Journal of Hazardous Materials 290 (2015) 26–33

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Journal of Hazardous Materials

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Efficient removal of uranium from aqueous solution by zero-valentiron nanoparticle and its graphene composite

Zi-Jie Li a, Lin Wanga, Li-Yong Yuana, Cheng-Liang Xiaob, Lei Meia, Li-Rong Zhengc,Jing Zhangc, Ju-Hua Yanga, Yu-Liang Zhaoa, Zhen-Tai Zhud,Zhi-Fang Chaib,∗, Wei-Qun Shia,∗

a Key Laboratory of Nuclear Radiation and Nuclear Energy Technology and Key Laboratory For Biomedical Effects of Nanomaterials and Nanosafety,Institute of High Energy Physics, Chinese Academy of Sciences, Beijing 100049, Chinab School of Radiological and Interdisciplinary Sciences (RAD-X), and Collaborative Innovation Center of Radiation Medicine of Jiangsu Higher EducationInstitutions, Soochow University, Suzhou 215123, Chinac Beijing Synchrotron Radiation Facility, Institute of High Energy Physics, Chinese Academy of Sciences, Beijing100049, Chinad State Key Laboratory of NBC Protection for Civilian, Beijing 102205, China

h i g h l i g h t s

• Uranium removal by ZVI-nps: inde-pendent of pH, the presence of CO3

2−,humic acid, or mimic groundwaterconstituents.

• Rapid removal kinetics and sorptioncapacity of ZVI-nps is 8173 mg U/g.

• Two reaction mechanisms: suffi-cient Fe0 → reductive precipitation asU3O7; insufficient Fe0 → hydrolysisprecipitation of U(VI).

• Fe/graphene composites: improvedkinetics and higher U(VI) reductionratio.

g r a p h i c a l a b s t r a c t

a r t i c l e i n f o

Article history:Received 23 October 2014Received in revised form 23 January 2015Accepted 9 February 2015Available online 11 February 2015

Keywords:Zero-valent iron nanoparticlesGraphene compositesUranium removalReductive precipitationHydrolysis precipitation

a b s t r a c t

Zero-valent iron nanoparticle (ZVI-np) and its graphene composites were prepared and applied in theremoval of uranium under anoxic conditions. It was found that solutions containing 24 ppm U(VI) couldbe completely cleaned up by ZVI-nps, regardless of the presence of NaHCO3, humic acid, mimic ground-water constituents or the change of solution pH from 5 to 9, manifesting the promising potential of thisreactive material in permeable reactive barrier (PRB) to remediate uranium-contaminated groundwater.In the measurement of maximum sorption capacity, removal efficiency of uranium kept at 100% untilC0(U) = 643 ppm, and the saturation sorption of 8173 mg U/g ZVI-nps was achieved at C0(U) = 714 ppm.In addition, reaction mechanisms were clarified based on the results of SEM, XRD, XANES, and chem-ical leaching in (NH4)2CO3 solution. Partially reductive precipitation of U(VI) as U3O7 was prevalentwhen sufficient iron was available; nevertheless, hydrolysis precipitation of U(VI) on surface would bepredominant as iron got insufficient, characterized by releases of Fe2+ ions. The dissolution of Fe0 cores

∗ Corresponding authors. Tel.: +86 10 88233968.E-mail addresses: zfchai@suda.edu.cn (Z.F. Chai), shiwq@ihep.ac.cn (W.Q. Shi).

http://dx.doi.org/10.1016/j.jhazmat.2015.02.0280304-3894/© 2015 Elsevier B.V. All rights reserved.

Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33 27

was assigned to be the driving force of continuous formation of U(VI) (hydr)oxide. The incorporation ofgraphene supporting matrix was found to facilitate faster removal rate and higher U(VI) reduction ratio,thus benefitting the long-term immobilization of uranium in geochemical environment.

© 2015 Elsevier B.V. All rights reserved.

1. Introduction

Uranium occurs in nature in primary deposits, U(IV)O2. Cur-rently, uranium is being used as a typical nuclear fuel to enhanceelectrical production ability. With uranium ore mining, process-ing, fuel manufacture, spent fuel reprocessing and other relatedactivities, more and more highly mobile U(VI) is released into theenvironment, making uranium a common contaminant to soils,surface and groundwater [1]. The concentration of uranium in someacid mine water and contaminated areas around waste disposalsites can even attain as high as several tens of ppm [2,3]. On theother hand, the allowed maximum level of uranium for drinkingwater is only 30 �g/L, recommended by EPA [4]. It is well knownthat intakes of uranium from food and/or drinking water can leadto internal irradiation and/or chemical toxicity. Long term expo-sure to uranium may result in cancer, kidney and liver damages,or all.

Permeable reactive barrier (PRB) technology has been suc-cessfully utilized to in situ remediate contaminated groundwater.Compared to conventional pump-and-treat, dig-and-treat, andcontainment technologies, PRB technology has many advantages,by which reactive media filled can adsorb, degrade, and/or pre-cipitate various pollutants as contaminant plumes flow throughthe subsurface treatment wall. Therefore, it has been regarded tobe technically attractive and cost effective [5,6]. Zero-valent iron,metallic iron (Fe0), as the reactive material, has been extensivelyinvestigated to remove heavy metals such as Pb2+, Cu2+, AsO4

3−,CrO4

2−, Ni2+, Zn2+, Cd2+, Ba2+ [7], radionuclides, e.g., TcO4− [8] and

UO22+ [7,9], and to degrade halogenated-hydrocarbon compounds

from contaminated areas. Removal mechanism of uranium by Fe0

was generally described as (i) Fe0 and structural Fe(II) ions reduceU(VI) ions to sparingly soluble UO2, precipitating on the surfaceof iron, (ii) physical and/or chemical adsorption of U(VI) on corro-sion products of iron [7,10], and (iii) probable formation of U(VI)(hydr)oxide precipitate [10].

Zero-valent iron nanoparticles (ZVI-nps) are believed to haveimproved performances because of increased specific surface areaand more reactive sites on the surface. Additionally, nanoscalemetal particles could be applied through direct injection of par-ticle suspensions to contaminated sediments and aquifers insteadof constructing metal walls [11]. Bare ZVI-nps tend to agglomerateinto larger aggregates due to magnetic properties. Incorporation ofnanoparticles into carbon or polymer matrix can prevent the aggre-gation and their susceptible oxidation, therefore activated carbon-,chitosan-, and bentonite-based hybrid materials have been devel-oped [12,13]. Graphene is a new generation of carbon materialand can be regarded as a single layer of graphite. Graphene oxide(GO) can be easily prepared by several classical methods fromcheap natural graphite, introducing oxygen-containing functionalgroups such as carboxyl and hydroxyl groups into carbon sheets.GO has been demonstrated to be a promising adsorbent to removeheavy metals such as uranium [14,15] from aqueous solution. Inthis regard, it is expected that the removal performance of ZVI-nps for uranium could be improved by attaching GO sheets. Jabeenet al. reported the successful synthesis of nanoscale iron-decoratedgraphene sheets and their applications in Cr(VI) removal. Highermaximum sorption capacity and higher reduction ratio of Cr(VI)adsorbed were achieved with the composites compared to that in

bare ZVI-nps [12]. Moreover, Fe/graphene composites were alsoused to decolorize methyl blue solution. Higher removal capacitiesof the composites are due to the increased sorption sites, whichoriginate from the inhibition of the particle aggregation and thereduction of the Fe particle size [16].

In this work, ZVI-np and its graphene composite were preparedsuccessfully and applied in the removal of uranium from aque-ous solution in anoxic atmosphere. The influences of solution pH,the presence of NaHCO3, synthetic groundwater constituents, andhumic acid on removal efficiencies of uranium were investigatedsystematically in order to evaluate the efficacy of Fe0 to remediateuranium-contaminated groundwater. Sorption kinetics and capac-ities were studied as well to clarify reaction mechanisms with thehelp of various analytical techniques, e.g., SEM, XRD, and XAS, ana-lyzing the reacted adsorbents after sorption processes.

2. Experimental

2.1. Reagents

All common chemicals used in this study were purchased fromAladdin (Shanghai, China) and are of analytical grade. A 10 mMuranium stock solution was prepared by dissolving appropriateamount of UO2(NO3)2·6H2O (Sigma–Aldrich Co.) in Milli-Q water(18.2 M� cm, Millipore Co.). Synthesized groundwater consistsof 0.29 mM Ca(NO3)2, 0.31 mM CaBr2, 0.53 mM MgSO4, 0.45 mMNa2SO4, 0.011 mM Na2CO3, 0.60 mM NaHCO3, and 0.43 mM KHCO3[17].

2.2. Synthesis of ZVI-nps, reduced GO(RGO), and Fe/RGOcomposites

Fe0 was obtained according to a NaBH4 reduction method [16]under the protection of N2. Single-layered GO sheets, preparedfrom graphite according to the previously described procedure [14],were used as supporting matrix. In the synthesis of Fe/RGO com-posites (containing ∼50% Fe by mass), a concentrated FeCl3 solution(724 mg FeCl3·6H2O) was dropwise added into 300 mL of 0.5 mg/mLGO solution with ultrasonication. The mixed solution was mag-netically stirred overnight, then NaBH4 powder (406 mg) wasincrementally added into the mixture to simultaneously reduceFe3+ and GO to Fe0 and RGO, respectively [12,16], resulting in ablack homogenous dispersion. After 30 min shaking, Fe/RGO com-posites were recovered by vacuum filtration, and the solid wastotally washed by deoxygenated ethanol prior to the formation ofa liquid meniscus so as to prevent iron rust [18]. Finally, the blacksolid was vacuum-dried at 70 ◦C for 4 h, broken up with a spatula,and stored in a common desiccator. The accurate Fe content wasdetermined to be 47.1% by dissolving the composites in 5% HNO3and measuring Fe2+ quantity. Additional composite materials with∼20% and ∼80% Fe loading were also attempted, and the morphol-ogy and removal performances for uranium are shown in Figs. S1and 2, (Supporting information).

ZVI-nps and RGO were prepared according to the aforemen-tioned method in the absence of GO and FeCl3, respectively. It wasnoted that newly formed ZVI-nps were prone to aggregating incomparison with the Fe/RGO composites.

28 Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33

2.3. Batch experiments

Anoxic experiments were carried out in an Ar-filled glovebox.A small amount of ingressive O2 from outside was monitored tobe around 30 ppm. All aqueous solutions used in sorption experi-ments were prepared in advance and placed in the anoxic gloveboxfor at least 24 h to examine the solution stabilities. Next, aque-ous solution (50 mL) was transferred into a glass bottle, in whichthe adsorbent had been weighted and added. Supernatant pH wasadjusted by negligible amounts of NaOH and HNO3 solution. Theglass bottle was then sealed with a screw cap and the system wasvigorously agitated by a magnetic stirrer to maintain the adsor-bent suspension. Upon completion, portions of the supernatantwere sacrificially filtered by 0.22 �m syringe filters, and the filtratewas acidified by 4% ultrahigh purity HNO3. Finally, concentrationsof uranium, iron, and other metals in the solutions were analyzedby inductively coupled plasma optical emission spectrometer (ICP-OES, Horiba JY2000-2, Japan). Sorption capacity (Q) was definedas Q = (C0 − Ce)V/m, where C0 and Ce denote the initial and equilib-rium concentrations of U(VI), respectively, in aqueous phase; V andm are the volume of the solution and the dry weight of adsorbentsused in the sorption experiments. Experiments were performed induplicates and the error bars associated with data shown in figuresrepresent the standard deviations of the two runs. Oxic experi-ments were performed in open laboratory using the screwed glassbottles in consideration that it is more useful for ex situ treatmentof contaminated water and/or industrial effluent. Finally, reactedadsorbents were collected and dried in the vacuum oven for char-acterization analyzes.

2.4. Characterizations of as-prepared materials and reactedadsorbents

The microcosmic morphology of adsorbents before and aftersorption was observed by scanning electron microscopy (SEM, S-4800, Hitachi) at an accelerating voltage of 10 kV. Nitrogen BETsurface analysis was performed using a Micromeritics ASAP 2020.X-ray diffraction (XRD) was carried out on a Bruker D8 Advance X-ray diffraction instrument (Cu K�, � = 1.5406 Å) with a step size of0.02◦, and the diffraction angel (2�) from 5◦ to 90◦ was scanned. X-ray absorption near edge spectroscopy (XANES) of U LIII-edge wascollected at the beamline 1W1B of Beijing Synchrotron RadiationFacility. Detailed experimental parameters and data treatmentshave been described previously [14].

3. Results and discussion

3.1. Characterizations of as-prepared adsorbent materials

SEM and XRD results of ZVI-nps, RGO, and Fe/RGO composites(47.1% Fe0, below omitted) are shown in Fig. 1. ZVI-nps present aspherical shape, aggregating into chain-like structures mainly dueto magnetic property, and RGO presents a layered morphology. Inthe composites, more discrete iron nanoparticles disperse on RGOsheets, indicating an effective inhibition of particle aggregation.The particle size ranges from 30 to 100 nm. As for the XRD pat-tern of ZVI-nps, a dominant diffraction peak located at 44.7–44.8◦

is the characteristic of poorly crystalline/amorphous metallic iron[19,20]. In the case of RGO, a characteristic diffraction peak of GO at10.9◦ remains, besides a broad diffraction peak (002) of graphite at∼23◦ and a sharp peak at 44◦, suggesting incomplete reduction ofGO [21]. XRD results of Fe/RGO composites reveal the presence ofFe0 and that the orderly layered stacking of RGO sheets can be pre-vented effectively. BET specific surface areas of ZVI-nps, RGO, and

Fig. 1. SEM photographs and XRD patterns of freshly prepared ZVI-nps (a), RGO (b),and Fe/RGO composites (c).

Fe/RGO were determined to be 9.7, 167.8, and 218.9 m2/g, respec-tively, highlighting the advantage of the composites.

It is known that in anoxic atmosphere, Fe0 corrodes accordingto the equation of Fe0 + 2H2O → Fe2+ + H2(g) + 2OH−. The pH-dependent iron dissolution was studied in the current sorptionsystem and found that at initial pH (pH0) larger than 3.6, the amountof Fe2+ released was less than 0.5% (Fig. S3, Supporting Information).

3.2. Removal kinetics of uranium by ZVI-nps and Fe/RGOcomposites

One practical requirement for reactive materials filled into PRBis that the kinetics of immobilization reactions must be fast. Aslow reaction rate would require a prohibitively thick barrier toextend the contaminant’s residence time [6]. Removal rates of ura-nium by ZVI-nps and Fe/RGO composites with the same iron dosagewere therefore examined systematically, and residual uranium (%)in solution as a function of contact time is plotted in Fig. 2a. Itcan be seen that at C0(U) = 24 ppm, the concentration of uraniumremaining in solution decreased rapidly, and was below the detec-tion limit of the instrument within 20 min. At C0(U) = 333 ppm, ittook almost 40 min for the complete removal of uranium by ZVI-nps. Such rapid kinetics suggests that even in column or in situbarriers, the primary reduction process occurs upon first contactof the adsorbents with U(VI) species. To distinguish the kinetics ofbare Fe0 and the composites, sorption systems were shaken gen-tly by hand. It was found that for ZVI-nps and Fe/RGO, it required60 and 40 min, respectively, to clean up the 24 ppm uranium solu-tion, showing that the removal rate by Fe/RGO is faster than thatby ZVI-nps. This could be attributed to the following aspects: i) theincreased surface area and reaction sites of Fe0 after combinationwith graphene; ii) residual oxygen-containing functional groups onthe surface of RGO, which can capture uranium rapidly; and iii) theimproved suspension stability of the composites.

Recorded pH and Fe2+ concentrations in solutions during sorp-tion processes are presented in Fig. 2b and c, respectively. Uponthe treatment of 24 ppm uranium solution with either ZVI-nps orFe/RGO, solution pH increased to 9.3, and faster reaction rate canlead to faster pH increases. In contrast, the treatment of 333 ppmuranium solution resulted in a final pH value (pHf) of 6.0. Irondissolution was negligible all along at C0(U) = 24 ppm, but Fe2+

concentration in the solution containing ZVI-nps and 333 ppm

Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33 29

0

25

50

75

100R

esid

ual ura

niu

m (

%)

24-U-Fe 24-U-Fe-hand

24-U-Fe/RGO 24-U-Fe/RGO-hand

333-U-Fe 24-U-Fe-oxic

333-U-Fe-oxic

a

4

6

8

10

pH

b

0 100 200 300 400 1000 1500

0

20

40

60

Fe

2+ r

ele

ase (

ppm

)

Contact time (min)

c

Fig. 2. (a) Removal kinetics of uranium by ZVI-nps and Fe/RGO in pure Milli-Q waterunder anoxic and oxic conditions; corresponding pH (b) and Fe2+ concentrations inaqueous phase (c) as a function of contact time. mZVI−nps = 4.0 mg, mFe/RGO = 8.5 mg,V = 50 mL, C0(U) = 24 or 333 ppm, and pH0 5.0.

uranium approached 44.4 ppm and maintained stable over theremainder of the sorption period.

In oxic atmosphere, uranium removal rates by ZVI-nps weremuch slower than those of corresponding anoxic groups, and therewas a re-dissolution of already adsorbed uranium. Dissolved oxy-gen has been reported to be able to re-oxidize bioreduced U(IV)to the soluble U(VI) species within several hours to days [22], andwas likely responsible for the re-dissolution here. Iron corrosion inthe presence of O2 was written as 2Fe0 + 2H2O + O2 → 2Fe2+ + 4OH−

and Fe2+ + 5/2H2O + 1/4O2 → Fe(OH)3(s) + 2H+. At both C0(U) = 24and 333 ppm, iron dissolution arose, followed by gradual decreasesin Fe2+ concentration. Crane et al. [20] reported a similar tendencyand assigned the fall after a rise to the formation of Fe(OH)3 withthe slow ingress of atmospheric oxygen.

Additionally, the removal rate of uranium by ZVI-nps andFe/RGO composites from 24 ppm uranium-containing syntheticgroundwater (pH0 7.9) was also examined and found to be veryfast as well (Fig. S4, Supporting Information).

3.3. Influences of several environmental factors on uraniumremoval by ZVI-nps

It’s well known that CO32− occurs ubiquitously in natural or

contaminated groundwater [23], and it can form stable com-plexes with uranium especially under neutral-basic conditions. Inreal groundwater, ternary calcium–uranyl–carbonato complexes,i.e., Ca2UO2(CO3)3 and CaUO2(CO3)3

2−, are prevalent, which areless efficient electron acceptors than uranium hydroxyl species[23,24]. Humic acid, a representative of natural organic matters ingroundwater, possesses organic functional groups such as carbonyl,

2 3 4 5 6 7 8 9 10

0

50

100

150

200

250

300

350

MilliQ-water

NaHCO3

Mimic groundwater

Humic acid

Fe2+

release in water

(right axis)

pH0

Sorp

tion c

apacity (

mg U

/g)

0

25

50

75

100

Fe

2+ re

lease (%

)

Fig. 3. Left axis: influence of solution pH on the uranium removal by ZVI-nps in Milli-Q water, 1.0 mM NaHCO3, mimic groundwater, and 10 ppm humic acid solutions,respectively; right axis: influence of solution pH on the Fe2+ release in the Milli-Qwater system. mZVI−nps = 4.0 mg, V = 50 mL, C0(U) = 24 ppm, and contact time = 24 h.

carboxylate, phenol, and hydroxyl groups, which are capable of cap-turing uranium. Recently, there have been several reports aboutreduction ability of humic acid [25,26]. Therefore, the effectivenessof ZVI-nps for the removal of uranium was investigated in 1.0 mMNaHCO3, synthetic groundwater, and 10 ppm humic acid matrices,and typical pH values (5–9) of groundwater were selected to studypH influences.

Sorption capacities of ZVI-nps for uranium are plotted in Fig. 3as a function of pH0. In pure Milli-Q water, uranium sorption onZVI-nps was recorded to be ∼300 mg U/g, corresponding to 100%removal of uranium, independent of solution pH0, except of theextremely acidic condition (pH0 2.1). Under that condition, 92.9%iron dissolved and little uranium was removed, in contrast at pH04.0–9.0, Fe2+ releases into solution were negligible. It was alsofound that the presence of NaHCO3, synthetic groundwater con-stituents, and humic acid could not affect the efficient removal ofuranium from water. Yan et al. [24] once reported the inhibitiveeffect of higher pH, the presence of HCO3

−, and the utilizationof the synthetic groundwater on the removal rate of uranium byZVI-nps. It is easily understood because the formation of either car-bonate or calcium–uranyl–carbonato complexes can stabilize U(VI)in aqueous phase, therefore suppressing U(VI) sorption and reduc-tion. However, complete removal of uranium could be expectedfrom a thermodynamics point of view since continuous liberationof UO2

2+ from U(VI) complexes occurred during the treatments, e.g.,Fe0(s) + UO2(CO3)3

4− + 6H+ → UO2(s) + Fe2+ + 3CO2(g) + 3H2O [20].Accordingly, it is concluded that ZVI-nps are promising candidatesfor the effective remediation of uranium-contaminated groundwa-ter. Additionally, the uranium-containing humic acid solution wasstable over the timescale considered, suggesting that humic acidwas incapable of reducing UO2

2+ under current experimental con-ditions. Czerwinski et al. also reported that no reduction of U(VI)occurred up to 40 days in the presence of aquatic humic acid [27].

It is interesting to note that unlike the Milli-Q water andhumic acid systems, the Fe0-treated NaHCO3 solution and syn-thetic groundwater presented blue-green, the characteristic colorof green rust. Green rust is a layered Fe(II)/Fe(III) hydroxide withhydrated anions such as CO3

2− and SO42− located in interlayers.

A green rust mineral morphology, pseudo-hexagonal form, wasthen observed in SEM of the reacted ZVI-nps (Fig. S5, Support-ing Information). Actually, green rust has been reported to be thecorrosion product of iron in an in situ remediation of radionuclide-contaminated groundwater [5].

30 Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33

3.4. Sorption capacities of ZVI-nps, RGO, and Fe/RGO composites

Maximum sorption capacities of ZVI-nps, RGO, and Fe/RGO com-posites for uranium were determined in Milli-Q water containing0–714 ppm uranium (pH0 5.0). The results are plotted in Fig. 4aas a function of initial uranium concentration. It was found that:(i) with the increase in C0(U), removal efficiencies of uranium byZVI-nps kept at 100% until C0(U) = 643 ppm, and decreased to 87% atC0(U) = 714 ppm, corresponding to the maximum sorption capacityof 8173 mg U/g; (ii) the sorption capacity of RGO increased gradu-ally with the increase in C0(U), attaining a saturation of 341 mg U/g,which is far smaller than that of ZVI-nps, but comparable to thepreviously reported value of this material [14]; (iii) the sorptiontendency of uranium on Fe/RGO was similar to that on ZVI-nps,achieving an equilibrium of 4174 mg U/g. This value is reason-able in light of the Fe0 percentage in composites. In the case of100 mg ZVI-nps, complete removal of uranium was achieved evenat C0(U) = 714 ppm. This agrees with the description that metalliciron has infinite sorption capacity for uranium so long as Fe0 is suf-ficient in the system, maintaining a favorable reducing condition.Gu et al. also observed this ‘infinite’ removal efficiency: interac-tions between UO2

2+ (0 − 1.8 × 104 ppm) and Fe0 (2.0 × 104 mg/L)resulted in 100% removal of uranium from solution (pH0 5.0). Thesolution pH increased from 5 to 10 in less than 30 min [28].

After the treatment with ZVI-nps, pHf increased to 9.2 and 5.4 atC0(U) = 24 and 714 ppm, respectively, as labeled in the figure. Fig. 4bgives Fe2+ releases from the iron-bearing materials as a function ofC0(U). In the case of 4.0 mg ZVI-nps, Fe2+ releases were not observedat C0(U) = 0 − 71 ppm; with the continuous increase in C0(U), Fe0

started to dissolve and Fe2+ concentration increased, the recordedmaximum dissolution was 58.8% of iron. The case of Fe/RGO is sim-ilar to that of ZVI-nps, and a more apparent dissolution stop (61.0%of iron) was concurrent with the saturation sorption of Fe/RGO for

0

2000

4000

6000

8000pH

f=6.6, 5.4

pHf=5.3, 5.4, 5.4

70% removal

pHf=9.2

Fe

Fe-100mg

Fe/RGO

RGO

Fe-Oxic

Sorp

tion c

apacity (

mg U

/g)

87%

removal

a

0 100 200 300 400 500 600 700 800

0

10

20

30

40

50

60

Fe

Fe-100mg

Fe/RGO

Fe-Oxic

Fe

2+ r

ele

ase (

ppm

)

C0(U) / ppm

b

Fig. 4. (a) Sorption capacities of ZVI-nps, Fe/RGO, and RGO for uranium as a func-tion of C0(U); (b) corresponding Fe2+ concentrations in solutions. m = 4.0 or 100 mg(indicated in figure), V = 50 mL, pH0 5.0, contact time = 24 h.

Fig. 5. SEM photographs of ZVI-nps prepared in a blank experiment (a), ZVI-nps (b) and Fe/RGO (c) treated at C0(U) = 24 ppm, ZVI-nps (d) and Fe/RGO (e) treated atC0(U) = 333 ppm, ZVI-nps (f) treated at C0(U) = 714 ppm, and reacted ZVI-nps obtained under oxic conditions (i); EDX spectra (g and h) for d and e, respectively. mZVI−nps = 4.0 mg,mFe/RGO = 8.5 mg, V = 50 mL, and pH0 5.0.

Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33 31

2θ (degree)

Blank

24-U-Fe

FeOOH (PDF 70-0714)

Maghemite (PDF 39-1346)

U3O

7(PDF 42-1215)

20 30 40 50 60 70 80 90

Fe-Oxic

714-U-Fe

333-U-Fe/RGO

333-U-Fe

238-U-Fe

24-U-Fe/RGO

Fig. 6. XRD patterns of the reacted adsorbents.

uranium. As for 100 mg ZVI-nps, pHf values were recorded to belarger than 9.0 and Fe2+ releases were all negligible.

Under oxic conditions, uranium removal by ZVI-nps was largelydecreased with the maximum sorption capacity of 1354 mg U/g.Fe2+ releases were inhibited as well.

3.5. Characterizations of adsorbents after uranium sorption

3.5.1. SEM and XRDSEM photographs and XRD patterns of reacted adsorbents are

shown in Figs. 5 and 6, respectively. A blank experiment withouturanium addition yielded blank ZVI-nps, which appearance (Fig. 5a)is not significantly different from the freshly prepared one and XRDanalyzes revealed that metallic iron is the predominant phase still.

After the treatment of 24 ppm uranium solution, bare ZVI-nps (Fig. 5b) and the nanoparticles decorated on RGO sheets(Fig. 5c) grew larger and surfaces became rougher, which could beattributed to a combination precipitation of corrosion products ofiron and U(IV) (hydr)oxide on iron surface [10]. Magnetite (Fe3O4)and/or maghemite (�-Fe2O3) was identified as the predominantcorrosion product of iron by XRD. XRD technique is unable to dis-tinguish Fe3O4 and �-Fe2O3 because structural Fe(II) ions locatedat octahedral sites of the oxide can be reversely oxidized andreduced in the same crystalline structure. During uranium removal,Fe(II)/Fe(III) mixture oxide was generally reported as the oxidativeproduct of iron [10,29,30]. The presence of Fe(III) ions confirmedthat structural Fe(II) ions can reduce U(VI) further [29,31,32]. Thereflection peak of residual Fe0 is obvious in XRD of treated ZVI-nps, while it is very weak for Fe/RGO composites, implying that Fe0

incorporated into the composites is more reactive. Unfortunately,there is no uranium-related diffraction peak, probably due to smallamounts of deposited uranium phase.

17140 17160 17180 17200 17220

0.0

0.5

1.0

1.5

UO2

UO2(OH)

2

24-U-Fe 238-U-Fe

333-U-Fe 333-U-Fe/RGO

714-U-Fe Fe-Oxic

µ(E

)

E (eV)

Fig. 7. XANES of the reacted adsorbents.

At C0(U) = 238 ppm, the presence of U3O7 (UO2.3) in reacted ZVI-nps was verified by the XRD analysis, illustrating a partial reductionof U(VI) removed by ZVI-nps. Iron oxidation to Fe3O4/�-Fe2O3 wasaccompanied. The partially reductive precipitation of U(VI) on thesurface of iron has been addressed: Scott et al. identified uraniumspecies associated with mild steel surface as U4O7 based on XPSresults [31]; Riba et al. observed a broad peak at 2� = 28◦ in XRD,which was assigned to UO2 although U(VI) was found invariablyfrom XPS [10].

After the treatment of 333 ppm uranium solution, not only ironparticles but also RGO sheets (Fig. 5d and e) were covered withunknown substance. The SEM coupled with energy dispersive X-ray (EDX) analyzes (Fig. 5g and h) revealed the presence of greatamounts of uranium. The treated ZVI-nps and Fe/RGO compositesare somewhat XRD amorphous, but it seems like that Fe/RGO is richin U3O7. This might be reasonable in consideration of the increasedreactivity of iron nanoparticles decorated on RGO sheets. After thetreatment of 714 ppm uranium solution, the particle morphologyof ZVI-nps disappeared (Fig. 5f) and the XRD reflection peaks aremore flattened.

The SEM microphotograph of reacted ZVI-nps obtained in air(Fig. 5i) displays irregularly spherical particles and thin plates,XRD analyzes demonstrate the concurrent occurrence of U3O7and Fe3O4/�-Fe2O3. Chemical reduction of U(VI) by Fe0 has beenreported to be possible in both anoxic and oxic atmosphere, butpromoted at a low oxygen level [20]. If appreciable oxygen is avail-able, uranium removal would be dominated by the sorption of U(VI)on corrosion products of iron [29]. Further analysis suggested thecoexistence of FeOOH, its thin plate form has already been observedin SEM. FeOOH was considered to be generated from Fe(OH)3 dehy-dration and the higher proportion of Fe(III) ions in an oxic systemwas attributed to greater amounts of dissolved O2.

3.5.2. XANES analyzesXANES spectra provide an important fingerprinting method for

judging oxidation states of investigated metal ions by comparingwith known and standard samples [33]. The energy of absorp-tion edge increases with increasing average oxidation state of anabsorbing atom, and in uranium spectra, a resonance shoulder at∼17180 eV is assigned to multi-scattering of axial oxygen atoms inthe uranyl moiety [34,35]. In this work, XANES were performed onthe above reacted adsorbents and the results are plotted in Fig. 7.It can be seen that the near edge absorption curves locate in theright side between those of U(IV)O2 and U(VI)O2(OH)2, illustrat-ing a mixture of U(IV) and U(VI) in reacted adsorbents. For 24 ppmuranium-treated ZVI-nps, the energy of absorption edge and theabsence of the resonance peak suggests that U(IV) is the predomi-nant oxidation state. Absorption edges of 24, 238, 333, and 714 ppm

32 Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33

Table 1(NH4)2CO3 leaching of the reacted adsorbents (V = 10 mL, leaching time = 30 min).

24 ppm U-Fe 24 ppm U-Fe/RGO 333 ppm U-Fe 333 ppm U-Fe/RGO 333 ppm U-Fe-pH0 6.5 714 ppm U-Fe

U(VI) leaching 0.7% 0.3% 34.0% 21.1% 30.4% 47.5%

uranium-treated ZVI-nps are arranged from low to high energy, andresonance features of O U O are increasingly apparent in the sameorder, which implies that the U(VI) proportion of uranium removedby iron increases with the increase in C0(U). The absorption curve of714 ppm uranium-treated ZVI-nps closely follows that of U(VI) pre-cipitate. Additionally, the absorption edge energy of treated Fe/RGOat C0(U) = 333 ppm is a little lower than that of the correspondingZVI-nps, and the Oax peak is more inhibited, confirming the higherU(VI) reduction ratio in Fe/RGO composites.

3.5.3. Desorption experimentsBased on the references [35,36], (NH4)2CO3 (0.1 M) was adopted

to desorb U(VI) ions associated with iron oxides and to dis-solve U(VI) minerals such as metaschoepite, therefore quantifyingU(VI)/U(IV) proportion. From Table 1, it can be seen that atC0(U) = 24 ppm, very small amounts of removed uranium onZVI-nps and Fe/RGO could be leached by the CO3

2− solution,confirming the U(IV) valence predominant in the adsorbents. AtC0(U) = 333 ppm, U(VI) ratio increased largely to 34.0% in reactedZVI-nps; the supporting material, RGO and higher pH0 enhancedU(VI) reduction ratio to some extents. At C0(U) = 714 ppm, nearlyhalf of uranium removed by ZVI-nps was ready to wash out. Themaximum sorption capacity of synthetic magnetite nanoparticleswas only 27 mg U/g [37] and Gu et al. [28] reported that the per-centage of U(VI) associated with iron corrosion products was in therange of 2.77–3.91%. It is therefore not reasonable to assign suchhigh proportion of U(VI) to mere U(VI) adsorption on iron oxides,hydrolysis precipitation of U(VI) on iron surface was thus broughtforward.

3.6. Reaction mechanisms

Important references studying removal mechanisms of Fe0 foruranium are summarized in Table S1 (Supporting Information), andchemical reactions written below are supposed to occur during thetreatment in anoxic atmosphere,

2 > Fe–OH + UO22+ → [(> Fe–O)2UO2]ads + 2H+ (1)

UO22+ + Fe0 + 2H2O → U(IV)(s) + Fe2+ + 4OH− (2)

UO22+ + 2Fe(II)(structural) + 2H2O → U(IV)(s)

+ 2Fe(III)(structural) + 4OH− (3)

Fe0 + 3Fe2O3(s) + H2O → 2Fe3O4(s) + Fe2+ + 2OH− (4)

UO22+ + 2OH− → UO2(OH)2(s) (5)

3.6.1. Reductive precipitation of U(VI) on iron surfaceAccording to Reaction (1), the oxide layers of Fe0 particles

contain hydroxyl groups, which are capable of adsorbing UO22+

via proton exchanges. Adsorbed U(VI) ions will be reduced tothe sparely soluble U(IV) precipitate by Fe0 cores (Reaction (2)).The predominant corrosion product of iron has been identifiedas Fe3O4/�-Fe2O3, revealing that structural Fe(II) ions can reduceU(VI) to U(IV) further (Reaction (3)). Reaction (4) presents the pro-cess that a Fe0 core transfers electrons to the surface Fe(III) oxide,leading to the re-generation of Fe(II) ions. As a result of these reac-tions, solution pH largely increases, leading to a total precipitationof released Fe2+ as Fe(OH)2 (Ksp(Fe(OH)2) = 8.0 × 10−16 [38]) [5].

3.6.2. Hydrolysis precipitation of U(VI)Hydrolysis precipitation of U(VI) on iron surface is described

subsequently, characterized by an inhibition of the pH increaseand releases of significant amounts of Fe2+ ions. The continuousprecipitation of U(IV) and iron (hydr)oxides on surface inevitablyreduces iron reactivity and limits UO2

2+ access to Fe0/Fe3O4(and/or�-Fe2O3) surfaces. At the high pH values, UO2

2+ ions would pre-cipitate out from solution according to Reaction (5), resulting inacidification of the solution, which could initiate the dissolutionof Fe0 cores along surface cracks. The generated OH− ions precipi-tate UO2

2+ as UO2(OH)2 continuously (Ksp(UO2(OH)2) = 3.5 × 10−23

[39]), accompanied with Fe2+ releases. Since UO22+ reduction by

aqueous Fe2+ ions is a rather slow process [32], Fe0 plays a pre-cipitant role at this time virtually. When Fe0 cores are depleted,the hydrolysis reaction of UO2

2+ will stop, and the so-called max-imum sorption of Fe0 for uranium will be established, just likethe data shown in Fig. 4. Riba et al. confirmed the occurrence ofmetaschoepite (UO3·2H2O) in treated ZVI-nps with 850 ppm ura-nium solution (pH 5.0–6.0) by XRD, and observed a simultaneousincrease in Fe2+ concentration [10]. Zhang et al. [19] reported asimilar phenomenon when using ZVI-nps to sequester Pb2+. Leadremoval was well correlated with a rise of Fe2+ concentration insolution. Over 90% of deposited lead was found to exist as Pb(II) ionsrather than Pb0 from XPS analyzes, and Fe0 cores were depleted,leaving behind empty iron oxide shells observed by TEM. The shellconstituent of Fe(OH)3 was finally assigned to dissolve Fe0, leadingto the formation of Pb(OH)2 precipitate.

Under oxic conditions, more Fe3+ is present and it’s known thatFe3+ ions more tend to precipitate out (Ksp(Fe(OH)3) = 4.0 × 10−38

[40]), therefore UO22+ hydrolysis could be prevented effectively.

This viewpoint can be supported by the characterization results ofreacted ZVI-nps obtained in air (Section 3.5).

3.6.3. Influence of supporting materials and higher pH0 on thepathway of uranium immobilization

At C0(U) = 333 ppm, Fe2+ releases accounted for 33.9% and 14.2%of iron in ZVI-nps and Fe/RGO composites, respectively, with pHfvalues of 6.3 and 6.5. In view of the above discussed correlationbetween the speciation of adsorbed uranium and pH changes, Fe2+

releases, U(IV) proportion among uranium adsorbed on Fe/RGOis expectedly higher than that on bare ZVI-nps, which has beenaffirmed by XRD, XANES, and the desorption experiments. Recently,Sheng et al. [13] observed that the reduction of U(VI) to U(IV) wasenhanced by using Na-bentonite as the support of ZVI-nps. Higherreduction ratio (60%) of Cr(VI) was achieved in Fe/RGO compared tothe 45% value in bare ZVI-nps. The authors held that iron decoratedgraphene composites have smaller size of iron nanoparticles andhigher surface area, which could increase catalytic and adsorptionsites of Cr(VI) [12]. These results prove that introducing a support-ing material can not only enhance sorption capacities, improveremoval rate but also increase reductive transformation of redoxactive heavy metal ions into less toxic precipitates.

After the treatment of 333 ppm uranium solution (pH0 6.5), thepercentage of iron dissolution decreased to 8.7% of ZVI-nps andpHf was 6.7. Higher reduction of U(VI) to U(IV) was demonstratedas well in the desorption experiments. pH0 higher than 6.8 wasreported to be a benefit for the reductive precipitation of U(VI)[7,24]. Riba et al. [10] addressed that after 12 min reaction, 30%more UO2 was formed on iron surface at pH0 6.5 and 7.0 in compar-ison with pH0 5.0, 5.5, and 6.0 based on XPS analyzes. The authors

Z.-J. Li et al. / Journal of Hazardous Materials 290 (2015) 26–33 33

attributed this to the proportion of (FeIIIOFeIIOH)0, a more effi-cient electron donor over the other surface species of (FeIIIOFeII)+,increasing with alkalinity [41].

4. Conclusions

ZVI-nps are very effective reactive materials for PRB techniqueto remediate uranium-contaminated subsurface area. Rapid reduc-tive precipitation of U(VI) to the sparingly soluble U(IV) speciesis responsible for uranium immobilization. However, corrosionproducts of iron and uranium precipitate formed during watertreatment decrease the reactivity of iron surface gradually. Therole of Fe0 changes from a reductant to a precipitant, leading toa formation of U(VI) (hydr)oxide and a release of Fe2+ ions. Thisis not favorable for long-term immobilization of uranium in geo-chemical environment. Composite materials such as Fe/RGO coulddecrease the particle size of iron nanoparticles and prevent theiraggregation, thus largely increasing specific surface area for reac-tion with uranium. Iron is therefore more fully utilized and U(VI)reduction ratio is enhanced. Additionally, neutral-basic groundwa-ter (pH0 6.5) was considered to benefit the reduction pathway ofU(VI) over hydrolysis precipitation. In extremely acidic solutionsuch as uranium-containing acid mine water, iron nanoparticlesdecorated on RGO sheets are much easier to be dissolved intosolution, decreasing the effectiveness of uranium removal (Fig. S6,Supporting Information). Therefore, ZVI-nps are relatively moreappropriate to such conditions.

Acknowledgements

This work was supported by National Natural Science Founda-tion of China (Grants 11205169, 91326202, 21261140335) and the“Strategic Priority Research program” of the Chinese Academy ofSciences (Grants.XDA030104). The financial support from the StateKey Laboratory of NBC Protection for Civilian (No.SKLNB2014-12)is also acknowledged.

Appendix A. Supplementary data

Supplementary data associated with this article can be found,in the online version, at http://dx.doi.org/10.1016/j.jhazmat.2015.02.028.

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