In Situ Preparation of Stabilized Iron Sulfide Nanoparticle-Impregnated

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Remediation and Control Technologies

In situ preparation of stabilized iron sulfide nanoparticlesimpregnated alginate composite for selenite remediation Jun Wu, and Raymond J Zeng Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b05861 • Publication Date (Web): 03 May 2018 Downloaded from http://pubs.acs.org on May 3, 2018

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Environmental Science & Technology

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In situ preparation of stabilized iron sulfide nanoparticles-impregnated alginate

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composite for selenite remediation

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Jun Wu1,2, Raymond Jianxiong Zeng1,2,*

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College of Resources and Environment, Fujian Agriculture and Forestry University,

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Fuzhou 350002, China

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2

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University of Science & Technology of China, Hefei 230026, PR China

Fujian Provincial Key Laboratory of Soil Environmental Health and Regulation,

CAS Key Laboratory for Urban Pollutant Conversion, Department of Chemistry,

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* Correspondence concerning this article should be addressed to:

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Raymond J. Zeng at [email protected]. Tel/Fax: +86 591 83303682

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ACS Paragon Plus Environment


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Abstract

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Iron sulfide (FeS) nanoparticles have been applied for selenite (Se(IV))

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remediation in recent decades. However, the easy aggregation and oxidization of FeS

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hamper their reactivity. In this study, in situ immobilization technology was applied to

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prepare FeS nanoparticles-impregnated alginate composite (FeS-SA) for Se(IV)

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remediation. FeS-SA removed 100% of the Se(IV) (0.13 mM), whereas pure

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non-stabilized FeS and sodium alginate (SA) beads only eliminated 27% and 20% of

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the Se(IV), respectively. The removal efficiency increased to 73% when pure

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stabilized FeS was used. Therefore, FeS-SA showed superior removal efficiency that

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was comparable with the joint effect of pure stabilized FeS and SA beads due to the

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homogeneous distribution of FeS in SA matrix. Furthermore, minor differences were

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established in the oxidation retardation effect of FeS exerted by SA beads under

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anoxic and oxic conditions. The biogenic regenerated FeS-SA still showed 40%

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removal efficiency for Se(IV) after 5 cycles due to the Fe leaching. XPS technique

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combined with the reference compounds and electron balance revealed that FeSe and

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metal selenium were the main selenium species after treatment. This in situ

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preparation of stabilized FeS-SA exhibited an excellent application prospect in the

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remediation of Se(IV).

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Introduction

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Iron sulfide (FeS), known also as mackinawite, is a tetragonal ferrous

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monosulfide that has been applied in the treatment of groundwater and soil

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contaminated with pollutants, such as mercury,1 arsenic,2 uranium,3 chromium,4 and

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selenite (Se(IV)).5 This application is due to its potential as a source of Fe(II) and

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S(−II) species, both of which can act as reducing agents and thus facilitate the

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removal of pollutants from water.6 However, the natural mineral FeS is

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thermodynamically stable under reducing and anoxic conditions. Fe(II) is unstable

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under oxidative conditions and is transformed into Fe(OH)3(s), goethite (α-FeOOH),

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and lepidocrocite, while S(−II) is oxidized to form polysulfide and elemental sulfur.7-9

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Thus, FeS can be easily oxidized after contact with aqueous fluids containing

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dissolved oxygen, which in practice limits its application for remediation of polluted

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water.7

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FeS nanoparticles, prepared by traditional methods, commonly through

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co-precipitation under anoxic environments in the aqueous phase, are not an

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exception.10 As a consequence, the diminishment of FeS specific surface area and

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reactive sites lead to low efficiency of pollutant removal. To achieve control of

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particle morphology and size distribution, some recent works have developed a new

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class of materials to stabilize FeS nanoparticles, such as carboxymethyl cellulose

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(CMC) and various biomaterials.11, 12 Moreover, due to the negative charges carried

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by sodium alginate (SA), it can also be used to stabilize the FeS nanoparticles and

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improve their reactivity.13

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SA, a biodegradable and biocompatible polysaccharide, is commonly used to

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immobilize a variety of materials.14, 15 The immobilization technology based on SA in

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the remediation of environmental pollutants has gained significant attention.

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Photocatalytic materials (e.g., TiO2)16 and iron compounds (e.g., magnetic Fe3O4 or

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γ-Fe2O3 and Fe(OH)3)17,

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pollutants.

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are usually immobilized into gels to remove various

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Although the stabilizing effects of SA used to control FeS nanoparticles

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aggregation have been previously demonstrated,13 it is still unclear whether this

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impact is still exerted after FeS immobilization into SA gel beads. Moreover, whether

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SA gel beads can create reducing conditions by limitation of oxygen transfer,

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facilitating the prevention of FeS oxidation by dissolved oxygen, has not yet been

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elucidated. The establishment of such conditions is of crucial importance to the

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pollutant removal capacity of FeS. Therefore, to extend our knowledge on the

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above-mentioned aspects, in this study, we employed SA to immobilize FeS. We

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hypothesized that simultaneously obtaining a stabilizing effect and antioxygenic

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properties can be achieved by the proposed technology. To confirm our hypothesis, a

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representative pollutant, selenium (Se), was used in our examination. Se is an

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essential trace chemical element that exerts crucial functions in the maintenance of

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biochemical processes in most animals and humans, but it can be toxic in cases of

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excessive uptake.19-21 Five different oxidation states of selenium exist in the

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2ଶି environment: selenate (SeOଶି ସ ), selenite (SeOଷ ), elemental Se (Se(0)), selenide (Se ),

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and organic Se.22, 23 Among them, Se(IV) is considered the most toxic form of Se.24, 25

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Traditionally, for its removal, adsorbent materials (e.g., iron compounds or other

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inorganic materials) are prepared, followed by immobilization into an organic

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matrix.26, 27 However, this approach is not appropriate for FeS due to its oxidability

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which hinders the immobilization process. Thus, it is vital to develop a new method

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for immobilization of FeS.

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The aim of this study was to use an immobilization technology to prepare in situ

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a FeS nanoparticles-impregnated alginate composite for Se(IV) remediation. We

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investigated whether SA immobilization contributed to the enhancement of the

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dispersity and reduction in the aggregation of FeS. The role of alginate gel in the

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retardation of the oxidation of FeS by O2 was also examined. Biostimulation was

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employed to regenerate the exhausted FeS under the sulfate reducing conditions.

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Analysis of the solids after Se removal was conducted using X-ray photoelectron

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spectroscopy (XPS) to determine the mechanism of removal. We also evaluated the

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effects of pH and different anions on the efficiency of Se(IV) removal. This study can

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provide an outstanding application of FeS-SA beads for environmental remediation.

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Materials and methods

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In situ preparation of FeS impregnated alginate beads

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Detailed information about chemicals can be found in Supporting Information (SI).

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Stock solutions of Na2S 9H2O and FeSO4 7H2O (113.6 mM) were first prepared with

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purged DI water (purged with > 99% N2 for 20 min). A 5-mL aliquot of the Na2S

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stock solution was diluted with 45 mL of purged DI water. An amount of 1.0 g of SA

.

.

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(2%) was added to the diluted Na2S solution under N2 purging to form a Na2S–SA

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mixture. The resulting solution was magnetically stirred until complete dissolution in

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an anaerobic glove box containing 5% hydrogen and 95% nitrogen. Likewise, a 5-mL

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aliquot of the FeSO4 stock solution was diluted with 45 mL of purged DI water, and

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1.5 g of CaCl2 (3%) was added to the solution. Subsequently, a syringe with a

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1-mm-diameter needle was used to absorb the Na2S–SA mixture, and 4 mL of the

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mixture was added into the FeSO4-CaCl2 solution. When the drops contacted the

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solution, black pellets were immediately formed. After crosslinking and reacting for

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12 h, the pellets were taken out and rinsed three times with purged DI water to

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remove the unreacted FeSO4 and Na2S. For comparison, calcium alginate gel beads

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without FeS were also prepared following the same procedure with no FeSO4 and

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Na2S used. SA stabilized FeS nanoparticles (Abbreviated as FeS*) were prepared

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through the approach developed in our previous research.13 Non-stabilized FeS

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followed the same procedure with no SA used.

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In the preliminary experiment, corrosion products of FeS contained α-FeOOH

112

(data not shown). α-FeOOH is known as a strong sorbent for Se(IV).28, 29 To evaluate

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the adsorption of α-FeOOH for Se(IV) in this study, α-FeOOH was immobilized into

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the SA to form α-FeOOH-SA then test its adsorption performance. The preparation of

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α-FeOOH could be found in the previous study,30 and TEM and XRD demonstrated

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that α-FeOOH was prepared successfully (Figure S1). The molar quantity of

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immobilized α-FeOOH was equal to FeS in FeS-SA. In addition, according to earlier

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literatures, FeS was unstable under oxidative conditions and was transformed into

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α-FeOOH.7, 8 Thus, FeS-SA was oxidized in the air firstly (abbreviated as FeS-SA*)

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then was used to adsorption Se(IV).

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Chemical analysis and characterization

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Se(IV) concentration was determined by the ascorbic acid spectrophotometric

123

method developed in a prior study.31 The detailed description is as follows. A volume

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of 600 µL of the filtrate obtained by filtration through a 0.22-µm membrane filter was

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mixed with 300 µL of HCl (4 M) and then with 600 µL of ascorbic acid (1 M) by

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vortexing. Further, after 10 min reaction, absorbance was measured at 500 nm.

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Ferrous ions were detected by spectrophotometry at 510 nm using o-phenanthroline.32

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The detailed instrumental characterization of scanning electron microscopy (SEM),

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transmission

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thermogravimetric (TG) analysis, XPS, and phase-contrast microscope can be found

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in SI.

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Batch experiments

electron

microscopy

(TEM),

X-ray

diffraction

(XRD),

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The amount of FeS immobilized in alginate beads was measured and detailed

134

method can be seen in SI. Six different experimental analyses were performed,

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namely, on FeS-SA beads, pure stabilized FeS nanoparticles suspension, pure

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non-stabilized FeS suspension, SA beads, α-FeOOH-SA beads, and FeS-SA. The

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reaction was initiated by adding a Se(IV) stock solution to the above four

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experimental sets to attain a Se(IV) concentration of 0.13 mM. While the Se(IV)

139

concentration used in XPS analysis still kept 0.63 mM, otherwise it will lead to the

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undetectable of Se(IV) by XPS. Then, glass vials were placed in a shaker (at 170 rpm)

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at room temperature (25 °C). The solution was sacrificially sampled at the

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predetermined time to measure the concentration of Se(IV). Unless otherwise stated,

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all of the above batch experiments were carried out in triplicate in 50-mL glass vials

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sealed with an aluminum cap and under anoxic conditions. The initial pH was

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adjusted to 6.0 ± 0.2 with 1 M HCl or NaOH solution. Unless otherwise stated, 20

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mM 4-(2-hydroxyethyl)-1-piperazineëthanesulfonic acid (HEPES) was added in the

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all batch experiments to control the solution pH. Individual buffer used did not cause

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different influence on Se(IV) by FeS-SA (data not shown) and previous studies have

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reported similar results.33,

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isotherm models, effects of pH and anions, column experiment can be found in SI.

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The experimental details about kinetic experiments,

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The reaction rate and rate-limiting diffusion in porous materials are significant

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messages for scaling this FeS-SA for water treatment. Understanding diffusion

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controlled mass transport mechanism may allow use of scaling techniques to reduce

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the time and associated costs as well as optimizing process design (e.g., empty bed

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contact time, column configurations). Therefore, both the pseudo-first-order and

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pseudo-second-order models were used to determine the reaction rate constants of

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Se(IV) adsorption. Weber-Morris model was used to determine the rate-limiting

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mechanism in porous materials.35-37 All these models were used to identify the

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removal isotherm, rate and rate controlling mechanism. The detailed description and

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equation can be seen in the SI.

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The regeneration process of exhausted FeS-SA

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The exhausted FeS-SA beads were recovered after the batch experiment for

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Se(IV) removal and rinsed three times with purged DI water. Next, the exhausted

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FeS-SA and SRB were added to the SRB medium (Detailed preparation method could

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be found in SI) simultaneously, and cultivated anaerobically in the serum bottles in a

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shaker (37 ℃, 170 rpm). After 24 h, the regenerated FeS-SA beads were taken out and

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rinsed three times with purged DI water. Se(IV) removal test using regeneration

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FeS-SA beads were conducted as the aforementioned method. The Fe concentration in

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the solution was determined during the batch test owing to the leaching of Fe from

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FeS-SA.

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Antioxygenic property

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Anoxic conditions rarely occur in the real world. Therefore, it is essential to

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investigate the effect of O2 on the removal efficiency of FeS-SA. The glass vials were

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exposed to air, whereas the remaining experimental conditions were identical to those

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used for the sealed glass vials. The change of Se(IV) concentration at the

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predetermined times was recorded. Furthermore, the changes of Fe(II) concentration

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of FeS-SA beads after exposure to air for 12 h were detected in the presence of air.

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The results could be used as a criterion to judge the effect of oxidation on the

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reducibility of FeS. For comparison, aforementioned same procedures were also

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conducted in pure stabilized FeS nanoparticles suspension.

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Results and discussion

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Characterization of FeS-SA beads

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SEM was used to determine and characterize the surface morphology of the

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prepared FeS-SA beads. The SEM image of the external surface of a FeS-SA bead

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indicated that it contained FeS particles surrounded by SA (Figure 1A). The XRD

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pattern of the FeS-SA beads showed intense peaks at 2θ = 17.6°, 30.1°, 39.0°, 50.3°,

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and 53.0°, which respectively corresponded to the (001), (101), (111), (112), and (201)

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planes of mackinawite (FeS) (JCPDS no. 89-2738) (Figure 1B). These results

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confirmed the successful preparation of FeS. The black FeS-SA beads with an

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approximate diameter of 1.3 mm can be seen in the photograph presented in Figure

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S2A.

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Optical microscopy and TEM were used to characterize the structure of FeS-SA

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after its cutting into slices. Many dendritic-like pores were evenly scattered inside the

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gel (Figure 1C). The network can be seen in the TEM image (Figure 1D). The pore

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diameter was approximately tens of microns. The magnification of the slice during the

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observation by TEM revealed that FeS nanoparticles with an approximate diameter of

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50 nm were dispersed homogeneously on the organic supporter (Figure 1E). The

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stabilized FeS nanoparticles approximately 80 nm in diameter were dispersed

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homogeneously (Figure S3A), whereas non-stabilized FeS formed large agglomerates

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(Figure S3B).

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Removal of Se(IV) by FeS-SA beads

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The final pH after batch experiments was measured and result showed that minor

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change occurred during the experiments owing to the HEPES buffer ability. The

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adsorption isotherm data of FeS-SA beads and two isotherm models are presented in

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Figure S4A. The maximum adsorption capacity was found to increase from 20.0 to

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128.9 mg/g as the initial concentration of Se(IV) rose from 0.025 to 0.25 mM. The

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parameters and coefficients of determination (R2) of Langmuir and Freundlich

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isotherm models are listed in Table S1. The adsorption intensity (1/n) and the

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distribution coefficient (KF) increased with the elevation of the initial Se(IV)

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concentration. This indicated the dependence of adsorption on the initial

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concentration.

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Based on the results of Fe(II) concentration, we determined the amount of FeS in

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FeS-SA, which was about 5.00 mg. TG analysis indicated that the weight difference

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after 700 °C was around 0.19% (Figure S5). The amount of FeS in FeS-SA was

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calculated to be approximately 4.94 mg based on the initial weight of FeS-SA (2.6 g).

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It was worth noting that above amount of FeS was based on FeS-SA beads from 10

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mL Na2S–SA mixture. In the batch test, only 4 mL Na2S–SA mixture was used to

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prepare FeS-SA. Therefore, the amount of FeS in the FeS-SA was about 2 mg and the

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concentration of FeS in batch experiment was 1.14 mM (20 mL solution). For

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comparison, 1.14 mM pure stabilized and non-stabilized FeS suspensions were

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prepared in 20 mL solution. The changes of the Se(IV) concentration in the

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experimental sets are depicted in Figure 2A. The concentration of Se(IV) in pure SA

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beads decreased only within the initial 100 min, and approximately 20% of the Se(IV)

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in the solution was removed. Almost no significant change happened after that. The

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concentration of Se(IV) in the pure non-stabilized FeS suspension shared a similar

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tendency with bare calcium alginate gels and about 27% Se(IV) was removed by this

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non-stabilized FeS. At the point of equilibrium, approximately 73% of the Se(IV) was

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229

still present in the aqueous solution. In the pure stabilized FeS nanoparticles

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suspension, the removal efficiency increased to 73% that apparently resulted from the

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higher dispersibility and reactivity (Figure S3). However, almost 100% of the Se(IV)

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was removed by FeS-SA within 800 min. The removal rate obviously decreased

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gradually until the complete removal of Se(IV). It is noteworthy that even the joint

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removal efficiency of pure non-stabilized FeS suspension and SA beads was not

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comparable with that of FeS-SA beads. Interestingly, if the overall removal efficiency

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of pure stabilized FeS nanoparticles and SA beads was taken into consideration, it was

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consistent with the removal efficiency of FeS-SA beads. Therefore, it could be

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concluded that the superior removal efficiency of FeS-SA beads benefited from the

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homogeneous distribution of FeS on the SA matrix.

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Although the adsorption of α-FeOOH for Se(IV) was demonstrated previously,28

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the removal efficiency of Se(IV) by α-FeOOH-SA was only 31% and therefore less

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than FeS-SA (Figure 2A). As for this 31% of removal efficiency, pure SA beads

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contributed 20%, thus sole α-FeOOH only removed 11% of Se(IV) (Table S2). On the

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one hand, the adsorption capacity of Se(IV) by α-FeOOH is low. Mustafa et al.

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showed that the removal capacity was only 2 mg/g.38 Similarly, 3.8 mg/g of capacity

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was also found in another study.28 On the other hand, earlier study has demonstrated

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that the mechanism of Se(IV) removal by α-FeOOH was attributed to the adsorption

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and no change in the Se(IV) oxidation state was found, but X-ray Absorption

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Near-Edge Structure (XANES) showed evidence of Se(IV) reduction in the presence

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of FeS.39 That is to say, except for the adsorption and reduction property of FeS, the

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oxidation products of FeS, such as α-FeOOH (see following analysis), could still

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adsorb Se(IV) after FeS was oxidized by Se(IV). Therefore, FeS-SA showed superior

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capacity for Se(VI) removal than α-FeOOH-SA. The oxidation product (mainly in the

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form of α-FeOOH) of FeS-SA by air (FeS-SA*) showed about 40% of removal

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efficiency. Obviously, FeS-SA* had a higher removal efficiency than α-FeOOH-SA. It

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was supposed that α-FeOOH was more homogeneously distributed in FeS-SA* than

257

α-FeOOH-SA, because the former α-FeOOH was generated in-situ in SA from air

258

oxidation, whereas the latter α-FeOOH was prepared ex-situ firstly then immobilized

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into SA. If the adsorption of SA for Se(IV) was not considered, sole α-FeOOH

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adsorption for Se(IV) in this study was only 11% (Table S2). Therefore, FeS-SA had

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superior Se(IV) removal properties compared to α-FeOOH sorbents in oxic systems.

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The commonly used pseudo-first-order and pseudo-second-order models were

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applied to simulate the kinetic data (Figure S4B and Figure S4C); the fitting

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parameters are shown in Table S3. The rate constant (K2) and equilibrium adsorption

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quantity (qe) were (4.5 ± 0.8) ×10-4g/(mg·min) and 98.0 ± 1.3 mg/g, respectively. For

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pure FeS*, FeS, and SA, however, irrespective of whether they were fitted by the

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pseudo-first-order or the pseudo-second-order models, the adsorption quantity and

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rate were apparently lower than those of the FeS-SA beads. Weber-Morris was used to

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describe the rate-limiting diffusion. As depicted in Figure S4D, the curve was

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separated three linear plots and it presents multi-linearity. The parameters of each

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linear plot are shown in Table S4. The initial sharper portion indicting external mass

272

transfer or instantaneous adsorption stage is followed by a linear portion, which is the

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gradual adsorption stage, where the intraparticle diffusion is rate-controlling step. The

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third plateau portion is final equilibrium stage where the intraparticle diffusion starts

275

to slow down due to extremely low solute concentrations in the solution. As shown in

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Figure S4D and Table S4, the Weber–Morris model could describe well the linear

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portion, which indicated that the intraparticle diffusion controlled the adsorption

278

process steps. However, the lines did not pass through the origin. Therefore, it

279

suggested that the mechanism of adsorption was complex and external and

280

intraparticle diffusion contributes to the actual Se(IV) removal process. In the Se(IV)

281

removal process, it is a complex process that involved not only adsorption by SA

282

beads but also reduction by FeS nanoparticles for the Se(IV) removal.

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The results of the column experiment showed the breakthrough curves of Se(IV)

284

through the fixed-bed column (Figure S6). At the initial stage, the Se(IV) was

285

completely removed. After 30 pore volumes, 80% of the Se(IV) was detected in the

286

effluent. However, it was not exhausted until 40 pore volumes were performed even

287

though the breakthrough had happened before that.

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The FeS-SA beads showed superior removal efficiency that was comparable with

289

the joint effect of pure stabilized FeS nanoparticles suspension and SA beads.

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Previous studies have also proposed various methods to eliminate the environmental

291

risks from Se(IV) contamination. A mixed adsorbent based on Fe-Mn hydrous oxides

292

have usually been applied for Se(IV) removal and its mechanism has been thoroughly

293

studied.40 However, an equilibrium adsorption quantity (qe) of only tens mg/g was

294

established in earlier examinations.41 In another investigation, although a stabilizer

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was used to stabilize these Fe-Mn binary oxide nanoparticles, its qe was still around

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100 mg/g.42 Our FeS-SA had a higher qe, which reached approximately 128.7 mg/g

297

based on Langmuir model (Table S1). The FeS nanoparticles were homogeneously

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distributed on the SA matrix, which reduced the aggregation of FeS and increased the

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quantity of its reactive sites. As depicted in aforementioned result (Figure 2A), the

300

adsorption of pure SA beads only removed 20% Se(IV). Thus, if we considered only

301

the adsorption of SA beads and non-stabilized FeS and the reduction ability of FeS,

302

their joint removal efficiency could be at least equal to the removal efficiency of

303

FeS-SA. However, the joint effect of pure non-stabilized FeS suspension and SA

304

beads obviously was not comparable with that of FeS-SA beads (Figure 2A). In fact,

305

the overall removal efficiency of pure stabilized FeS nanoparticles and SA beads was

306

consistent with the removal efficiency of FeS-SA beads (Figure 2A). Therefore, it

307

could be assumed that the stabilized and uniformly distributed FeS was responsible

308

for enhancing the removal efficiency, which was also confirmed by the TEM image

309

(Figure 1E). Consequently, the efficiency of Se(IV) removal was improved

310

Effect of FeS Oxidation

311

The above-described experiments with Se(IV) removal were all conducted under

312

anoxic conditions, and thus it was essential to evaluate the removal efficiency of

313

Se(IV) containing dissolved oxygen. As depicted in Figure 2B, the removal efficiency

314

and rate of Se(IV) by FeS-SA beads in the absence and presence of air did not display

315

a significant difference. When complete removal was achieved in the sealed glass vial,

316

only 10% of the Se(IV) was left in the solution exposed to air. However, the removal

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317

efficiency decreased from 73% to 36% in stabilized FeS nanoparticles suspension in

318

the presence of air. It indicates that O2 exerts negative influence on the removal

319

efficiency of Se(IV) by pure stabilized FeS nanoparticles suspension. The changes of

320

Fe(II) concentration in FeS-SA beads and pure stabilized FeS nanoparticles

321

suspension are shown in Table 1. The initial concentrations of Fe(II) were all

322

approximately 1.17 mM. Nevertheless, this concentration decreased to 1.07 mM in

323

the FeS-SA beads, and 91% of Fe(II) was not oxidized after exposure to air for 12 h.

324

However, only 16% of Fe(II) was found in the stabilized FeS nanoparticles

325

suspension; obviously, the remaining 84% of Fe(II) was oxidized by dissolved

326

oxygen.

327

To the best of our knowledge, this is the first time that immobilization

328

technology has been employed to form an anoxic condition, thus decreasing the FeS

329

oxidation by dissolved oxygen. We found minor differences between the removal

330

efficiency of FeS-SA beads under anoxic and aerobic conditions. Although some good

331

results were achieved in removing Se(IV) by pure FeS in previous research,5, 43 all

332

experiments have been conducted under anoxic conditions due to the easy oxidation

333

of FeS. To our knowledge, reducing compounds cannot exist long in oxic waters

334

unless kinetic or physical constraints can keep oxygen from quickly reacting with FeS.

335

Therefore, we cannot ignore this problem if FeS is applied in real practice. In this

336

study, FeS was immobilized into the calcium alginate gel beads. This form of

337

application contributed to the retardation of the oxidation of FeS by dissolved oxygen.

338

Thus, the removal efficiency of FeS-SA beads under anoxic conditions was only 10%

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higher than under aerobic conditions. The electron utilization ratio of FeS, describing

340

the ratio of the electrons that target Se(IV) to the total electrons provided by FeS, was

341

improved.44,

342

between FeS and Se(IV). The dilute Se(IV) solution could be concentrated inside of

343

SA beads by adsorption. Consequently, reduction of Se(IV) proceeded in the gel

344

beads.

345

Regeneration of FeS-SA

45

Besides, SA beads provided the necessary space for the reaction

346

The exhausted FeS-SA beads were regenerated in the sulfate reducing conditions.

347

It could be found that yellow exhausted FeS-SA beads at the beginning changed color

348

to brown (12 h) and then black (Figure S7). The regenerated FeS-SA was used to test

349

the removal performance for Se(IV). As shown in Figure 3, in the first cycle, almost

350

100% Se(IV) was removed. However, the removal efficiency gradually decreased as

351

the FeS-SA beads were regenerated. After 5 cycles, there was only 40% removal for

352

Se(IV). It was speculated that decreased removal efficiency was attributed to the

353

leaching of Fe from SA, thus the amount of Fe in the solution was measured. The

354

amount of Fe leaching increased with the regenerating of FeS-SA and reached to 1.02

355

mg after 5 cycles. It was mentioned earlier that 2 mg FeS was immobilized into the

356

FeS-SA, thus the amount of Fe was about 1.27 mg. 80% Fe leaked from FeS-SA after

357

5 cycles regeneration and only 20% Fe left in the matrix of SA. Interestingly, if 20%

358

Fe corresponded to 20% removal efficiency of Se(IV) by FeS-SA, the overall removal

359

efficiency of pure SA beads (pure SA was responsible for 20% Se(IV) removal

360

(Figure 2A)) and sole 20% Fe was consistent with the removal efficiency (40%) of

17



361

FeS-SA beads after 5 cycles. Although the leaching of Fe from FeS-SA occurred, the

362

FeS-SA still possessed a remarkable regenerated performance.

363

In this study, the regenerated FeS-SA under the sulfate reducing conditions

364

showed 40% removal efficiency for Se(IV) after 5 cycles due to the Fe leaching.

365

Although FeS possesses excellent reducing ability, it is difficult to regenerate once it

366

is oxidized to form iron oxide in oxic condition. Actually, the regeneration of

367

exhausted FeS-SA by SRB is assigned to sulfidization of iron oxides. Previous study

368

demonstrated that adsorbed U(VI) during sulfidization of lepidocrocite and hematite

369

could be desorbed and released to solution, then reduced.46 Moreover, arsenic could

370

also be repartitioned during biogenic sulfidization and transformation of ferrihydrite.47

371

Therefore, it can be supposed that the Se (mainly in the form of Se(0) and FeSe) can

372

detach from α-FeOOH during sulfidization, so that the removal capacity for Se(IV)

373

can be recovered. However, it does not mean that Se will release to the SRB medium

374

solution, because the detached Se can still stay in the SA matrix. In order to

375

demonstrate this point, the total Se in the SRB medium solution was measured during

376

sulfidization, but only quite a few (0.02 mM) was recorded, which is beneficial to

377

following recovery of Se due to the its retain in SA. Most earlier applications of FeS

378

are considered from the point of view of groundwater remediation in which FeS can

379

be generated in situ in anoxic environment. In addition, sulfate-reducing bacteria

380

(SRB) have the ability of producing FeS in the form of mackinawite,48 and such

381

biogenically produced FeS has been shown to be reactive towards contaminants such

382

as arsenic.49 Therefore, it is valuable to employ biostimulation to regenerate the

18


Page 18 of 42

Page 19 of 42


383

exhausted FeS in the sulfate reducing conditions. This FeS-SA had excellent

384

regeneration performance, thus it is a good candidate for environmental remediation.

385

Effects of pH and anions on Se(IV) removal

386

The levels of pH and anions are vital factors to the successful treatment of

387

groundwater due to the complexity of water sources27, 50. We found that the efficiency

388

of Se(IV) removal under neutral and slightly acidic conditions was higher than under

389

alkaline. Higher pH values resulted in lower removal efficiencies as the pH increased

390

from 4 to 10 (Figure S8A). The effects of various co-existing anions (phosphate,

391

carbonate, and chloride) at five concentration levels (0, 5, 40, 80, and 120 mg/L) on

392

the adsorption of Se(IV) was also investigated, the results of which are illustrated in

393

Figures S8B-D. The low concentrations of anions exerted minor effects on Se(IV)

394

removal. However, the removal efficiency apparently decreased with the increase of

395

anion concentration, and only about 40% of the Se(IV) was removed. Therefore, the

396

co-existing anions might exert negative influence on the Se(IV) removal, which led to

397

a considerable reduction in the removal performance of the FeS-SA beads.

398

Analysis of final products

399

As can be seen from the TEM image of a FeS-SA slice after the reaction with

400

Se(IV) (shown in Figure 1F), some needle-like products and particles were attached

401

on the surface of the organic slice. The color also changed from black to

402

yellowish-brown (Figure S2B). As depicted in Figure 4A, the morphology of FeS-SA

403

beads after Se(IV) removal had a smooth surface with some particles attached on it.

404

Elemental mapping indicated that the loading of Fe and Se was distributed uniformly

19



405

on the surface of FeS-SA (Figure 4B and Figure 4D). The big highlight of sulfur dots

406

were more obvious than Fe and Se (Figure 4C) that ascribed to the sulfide oxidation

407

to elemental sulfur.51, 52

408

As depicted in aforementioned result (Figure 2A), the adsorption of pure SA gel

409

only removed 20% Se(IV), thus it could be supposed that adsorption and reduction of

410

Se(IV) by FeS in SA was responsible for the significant Se(IV) removal. To better

411

understand the removal mechanism of Se(IV) by FeS-SA, Fe 2p, 3p, and Se 3d XPS

412

spectra were obtained and characterized before and after the Se(IV) treatment. The Fe

413

2p3/2 spectra are fitted in Figure S9, and the detailed analytical results are provided in

414

Table S5. Fe was present mainly in the form of Fe(II)–S species (90.2%; signals

415

centered at 707.4 and 710.3 eV). Only a small amount of the Fe was in the form of

416

Fe(II)–O and Fe(III)–S species before the reaction (Figure S9A). Fe(III)–O species,

417

including α-FeOOH (accounting for 68.9%; signals centered at 711.5 eV and 713.0

418

eV) and FeSe (26.8%, centered at 710.9 eV), were the main forms present in the final

419

reaction products (Figure S9B). The sulfur XPS spectrum showed that almost all S(-II)

420

was oxidized to Sn (Figure S10 and Table S5). The analysis of Se on FeS-SA by this

421

technique indicated that all the photoelectrons of Se appeared to interfere with the

422

photoelectrons of Fe.53, 54 We thus chose to study the Se 3d spectra and deconvolute

423

the signal in two contributions: the Se 3d and the Fe 3p peaks. Se(IV) (SeOଶି ଷ ),

424

elemental Se (Se(0)), and selenide (Se(-II) were the main possible forms of Se species

425

after the treatment. The Na2SeO3, Se(0), and FeSe reference compounds were

426

characterized via Se 3d XPS. Their Se 3d signals exhibited peaks at 58.55, 55.5, and

20


Page 20 of 42

Page 21 of 42


427

55.0 eV, respectively (Figure 5). These energy are almost in agreement with the ones

428

given by previous studies.55,

429

deconvolute due to the asymmetry of the Fe signal, the resulting peak reflected the

430

combined effect of Se 3d and Fe 3p. Therefore, the combined contribution of Se 3d

431

and Fe 3p was considered. Similarly, we chose to characterize Fe(OH)3 by Fe 3p XPS

432

due to the existence of Fe (III)-O as the main Fe species based on Fe 2p XPS spectra

433

(except for FeSe). The Fe 3p peak of the Fe(OH)3 reference compound was centered

434

at 55.9 eV. The final products were deconvoluted based on the reference compounds

435

(e.g. 55.0, 55.5, 55.9, and 58.55 eV). The results are fitted in Figure 5, and detailed

436

analytical results are shown in Table 2. FeSe, Se (0), and Fe(OH)3 constituted 27.9%,

437

30.6%, and 38.4% of the final products, respectively. Furthermore, FeSe, Se (0), and

438

Na2SeO3 were the main forms of Se (45.3%, 49.7%, and 5.0%). From the above XPS

439

analysis results, it could be concluded that, Se(IV) in the final products was mainly

440

reduced to FeSe and Se (0). Earlier studies using advanced XANES also demonstrated

441

that FeSe and Se (0) were the main reduced products when contacting with FeS.5, 43, 57

442

We combined the electron balance with XPS data to make the quantitative

443

fraction of Se more reliable. In this study, Fe(II) and S（-II）are electron donor, and

444

Se(IV) is electron acceptor. According to the XPS analytic results (Table S5), 65.5 %

445

Fe(II) (denoted as FFe(II)) and 96.9 % S(-II) (FS(-II)) in the initial solution were oxidized

446

to Fe(III) and Sn(0), so the amount of donated electrons (Nd) can be calculated by eq1.

447 448

56

Even though the FeSe signal was difficult to

Nd= FFe(II) × NFe(II) + FS(-II) × 2NS(-II)

(1)

where NFe(II) and NS(-II) are the amounts of Fe(II) and S(-II). The Se(IV) reduced

21



Page 22 of 42

449

products are Se(0) and Se(-II) based on the Se 3d analysis (Figure 5 and Table 2). It is

450

worth noting that the relative fraction of Se in Table 2 included the Fe 3p contribution,

451

so the relative fraction of Se(0) (FSe(0)) and Se(-II) (FSe(-II))in the final Se products are

452

49.7% and 45.3 %, respectively. Therefore, the amount of accepted electrons (Na) can

453

be calculated by eq2.

454

Na= FSe(0) × 4NSe(0) + FSe(0) × 6NSe(-II)

(2)

455

where NSe(0) and NSe(-II) are the amounts of Se(0) and Se(-II). It is worth noting that

456

the initial 50 mL solution contained 1.14 mM FeS and 0.63 mM Se(IV). Therefore,

457

the Nd is 0.147 mmol and Na is 0.148 mmol. Nd is almost equal to Na, which further

458

indicated the quantitative fraction of Se from XPS data is credible.

459

Even though the final products of the removal of Se(IV) by FeS had been

460

previously studied by advanced XANES, we achieved similar results by using the

461

traditional XPS technique that is commonly available to researchers. The Fe

462

compounds were first analyzed by Fe 2p XPS to confirm the Fe phase. Then, the Fe

463

reference compounds and possible Se forms reference compounds were characterized

464

by Se 3d XPS. If Se compounds were considered separately, it was difficult to obtain

465

the fraction of each Se compound due to the overlap of Fe 3p. In our study, however,

466

Fe reference compound (especially Fe(OH)3) was considered as one of Se compounds.

467

The individual contribution of Fe 3p for the Se 3d was deconvoluted. Therefore, the

468

Se was deconvoluted successfully even though the photoelectrons of Se interfered

469

with the photoelectrons of Fe. Previous study also employed XPS technique and used

470

reference compounds to analyze the solids after Se sorption by pyrite and chalcopyrite,

22


Page 23 of 42


471

then XANES was used to demonstrate related conclusion.55 Therefore, it was reliable

472

to utilize XPS technique to obtain the fraction of each Se.

473

In situ immobilization and strength property

474

FeS is easily oxidized during the preparation and immobilization processes. In

475

this study, FeS-SA was in situ prepared, which means preparation and immobilization

476

of FeS were carried out simultaneously. Consequently, this approach facilitated the

477

effective avoidance of the oxidation of FeS during the preparation and immobilization

478

processes. Traditionally, adsorbent materials were first prepared and then immobilized

479

into

480

Fe3O4@Zr(OH)4-impregnated chitosan beads were synthesized by mix-shaping of

481

previously prepared Fe3O4@Zr(OH)4 and chitosan slurry.27 A biosorbent was prepared

482

by coating prior-prepared ceramic alumina with a natural biopolymer, chitosan, using

483

a dip-coating process for the removal of arsenic (III) and arsenic (V).58 In our

484

examination, SA was initially mixed with a Na2S solution, and CaCl2 was mixed with

485

a FeSO4 solution. When the mixed solution of Na2S–SA was dropwise added to the

486

mixed solution of FeSO4-CaCl2, the formation of FeS and calcium alginate gel

487

occurred simultaneously. There are two advantages in using in situ one-step

488

preparation of FeS-SA beads. On the one hand, it could prevent the oxidation of FeS

489

during the processes of preparation and immobilization. FeS is easy to oxidize during

490

the operational process if it is first prepared and then immobilized into the SA matrix.

491

On the other hand, the Na2S–SA and FeSO4-CaCl2 solutions were uniformly mixed

492

when the gel was formed. Hence, the synthesized FeS nanoparticles could

the

organic

matrix.

For

example,

23


the

magnetic

particles


Page 24 of 42

493

immediately combine with gel network and uniformly distribute on the gel. This

494

outcome could enhance the reactivity of FeS nanoparticles due to reducing the

495

aggregation.11, 59 The mechanical strength-test was carried out with the rheometer to

496

estimate the shear resistance of FeS-SA beads. Prior to reaching to the critical point,

497

the initial beads before Se(IV) treatment displayed splendid performance and the

498

strength kept the high value namely of high elastic modulus and viscous modulus

499

(Figure S11A). No obvious change occurred after treating Se(IV) (Figure S11B),

500

which indicated the long term usability of such beads in water treatment is suitable. In

501

addition, although chitosan, polyvinyl alcohol and polyacrylamide can also be used as

502

immobilization materials in earlier studies,60-62 the advantages of SA as

503

immobilization materials are apparent including simple preparation process,

504

economical efficiency, environment-friendly, reliable strength, etc. Certainly, the

505

composite of SA and other immobilization materials may be better. Serp et al. found

506

the mechanical resistance of alginate beads can be doubled by treatment with 5–10

507

kDa chitosan.63 María studied microencapsulation of a probiotic and prebiotic in

508

alginate-chitosan capsules that improved survival in simulated gastro-intestinal

509

conditions.64 Fe0-Fe3O4 nanocomposites embedded polyvinyl alcohol/sodium alginate

510

beads showed high stability, Cr(VI) removal efficiency and reusability.65

511 512 513 514

Implication This study applied in situ immobilization technology to prepare stabilized FeS nanoparticles-impregnated

alginate

composite

(FeS-SA)

24


beads

that

possess

Page 25 of 42


515

antioxidation property for successful Se(IV) remediation. Dissolved O2 in the shallow

516

wastewater is commonly found that apparently makes the FeS passivated and limits

517

FeS application. Therefore, SA gel endowed FeS with an antioxidation property that is

518

vital if FeS is applied for the real wastewater. In addition, the oxidation of FeS in the

519

process of preparation, storage, and transport can be effectively avoided. If these

520

FeS-SA beads are used in the underground water, dissolved O2 may not be a limiting

521

factor for the FeS. However, SRB can produce sulfide that reacts Fe(III) oxides to

522

form FeS under anaerobic conditions. Consequently, the exhausted FeS can be

523

regenerated in the interior of SA and used again till complete leaching of Fe from SA.

524

Subsequent recovery after pollutants treatment can be easily achieved by a sifter with

525

fixed diameter, although this troublesome problem is commonly ignored in some

526

investigations. Certainly, this in situ immobilization method is applicable to not only

527

FeS nanoparticles but also other reductive nanoparticles, for example, nanoscale

528

zero-valent iron obtained by the simple preparation technique.

529 530

Supporting information

531

Chemicals, the cultivation of SRB, instrumental characterization, determination

532

of FeS amount in SA beads, kinetic and isotherm models, experimental details about

533

the batch and column experiment, TEM and XRD patterns of the α-FeOOH,

534

photograph of FeS-SA, TEM of stabilized and non-stabilized FeS, TG curve of

535

FeS-SA and SA, breakthrough curves of Se(IV) through fixed-bed column,

536

photograph of the process of FeS-SA regeneration, the effects of pH and anions on the

25



537

reactivity of FeS-SA, Fe and S 2p3/2 XPS spectra of FeS-SA, strength analysis of

538

FeS-SA, Se(IV) removal efficiency in various species and separated component,

539

fitting results of model, XPS analytic results.

26


Page 26 of 42

Page 27 of 42


540

Acknowledgments

541

The authors would like to acknowledge the financial support of Collaborative

542

Innovation Center of Suzhou Nano Science and Technology, the Program for

543

Changjiang Scholars and Innovative Research Team in University, and the

544

Fundamental Research Funds for the Central Universities.

27



545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560 561 562 563 564 565 566 567 568 569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587 588

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633 634 635 636 637 638 639 640 641 642 643 644 645 646 647 648 649 650 651 652 653 654 655 656 657 658 659 660 661 662 663 664 665 666 667 668 669 670 671 672 673 674 675 676

reductase in periplasm. Sci. Rep. 2014, 4. 32. Leupin, O. X.; Hug, S. J.; Badruzzaman, A. B. M., Arsenic removal from Bangladesh tube well water with filter columns containing zerovalent iron filings and sand. Environ. Sci. Technol. 2005, 39, (20), 8032-8037. 33. Matheson, L. J.; Tratnyek, P. G., Reductive Dehalogenation of Chlorinated Methanes by Iron Metal. Environ. Sci. Technol. 1994, 28, (12), 2045-2053. 34. Liu, Y. Q.; Lowry, G. V., Effect of particle age (Fe-o content) and solution pH on NZVI reactivity: H-2 evolution and TCE dechlorination. Environ. Sci. Technol. 2006, 40, (19), 6085-6090. 35. Juang, R. S.; Tseng, R. L.; Wu, F. C.; Lee, S. H., Adsorption behavior of reactive dyes from aqueous solutions on chitosan. J. Chem. Technol. Biot. 1997, 70, (4), 391-399. 36. Zhu, H. J.; Jia, Y. F.; Wu, X.; Wang, H., Removal of arsenic from water by supported nano zero-valent iron on activated carbon. J. Hazard. Mater. 2009, 172, (2-3), 1591-1596. 37. Chatterjee, A.; Schiewer, S., Multi-resistance kinetic models for biosorption of Cd by raw and immobilized citrus peels in batch and packed-bed columns. Chem. Eng. J. 2014, 244, 105-116. 38. Sharrad, M. O. M.; Liu, H. J.; Fan, M. H., Evaluation of FeOOH performance on selenium reduction. Sep. Purif. Technol. 2012, 84, 29-34. 39. Mitchell, K.; Couture, R. M.; Johnson, T. M.; Mason, P. R. D.; Van Cappellen, P., Selenium sorption and isotope fractionation: Iron(III) oxides versus iron(II) sulfides. Chem. Geol. 2013, 342, 21-28. 40. Chubar, N.; Gerda, V.; Szlachta, M., Mechanism of Selenite Removal by a Mixed Adsorbent Based on Fe-Mn Hydrous Oxides Studied Using X-ray Absorption Spectroscopy. Environ. Sci. Technol. 2014, 48, (22), 13376-13383. 41. Szlachta, M.; Chubar, N., The application of Fe-Mn hydrous oxides based adsorbent for removing selenium species from water. Chem. Eng. J. 2013, 217, 159-168. 42. Xie, W. B.; Liang, Q. Q.; Qian, T. W.; Zhao, D. Y., Immobilization of selenite in soil and groundwater using stabilized Fe-Mn binary oxide nanoparticles. Water Res. 2015, 70, 485-494. 43. Breynaert, E.; Bruggeman, C.; Maes, A., XANES-EXAFS analysis of se solid-phase reaction products formed upon contacting Se(IV) with FeS2 and FeS. Environ. Sci. Technol. 2008, 42, (10), 3595-3601. 44. Liu, H.; Wang, Q.; Wang, C.; Li, X. Z., Electron efficiency of zero-valent iron for groundwater remediation and wastewater treatment. Chem. Eng. J. 2013, 215, 90-95. 45. Liu, Y. Q.; Majetich, S. A.; Tilton, R. D.; Sholl, D. S.; Lowry, G. V., TCE dechlorination rates, pathways, and efficiency of nanoscale iron particles with different properties. Environ. Sci. Technol. 2005, 39, (5), 1338-1345. 46. Alexandratos, V. G.; Behrends, T.; Van Cappellen, P., Fate of Adsorbed U(VI) during Sulfidization of Lepidocrocite and Hematite. Environ. Sci. Technol. 2017, 51, (4), 2140-2150. 47. Kocar, B. D.; Borch, T.; Fendorf, S., Arsenic repartitioning during biogenic 30


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sulfidization and transformation of ferrihydrite. Geochim. Cosmochim. Ac. 2010, 74, (3), 980-994. 48. Vaughan, D. J.; Lennie, A. R., The Iron Sulfide Minerals - Their Chemistry and Role in Nature. Sci. Prog. 1991, 75, (298), 371-388. 49. Vannela, R.; Adriaens, P.; Hayes, K. F., Reactivity studies of biogenic nanosized iron sulfide minerals for As(III) removal. 234th National Meeting and Exposition of the American Chemical Society 2007. 50. Abollino, O.; Aceto, M.; Malandrino, M.; Sarzanini, C.; Mentasti, E., Adsorption of heavy metals on Na-montmorillonite. Effect of pH and organic substances. Water Res. 2003, 37, (7), 1619-1627. 51. Finck, N.; Dardenne, K., Interaction of selenite with reduced Fe and/or S species: An XRD and XAS study. J. Contam. Hydrol. 2016, 188, 44-51. 52. Charlet, L.; Kang, M. L.; Bardelli, F.; Kirsch, R.; Gehin, A.; Greneche, J. M.; Chen, F. R., Nanocomposite Pyrite-Greigite Reactivity toward Se(IV)/Se(VI). Environ. Sci. Technol. 2012, 46, (9), 4869-4876. 53. Briggs, D. S., M. P., Practical Surface Analysis, 2nd ed. John Wiley and Sons: New York 1993, 1. 54. Wagner, C. D. R., W. M.; Davis, L. E.; Moulder, J. F., Handbook of X-ray photoelectron spectroscopy. Muilenberg G.E., Perkin-Elmer Corporation: Eden Prarie, MN 1979. 55. Naveau, A.; Monteil-Rivera, F.; Guillon, E.; Dumonceau, J., Interactions of aqueous selenium(-II) and (IV) with metallic sulfide surfaces. Environ. Sci. Technol. 2007, 41, (15), 5376-5382. 56. Shenasa, M.; Sainkar, S.; Lichtman, D., Xps Study of Some Selected Selenium-Compounds. J. Electron. Spectrosc. 1986, 40, (4), 329-337. 57. Scheinost, A. C.; Kirsch, R.; Banerjee, D.; Fernandez-Martinez, A.; Zaenker, H.; Funke, H.; Charlet, L., X-ray absorption and photoelectron spectroscopy investigation of selenite reduction by Fe-II-bearing minerals. J. Contam. Hydrol. 2008, 102, (3-4), 228-245. 58. Boddu, V. M.; Abburi, K.; Talbott, J. L.; Smith, E. D.; Haasch, R., Removal of arsenic(III) and arsenic(V) from aqueous medium using chitosan-coated biosorbent. Water Res. 2008, 42, (3), 633-642. 59. Xiong, Z.; He, F.; Zhao, D. Y.; Barnett, M. O., Immobilization of mercury in sediment using stabilized iron sulfide nanoparticles. Water Res. 2009, 43, (20), 5171-5179. 60. Li, X. L.; Li, Y. F.; Ye, Z. F., Preparation of macroporous bead adsorbents based on poly(vinyl alcohol)/chitosan and their adsorption properties for heavy metals from aqueous solution. Chem. Eng. J. 2011, 178, 60-68. 61. Yamani, J. S.; Lounsbury, A. W.; Zimmerman, J. B., Adsorption of selenite and selenate by nanocrystalline aluminum oxide, neat and impregnated in chitosan beads. Water Res. 2014, 50, 373-381. 62. Yang, S. B.; Hu, J.; Chen, C. L.; Shao, D. D.; Wang, X. K., Mutual Effects of Pb(II) and Humic Acid Adsorption on Multiwalled Carbon Nanotubes/Polyacrylamide Composites from Aqueous Solutions. Environ. Sci. Technol. 2011, 45, (8), 3621-3627. 31



721 722 723 724 725 726 727 728 729 730

63. Serp, D.; Cantana, E.; Heinzen, C.; von Stockar, U.; Marison, I. W., Characterization of an encapsulation device for the production of monodisperse alginate beads for cell immobilization. Biotechnol. Bioeng. 2000, 70, (1), 41-53. 64. Chavarri, M.; Maranon, I.; Ares, R.; Ibanez, F. C.; Marzo, F.; Villaran, M. D., Microencapsulation of a probiotic and prebiotic in alginate-chitosan capsules improves survival in simulated gastro-intestinal conditions. Int. J. Food Microbiol. 2010, 142, (1-2), 185-189. 65. Lv, X.; Jiang, G.; Xue, X.; Wu, D.; Sheng, T.; Sun, C.; Xu, X., Fe(0)-Fe3O4 nanocomposites embedded polyvinyl alcohol/sodium alginate beads for chromium (VI) removal. J. Hazard. Mater. 2013, 262, 748-58.

32


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Page 33 of 42


731

Firgure captions

732

Figure 1. Characterization of FeS-SA. (A) SEM photomicrograph of the external

733

surface of a FeS-SA bead; (B) XRD patterns of FeS-SA and standard FeS

734

(Mackinawite, JCPDS: 89-2738); (C) optical photomicrograph of FeS-SA; TEM

735

images of (D) araneose FeS-SA; (E) FeS particles formed homogeneously in the

736

FeS-SA; (F) FeS-SA after the reaction with Se(IV).

737

Figure 2. (A) Profiles of Se(IV) removal under different conditions; (B) The removal

738

efficiency of Se(IV) by FeS-SA and FeS* in the absence and presence of air, square

739

denotes FeS-SA and triangle denotes FeS*; Initial Se = 0.13 mM, initial pH = 6.0 ±

740

0.2.

741

Figure 3. The regeneration profiles of FeS-SA for Se(IV) removal in presence of SRB

742

and leaching of Fe from FeS-SA (Errors given as standard deviation among triplicate

743

samples).

744

Figure 4. (A) SEM image of FeS-SA bead and the distribution of (B) Fe, (C) S, and

745

(D) Se in the FeS-SA bead after Se(IV) removal.

746

Figure 5. Se 3d and Fe 3p XPS spectra of reference compounds (FeSe, Na2SeO3, Se,

747

and Fe(OH)3 ) and the sample of FeS-SA beads after Se(IV) removal.

33



748 749

Figure 1. Characterization of FeS-SA. (A) SEM photomicrograph of the external

750

surface of a FeS-SA bead; (B) XRD patterns of FeS-SA and standard FeS

751

(Mackinawite, JCPDS: 89-2738); (C) optical photomicrograph of FeS-SA; TEM

752

images of (D) araneose FeS-SA; (E) FeS particles formed homogeneously in the

753

FeS-SA; (F) FeS-SA after the reaction with Se(IV).

34


Page 34 of 42

Page 35 of 42


754

755

Figure 2. (A) Profiles of Se(IV) removal under different conditions; (B) The removal

756

efficiency of Se(IV) by FeS-SA and FeS* in the absence and presence of air, square

757

denotes FeS-SA and triangle denotes FeS*; Initial Se = 0.13 mM, initial pH = 6.0 ±

758

0.2. Here, SA stabilized FeS nanoparticles suspension is abbreviated as FeS* and

759

non-stabilized FeS is abbreviated as FeS; Oxidized FeS-SA by air is abbreviated as

760

FeS-SA* (Errors given as standard deviation among triplicate samples).

35



761 762

Figure 3. The regeneration profiles of FeS-SA for Se(IV) removal in presence of SRB

763

and leaching of Fe from FeS-SA (Errors given as standard deviation among triplicate

764

samples).

36


Page 36 of 42

Page 37 of 42


765 766

Figure 4. (A) SEM image of FeS-SA bead and the distribution of (B) Fe, (C) S, and

767

(D) Se in the FeS-SA bead after Se(IV) removal.

37



768

Figure 5. Se 3d and Fe 3p XPS spectra of reference compounds (FeSe, Na2SeO3, Se,

769

and Fe(OH)3 ) and the sample of FeS-SA beads after Se(IV) removal.

38


Page 38 of 42

Page 39 of 42


770

Table lists

771

Table 1. The change of Fe(II) concentration after exposure to air.

772

Table 2. XPS analytic results based on the curve fitting for Se 3d and Fe 3p peaks in

773

Figure 5.

39



Page 40 of 42

774

Table 1. The change of Fe (II) concentration after exposure to air (the error data

775

represent standard deviation among triplicate). FeS* (mM)a

FeS-SA (mM)

776

0h

1.17 ± 0.02

1.17 ± 0.02

12 h

1.07 ± 0.04

0.19 ± 0.02

C/C0

0.91 ± 0.05

0.16 ± 0.02

a-FeS* is the abbreviation of stabilized FeS nanoparticles suspension.

40


Page 41 of 42


777

Table 2. XPS analytic results based on the curve fitting for Se 3d and Fe 3p peaks in

778

Figure 5. Element

Se 3d

Fe 3p

Binding energy (eV)

species

Relative fraction (%)

55.0

FeSe

27.9 ± 2.8

58.55

Na2SeO3

3.1 ± 1.2

55.5

Se

30.6 ± 3.2

55.9

Fe(OH)3

38.4 ± 2.5

779

41



84x49mm (144 x 144 DPI)


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In Situ Preparation of Stabilized Iron Sulfide Nanoparticle-Impregnated

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