Experimental determination and modeling of arsenic complexation with humic and fulvic acids

Journal of Hazardous Materials 279 (2014) 569–578

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Experimental determination and modeling of arsenic complexation with humic and fulvic acids Hoda Fakour, Tsair-Fuh Lin ∗ Department of Environmental Engineering and Global Water Quality Research Center, National Cheng Kung University, Tainan City, Taiwan

h i g h l i g h t s • • • • •

A modeling approach for arsenic complexation with humic and fulvic acids was proposed. Both arsenic species form organic complexes, with higher affinity for arsenate. The two-site binding model successfully simulates the As complexation with both HA and FA. Number of binding sites is proportional to HA concentration for both arsenic species. The apparent stability constants are independent of the HA concentrations.

a r t i c l e

i n f o

Article history: Received 1 April 2014 Received in revised form 17 July 2014 Accepted 18 July 2014 Available online 27 July 2014 Keywords: Arsenic Complexation Fulvic acid Humic acid Iron based adsorbent Ligand binding

a b s t r a c t The complexation of humic acid (HA) and fulvic acid (FA) with arsenic (As) in water was studied. Experimental results indicate that arsenic may form complexes with HA and FA with a higher affinity for arsenate than for arsenite. With the presence of iron oxide based adsorbents, binding of arsenic to HA/FA in water was significantly suppressed, probably due to adsorption of As and HA/FA. A two-site ligand binding model, considering only strong and weak site types of binding affinity, was successfully developed to describe the complexation of arsenic on the two natural organic fractions. The model showed that the numbers of weak sites were more than 10 times those of strong sites on both HA and FA for both arsenic species studied. The numbers of both types of binding sites were found to be proportional to the HA concentrations, while the apparent stability constants, defined for describing binding affinity between arsenic and the sites, are independent of the HA concentrations. To the best of our knowledge, this is the first study to characterize the impact of HA concentrations on the applicability of the ligand binding model, and to extrapolate the model to FA. The obtained results may give insights on the complexation of arsenic in HA/FA laden groundwater and on the selection of more effective adsorption-based treatment methods for natural waters. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Arsenic is a semi-metallic element widely distributed in the earth’s crust and introduced into ground water through the dissolution of minerals and ores. Due to its toxicity and wide occurrence, the element is today recognized as one of the most dangerous inorganic pollutants and threats to drinking water supplies [1]. Therefore, to protect human health, the World Health Organization (WHO) recommended the maximum concentration of 10 ␮g/L arsenic in drinking water [2]. Arsenic mobility in the environment is controlled primarily by sorption onto metal-oxide surfaces. One potentially important factor influencing stability of arsenic could be the presences of other dissolved substances. Natural organic matter (NOM), which comprises a prevalent constituent in natural waters, is highly reactive with both metals and surfaces, and is thus an apparent nominee to influence arsenic mobility and bioavailability [3–6]. In natural water, NOM can be classified into two

∗ Corresponding author. Tel.: +886 6 2364455; fax: +886 6 2752790. E-mail addresses: tfl[email protected], [email protected] (T.-F. Lin). http://dx.doi.org/10.1016/j.jhazmat.2014.07.039 0304-3894/© 2014 Elsevier B.V. All rights reserved.

groups, humic and non-humic fractions. The former fraction is more hydrophobic in character and comprises humic and fulvic acids, while the latter is less hydrophobic and includes hydrophilic acids, proteins, amino acids, and carbohydrates [7]. Humic acid (HA) represents 50–90% of dissolved organic carbon in aquatic and terrestrial soil systems [8]. It may complex with both metals and metalloids, adsorb onto mineral surfaces, participate in redox and photochemical reactions, and form coatings on mineral surfaces [9]. A number of studies have demonstrated strong correlations between organic carbon content and As distribution in water systems, suggesting NOM plays an important role in controlling As transport [4,10,11]. Fulvic and humic acids have also been reported to form stable complexes with mineral surfaces and effectively block arsenic from adsorption on iron oxide [12]. Moreover, it has been shown that NOM may form NOM–As complexes in aquatic environments, facilitating the release of As from natural and contaminated environments [3,4]. The formation of organically As complexes could be a result of either direct association of As with NOM, through ligand exchange with the functional groups such as amine in NOM, or inherent metal content of NOM samples [6]. Although studies regarding the interactions between As and NOM are available (e.g. [3,13–15]), the complexation of As with NOM is not well understood and is still difficult to be quantified. A few parameters have been proposed to quantitatively

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Table 1 Properties of the tested adsorbents. Properties

IOCDa

A33E

Media type

Hematite-coated diatomite 0.17 93 8±1 35.3

Hybrid ion exchanged resin with FeO(OH) 0.3–1.2b 82b 8.88 ± 0.18c 21d

Particle diameter (mm) BET surface area (m2 /g) pHzpc Fe content (%) a b c d

[17]. [57]. [58]. Product information, Purolite, USA.

describe such interaction, including the conditional distribution coefficients (KD ) [14] and the apparent stability constant (ks ) [13] of the resulting As–NOM complexes. To describe a system of interacting As and NOM, ligand-binding models have been combined with the apparent stability constant to quantitatively describe the complexation [14–16]. In these studies, experiments were conducted at different pHs with only one humic acid concentration. The As–NOM complexation data were then modeled with the two-site ligand model. However, the impact of NOM concentrations on the model applicability and the model parameters has not been systematically examined. Moreover, these studies only focused on HA as the effective fraction of NOM and the applicability of the model on another important NOM fraction, FA, has not been studied. Given the lack of information available, this study aimed to quantitatively determine and model the complexation of arsenic with HA and FA under different concentrations. A ligand-binding type model was employed to describe the distribution of arsenic species in water and in complexed form. The extent of complexation is also studied in presence and absence of iron-based adsorbent (IBA) as the solid surface. The results of this study can help to advance the current understanding of As contamination in NOM-laden groundwater.

2. Materials and methods 2.1. Materials Two IBAs, including an iron-oxide-coated diatomite (IOCD), which was recently developed in our laboratory [17], and a commercialized iron oxide-loaded porous anion resin, FerrIX A33E (Purolite, USA), were used as model solid surface in this study. 0.11 mm-sized diatomite coated twice with iron oxide (IOCD-20.11 ) was used due to its high efficiency in arsenic removal based on the earlier results [17]. Detailed preparation method for IOCD can be found in Pan et al. [17]. The physical and chemical properties of the model adsorbents are presented in Table 1. Sodium arsenate (Na2 HAsO4 ·7H2 O, KR Grade, Aldrich, USA) and sodium meta-arsenite NaAsO2 (GR Grade, Sigma, USA) were respectively used to represent arsenate and arsenite in this study. To prepare the stock solution of HA, commercially available HA (Aldrich Chemical, Switzerland) was dissolved in ultrapure water (>18.1 MX cm) followed by solution filtration through 0.45-␮m acetate cellulose membranes (Advantec, Japan). HA stock solution was stored in dark in a glass bottle at 4 ◦ C until used in the experimental procedures, all of which were preformed within 3 weeks after the stock solution was prepared. During the experiments, samples were controlled at the predetermined pH by adding 1 M HNO3 or NaOH. The sample pH was checked and adjusted every 6 h for the initial 36 h and 12 h for the rest of the experiments. 2.2. Complexation experiments To study the interaction of arsenate and arsenite with HA in water, complexation experiments were performed at constant pH (7.5) and various HA concentrations (5–30 mg/L) using 100 ml polyethylene (PE) bottles. Another set of experiments were also conducted for FA (Aldrich Chemical, Switzerland) at 30 mg/L. During the experiments, small volumes of concentrated HNO3 or NaOH solutions were used for pH maintenance. The reaction bottles were

put into a 360◦ rotator (TCLP-601P, Taiwan) for 144 h in darkness with a rotating speed of 27 ± 1 rpm at room temperature. Control samples were also prepared by simply conducting the experiments without adding HA or FA in the systems. All samples were dark stored at 4 ◦ C until analysis. 2.3. Analytical methods Total arsenic in solutions, including dissolved and that bound to HA/FA, was determined using an inductively coupled plasma–optical emission spectrometer (ICP–OES, Ultima 2000, USA). All samples were filtered before analysis and the instrument was calibrated each time before use. Each sample was analyzed three times, and the relative standard deviations (RSD) of the triplicate analyses were all within 5%. For the measurement of free arsenic species, a high performance liquid chromatography with on-line hydride generation and atomic fluorescence spectrometry (HPLC-HG–AFS) (PS Analytical, UK) was used. Organically complex As was determined by calculating the difference between total and free As (Supplementary Fig. 1S). Solution pH was measured using a pH meter (Suntex SP-2200, Taiwan) and all experiments were at least duplicated. Supplementary Fig. 1S related to this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat.2014.07.039. Strict quality control procedures were followed throughout the experiments, sample analyses, and data analyses. The control experiments suggested that the adsorption of As species by the PE bottles used in the experiments were negligible (recoveries for total As determination within 97–103%). Preliminary experiments showed that shaking for 144 h was sufficient to reach equilibrium, for which all experiments were conducted at least in duplicate. The relative percent difference between duplicates was always less than 10%, and averages are reported in this study unless otherwise stated. 2.4. FTIR spectroscopy and elemental analysis A Fourier transform infrared (FTIR) spectrometer (Model Spectrum One, Perkin Elmer) was employed to characterize the functional groups present in the original and As-complexed HA and FA. Combined with the KBr-pellet technique [18], 0.3 mg sample to 30 mg KBr, the FTIR spectrum was obtained in the range of 4000–400 cm−1 with a resolution of 4 cm−1 . Preparation of the studied samples followed those reported in Chang-Chen et al. [19] and Asing et al. [20]. For the HA sample, the solution with 30 mg/L HA was first acidified with 6 M HCl to pH <2 to form HA precipitate. The precipitate was then separated from the solution by centrifugation (Centrifuge Z 206 A, Hermle Labortechnik, Germany) at 6000 rpm for 15 min. After oven-dried at 40 ◦ C for 24 h, the samples were stored in a desiccator before analysis. For the As-complexed samples, the solution of 30 mg/L HA was first reacted with 5 mg/L of As(III) or As(V) for 96 h to allow equilibrium. Then, the solution was acidified, precipitated and centrifuged following the same procedures for the HA sample. For original and As-complexed FA samples, the solutions were first frozen (at −4 ◦ C) for 24 h and freeze-dried for 48 h in Freeze Dryer apparatus (FD-1000 – EYELA, Japan) to obtain amorphous particles. Then the particles were ground in a glass mortar. The ground fine powder was used for analysis. To remove moisture and to avoid unexpected water molecules associated O H peak in the sample [21], all samples were kept in a desiccator before analysis. An elemental analyzer (Vario EL III, Elementar, Germany) was also used to determine the elemental composition of the studied HA and FA. The procedures of the analysis followed that prescribed in Lu et al. [22].

H. Fakour, T.-F. Lin / Journal of Hazardous Materials 279 (2014) 569–578 Table 2 Composition and elemental analysis data of humic and fulvic acid.

2.5. Conditional distribution coefficient and apparent stability constant One of the fundamental parameters in a quantitative description of organic matter interactions with As is conditional distribution coefficient. It is used to characterize the capacity of HA to bind As from the mass perspective and is useful for estimating the amount of As bound by humic substances. The conditional distribution coefficient of As binding to HA (KD , L/kg) was calculated as follows [14]: KD =

[As]b [As]f × [HA]

CbAs CfAs × CNOM

(2)

where CfAs and CbAs are free and bound As concentrations (M), respectively, and CNOM is the molar concentration of effective ligands contained in NOM (HA or FA in this study). Since the molecular structures of HA and FA are not clear, CNOM is usually unavailable [15]. The apparent stability constant, ks , is therefore usually determined using a Scatchard plot [23,24] and ligand binding model. The Scatchard plot, as the most common method used for the presentation of ligand binding data, has been applied to the determination of metal humate stability constants [23–25]. A Scatchard plot plots the concentration ratios of bound ligands to unbound ligands versus the bound ligand concentration. If the binding sites are identical and independent, a straight line is obtained, as listed below. CbAs CfAs

= −ks × CbAs + ks × Bmax

Organic matter fraction

Elemental analysis (%)

Reference

C

H

N

O

Humic acid

59.4 60.7 55.8

3.8 3.7 4.6

1.2 1.5 0.6

35.6 34.1 38.9

Current study [27] [28]

Fulvic acid

50.3 45.7 48.71

3.1 5.4 4.3

1.5 2.1 2.7

45.1 44.8 43.3

Current study [29] [30]

(1)

where [As]f and [As]b are free and HA-bound As concentrations in mg/L; and [HA] is the concentration of HA in solution in kg/L. The stability of As(III)/(V)–HA/FA complex in solution was estimated using an apparent stability constant (ks , M−1 ) [13]. ks =

571

presence of C O stretch in carbohydrates (900–1050 cm−1 ), presence of aromatic C C skeletal vibrations, asymmetric stretching of C O of quinones and ketones, symmetric stretching of COO , C O stretching of amide I band (1600 cm−1 ), and presence of H-bonded O H groups and partially N H stretch (∼3600 cm−1 ). Fig. 1b (for FA) shows presence of Amide V (N H out-of-plane bend) (620 cm−1 ), ethers Ph O C group with C O C bond (1040 cm−1 ), representing C O stretch in COOH, CO stretching and OH deformation of COOH and phenolic groups (1200 cm−1 ), ketones structure as CO C C OH in C O (bond double peak at 1500–1600 cm−1 ), and H-bonded O H groups and partially N H stretch (3400 cm−1 ). The overall FTIR spectra suggested the presence of carboxyl, ketone, phenol, and hydroxyl functional groups in the studied

(3)

where Bmax (M) is the maximum binding capability of HA toward As. If there is interaction between the binding sites, or there are several classes of independent sites, the plot is not linear [26]. After checking the binding sites’ homogeneity present in NOM, the ligand binding model may be used to fit the experimental data and calculate ks,i and Bmax,i (Eq. (4)). CbAs = ˙i

Bmax,i × CfAs (1/ks,i ) + CfAs

(4)

where i is the number of binding site types (i = 1, 2, . . .). 3. Results and discussion 3.1. Characteristics of humic and fulvic acid The elemental compositions of the studied HA and FA are provided in Table 2. The table shows the percentage distribution of C, H, N and O contents of the model humic substances. Compared with those reported in the literature [27–30], the elemental compositions of the studied HA and FA are similar (Table 2). Fig. 1 shows the FTIR spectra for the studied HA and FA before and after complexing with As. As suggested by Stevenson [31] and Silverstein and Webster [32], the peaks, bands, and/or shoulders appeared on FTIR spectra may be linked to the functional groups present on the studied organic materials. Fig. 1a indicates presence of aromatic ring and Me OH groups (470 cm−1 ), presence of C O C bonding and deformation in aromatics (540 cm−1 ),

Fig. 1. FTIR spectra of the studied natural organic matter before and after interaction with two arsenic species, where (a) is for humic acid and (b) is for fulvic acid.

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humic substances. Presence of these functional groups has been linked to the high affinity of humic substances for trace elements in the aquatic environment [33]. As different functional groups are expected to have different binding energies [31], trace elements (such as arsenic) may form complexes with different types of adsorption sites with diverse binding energies on HA and FA [34]. It has been estimated that <10% of total binding sites of humic substances are strong binding sites. These sites may have a wide range of binding energies such as strongly complexing N functional group containing sites. The rest of binding sites, >90%, are assumed to be weak, representing the majority of the complexation sites including carboxylic and phenolic OH groups [34]. Fig. 1 also illustrates the change of FTIR spectra for both HA and FA after interacting with As. Reaction between As and COOH functional groups resulted in the appearance of new bands at about 1230–1240 cm−1 region, which may be attributed to the formation of a metal ion complex [35]. This band is assigned to symmetrical bending of coordinated N H groups in the N-bonded biuret complex [36], suggesting partial complexation through nitrogen groups in this humic–As complex. The fairly small peak at 1700 cm−1 generally assigned to C O stretching of COOH and other carbonyl groups seems to have contributed to the band intensity of As–HA complex formation [37]. The wide peak at ∼3400 cm−1 may be ascribed to vibration of the phenolic groups of the HA–As structure [38]. This band shown on the spectra formed for HA complexed with both As species are all linked to the occurrence of hydration water, confirming the aqua-complexes formation of As–HA [39]. FT-IR spectra of the FA–As complex exhibited several specific peaks in the fingerprint region. The new peak at 866 cm−1 appeared for FA–As(V) system is attributed to C H ring stretch in carbohydrates [36]. Peaks at 1000–1200 cm−1 disappeared for both As species and shift to ∼1400 cm−1 , suggesting the formation of FA–As complex through C C stretch (aromatic ring) and/or O H bending (phenols) [35]. The spectra also illustrated a broad wide peak at ∼3490 cm−1 , which is associated with the phenolic groups of the FA–As structure [38]. Generally, oxygen-containing functional groups were responsible for As ion binding due to their capability of electron pair donation include carboxylic acids, phenols, alcohols, carbonyls, ethers, and esters [40]. Oxygen-containing functional groups may thus play a significant role in the biogeochemical cycling of As species in aquatic and terrestrial environments. 3.2. As binding to HA Analysis of the total and free arsenic revealed that a considerable portion of As was complexed with HA at different initial As concentrations, with higher proportions at low As concentrations (Fig. 2). At 5 mg/L HA, the amounts of As(III) and As(V) complexed with HA were 0.54–3.1% and 4.1–7.0% to that of As(III) and As(V) spiked, respectively. The HA-complexed As increased with increasing HA concentrations, with 1.9–11.0% and 11.0–22.1% of As(III) and As(V) spiked for HA = 15 mg/L, and respectively 4.7–25.0% and 25.2–45.0% for HA = 30 mg/L. Fig. 2 also illustrates that for all HA concentrations examined, the ratios of HA-complexed As decreased with increasing initial As concentrations. The binding of arsenic to NOM has been linked to different functional groups such as amine structures present in HA through ligand exchange [6,41]. In addition, the presence of As–HA complexes could be a result of direct association of As with HA through the metal present in HA, which may also bridge arsenic and NOM [42]. The chemical composition of humic substances depends largely on its origin and treatment, but even purified HA may still have small quantities of trace elements that may act as bridging metals in complexation. Studies have revealed that trace amounts of metals including Fe, Si, Ba, Cr, Mg, and Mn are present in HA [43,44]. In addition, Redman et al. [3] reported a great variability in

complexation behavior of As(III) and As(V) onto diverse NOM samples and indicated that cationic metals might be involved in a ternary complexation mechanism: humic acid–cation–arsenic. As(III) is present in neutral form when pH is <9.0 and forms stable neutral hydroxo-complexes, such as As(OH)3 . Since HAs contain phenolates as functional entities, a ligand exchange reaction may thus occur [14]. In contrast to As(III), the inorganic As(V) species, H2 AsO4 − and HAsO4 2− , are negatively charged in the studied pH here (7.5). Unlike cations binding with humic substances [45], anions binding by humics have so far not been comprehensively investigated. Due to the generally negatively charged properties of HA in water at the studied pH, only weak binding between As(V) and HA was expected [45]. Interestingly, in the current study, stronger As(V) binding compared to As(III) was observed, as also reported by other studies [13,46]. The possible explanations for this could be due to the higher charge at the As(V) center and additional chelation and stabilizing effects. For coordination numbers <6, an associative ligand exchange mechanism at positively charged metal centers may take place [14,47]. Since the arsenate center has a formal charge of +V, adding a phenolate entity at the electrophilic center followed by protonation and water release might occur. Even though the charge is generally negative for both compounds, the driving force is probably stabilization through phenolate donor characteristics, additional chelation by other functional groups, and/or H-bridges [14]. The binding of the negatively charged As(V) and negatively charged humic substances has been also suggested to occur through ternary complexation with cationic metals inherently present in the NOM samples [48] and the higher the metal content the greater the extent of complexation [3]. Regarding environmentally relevant conditions, our finding of As(III) and As(V) binding to HAs is probable because it has been shown that humics have carboxylic, phenolic, and amine functional groups (Fig. 1a and b). Based on the experimental results, it may be concluded that HA in water may complex with As(III) and As(V). Similar to our observation, Chen et al. [49] found that the complexation of As with inorganic colloids and dissolved organic matter (DOM) could facilitate As transport by soil-derived dissolved substances. Other studies have also found that As distribution is greatly correlated by DOM concentration in soil or groundwater [4,11]. 3.3. Conditional distribution coefficients of As(III) and As(V) binding to HA Fig. 3 shows the conditional distribution coefficients (KD ) of arsenite and arsenate with HA/FA. Increasing HA concentration seems not to significantly influence the KD values, as shown in the figure. At low As concentrations (<∼1 mg/L), KD decreased rapidly with increasing As concentrations for both As species and all HA and FA concentrations. Buschman et al. [14] showed that for both As(III) and As(V), KD values increased at lower As/DOC ratios, which is similar to our results. At higher As concentrations, KD values become almost constant for all the studied conditions. The figure also shows that compared with those for HA, KD values for FA at 30 mg/L are slightly lower. The observation is in accordance with Young et al. [50], in which they suggest that higher molecular weight HAs generally bound heavy metals more strongly than lower molecular weight humics. For the two As species tested, Fig. 3 shows that KD values are higher for As(V) which is similar to that reported in Buschman et al. [14]. They suggested that As(V) was bound more strongly on DOC than As(III). 3.4. Apparent stability constants of As–NOM complexes The Scatchard plot ((CbAs /CfAs )∼CbAs at different As concentrations) showed the non-linearity of the datum points for all HA and

H. Fakour, T.-F. Lin / Journal of Hazardous Materials 279 (2014) 569–578

573

Fig. 2. Concentrations of free and bound (a–c) arsenite, and (d–f) arsenate in humic acid solutions.

Fig. 3. Conditional distribution coefficients (KD , y axis) of (a) arsenite and (b) arsenate between water and humic acid/fulvic acid as a function of free As (x axis). Data points represent experimental results. Solid lines are fitted lines with non-linear regression.

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Fig. 4. Scatchard plot showing two types of complexation between (a–d) arsenite and humic acid/fuvic acid and (e–h) arsenate and humic acid/fuvic acid at pH 7.5. Data points represent experimental results. Long dashed lines show Scatchard lines of bound As/free As versus bound As using Bmax (maximum binding capability of HA/FA toward As) and ks (apparent stability constant of As–HA/FA) values determined by non-linear regression.

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575

Table 3 Apparent stability constants (ks , M−1 ), and maximum binding capabilities (Bmax , ␮M) of As(III) and As(V) to NOM fractions. Non-linear regression was performed by fitting experimental data with the two-site ligand binding model. S1 and S2 represent strong and weak binding sites, respectively. As species

HA/FA conc.a

ks,i

R2

Bmax,i

S1

S2

S1

S2

As(III)

5 HA 15 HA 30 HA 30 FA

2.55 2.61 2.54 1.53

± ± ± ±

0.80 0.50 0.52 0.20

0.019 0.020 0.020 0.012

± ± ± ±

0.001 0.002 0.001 0.002

0.058 0.27 0.61 0.38

± ± ± ±

0.008 0.08 0.11 0.09

As(V)

5 HA 15 HA 30 HA 30 FA

3.81 3.96 3.77 2.54

± ± ± ±

0.30 0.07 0.12 0.09

0.028 0.031 0.033 0.020

± ± ± ±

0.005 0.001 0.001 0.003

0.11 0.37 0.92 0.51

± ± ± ±

0.04 0.07 0.11 0.06

± ± ± ±

0.04 0.71 0.50 0.42

0.99 0.98 0.98 0.97

2.15 6.57 13.51 11.67

± ± ± ±

0.21 0.48 0.90 0.81

0.99 0.99 0.99 0.98

HA, humic acid; FA, fulvic acid; and the number represents concentration in mg/L.

FA concentrations (Fig. 4). The results illustrated in Fig. 4 indicate the heterogeneity of the binding sites present in both model organic matters. Since there appeared to be at least two distinct slopes for the plot, the ks could not be determined directly from the slope of the Scatchard plot; consequently, the experimental data were thus fitted into the ligand binding model (Eq. (4)). In fitting the model to the data, a two-site ligand binding model was employed to describe the complexation of As with the two NOM fractions. In the model, two types of binding sites were assumed, one with high stability (strong complexation sites, S1 sites), and the other with lower stability (weak complexation sites, S2 sites). Table 3 shows that the proposed model fits with the experimental data very well, with all correlation coefficients (R2 ) higher than 0.97. Theoretically, in the two-site model, the maximum binding capacity (Bmax,i ) is expected to be proportional to the HA concentrations added into the system, while the stability constant (ks,i ) is expected to be independent of the HA concentrations. Table 3 clearly demonstrates that the extracted ks ’s for S1 and S2 are independent of HA concentrations for both arsenic species, suggesting that the ks ’s are almost constant for a specific site/arsenic speciation combination. As also expected, the ks,1 values of the strong sites (S1 sites) were larger than those of the weak sites (ks,2 for S2 sites). As to the maximum binding capacity, Bmax,1 for S1 sites is lower than for S2 sites (Bmax,2 ), indicating that the strong site numbers were far fewer than the weak ones in HA/FA. To further understand the relationship between the maximum binding capacity and humic acid concentrations, the extracted Bmax ’s were plotted with their corresponding HA concentrations. Fig. 5 clearly

S1 (AsIII) y = 0.0221x - 0.0561 R² = 0.9997

1 0.9

16 14

S1 (AsV) y = 0.0327x - 0.0789 R² = 0.9921

0.8 0.7

Bmax (S1)

demonstrates that the Bmax is linearly proportional to HA concentrations for both the S1 and S2 sites of both arsenic species, with R2 > 0.98. The constant ks ’s and proportionality between Bmax and HA concentrations observed in this study suggest that the two-site binding model is appropriate for used in describing the complexation of arsenic onto HA. Although the two-site model has been applied in describing complexation of As and NOM in different pHs [15], this study further extend the applicability of the model to different HA concentrations and to a system with FA. Comparing the ks and Bmax values for both arsenic species, the results listed in Table 3 demonstrate the higher affinity and binding capability of HA toward arsenate than arsenite, indicating that there is a different binding mechanism for the two species. Warwick et al. [13] also showed that the binding structure for As(V) showed double layer formation of the complex which was not seen with As(III) illustrating that there was a different mechanism of binding taking place. Table 3 also indicates lower affinity and binding capacity of FA compared with those of HA, which may be caused by the differences in sizes and complexity between the two humic substances. The substantially larger molecules and the more complex structure of HA were suggested as the reasons for more binding sites and higher binding capacity as compared to FA [51]. Although both HA and FA are complex mixtures of polyfunctional organic acids containing many different types of sites with different binding energies [25], this study suggest that two-site binding model is able to simulate the complexation of arsenic onto both HA and FA. The observation can also be linked to those found in Fig. 3 for conditional distribution coefficients (KD ) and those

12

y = 0.1691x - 0.3384 S2 (AsIII) R² = 0.9839

0.6 0.5

10 8

y = 0.4551x - 0.1742 S2 (AsV) R² = 0.9998

0.4

6

0.3

4

0.2

2

0.1 0

Bmax (S2)

a

0.32 2.51 4.61 3.72

0

5

10

15

20

25

30

35

0

HA concentration (mg/L) Fig. 5. Relation between maximum adsorption capacity on the binding sites and different concentrations of humic acid.

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Fig. 6. Distribution of bound, free and sorbed arsenite and arsenate in humic acid solutions in the presence of (a–f) IOCD and (g–l) A33E.

shown in the Scatchard plot (Fig. 4) for determining ks ’s. In the As–NOM system, under low As concentration, less As is complexed by NOM and thus the As should be preferably bound to stronger sites, resulting in higher ks and KD values. In contrast, at higher As concentration, more As is complexed by NOM and a large fraction would be bound to weaker sites after the stronger sites are occupied, yielding lower ks and KD values [15]. For that reason, the calculated ks of the As–NOM complex would be expectedly low at high As concentrations, as reported by Warwick et al. [13] and Liu and Cai [15].

broken in the presence of a competitor ligand, such as a sorptive surface. It is noteworthy that the complex abilities between As species and HA/FA could also influence the underlying mechanisms involved in the interaction among the ternary system (As, HA/FA and IBA). Besides the presence of NOM as the competitive ligand for As adsorption, humic substances may also lead to As release into the solute phase by competition between As and NOM anions for mineral sorption sites [56]. Therefore, when studying arsenic speciation, bioavailability, and treatment in the aquatic environment, As binding to dissolved humics should be taken into account.

3.5. As binding to HA in presence of adsorbent 4. Conclusion In order to estimate the As complexation with organic matter in the presence of adsorbent, bound As was measured in adsorption equilibrium experiments (Fig. 6). The results revealed that the extent of As binding to HA is much less than those of the aqueous complexation experiments (Fig. 2). At the highest HA dosages studied (30 mg/L), HA bound As was only 4 and 2% on average, with the maximum at 12 and 6% of initial As(V) for IOCD and A33E, respectively. This could be due to three possible reasons: (i) As adsorption onto solid surfaces causes less free As available; (ii) NOM adsorption onto solid surfaces causes less HA mass available; and (iii) As–HA complex adsorption onto solid surfaces. It has been shown that sorption of NOM by metal hydroxides constitutes an important process through which NOM is retained and stabilized in natural environments [52–54]. Davranche et al. [55] also showed that the organic complex formation involves weak binding which are easily

The present work shows that significant amount of arsenic may complex with HA and FA in water. Under the experimental conditions tested, HA = 5–30 mg/L and FA = 30 mg/L, 0.2–5 mg/L of arsenite and arsenate were observe to bind with FA and HA. With presence of IBAs, HA or FA-complexed arsenic in water was reduced to 2–12%, probably due to stronger adsorption of arsenic and HA/FA on the adsorbent surface. The two-site ligand binding model simulated all the experimental complexation data very well and proved that two classes of binding sites were involved in the complexation of As with the studied organic matter. Model results confirmed that the numbers of weak binding sites are more than 10 times those of strong binding sites. The stability constants, used to describe the extent of interaction between arsenate/arsenite and organic matter were proved to be almost constant for all the HA concentrations

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