Science of the Total Environment 476–477 (2014) 20–28

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Nanoscale zero-valent iron for the removal of Zn2 +, Zn(II)–EDTA and Zn(II)–citrate from aqueous solutions Nina Kržišnik a, Ana Mladenovič a, Andrijana Sever Škapin a, Luka Škrlep a, Janez Ščančar b, Radmila Milačič b,⁎ a b

Slovenian National Building and Civil Engineering Institute, Dimičeva ulica 12, 1000 Ljubljana, Slovenia Department of Environmental Sciences, Jožef Stefan Institute, Jamova 39, 1000 Ljubljana, Slovenia

H I G H L I G H T S • • • • •

Removal of ionic and complexed Zn(II) species from aqueous solutions by nZVI was studied. Conditions for mutual removal of Zn2 +, Zn(II)–EDTA and Zn(II)–citrate were examined. Untreated, silica-fume supported, and choline dispersed nZVI were used at pH 5–7. Zn2 +, Zn(II)–EDTA and Zn(II)–citrate were effectively removed by untreated nZVI. At pH 5 simultaneous removal of ionic and complexed Zn(II) species by nZVI is possible.

a r t i c l e

i n f o

Article history: Received 31 October 2013 Received in revised form 26 December 2013 Accepted 26 December 2013 Available online 22 January 2014 Keywords: Nanoparticles of zero-valent iron Remediation Aqueous solutions Zn2 + Zn(II)–EDTA Zn(II)–citrate

a b s t r a c t The parameters which influence the removal of different zinc (Zn) species: Zn2+, Zn(II)–EDTA and Zn(II)–citrate from aqueous solutions by nanoparticles of zero-valent iron (nZVI) were investigated at environmental relevant pH values. Untreated, surface modified and silica-fume supported nZVI were applied at different iron loads and contact times to Zn solutions, which were buffered to pH 5.3, 6.0 and 7.0. The results revealed that pH, the type of nZVI, the iron load, the contact time, and the Zn species all had a significant influence on the efficiency of removal. Zn2+, Zn(II)–EDTA and Zn(II)–citrate were the most effectively removed from aqueous solutions by untreated nZVI. Zn2+ removal was governed mainly by adsorption onto precipitated iron oxides. Complete removal of Zn2+ and Zn(II)–citrate was obtained at all pH values investigated. The removal of strong Zn(II)–EDTA complex was successful only at acidic pH, which favored degradation of Zn(II)–EDTA. Consequently, the released Zn2+ was completely removed from the solution by adsorption onto iron oxides. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Worldwide, there is an increasing need for the technically and economically effective remediation of polluted surface or ground waters. Water is polluted in huge quantities by everyday municipal, agricultural and industrial activities. Also, uncountable burdens from the past, i.e. sites requiring clean-up, exist, which continually or spasmodically release pollutants into the environment, or are for time dormant, representing a latent hazard under changed environmental conditions (Prokop et al., 2000). Numerous remediation technologies are available for the cleaning up of contaminated waters, but not all approaches are equally effective. Their efficiency depends mostly on the target contaminant and on the hydrogeological conditions (Mueller and Nowack, 2010). Regardless of this fact it can

⁎ Corresponding author. Tel.: +386 1 477 3560, fax: +386 1 2519 385. E-mail address: [email protected] (R. Milačič). 0048-9697/$ – see front matter © 2014 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.scitotenv.2013.12.113

be observed that ex-situ technologies are on the decrease, and that in-situ technologies are on the increase (Karn et al., 2009; Parbs and Birke, 2005). Several types of engineered nanomaterials have been tested for in-situ remediation purposes, utilizing their better reactivity and adsorption capacity when compared to bulk material (Theron et al., 2008; Zhang, 2003; Hua et al., 2012). For the time being, among all nanomaterials, nanoparticles of zero-valent iron (nZVI) are the most commonly used for soil and groundwater remediation (Karn et al., 2009). The reasons for their use lie in their high efficiency for the removal/destruction of pollutants with iron corrosion products (primarily by adsorption and co-precipitation, as well as by highly effective reduction (Noubactep, 2009)), their simple synthesis, and their suitability for both cation and anion sorptions (Deliyanni et al., 2004). Among all types of nanomaterials, nZVI, because of their fast oxidation and consequent growth to micrometer size aggregates, thus losing their nanocharacter, have probably resulted in the lowest degree of controversy and public opposition due to the fear of possible health threats to the

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environment and biota (Karn et al., 2009; Tratnyek and Johnson, 2006; Deliyanni et al., 2004). When nZVI are applied to the problems of in-situ environmental remediation, the critical property that should be taken into account is the colloidal instability and limited lifetime of aqueous nZVI dispersions, and the consequent formation of agglomerates (Phenrat et al., 2007; Kim et al., 2010). In order to enhance their colloidal stability, several modifications of nZVI have been tested in recent years (Phenrat et al., 2007). Among them, surface modification is the most common (Frost et al., 2010; Wang and Zhang, 1997; Li et al., 2011a; Lin et al., 2010). A major part of previous investigations in the field of the nanoremediation of metal contaminated water has been related to two or three valent positively charged ions, i.e. Cr, Pb, As, Cu, Cd, Ni, and Zn (Ponder et al., 2000; Kanel et al., 2005), and hexavalent Cr (Li et al., 2011b; Ponder et al., 2001). Experiments have been performed either at pH values which are characteristic for the target environment (Dickinson and Scott, 2010), or at variable pH values (Scott et al., 2011; Deliyanni et al., 2007; Kanel et al., 2005). However, in environmental and wastewater samples, metals are present not only in cationic species, but are also complexed by available inorganic anions and organic ligands (Collins, 2004; Noradoun and Cheng, 2005; Gylienė et al., 2008). The distribution of metal species depends on the metal concentrations, the concentrations of inorganic and organic ligands, and their corresponding stability constants with metals, as well as their pH and temperature. Naturally occurring organic compounds which form stable complexes with metal ions are high-molecular mass humic and fulvic acids, and low-molecular mass phenolic compounds, sugars and carboxylic acids (Collins, 2004). In industrial applications ethylenediaminetetraacetic acid (EDTA) is widely used in pharmacy, agriculture and the household (Noradoun and Cheng, 2005; Gylienė et al., 2008). Being a strong chelating agent, EDTA is also often applied for the washing of metal contaminated soils (Voglar and Leštan, 2013). Due to its extensive use, EDTA is frequently present as a pollutant in wastewater effluents. It represents a strong complexing agent which competes for metal ions in industrial and environmental water samples with other manufactured or naturally present agents, which form negatively or neutrally charged metal complexes. Among them citrate is a common naturally occurring ligand and present also as a consequence of various remediation activities (Kwan and Chu, 2007; Gylienė et al., 2007). Despite the fact that in contaminated waters metal ions are mostly complexed, information on the nano-remediation of metal ion complexes is rather scarce. As a consequence of industrial activities and mining, zinc (Zn) is often present as a pollutant in industrial wastewaters and environmental waters, from which it needs to be cleaned up. In the present work parameters that influence efficiency and kinetics of removal of Zn2+, Zn(II)–citrate and Zn(II)–EDTA from aqueous solutions by untreated, choline dispersed and silica-fume supported nZVI were investigated with the aim for their simultaneous removal from polluted waters. Zn(II)–citrate represents a naturally occurring and manufactured complex, whereas Zn(II)–EDTA is a species which is frequently present as a pollutant in wastewater effluents. All of the trials were performed at environmentally relevant pH (5.3, 6.0 and 7.0). 2. Materials and methods 2.1. Preparation of nanoparticles of zero-valent iron Untreated nZVI was synthesized following the method first described by Wang and Zhang (1997), where sodium borohydride is used to reduce ferric iron to its metallic state. First, sodium borohydride (NaBH4) (Sigma-Aldrich) solution was prepared by dissolving 0.95 g of NaBH4 in 10 mL of deionized water, and a ferric solution was prepared by dissolving 0.81 g of iron(III) chloride (FeCl3) (Acros Organics) in 40 mL of deionized water. The equivalent of iron in the ferric solution was 0.27 g of dissolved Fe(III). Nanoparticles were then synthesized

21

by the dropwise addition of sodium borohydride solution to the ferric solution, mixing both together vigorously under ambient temperature conditions. Black solid particles of nZVI appeared immediately after the addition of the first drop of sodium borohydride solution as was also observed by Yuvakkumar et al. (2011). After addition of the sodium borohydride solution the mixture was stirred for another 30 min to complete the Fe3+ reduction. Finally, nZVI were isolated by centrifuging, separated with a magnet, and thoroughly washed with deionized water. Surface modified nZVI with octa(cholinium)-polyhedral oligomeric silsesquioxane (choline-POSS) were prepared as untreated particles with the addition of 4 g of a 10% solution of choline-POSS during nZVI synthesis. Choline-POSS was synthesized according to procedure reported by Asuncion et al. (2005), with the exception that, as a silica source, silica-fume was used instead of rice hull ash. Silica-fume supported nZVI were prepared according to the procedure described by Li et al. (2011a). It should be pointed out that the nZVI were prepared immediately before their intended use. To minimize the oxidation of Fe, the nZVI were added to the samples of Zn solutions in the form of 1% aqueous dispersion. 2.2. Preparation of the Zn2+, Zn(II)–citrate and Zn(II)–EDTA solutions For the preparation of the samples and standard solutions Merck (Darmstadt, Germany) chemicals and ultrapure 18.2 MΩ cm water obtained by the Direct-Q 5 Ultrapure water system (Millipore, Watertown, MA, USA) were used. A stock Zn2+ solution (1000 mg Zn L−1), which was used for the preparation of stock solutions of Zn2+, Zn(II)–citrate and Zn(II)–EDTA, was prepared from an appropriate amount of Zn(NO3)2 × 4H2O salt dissolved in water. From the stock solution 10 and 50 mg L− 1 of Zn2 + solutions were prepared by dilution with water. To prevent hydrolysis of Zn2+, 100 μL of nitric acid was added to 1 L of the solution. Zn(II)–citrate and Zn(II)–EDTA (50 mg Zn L−1) were made by mixing the stock Zn2 + solution with an appropriate amount of citric acid monohydrate (C6H8O7 × H2O) or EDTA sodium salt (C10H14N2Na2O8 × 8H2O), so that the Zn to citric acid, or the Zn to EDTA molar ratio, was 1:3. Zn(II)–citrate and Zn(II)–EDTA complexes were formed within 24 h. Speciation analysis data confirmed that the stock Zn(II)–citrate and Zn(II)–EDTA solutions were stable for at least 4 weeks when stored at 4 °C (Milačič et al., 2012). In order to study the efficiency of Zn removal by the applied nanoparticles, working solutions of the different Zn species were prepared within the environmentally relevant pH range, from 5 to 7. The alkaline pH region was not investigated since at pH values higher than 7, Zn2+ tends to precipitate forming sparingly soluble zincite (Deliyanni et al., 2007). Consequently, in the case of alkaline pH values it would not be possible to evaluate the extent of removal of Zn2+ by nanoparticles. In order to maintain nanoremediation at the selected pH, the use of buffers was mandatory. Otherwise, after the nano-treatment, the pH of samples increased to a value of 9. The increase in pH is the consequence of the rapid consumption of oxygen, followed by the production of hydrogen and the release of hydroxide anions into the solution (Zhang, 2003). Thus, Zn solutions were buffered prior to the addition of the nanoparticles. For the adjusting of the pH, 2-(N-morpholino)ethanesulfonic acid (MES) (pH 5.3 and 6) and 4-(2-hydroxyethyl)-1-piperazineethanesulfonic acid (HEPES) buffers, which do not form complexes with Zn, were applied (Milačič et al., 2012). Working solutions of the Zn species at the selected pH values of 5.3, 6 or 7 were prepared by mixing equal volumes of Zn2+ (10 or 50 mg Zn L−1) and Zn(II)–citrate or Zn(II)–EDTA solutions (10 mg Zn L−1) with 0.2 mol L−1 MES or HEPES buffers, so that the final concentration of the Zn species in the 0.1 mol L−1 buffer solutions was 5 or 25 mg Zn L−1, respectively. Before use, the Zn(II)– citrate and Zn(II)–EDTA working solutions were left for 24 h to allow equilibration of the Zn complexes at the investigated pH. The equilibration time for the preparation of the working solutions of Zn(II)–citrate

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and Zn(II)–EDTA, as well as the stability of the Zn complexes, was verified by speciation analysis, as in our previous work (Svete et al., 2001; Milačič et al., 2012). It is necessary to stress that in investigations of the efficiency of removal of Zn from contaminated solutions at various pH values, the use of buffers, instead of adjusting pH with acid or base, is mandatory. In our previous work (Milačič et al., 2012) it was clearly proven by speciation analysis that appropriate buffers for the pH range between 5 and 7, which do not interact with Zn(II) species are HEPES, MES and 3-(Nmorpholino)propanesulfonic acid (MOPS). It was also demonstrated that only by the use of adequate buffers for adjusting the pH, speciation of Zn is not compromised in the samples investigated. The experimental data further revealed that in preparation of Zn complexes it is essential to consider also the slow kinetics of their formation. So, before use, the mixtures of Zn(II)–nitrate with citrate or EDTA should be kept for at least 24 h, to ensure complete Zn(II)–citrate and Zn(II)–EDTA complex formation (Milačič et al., 2012; Svete et al., 2001). These requirements were fulfilled in the present study, but were overlooked by other researchers who modeled the sorption of metals from aqueous solutions on goethite fixed-beds (Lehmann et al., 2001). The authors adjusted pH by adding acid or base, prepared Zn(II)–EDTA complex by the addition of stoichiometric amounts of EDTA to Zn(II)–nitrate solutions, and followed the complex sorption by α-FeOOH at different pH values. If Zn(II)–EDTA complex would be formed, the adsorption of negatively charged Zn(II)–EDTA at pH 3.5 should be, at positively charged surface of α-FeOOH, much higher than that for Zn2 + (Lehmann et al., 2001).

Fe0

Intensity (a.u.)

22

20

30

Fe0

Fe0

FeO

40

50

60

70

80

90

Diffraction angle 2θ (°) Fig. 1. XRD pattern of the untreated nZVI.

2.3. Experimental set-up for investigation of nano-remediation Nano-remediation by untreated nZVI, silica-fume supported nZVI, and nZVI dispersed with choline, was investigated on the laboratory scale. The experiments were carried out at room temperature (22 °C), in duplicates. 50 mL of aqueous solutions containing the different Zn species (5 or 25 mg Zn L− 1), buffered to pH values of 5.3, 6.0 and 7.0, was transferred into 55 mL polyethylene bottles. The solutions were treated with different amounts of a 1% aqueous dispersion of nZVI. The bottles were capped and shaken for 5 h on a horizontal shaker, and then transferred to a rotating shaker. The samples were taken at different time intervals. In each experiment 5 mL of the duplicate samples was blended. The 10 mL composite sample was centrifuged at 4000 rpm for 20 min, and the supernatant was filtered through 0.45 μm Minisart filters (Sartorius Stedim Biotech GmbH, Goettingen, Germany). In order to evaluate the efficiency of the remediation procedure, various concentrations of Zn, solutions, prepared according to the same protocol but without the addition of nZVI, were also monitored during the course of the experiments. In the filtered samples the pH was measured using a WTW (Weilheim, Germany) 310 pH meter, and the concentrations of Zn were determined by flame atomic absorption spectrometry, using a Varian (Mulgrave, Victoria, Australia) SpectrAA 110 atomic absorption spectrometer. 2.4. Methods of characterization of the nanoparticles and their precipitates The mineral composition of the synthesized nZVI and the nZVI precipitates was determined by X-ray powder diffraction (XRD), using a D4 Endeavor, Bruker AXS diffractometer (Karlsruhe, Germany) with Cu Kα radiation (λ = 0.154 nm), and a Sol-X energy-dispersive detector within the 2θ angular range from 20 to 90°, a step size of 0.02°, and a collection time of 5 s. Qualitative analyses of powder diffraction patterns were performed by using the X'Pert HighScore Plus Ver. 2.2c program equipped with a database PDF-2, release 1995 (International Centre for Diffraction Data, Newtown Square, PA). Diffractograms were identified with reference to the ICDD-PDF files for iron (ferrite)

Fig. 2. FESEM image of A: untreated nZVI, B: choline modified nZVI, and C: silica-fume supported nZVI.

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(card No. 00-006-0696), FeO (wustite) (card No. 00-006-0615), franklinite (card No. 00-022-1012) and lepidocrocite (card No. 00008-0098). The specific surface area of the nZVI was determined by a Singlepoint Brunauer–Emmett–Teller (BET) surface area analyzer Micromeritics ASAP 2020 (Micromeritics, Germany GmbH). The dried nZVI samples were evacuated at 150 °C until a final vacuum of 2 Pa was achieved. Nitrogen was used as the adsorptive agent for the analysis. Morphological analysis of nZVI was performed by a field-emission scanning electron microscopy (FESEM, Supra 35 VP, Carl Zeiss, Germany). The samples for FESEM observations were prepared by putting a droplet of the nZVI, in an absolute ethanol dispersion, on a gold-coated pore polycarbonate membrane.

23

3. Results and discussion 3.1. Characterization of the nanoparticles of zero-valent iron The mineral composition of the untreated nZVI was determined by XRD. The corresponding XRD pattern is shown in Fig. 1. The XRD pattern contains reflections which are typical for Fe0. The high background indicates a partly amorphous sample, whereas the width of the reflections suggests nano-dimensioned particles. Similar observations have been reported by Li et al. (2011a). The reflection at diffraction angle 2θ 35.8° indicates the presence of iron oxide (FeO), which is in accordance with the observations of Sun et al. (2006). BET analysis of the nZVI revealed that the specific surface area was 26.4 m2 g−1 in the case of untreated nZVI, 303.1 m2 g−1 in the case of

Fig. 3. Removal of Zn2+ (5 mg L−1) by nZVI (0.4 g Fe L−1) as a function of reaction time at different pH values. A1: control (Zn2+ without the addition of nZVI), B1: untreated nZVI, C1: nZVI dispersed with choline, D1: silica-fume supported nZVI, and variation of the corresponding pH values (A2–D2).

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choline modified nZVI, and 104.5 m2 g−1 in the case of silica-fume supported nZVI, respectively. However, as can be seen from the FESEM image in Fig. 2, the nZVI tend to form larger aggregates, which may decrease their tendency to remain dispersed. The image in Fig. 2A shows that untreated nZVI form chains with an average diameter of about 50 nm, which is in good agreement with results of the BET analysis (the smallest surface area). Despite the fact that the choline supported nZVI have the largest specific surface area, the choline surfactant overlaps the nZVI, so that a clear FESEM image cannot be obtained (Fig. 2B). Nevertheless, it can be seen from Fig. 2 that the nZVI dispersed with choline have a smaller particle size than the untreated nZVI particles. In Fig. 2C only silica-fume can be seen. The reason for this is that, for successful synthesis, silica-fume has to be added in excess in comparison with Fe. 3.2. Optimization of the parameters which influence the efficiency of the removal of Zn2+, Zn(II)–citrate and Zn(II)–EDTA from aqueous solutions by zero-valent iron nanoparticles It is well documented that pH has a significant influence on the rate and efficiency of remediation in the case of iron-based agents (Deliyanni et al., 2004; Scott et al., 2011). In addition to pH, the efficiency of nanoremediation further depends on the type of the nanoparticles used, and their concentrations (Ponder et al., 2001; Dickinson and Scott, 2010; Scott et al., 2011; Crane et al., 2011), as well as on the concentration and oxidation state of the element being removed (Ponder et al., 2000). Since, in environmental and industrial waters, Zn ions are present not only as ionic species, but are also complexed with available ligands, it is of crucial importance to also consider the chemical speciation of Zn when the nano-remediation is applied. In the present work, investigation was performed on the parameters which influence the efficiency of removal of Zn2+, Zn(II)–citrate and Zn(II)–EDTA by nZVI with the aim to optimize the nano-remediation process. 3.2.1. The influence of pH on the efficiency of removal of Zn2+ from aqueous solutions by different zero-valent iron nanoparticles In the first set of experiments the influence of pH on the efficiency of the removal of Zn2+ by different nZVI was investigated, within the environmentally relevant pH range. For this purpose Zn2 + samples (5 mg L−1) were prepared in buffer solutions at pH values of 5.3, 6.0 and 7.0. To each of Zn2+ solutions an aqueous dispersion of nZVI, containing 0.4 g Fe L−1, was added, and the remediation procedure was performed as described in the Experimental set-up for investigation of nano-remediation section, applying different contact times. The results of these experiments are presented in Fig. 3. As can be seen from Fig. 3, the concentration of Zn2+ in the control sample without nZVI, and the pH investigated remained constant during the course of the experiment. Maintaining the pH at a constant value enabled better insight into the processes which occur during contaminant removal (Scott et al., 2011). From Fig. 3 it is also evident that nano-remediation was the most effective at a pH of 7.0, regardless of the type of nZVI applied. When the pH was lowered the efficiency of the remediation decreased. Remediation by nZVI was achieved within 24 h, and remained almost constant in the case of longer contact times (slight variations occurred most probably due to the partial remobilization of Zn2+). After 70 days, about 100% of the Zn2 + was removed at a pH of 7.0 by all the different types of nZVI, whereas at a pH of 6.0 the efficiency of removal was about 90%, 85% and 60% for untreated nZVI, for silica-fume supported nZVI, and for choline dispersed nZVI, respectively. However, at a pH of 5.3 and iron load 0.4 g Fe L−1, only about 50% of the Zn2+ was removed by untreated nZVI, and 15% by silica-fume supported nZVI, whereas, in the case of choline dispersed nZVI, at this pH no remediation occurred. This influence of different pH values on the removal of Zn2+ was expected, since adsorption and metal hydroxide precipitation are favored in the case of higher pH by the iron corrosion process. In the

neutral pH range, the relatively low ξ potential of nZVI in solution supports the particle aggregation, while by lowering pH, ξ potential is gradually increased (Zhang and Elliott, 2006). Due to more positive charge of nZVI at pH 5.3, the sorption of positively charged Zn2 + is inhibited. Deliyanni et al. (2007) observed similar trends in the case of the removal of Zn2+ by nanocrystalline akaganeite. They found out that, at a pH of 5.3, nano-remediation was not effective, whereas at a pH of 7.5 the removal rate of Zn2+ was about 95%. From the data of the above-described experiments it can be concluded that the most effective remediation of Zn2+ is achieved by means of untreated nZVI. In order to better understand the mechanisms of the removal of Zn2+ by untreated nZVI from an aqueous solution, the samples collected at the end of the experiment (after 70 days) were centrifuged, the supernatants discarded, and the precipitates dried and characterized by XRD analysis. The XRD patterns of the untreated nZVI precipitates at different pH values are presented in Fig. 4. The XRD patterns from Fig. 4 demonstrate that, at pH 5.3 and pH 6.0, the final product of remediation was lepidocrocite (γ-FeOOH), a common phase often found as a result of the iron corrosion process by dissolved oxygen (Kamolpornwijit et al., 2004). By aging more lepidocrocite is formed, so at longer reaction times, adsorption of Zn2 + was possible also at pH 5.3. Nevertheless, at iron load 0.4 g Fe L− 1, only about 50% of the Zn2 + was removed by untreated nZVI. The X-ray diffractogram of the untreated nZVI precipitate at a pH of 7.0 reveals sharp reflections of two phases: lepidocrocite and franklinite. The presence of franklinite indicates the incorporation of Zn2+ in a spinel franklinite lattice, which additionally contributes to the most effective and rapid remediation at pH 7.0 (see Fig. 3B1). 3.2.2. The influence of different iron loads on the efficiency of removal of Zn2+ in aqueous solutions by zero-valent iron nanoparticles In the continuation of work the influence of different iron loads of nZVI on the efficiency of removal of Zn2+ (5 mg L−1) was investigated at pH 7.0, 6.0 and 5.3 over a time span of 24 h. The experiments were carried out with untreated and silica-fume supported, the two most perspective nZVI, containing 0.1, 0.25, 0.4, 0.7, 1 and 3 g Fe L− 1. The results are presented in Fig. 5. From the data presented in Fig. 5 it can be seen that the iron load plays a crucial role in remediation, particularly at lower pH values. These observations are based on the fact that Zn2+ removal is governed by adsorption onto precipitated iron oxides. With an increasing iron load the reactive sites of nZVI for adsorption are increased, too (Singh et al., 2012). As can be seen from Fig. 5, for the complete removal of 5 mg Zn2 + L− 1 at least 3 g Fe L−1 of untreated nZVI is needed at pH 5.3. An appreciably lower amount of untreated nZVI is needed to

L: Lepidocrocite FeO(OH) F: Franklinite ZnFe2O4

F L L L

Intensity (a.u.)

24

F

F

L

L,F L,F

F

L

L,F

pH 7

F L

L

L L L

L

L

L L

L

pH 6

L L

L L

L

20

L LL

L

L

10

L

L

30

40

L

50

L

60

L

L L L

pH 5.3 70

Diffraction angle 2θ (°) Fig. 4. XRD patterns of the untreated nZVI precipitate 10 weeks after exposure to Zn2+ solutions at different pH values.

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25

Fig. 5. Removal of Zn2+ (5 mg L−1) by nZVI as a function of the iron load and the reaction time at different pH values. A1–A3: untreated nZVI at pH 7.0, 6.0 and 5.3, respectively, B1–B3: silica-fume supported nZVI at pH 7.0, 6.0 and 5.3, respectively.

remove Zn2 + from the solution at pH 6.0 (0.4 g Fe L−1), whereas at pH 7.0 the sufficient amount of untreated nZVI is only 0.25 g Fe L−1. Similarly, the iron load importantly influences the efficiency of removal of Zn2 + by silica-fume supported nZVI. Small quantities are needed (0.25 g Fe L− 1) to completely remove Zn2 + at pH 7.0. However, at pH 5.3 and 6.0 Zn2+ removal is far less effective than if untreated nZVI is used, even if 3 g Fe L−1 are applied for remediation. From the data presented in Fig. 5 it can also be seen that by using untreated nZVI for remediation and applying a contact time of 5 h, Zn2+ (5 mg L−1) can be effectively removed from the solution at all the investigated pH values. In order to find out whether the increasing amount of Zn2 + also requires a higher iron load, the Zn2+ concentration was increased to 25 mg L−1 and different iron loads were applied using untreated nZVI for remediation at pH 7.0, 6.0 and 5.3 over a time span of 24 h. The initial iron concentrations were the same than those which were found to be sufficient for the complete removal of 5 mg Zn2 + L−1. They were 0.25 g Fe L−1, 0.4 g Fe L−1 and 3 g Fe L−1 for pH 7.0, 6.0 and 5.3, respectively (Fig. 5A1–A3). To find the optimal iron concentration needed for the successful removal of 25 mg L−1 Zn2+, these iron loads were

increased by factors of two and five. The data corresponding to these experiments are presented in Fig. 6. By comparing the data from Fig. 5A1–A3 with the data from Fig. 6 it can be seen that, in general, the same iron loads are sufficient to remediate 5 or 25 mg Zn2+ L−1, but longer times (e.g. 24 h) are necessary for the efficient removal of higher Zn2+ concentrations. 3.2.3. Removal of Zn(II)–EDTA and Zn(II)–citrate from aqueous solutions by zero-valent iron nanoparticles Due to heavy industrial activities and mining, Zn is often present as a pollutant in the aquatic and terrestrial environments. Although in contaminated waters Zn is mostly complexed, the importance of chemical speciation in the remediation of Zn complexes with nZVI has often been overlooked. In order to optimize the remediation of Zn complexes by nZVI, both Zn(II)–EDTA, which is frequently found as a pollutant because of anthropogenic activities (Noradoun and Cheng, 2005; Gylienė et al., 2008; Chen et al., 2008), and Zn(II)–citrate being a naturally occurring Zn complex and present also due to different citrate-related remediation

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1.25 g/L Fe 0.5 g/L Fe 0.25 g/L Fe

2 g/L Fe 0.8 g/L Fe 0.4 g/L Fe

A

B

C 15 g/L Fe 6 g/L Fe 3 g/L Fe

(104.5 m2 g−1). This phenomenon was related to overlay of nZVI with choline surfactant in choline modified, and with silica in silica-fume supported nZVI, which hinder direct interaction of nZVI with Zn species, and therefore decreased the remediation efficiency. From Fig. 7 it is further evident that 7 days after the addition of untreated nZVI, the degree of Zn(II)–citrate removal was higher than 90% for all the pH values investigated (Fig. 7B1), whereas in the case of Zn(II)–EDTA efficient removal (higher than 95%) was achieved only for pH 5.3, and decreased with increasing pH (Fig. 7A1). The latter phenomenon is associated with the stability of Zn complexes, which are more stable in neutral, then in acidic pH range and degradation of organic ligands by nZVI, which is more efficient at acidic pH values. In last decade researchers intensively investigated mechanisms of the degradation of EDTA and citrate by nZVI (Noradoun and Cheng, 2005; Gylienė et al., 2008; Chen et al., 2008), which go as follows: First, the reaction of nZVI with dissolved oxygen leads to the formation of Fe2+ and Fe3+ ions. Fe2+ and Fe3 + form strong complexes with EDTA (the corresponding log K at 25 °C and ionic strength 0.1 are 14.27 and 25.0, respectively (Martell and Smith, 1974)), while Fe3+ forms strong complex with citrate (the corresponding log K at 25 °C and ionic strength 0.1 is 11.50 (Martell and Smith, 1977)). These newly formed Fe-complexes generate the superoxide radical O•− 2 , which further reacts with Fe-complexes producing H2O2 that is a source for the Fenton reaction to yield a reactive hydroxyl radical OH•: 0

2þ

Fe →Fe 2þ

Fe

2þ

ð2Þ

þ L→Fe L

ð3Þ

3þ

ð4Þ

2þ

L þ O2

2þ

L þ H2 O2 →Fe L þ OH þ OH

Fe

3þ

•−

L þ O2 →Fe

Fe

(Kwan and Chu, 2007; Gylienė et al., 2007), were investigated. Zn forms negatively charged complexes with EDTA and citric acid within the pH range from 5.0 to 7.0. The [Zn(II)–EDTA]2− complex is very stable, with the log K of 16.5 (25 °C, ionic strength 0.16) (Martell and Smith, 1974), whereas Zn-citrate exists as a moderately stable complexes. In the pH range from 5 to 7 [Zn(II)(Cit)]− and [Zn(II)(Cit)2]4− complexes coexist. At pH 5 [Zn(II)(Cit)]− is the dominant species, while at pH 7 [Zn(II)(Cit)2]4− prevails (Powell and Pettit, 1997; Milačič et al., 2012). The corresponding log K (at 25 °C and ionic strength 0.16) are 4.98 and 5.90 for [Zn(II)(Cit)]− and [Zn(II)(Cit)2]4− complexes, respectively (Martell and Smith, 1977). The experiments were carried out using Zn(II)–EDTA and Zn(II)– citrate solutions (5 mg Zn L−1). Untreated, choline dispersed and silica-fume supported nZVI were added at iron loads of 0.25 g Fe L− 1, 0.4 g Fe L−1 and 3 g Fe L−1 for pH values of 7.0, 6.0 and 5.3 (the iron loads were selected based on our previous studies for Zn2+). The results of these experiments, which were performed over a time span of 35 days, are presented in Fig. 7. The results shown in Fig. 7 demonstrate that the rate of Zn removal reached a constant value 24 h after the addition of the silica-fume supported nZVI, and 7 days after the addition of untreated or choline dispersed nZVI. The efficiency of removal was higher for the Zn(II)– citrate than for the Zn(II)–EDTA. Likewise, in the case of Zn2 + the most effective removal of Zn complexes was achieved with untreated nZVI, despite its smallest surface area (26.4 m2 g−1) in comparison to choline modified (303.1 m2 g− 1) and silica-fume supported nZVI

ð1Þ

2þ

Fe

Fig. 6. Removal of Zn2+ (25 mg L−1) by untreated nZVI as a function of the iron load and the reaction time at different pH values. A: pH 7.0, B: pH 6.0 and C: pH 5.3, respectively.

−

þ 2e

•−

L þ O2 þ

þ 2H →Fe L þ H2 O2 3þ

•

−

ð5Þ

In the above Eqs. (1)–(5) L represents an organic ligand. The formation of hydroxyl radical is favored at acidic pH. It was reported that hydroxyl radical, which is a strong oxidant, is being able to degrade EDTA and citrate (Noradoun and Cheng, 2005; Gylienė et al., 2008; Chen et al., 2008; Keenan and Sedlak, 2008; Zhou et al., 2010). Similarly, the mechanism for Zn(II)–citrate and Zn(II)–EDTA degradation is most likely governed through a series of ligand mediated reactions, which resulted in release of Zn2+. Since [Zn(II)(Cit)]− and [Zn(II)(Cit)2]4− are both complexes of moderate stability, different distributions of these two Zn complexes within pH range 5 to 7, did not importantly influence their degradability by nZVI. Zn(II)–citrate complexes decomposed within 7 days of exposure to untreated nZVI at all pH values investigated, and Zn2+ was then effectively removed by adsorption on the iron corrosion products. In contrast, at close to neutral pH range, it was not possible to effectively remove Zn(II)–EDTA, due to the high stability of the negatively charged complex. In slightly acidic solutions (pH 5.3), the Zn(II)–EDTA complex was most probably degraded by the untreated nZVI. In addition, the iron dissolution reaction by the presence of oxygen, yielded formation of amorphous precipitate FeOOH, which acted as good sorbent (Gylienė et al., 2008). 2þ

2Fe

þ

þ ð1=2ÞO2 þ 3H2 O→2FeOOH þ 4H

ð6Þ

In our experiments lepidocrocite (γ-FeOOH) was confirmed as a prevailing corrosion product of reaction of nZVI with oxygen at pH 6 and 5.3 (Fig. 4). After the degradation of Zn(II)–EDTA complex, the released Zn2 + species were adsorbed on the surface of lepidocrocite, and were completely removed from the solution, but higher iron loads and longer reaction times were necessary. Similar observations were reported also by other researchers who found that the degradation of

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27

Fig. 7. Removal of A: Zn(II)–EDTA and B: Zn(II)–citrate (5 mg Zn L−1) by nZVI as a function of reaction time at different pH values. A1, B1: untreated nZVI, A2, B2: nZVI dispersed with choline, A3, B3: silica-fume supported nZVI. Iron loads: 0.25 g Fe L−1, 0.4 g Fe L−1 and 3 g Fe L−1 for pH 7.0, 6.0 and 5.3, respectively.

organic complexes is enhanced under the acidic pH, presence of oxygen, longer contact times and higher iron loads (Gylienė et al., 2008; Chen et al., 2008). The results of the present study showed that before an optimal in situ remediation process is designed, mutual optimization of the parameters involved, such as the type of nZVI, the iron load, and the reaction time, must be performed. These parameters should be optimized for the selected pH value of remediation as well as for different concentrations of the treated contaminant. The chemical speciation of Zn must also be considered. The data of the present investigation revealed that by careful optimization of parameters it is possible to find conditions that enable remediation of complex Zn solutions, containing positively charged ionic Zn species as well as negatively charged Zn complexes, which are frequently present in polluted industrial and environmental water samples. 4. Conclusions The potential of nZVI of different reactivities and stabilities was investigated for the removal of Zn2 +, Zn(II)–EDTA and Zn(II)–citrate from aqueous solutions. The efficiency of Zn removal was studied within the environmentally relevant pH range, between 5 and 7, by applying untreated nZVI, silica-fume supported nZVI, and nZVI dispersed with choline. At a pH of 7, rapid and efficient removal of 5 mg Zn2 + L− 1

was achieved by all of the studied types of nZVI (iron load 0.4 g Fe L−1, contact time 1 h), whereas, at lower pH values, remediation by untreated nZVI was more effective than by silica-fume supported and choline dispersed nZVI. Nevertheless, to efficiently remove Zn2+ by untreated nZVI at lower pH values, higher iron loads and longer contact times (5 h) were necessary. For the effective removal of higher Zn2+ concentrations (25 mg L−1), the same iron loads were sufficient, but longer times (24 h) were necessary. Zn2+ removal was achieved mostly by adsorption onto precipitated iron oxides. Zn(II)–EDTA and Zn(II)–citrate were also the most efficiently removed from aqueous solutions by untreated nZVI. First, complexes were degraded by nZVI and then after the released Zn2+ was adsorbed onto iron oxide precipitate. The degradation of moderately stable Zn(II)–citrate and subsequent removal of Zn2+ was obtained at all pH investigated, but the degradation of strong Zn(II)–EDTA complex was successful only at pH 5.3. As a result, the released Zn2+ species were completely removed from the solution by adsorption onto FeOOH. Data from our study demonstrated the significance of Zn speciation that should be considered when remediation by nZVI is planned. In order to obtain the efficient removal of Zn species it is important to choose the most appropriate type of nZVI, the iron load and contact time. The results obtained offer useful information for the establishment of protocols for the remediation of Zn by nZVI in polluted natural and industrial waters with neutral and acidic pH, which so far have not been frequently studied.

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