) examined natural schwertmannite, which inherently contained significant amounts of carbon. Further, they failed to take into account pHdependent changes in specific surface area. Commercial HA, which we used, is well characterized, but differs significantly from NOM occurring at natural environments by having a higher molecular weight, higher aromaticity and lower oxygen content than for example, the aquatic HAs and FAs of the International Humic Substances Society (O’Loughlin et al., 2000). Due to the higher molecular weight of the HA which we used, a slight increase in the adsorption was expected.
3.5.
Particle size effects
Adsorption of HA on iron oxides has been shown to increase with increasing ionic strength (Saito et al., 2004; Ille´s and Tomba´cz, 2004; Weng et al., 2006). Particle size of iron oxides increases simultaneously with ionic strength (Ille´s and Tomba´cz, 2006). Primary particles of schwertmannite are reported to be 2–4 nm thick and 60–90 nm in length, but aggregated densely to larger-sized, 200–500 nm, particles (Bigham et al., 1994). In this study, minimum particle size of initial schwertmannite and goethite in water, determined by laser diffraction, was 270 and 30 nm, respectively (Fig. 7). After batch adsorption experiment, goethite particles formed
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Fig. 6 – Specific surface area (initial SSA) weighted humic acid adsorption on (a) schwertmannite and (b) goethite surfaces in different initial humic acid concentrations and pH, whereby pH is final pH after batch adsorption experiment. Continuous lines with filled symbols represent data calculated using initial surface area, and dashed lines with empty symbols represent data calculated using surface area of schwertmannite determined after batch adsorption experiment in 0.1 mg/L C humic acid solution.
4.
Conclusions
This paper examined the adsorption of HA on iron oxide surfaces, the changes in surface charge, the effect of surface area and particle size to evaluate the colloidal stability of iron oxides in variable pH and in the presence of sulphate and HA. The following conclusions were drawn:
Fig. 7 – Volumetric particle size distribution of pure synthetic schwertmannite (SS) and goethite (SG) resuspended in water, stirred with 1500 rpm (thick line) and 1500 rpm using ultrasonic (thin line). Schwertmannite is represented by the continuous line and goethite by the dashed line.
non-transparent suspension in high (100 mg/L C) or low (0.1 mg/L C) HA concentration, which was expected, based on highly negative or positive zeta potentials (Fig. 1f). In contrast, schwertmannite suspensions were transparent after batch adsorption experiments, and settled despite the strongly negative zeta potentials in moderately acid pH (Fig. 1e). Due to continuous release of sulphate from schwertmannite (Section 3.3), sulphate is always present in small amounts in a suspension consisting of schwertmannite, which may lead to aggregation and settling. As with pure schwertmannites, schwertmannites synthesized in the presence of HA aggregated rapidly and settled. Aggregates had the size of several hundred nanometres to several tens of micrometres. Further investigations are been carried out to determine particle size effects of sulphate on schwertmannite and goethite colloids in acid mine drainage environments.
(1) The formation of schwertmannite can also occur in the presence of HA. Adsorbed or coprecipitated HA retards outward diffusion of sulphate from schwertmannite. (2) Sulphate similarly decreases schwertmannite and goethite surface charges, inducing coagulation. Release of sulphate from schwertmannite increases with increasing pH (up to pH 7). Sulphate release from schwertmannite surfaces leads to an increase in the surface area. (3) In reasonably high HA concentrations (10–100 mg/L C) with moderately acid conditions (pH 43.5), both schwertmannite and goethite surfaces are strongly negatively charged (zeta potential o30 mV), thus promoting colloid stabilization and colloidal transport. (4) The specific surface area weighted HA adsorption is similar for schwertmannite and goethite. The presence of sulphate on schwertmannite surfaces or in solution acts as a competitive ion, which decreases HA adsorption. However, due to its much larger surface area, schwertmannite is a more efficient adsorbent of HA than goethite. (5) Despite the highly negative surface charges of schwertmannite in moderately acid environments (pH 43.5), large particle size of schwertmannite and presence of sulphate on schwertmannite surfaces seem to diminish the risk of colloid stabilization and particle transport.
Acknowledgements This study was completed with the support of Ernst Mach stipend granted by the Austrian Federal Ministry of Education,
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Science and Culture (BMBWK) and stipend granted by the Kone Foundation. We thank two anonymous reviewers for their valuable comments and Dr. John Plant, Vienna, for checking the language. R E F E R E N C E S
Alcacio, T.E., Hesterberg, D., Chou, J.W., Martin, J.D., Beauchemin, S., Sayers, D.E., 2001. Molecular scale characteristics of Cu(II) bonding in goethite–humate complexes. Geochim. Cosmochim. Acta 65, 1355–1366. Ali, M.A., Dzombak, D.A., 1996. Interactions of copper, organic acids, and sulfate in goethite suspensions. Geochim. Cosmochim. Acta 60, 5045–5053. Au, K.-K., Penisson, A.C., Yang, S., O’Melia, C.R., 1999. Natural organic matter at oxide/water interfaces: complexation and conformation. Geochim. Cosmochim. Acta 63, 2903–2917. Bigham, J.M., Schwertmann, U., Carlson, L., Murad, E., 1990. A poorly crystallized oxyhydroxysulfate of iron formed by bacterial oxidation of Fe(II) in acid mine waters. Geochim. Cosmochim. Acta 54, 2743–2758. Bigham, J.M., Carlson, L., Murad, E., 1994. Schwertmannite, a new iron oxyhydroxysulphate from Pyha¨salmi, Finland, and other localities. Mineralogic. Mag. 58, 641–648. Bigham, J.M., Schwertmann, U., Traina, S.J., Winland, R.L., Wolf, M., 1996. Schwertmannite and the chemical modeling of iron in acid sulfate waters. Geochim. Cosmochim. Acta 60, 2111–2121. Brunauer, S., Emmett, P.H., Teller, E., 1938. Adsorption of gases in multimolecular layers. J. Am. Chem. Soc. 60, 309–319. Cornell, R.M., Schwertmann, U., 1979. Influence of organic anions on the crystallization of ferrihydrite. Clays Clay Miner. 27, 402–410. Evans, D.F., Wennerstro¨m, H., 1999. The Colloidal Domain: Where Physics, Chemistry, Biology, and Technology Meet, second ed. Wiley-VCH, New York, 632p. Ferris, F.G., Tazaki, K., Fyfe, W.S., 1989. Iron oxides in acid mine drainage environments and their association with bacteria. Chem. Geol. 74, 321–330. Filius, J.D., Meeussen, J.C.L., Lumsdon, D.G., Hiemstra, T., van Riemsdijk, W.H., 2003. Modelling the binding of fulvic acid by goethite: the speciation of adsorbed FA molecules. Geochim. Cosmochim. Acta 67, 1463–1474. Gaboriaud, F., Ehrhardt, J.-J., 2003. Effects of different crystal faces on the surface charge of colloidal goethite (a-FeOOH) particles: an experimental and modeling study. Geochim. Cosmochim. Acta 67, 967–983. Gagliano, W.B., Brill, M.R., Bigham, J.M., Jones, F.S., Traina, S.J., 2004. Chemistry and mineralogy of ochreous sediments in a constructed mine drainage wetland. Geochim. Cosmochim. Acta 68, 2119–2128. Gu, B., Schmitt, J., Chen, Z., Liang, L., McCarthy, J.F., 1994. Adsorption and desorption of natural organic matter on iron oxide: mechanisms and models. Environ. Sci. Technol. 28, 38–46. Gu, B., Schmitt, J., Chen, Z., Liang, L., McCarthy, J.F., 1995. Adsorption and desorption of different organic matter fractions on iron oxide. Geochim. Cosmochim. Acta 59, 219–229. Herrera Ramos, A.C., McBride, M.B., 1996. Goethite dispersibility in solutions of variable ionic strength and soluble organic matter content. Clays Clay Miner. 44, 286–296. Ille´s, E., Tomba´cz, E., 2004. The role of variable surface charge and surface complexation in the adsorption of humic acid on magnetite. Colloids Surf. A: Physicochem. Eng. Aspects 230, 99–109. Ille´s, E., Tomba´cz, E., 2006. The effect of humic acid adsorption on pH-dependent surface charging and aggregation of magnetite nanoparticles. J. Colloid Interface Sci. 295, 115–123.
2059
Jo¨nsson, J., 2003. Phase transformations and surface chemistry of secondary iron minerals formed from acid mine drainage. Doctoral Thesis, Department of Chemistry, Inorganic Chemistry, Umea˚ University, Sweden, 71p. ISBN 91-7305-552-2. Jo¨nsson, J., Persson, P., Sjo¨berg, S., Lo¨vgren, L., 2005. Schwertmannite precipitated from acid mine drainage: phase transformation, sulphate release and surface properties. Appl. Geochem. 20, 179–191. Jo¨nsson, J., Sjo¨berg, S., Lo¨vgren, L., 2006. Adsorption of Cu(II) to schwertmannite and goethite in presence of dissolved organic matter. Water Res. 40, 969–974. Karlsson, S., Allard, B., Ha˚kansson, K., 1988. Characterization of suspended solids in a stream receiving acid mine effluents, Bersbo, Sweden. Appl. Geochem. 3, 345–356. Kretzschmar, R., Sticher, H., 1997. Transport of humic-coated iron oxide colloids in a sandy soil: influence of Ca2+ and trace metals. Environ. Sci. Technol. 31, 3497–3504. Kumpulainen, S., Carlson, L., Ra¨isa¨nen, M.-L., 2007. Seasonal variations of ochreous precipitates in mine effluents in Finland. Appl. Geochem. 22, 760–777. Loan, M., Cowley, J.M., Hart, R., Parkinson, G.M., 2004. Evidence on the structure of synthetic schwertmannite. Am. Mineral. 89, 1735–1742. Lyklema, J., 1984. Points of zero charge in the presence of specific adsorption. J. Colloid Interface Sci. 99, 109–117. Majzlan, J., Navrotsky, A., Schwertmann, U., 2004. Thermodynamics of iron oxides: part III. Enthalpies of formation and stability of ferrihydrite (Fe(OH)3), schwertmannite (FeO(OH)3/4(SO4)1/8), and e-Fe2O3. Geochim. Cosmochim. Acta 68, 1049–1059. Matijevic, E., 1980. Colloid chemical aspects of corrosion of metals. Pure Appl. Chem. 52, 1179–1193. Murphy, E.M., Zachara, J.M., 1995. The role of sorbed humic substances on the distribution of organic and inorganic contaminants in groundwater. Geoderma 67, 103–124. Nordstrom, D.K., Alpers, C.N., 1999. Geochemistry of acid mine waters. In: Plumlee, G.S., Logsdon, J.J. (Eds.), The Environmental Geochemistry of Mineral Deposits, Part A. Processes, Techniques, and Health Issues, vol. 6A. Society of Economic Geologists, Reviews in Economic Geology, pp. 133–160. O’Loughlin, E.J., Traina, S.J., Chin, Y.-P., 2000. Association of organotin compounds with aquatic and terrestrial humic substances. Environ. Toxicol. Chem. 19, 2015–2021. Peak, D., Ford, R.G., Sparks, D.L., 1999. An in situ ATR-FTIR investigation of sulfate bonding mechanisms on goethite. J. Colloid Interface Sci. 218, 289–299. Peiffer, S., Walton-Day, K., Macalady, D.L., 1999. The interaction of natural organic matter with iron in a wetland (Tennessee Park, Colorado) receiving acid mine drainage. Aquat. Geochem. 5, 207–223. Persson, P., Lo¨vgren, L., 1996. Potentiometric and spectroscopic studies of sulfate complexation at the goethite–water interface. Geochim. Cosmochim. Acta 60, 2789–2799. Regenspurg, S., Brand, A., Peiffer, S., 2004. Formation and stability of schwertmannite in acidic mining lakes. Geochim. Cosmochim. Acta 68, 1185–1197. Ren, J., Packman, A.I., 2005. Coupled stream—subsurface exchange of colloidal hematite and dissolved zinc, copper and phosphate. Environ. Sci. Technol. 39, 6387–6394. Rietra, R.P.J.J., Hiemstra, T., van Riemsdijk, W.H., 2001. Comparison of selenate and sulfate adsorption on goethite. J. Colloid Interface Sci. 240, 384–390. Russell, J.D., Paterson, E., Fraser, A.R., Farmer, V.C., 1975. Adsorption of carbon dioxide on goethite (a-FeOOH) surfaces, and its implications for anion adsorption. J. Chem. Soc., Faraday Trans. I 71, 1623–1630. Saito, T., Koopal, L.K., van Riemsdijk, W.H., Nagasaki, S., Tanaka, S., 2004. Adsorption of humic acid on goethite: isotherms,
ARTICLE IN PRESS 2060
WAT E R R E S E A R C H
42 (2008) 2051– 2060
charge adjustments, and potential profiles. Langmuir 20, 689–700. Salminen, R., Batista, M.J., Bidovec, M., Demetriades, A., De Vivo, B., De Vos, W., Duris, M., Gilucis, A., Gregorauskiene, V., Halamic, J., Heitzmann, P., Lima, A., Jordan, G., Klaver, G., Klein, P., Lis, J., Locutura, J., Marsina, K., Mazreku, A., ˚ ., Ottesen, R.-T., Petersell, V., Plant, O’Connor, P.J., Olsson, S.A J.A., Reeder, S., Salpeteur, I., Sandstro¨m, H., Siewers, U., Steenfelt, A., Tarvainen, T., 2005. Geochemical Atlas of Europe. Part 1: Background Information, Methodology and Maps. Geological Survey of Finland, 525p. Schemel, L.E., Kimball, B.A., Bencala, K.E., 2000. Colloid formation and metal transport through two mixing zones affected by acid mine drainage near Silverton, Colorado. Appl. Geochem. 15, 1003–1018. Schudel, M., Behrens, S.H., Holthoff, H., Kretzschmar, R., Borkovec, M., 1997. Absolute aggregation rate constants of hematite particles in aqueous suspension: a comparison of two different surface morphologies. J. Colloid Interface Sci. 196, 241–253. Schwertmann, U., Cornell, R.M., 2000. Iron Oxides in the Laboratory, second ed. VCH, Weinheim, 188p. Slowey, A.J., Johnson, S.B., Rytuba, J.J., Brown Jr., G.E., 2005. Role of organic acids in promoting colloidal transport of mercury from mine tailings. Environ. Sci. Technol. 39, 7869–7874. Stjernman Forsberg, L., Ledin, S., 2003. Effects of iron precipitation and organic amendments on porosity and penetrability in sulphide mine tailings. Water Air Soil Pollut. 142, 395–408. Stumm, W., 1992. Chemistry of the Solid–Water Interface. Processes at the Mineral–Water and Particle–Water Interface in Natural Systems. Wiley, New York, 428p.
Suteerapataranon, S., Boyby, M., Geckeis, H., Fangha¨nel, T., Grudpan, K., 2006. Interaction of trace elements in acid mine drainage solution with humic acid. Water Res. 40, 2044–2054. Tessier, A., Fortin, D., Belzile, N., DeVitre, R.R., Leppard, G.G., 1996. Metal sorption to diagenetic iron and manganese oxyhydroxides and associated organic matter: narrowing the gap between field and laboratory measurements. Geochim. Cosmochim. Acta 60, 387–404. Tipping, E., 1981. Adsorption by goethite (a-FeOOH) of humic substances from three different lakes. Chem. Geol. 33, 81–89. Vermeer, A.W.P., van Riemsdijk, W.H., Koopal, L.K., 1998. Adsorption of humic acid to mineral particles. 1. Specific and electrostatic interactions. Langmuir 14, 2810–2819. van Geen, A., Robertson, A.P., Leckie, J.O., 1994. Complexation of carbonate species at the goethite surface: implications for adsorption of metal ions in natural waters. Geochim. Cosmochim. Acta 58, 2073–2086. Weidler, P.G., 1997. BET sample pretreatment of synthetic ferrihydrite and its influence on the determination of surface area and porosity. J. Porous Mater. 4, 165–169. Weng, L., van Riemsdijk, W.H., Koopal, L.K., Hiemstra, T., 2006. Adsorption of humic substances on goethite: comparison between humic acids and fulvic acids. Environ. Sci. Technol. 40, 7494–7500. Za¨nker, H., Moll, H., Richter, W., Brendler, V., Hennig, C., Reich, T., Kluge, A., Hu¨ttig, G., 2002. The colloid chemistry of acid rock drainage solution from an abandoned Zn–Pb–Ag mine. Appl. Geochem. 17, 633–648. Zuyi, T., Taiwei, C., Jinzhou, D., XiongXin, D., Yingjie, G., 2000. Effect of fulvic acids on sorption of U(VI), Zn, Yb, I and Se(IV) onto oxides of aluminium, iron and silicon. Appl. Geochem. 15, 133–139.