Remediation of hexavalent chromium from

Applied Surface Science 357 (2015) 1244–1250

Contents lists available at ScienceDirect

Applied Surface Science
journal homepage: www.elsevier.com/locate/apsusc

Remediation of hexavalent chromium from aqueous solution using
clay mineral Fe(II)–montmorillonite: Encompassing anion exclusion
impact
Mirle Vinuth a , Halehatty Seethya Bhojya Naik a,∗ , Jayappa Manjanna b
a
b

Department of Industrial Chemistry, Kuvempu University, Shankaraghatta 577451, India
Department of Chemistry, Rani Channamma University, PB NH-4, Belagavi 591156, India

a r t i c l e

i n f o

Article history:
Received 16 May 2015
Received in revised form
14 September 2015
Accepted 19 September 2015
Available online 25 September 2015
Keywords:
Hexavalent chromium
Remediation by reduction
Fe(II)–montmorillonite
Anion exclusion impact

a b s t r a c t
We have explored the highly efficient and environmentally benign clay mineral, Fe(II)–montmorillonite,
for the reduction of Cr(VI) in aqueous solution. Fe(II)–Mt was treated with K2 Cr2 O7 solution at different
pH, temperature and solid-to-liquid ratio. The [Cr2 O7 ]2− was estimated by UV–vis spectra with a correction for anion exclusion impact. In general, the Cr(VI) reduction was rapid at acidic pH and increased with
temperature up to 50 ◦ C. A complete reduction occurred in about 5 min at pH 3–5. The time taken for
complete reduction at 0 ◦ C, RT (30 ◦ C) and 40 ◦ C are 12 min, 8 min and 5 min, respectively. The reduction followed by immobilization of Cr(III) on the spent clay mineral was well characterized by EDX
and chemical extraction analysis. This remediation process could be easily scaled-up for real system
applications.

© 2015 Elsevier B.V. All rights reserved.

1. Introduction
Chromium is an extensively used in various industries such as
steel, paint, leather and ceramics. It exits at high concentration
in the effluents of electroplating, chromium tanning and paper
industries [1]. The hexavalent state of chromium is a well known
carcinogenic element which is highly toxic, soluble and mobile; this
was commonly found in soil and wastewater released from various
industries.
The Cr(VI) is highly toxic to humans, animals, plants and
microorganisms and is associated with the development of various chronic health diseases including organ damage, dermatitis
and respiratory impairment [2]. It is well known that Cr(VI) is more
toxic than Cr(III) as it leads to cancer and kidney damage because
of its high oxidizing potential, and it can easily penetrate biological membranes [3]. Given the potential magnitude of the problem,
it is obvious that Cr(VI) contamination of surface or groundwater
possess a significant threat to human health and the environment
[4].
The remediation by reduction, Cr(VI) → Cr(III), is the potentially
useful process to clean up the contaminated sites because Cr(III) is

less toxic and can be immobilized with solid phase [5] and became

∗ Corresponding author.
E-mail address: hsb naik@rediffmail.com (H.S. Bhojya Naik).
http://dx.doi.org/10.1016/j.apsusc.2015.09.167
0169-4332/© 2015 Elsevier B.V. All rights reserved.

bioavailable to microorganisms and plants. Accordingly, Fe(II) containing oxide surfaces like Fe3 O4 , mixed ferrites, etc. are commonly
used for the reduction/immobilization. Recently, the biogeochemical transformation i.e., Cr(VI) reduction by the naturally occurring
bacteria, is also being explored [6–9]. In most of the reduction
processes, the kinetics of Cr(VI) reduction was not only slow but
stoichiometrically inefficient due to the lack of fresh reactive sites
and/or diffusion controlled pathways for reactants [10–15]. In the
literature, H2 S [15,16], SO2 [11], H2 O2 [17], ferrous iron [13,18] are
reported for the chemical reduction of Cr(VI) → Cr(III). The SO2 and
H2 S, themselves show toxicity and create additional environmental problems. On the other hand, ferrous iron and glycerol are not
effective in basic medium [2].
Carbonaceous adsorbents such as activated carbons and oxidized activated carbons are also used for the removal of Cr(VI)
ions [19,20] because they are cheap, corrosion resistant and have
shown enhanced adsorption capacity for heavy metals. However,

the problem in utilizing these materials is their separation by conventional methods such as filtration and centrifugation, which are
time consuming and likely to lose adsorbents in small amounts [20].
Cr(VI) reduction is also reported by using magnetite, Fe3 O4
[21], green rust [22,23], ferrous sulfate–sodium dithionite [24],
granular zero-valent iron, ZVI [25–28] and nanoscale ZVI [29,30].
Although these heterogeneous reductants provide high surface
area for adsorption and precipitation, these are effective only in
lower pH [31] and may not be suitable for real system applications

M. Vinuth et al. / Applied Surface Science 357 (2015) 1244–1250

1245

(Si3.949 Al0.051 )tet O10 (OH)2 nH2 O is used here [47]. Aqueous solution of K2 Cr2 O7 was chosen as the model hexavalent chromium
contaminant. Double distilled water was used throughout this
study. The concentration of Cr(VI) was estimated from its optical
density at max = 350 nm using UV-Vis spectrophotometer.
2.1. Preparation of Fe(II)–Mt

Fig. 1. Schematic diagram of 2:1 dioctahedral smectite clay mineral, where Mn+

indicates the interlayer/exchangeable cation, for e.g., Fe2+ ions in Fe(II)–Mt.

such as water and soil treatments wherein large amounts of these
adsorbents/reductants are required. Therefore, it is imperious to
look for effective reductant and/or adsorbent for Cr(VI) in the wide
range of pH and temperature. Accordingly, we have explored the
use of Fe-based clay mineral for reduction and immobilization of
Cr(VI), and was found to be highly effective and feasible in terms of
stoichiometry and reaction kinetics.
Clay minerals are important class of materials which are readily
available in nature. These are used as very good adsorbents for
toxic elements such as arsenate [32,33], decolouration agents, ion
exchangers, molecular sieves, catalysts and also used in brick manufacturing industries [34,35]. There are few reports on natural and
modified clay minerals used for Cr(VI) reduction [36–41]. Moreover Tunisian clay, El Haria clay and raw/aluminum pillared clays
are used for removal toxic elements like Pb2+ , As(III), Cu2+ and Hg2+
ions in aqueous solution through adsorption processes [42–45].
Eloussaief et al. have investigated the efficiency of three different
clay materials such as raw, acid-activated and aluminum-pillared
Tunisian smectite (RSM, ASM, and Al-SM) for the adsorption of
Pb(II), Zn(II) and Cd(II) in single and multi-element systems [46].

From the above experimental results revealed that natural and
modified clay minerals act as effective adsorbent for removal of
toxic elements in aqueous solution.
Montmorillonite (Mt) is a 2:1 dioctahedral smectite group clay
mineral having a layered structure, Fig. 1. The octahedral alumina
sheet is sandwiched between tetrahedral silicate sheets. The negative charge is created on the clay mineral due to the isomorphic
substitution in the octahedral sites (by Mg, Fe, and Ti) and tetrahedral sites (by Al, Fe). Such a permanent negative layer charge is
balanced by exchangeable cations like Ca2+ , Na+ , etc. at the interlayer. Thus, the cation exchange capacity (CEC) of clay mineral
depends on the net elemental composition, which varies with the
geographical availability in the environment.
The properties and uses of this clay mineral can be modified
not only by altering the structural Fe(II)/Fe(III) ratio [36], but also
by replacing the interlayer cations with a variety of inorganic and
organic cations [32,33]. Further, the availability as well as the
amount and/or access of structural Fe(II) for redox reactions is very
limited. Therefore it is rational to make use of Fe(II)–Mt, i.e., the
redox sensitive Fe(II) ions placed in the interlayer of clay mineral
to augment the real field applications. Hence, it is possible to make
use of such an important redox sensitive clay mineral for the reduction of Cr(VI). Thus, in the present study, Fe(II)–Mt is used as an
effective reductant for Cr(VI) in aqueous solution followed by its

immobilization. The reduction reaction was carried out at different pH and temperature by varying the amount of Fe(II)–Mt. Anion
exclusion impact encountered in this study is also investigated.
2. Materials and methods
Na-montmorillonite (Kunipia F, Japan) with a CEC of
about 113 meq/100 g having approximate chemical composition,
(Na0.431 K0.002 Ca0.002 )
(Al1.56 Mg0.305 Fe0.099 Ti0.007 )oct

In the first step, Fe(III)–Mt was obtained by the cation exchange
of above raw clay mineral in 0.4 M FeCl3 solution. Such an Fe(III)–Mt
was treated with ascorbic acid to reduce interlayer Fe(III) to Fe(II)
ions [47]. For comparison, Na(I)– and Ca(II)–Mt were also prepared
by the conventional cation exchange method with 1 M solutions of
NaCl and CaCl2 , respectively.
In order to estimate the interlayer iron, Fe(II)–Mt was subjected for cation exchange with 0.05 M H2 SO4 for about 24 h. The
ratio of ferrous to ferric ions (Fe2+ /Fetotal where Fetotal = Fe2+ + Fe3+ )
released was determined by 1,10-phenanthroline method [48]
using UV-Vis spectrophotometer (max = 510 nm).
The X-ray diffraction pattern (XRD) of the samples were
recorded using D2 phaser XRD (Bruker AXS GmbH, Germany) with

Ni-filtered Cu K␣ radiation,  = 1.5417 nm. Infrared spectra of the
samples were recorded by KBr pellet method using IR analyzer
(FT-IR8600PC, Shimadzu Corporation, Japan). The micrographs of
freshly prepared and spent Fe(II)–Mt was recorded using field
emission scanning electron microscope (Nova NanoSEM 600, FEI
Company, Netherlands) along with energy dispersive X-ray (EDX)
analysis for elemental composition.
2.2. Reduction of Cr(VI) → Cr(III) by Fe(II)–Mt in aqueous solution
In a typical stoichiometric case, freshly prepared Fe(II)–Mt
(0.35 g) was added to the K2 Cr2 O7 solution (1 mM, 100 mL) and kept
stirring magnetically. The reactions were also carried out at different solid-to-liquid ratio, pH (adjusted with dil. HCl and NaOH) and
temperature. The reaction mixture was withdrawn periodically by
using syringe tube and then filtered through 0.2 ␮m membrane filter to remove dispersed clay particles. The decrease in the [Cr(VI)]
concentration was estimated from its optical density. It is important to note that the absorbance values here were corrected for
anion exclusion impact i.e., equivalent to that observed with typical
divalent clay mineral, Ca(II)–Mt (obtained in a separate experiment
with identical conditions). The decrease in K2 Cr2 O7 concentration,
due to reduction of Fe(II)–Mt, was expressed here in terms of %
reduction as function of time.
The spent or oxidized Fe(II)–Mt was separated by centrifugation, washed thoroughly with water and vacuum dried at room

temperature for further characterization using FESEM/EDX and
FT-IR. The adsorbed (immobilized) chromium on spent/oxidized
Fe(II)–Mt was extracted using different reagents viz., 0.05 M H2 SO4 ,
0.5 M (NH)4 C2 O4 , 1 M NaCl, 0.05 M Na2 EDTA. In each case, a known
amount (≈0.2 g) of the spent/oxidized Fe(II)–Mt was dispersed in
50 mL of reagents for 24 h. After the extraction, the supernatant
liquid, filtered through 0.2 ␮m membrane filter, was analyzed for
chromium by inductively coupled plasma optical emission spectroscopy (ICP-OES; PerkinElmer, Optima-7000DV, USA).
3. Results and discussion
3.1. Formation of Fe(II)–Mt
The ratio of ferrous to ferric ions (Fe2+ /Fetotal where
Fetotal = Fe2+ + Fe3+ ) released from the freshly prepared Fe(II)–Mt
on cation exchange with 0.05 M H2 SO4 was found to be close to
unity. This shows that most of the interlayer iron ions are in ferrous state. Also, the CEC of the freshly prepared Fe(II)–Mt was found

1246

M. Vinuth et al. / Applied Surface Science 357 (2015) 1244–1250

(003)

Q

(002)

Q

20
Scattering angle, 2θ

30

3429
3622

2925
2846

1381
1628

3429
3622

2925
2846

1381
1628

1046

692
469

3429
3622

2846
2925

1628

920

1381

1046

520

1.0

0.0
200

250

300

350

400

450

500

Fig. 4. UV–vis absorption spectra of K2 Cr2 O7 solution with and without Ca(II)–Mt
and Fe(II)–Mt [the anion exclusion impact is clearly seen with Ca(II)–Mt].

to adsorbed water and 3429 cm−1 for water present at the interlayer.
3.2. Cr(VI) reduction by Fe(II)–Mt

pure Fe(II)-mont

920

520

469

1046

692

920

520

Na(I)-mont

Spent Fe(II)-mont

692

= (1.19−∆a)

Wavelength, nm

Fig. 2. Powder XRD patterns of Fe(II)–Mt in comparison with Ca(II)– and Na(I)–Mt
at relative humidity of 40%.

469

Actual OD here

0
2

0.5

10

Transmittance (a.u)

1.5

Fe(II)-mont

d001=15.2

500

1

∆a

Abs (-)

(002)

Q

Q

0: Cr(VI) solution (1 mM, 100 ml), pH 5.5
1: after reacting with Ca(II)-mont (0.1 g)
2: after interacting with Fe(II)-mont (0.1 g)

2.0

Na(I)-mont

d001=12.1

(004)

Intensity (a.u)

2.5

Ca(II)-mont

(005)

RH=40%

d001=14.9

1000 1500 2000 2500 3000 3500 4000

Wavenumber (cm-1)
Fig. 3. FT-IR spectra of freshly prepared Fe(II)–Mt compared with Na(I)–Mt as well
as spent/oxidized Fe(II)–Mt.

to be 97 meq/100 g. In addition to the XRD and FT-IR results also
confirmed the formation of Fe(II)–Mt.
The XRD patterns of Fe(II)–Mt in comparison with Ca(II)–Mt and
Na(I)–Mt at relative humidity (RH) of 40% is shown in Fig. 2. The
˚ Ca(II)–Mt,
basal spacing (d0 0 1 ) are as follows: Fe(II)–Mt, 15.2 A;
˚ It is clear that the divalent cation
14.9 A˚ and Na–Mt, 12.1 A.
exchanged clay minerals showed higher d0 0 1 due to larger layer
of hydration when compared to monovalent cation exchanged clay
mineral. These values are in good agreement with the previously
reported values [49,50].
As shown in Fig. 3, the FT-IR spectra of freshly prepared Fe(II)–Mt
is similar to that of Na(I)–Mt. The basic structure of clay mineral
has not undergone any changes. For instance, the bending vibration
bands at ∼520 cm−1 for Si O Al, and 920 cm−1 for Al2 OH are intact.
However the stretching vibrations of Si O group ∼1046 cm−1 are
slightly broadened. The vibration bands at 1628 cm−1 corresponds

3.2.1. Anion exclusion impact (AEI)
It is well-established that the permanent negatively charged layered clay mineral such as montmorillonite exhibit strong anion
exclusion impact (AEI). These effects have an important impact
on the adsorption and diffusion of anion. The diffusion of Cl−
ions in compacted montmorillonite has been studied for better
understanding and modeling of engineered barrier system for the
geological disposal of nuclear waste [51]. In view of this, it has been
suggested to correct the AEI for Cr2 O7 2− ions to arrive at the proper
reduction level of hexavalent chromium by Fe(II)–Mt. Hence, we
treated a typical divalent cation-exchanged clay mineral, Ca(II)–Mt
with K2 Cr2 O7 solution under identical condition before subjecting
the actual clay mineral, Fe(II)–Mt. As shown in Fig. 4, the overall
absorption value of K2 Cr2 O7 solution has increased significantly in
presence of Ca(II)–Mt (OD:1) when compared to absorption spectra of K2 Cr2 O7 solution alone (OD:0). In the presence of Fe(II)–Mt,
although a similar AEI is applicable, we could see a decrease in
absorbance value (OD:2) of K2 Cr2 O7 solution due to redox reaction (Cr6+ + 3Fe2+ → Cr3+ + 3Fe3+ ). However, the absorbance value
here has been influenced by the AEI. In order to obtain the actual
decrease in absorbance value with Fe(II)–Mt, we must subtract by
a value which is equal to that enhanced value observed in case
of Ca(II)–Mt. It is important to note that the AEI varied with different parameters viz., contact time, pH, solid-to-liquid ratio and
temperature. Therefore here, all the absorption values have been
corrected using the corresponding OD:1 and OD:2 values while the
OD:0 remained almost the same.
3.2.2. Effect of stoichiometry on redox reaction
Fig. 5(a–c) shows the % reduction of [Cr2 O7 ]2− at different pH as
a function of time for different solid-to-liquid ratio viz., (a) oxidant
and reductant are in stoichiometric amounts, (b) reductant is in
excess and (c) oxidant in excess. In general, [Cr2 O7 ]2− reduction
by Fe(II)–Mt is a rapid process. In each case, we see a two-stage
reduction of Cr(VI) by Fe(II)–Mt: A rapid first stage followed by a
slow second stage. A complete reduction occurred in about 5 min
at pH 3–5 when stoichiometric amount of Fe(II)–Mt was present.
At neutral pH and above, the reduction was relatively slow. For
instance, at pH 8 there was about 80% reduction within 15 min and
thereafter a gradual reduction has occurred.

M. Vinuth et al. / Applied Surface Science 357 (2015) 1244–1250

100
Reductant & oxidant are in stoichiometric amounts
2i.e. [Fe(II)-mont] : [Cr2O7 ]

80

0.35 g

: 1 mM, 100 mL

60

pH 3
pH 5
pH 6
pH 8

40
20

(a)
60

0 °C
30 °C
40 °C
50 °C

40
20

0

0

0

5 10 15 20 25 120
Time (min)

150

0

180

2

: 1 mM, 100 mL

60
40

pH 5
pH 6

20

80

Cr(VI) reduction (%)

(b)

1

2

10

12

14

16

3
Time (min)

4

5

: 1 mM, 100 mL

(b)

40
20

0 °C
30 °C

0

6

100

0.5 g

60

0
0

8

Reductant is in excess @ pH = 5
2i.e. [Fe(II)-mont] : [Cr2O7 ]

2i.e. [Fe(II)-mont] : [Cr2O7 ]
0.5 g

6

100

Reductant is in excess

80

4

Time (min)

100

Cr(VI) reduction (%)

Stoichiometric amounts of reductant & oxidant @ pH = 5
2i.e. [Fe(II)-mont] : [Cr2O7 ]
0.35 g : 1 mM, 100 mL

80

(a)

Cr(VI) reduction (%)

Cr(VI) reduction (%)

100

0

1

2
3
Time (min)

4

5

100
Oxidant is in excess

0.2 g

: 1 mM, 100 mL

Cr(VI) reduction (%)

2i.e. [Fe(II)-mont] : [Cr2O7 ]

80

Cr(VI) reduction (%)

1247

(c)

60
40

pH 5
pH 6

20

Oxidant is in excess @ pH = 5
2i.e. [Fe(II)-mont] : [Cr2O7 ]
0.2 g : 1 mM, 100 mL

80

(c)
60
40

30 °C
0 °C

20
0

0
0

5

10

15

20

25

30

35

40

45

Time (min)
Fig. 5. Reduction of Cr(VI) by (a) stoichiometric amount of Fe(II)–Mt at different
pH. (b) Excess amount of Fe(II)–Mt at pH 5 and 6. (c) Sub-stoichiometric amount of
Fe(II)–Mt at pH 5 and 6.

It is well-known that the pH has a significant effect on the Cr(VI)
reduction. For instance, Xiang-Rong et al. [2] have shown that the
reduction of Cr(VI) by ascorbic acid under acidic pH is faster than in
neutral pH and slower in alkaline pH. A similarly observation was
made for Cr(VI) reduction using magnetite [52]. There was >90%
removal of Cr(VI) by magnetic nanoparticles at pH 2–4 whereas it
was 55% at pH 4–7 and only 40% at pH 7–10 [53]. Using nanoscale
zero valent iron supported on mesoporous silica (nZVI@MCM-41)
[54], a complete reduction was achieved at pH 3 in about 9 h and
it was decreased to 50% at pH 5. But at neutral and higher pH,

0

20

40

60
80
Time (min)

100

120

Fig. 6. Reduction of Cr(VI) at pH 5 by (a) stoichiometric amount of Fe(II)–Mt at
different temperatures. (b) Excess amount of Fe(II)–Mt at 0 ◦ C and RT. (c) Substoichiometric amount of Fe(II)–mont at 0 ◦ C and RT.

there was no reduction of Cr(VI) using nZVI@MCM-41. Bentonitesupported nZVI is also used for removal Cr(VI) from wastewater. At
pH 2, almost complete reduction occurred within one min, but at
pH 8 only 27% reduction was observed even after 20 min [40]. However, in the present study, we could obtain a significant reduction
of Cr(VI) even at pH 8.
Kadu et al. [55] have reported the remediation of Cr(VI) from
simulated water streams using Fe–Ni bimetallic nanoparticles
(Fe–Ni NPs) and their nanocomposites prepared with montmorillonite clay. Batch experiments with a 25 mg L−1 Cr(VI) solution and
2 g L−1 Fe–Ni NPs exhibited complete reduction of Cr(VI) within

1248

M. Vinuth et al. / Applied Surface Science 357 (2015) 1244–1250

10 min that followed first order reaction kinetics. Amongst 25%,
50%, 75% in situ and loaded nanocomposites, 75% compositions
showed better activity with enhanced reduction capacity below pH
4 due to the generation of reactive H• species.
Among the clay minerals used for Cr(VI) reduction, Fe(II)bearing phyllosilicates such as iron-rich montmorillonite, chlorite
and a regularly interstratified chlorite-smectite (corrensite) have
been studied at acidic pH 3 [37]. Chlorite and corrensite, owing to
the their high Fe(II)/Fe(III) ratio, showed rapid reduction of Cr(VI).
The oxidation of structural Fe(II) to Fe(III) was confirmed by Fe Kedge changes in the X-ray absorption spectra. Similarly, iron-rich
clay mineral (ferrous saponite from Deccan region of India) was
shown to reduce Cr(VI) gradually [38]. There are not many reports
on natural or modified clay minerals for Cr(VI) reduction. Further,
the availability as well as the amount and/or access of structural
Fe(II) for redox reactions is very limited. Hence, this study is having
a great significance because Fe(II)–Mt could be prepared in large
amounts to augment the real field applications.
Although there was stoichiometrically less amount of Fe(II)–Mt
(0.2 g ≈ 65%) in Fig. 5c, it is interesting to see a complete reduction of Cr(VI) in about 30 min. This must be due to the difference
in the absorption value while correcting the AEI because the entire
Fe(II)–Mt is consumed (oxidized) within 30 min, thereby ceasing
the AEI value close to zero. There are some reports on the solvent extraction of Cr(VI) with tetra butyl ammonium bromide from
aqueous solution [56] which showed that the efficiency decreased
considerably with increasing pH and ceased to zero at pH ∼ 6. However, in the present study we see the efficient reduction of Cr(VI)
even at near neutral pH.

3.2.3. Effect of temperature on Cr(VI) reduction by Fe(II)–Mt
The reduction of Cr(VI) by Fe(II)–Mt was carried out in different
temperatures (0–50 ◦ C) at pH 5, Fig. 6(a–c). In general, the Cr(VI)
reduction increased with temperature up to 40 ◦ C. The time taken
for complete reduction at 0 ◦ C, RT (30 ◦ C) and 40 ◦ C are 12 min, 8 min
and 5 min, respectively. When there was an excess of Fe(II)–Mt
(Fig. 6b), it took just 3 min for complete Cr(VI) reduction at 0 ◦ C and
RT. However, when Fe(II)–Mt (Fig. 6c) was stoichiometrically less,
at RT it took about 40 min for complete Cr(VI) reduction whereas
at 0 ◦ C, the reduction was about 95% even after 1 h.
Xiang-Rong et al. [2] have reported that the temperature
dependent reduction of Cr(VI) by ascorbic acid in the range of
5–40 ◦ C at pH 7 took 30 min for completion. A significant effect on
the reduction of Cr(VI) was observed when the temperature was
5–25 ◦ C. In vitro studies of Cr(VI) reduction by cell free extracts
of chromate-reducing bacteria have shown the maximum reduction at ambient temperature, 28 ◦ C [57]. Bentonite-supported nZVI

Table 1
Amounts of Fe and Cr released from spent or oxidized Fe(II)–Mt in different reagents.
Extraction reagents

Fe present in (mM)

Cr present in (mM)

0.05 M H2 SO4
0.5 M (NH)4 C2 O4
1 M NaCl
0.05 M Na2 EDTA

3.45
2.42
1.29
1.22

3.65
3.39
2.0
8.51

used for the removal Cr(VI) from waste water [40] showed about
74% reduction at 25 ◦ C and 82% reduction 40 ◦ C. On the contrary,
when Fe3 O4 -stabilized Fe0 nanoparticles were used for reduction
of Cr(VI) from aqueous solution [41], the rate of reduction was high
at lower temperature, for instance, about 90% reduction was seen at
25 ◦ C while it is 79% at 40 ◦ C. However, in the present study, we see
the efficient reduction of Cr(VI) by Fe(II)–Mt in all the temperatures
from 0 to 50 ◦ C.
3.3. Examination of spent or oxidized Fe(II)–Mt
In order to understand the reduction of Cr(VI) and immobilization of Cr(III) on the clay mineral, it is important to examine the
spent or oxidized Fe(II)–Mt. Fig. 7 shows the photograph of a dry
clay mineral Fe(II)–Mt before and after treating with 1 mM K2 Cr2 O7
solution at pH 5. The spent Fe(II)–Mt was analyzed by FESEM/EDX
spectra to observe any morphological changes upon Cr(VI) reduction. As shown in Fig. 8 there was no appreciable change in the
microstructure of Fe(II)–Mt. However, EDX confirmed the presence of immobilized Cr present in the spent/oxidized Fe(II)–Mt. As
revealed by FTIR spectra (Fig. 3), the spent clay mineral is intact
when compared to fresh Fe(II)–Mt in all respects.
The following reactions may be written for the oxidation of
interlayer Fe(II) ions and their precipitation to neutral species (if
any) and the Cr(VI) reduction followed by immobilization as Cr(III)
species: 6 (Fe2+ → Fe3+ + e− ); 3Fe3+ + 9H2 O → 3Fe(OH)3 + 9H+ ;
(Cr2 O7 )2− + 8H+ + 6e− → 2Cr(OH)3 + H2 O.
Also, there is no indication about the presence of additional
Fe–Cr oxide phase probably due to their small fraction. In order to
quantitatively estimate the adsorbed Cr and iron in the spent clay
mineral, Fe(III)–Mt, the chemical extraction was done with different reagents viz., 0.05 M H2 SO4 , 0.5 M (NH)4 C2 O4 , 1 M NaCl, 0.05 M
Na2 EDTA. In each case, a known amount (0.2 g) of the spent clay
mineral was dispersed in 50 mL of reagents for 24 h and the samples were analyzed by ICP-OES. As shown in Table 1, Na2 EDTA was
found to be the effective reagent to extract these metals due to
its chelating ability. The Fe and Cr content were slightly smaller
than the expected values. This is probably due to their existence as

Fig. 7. Photographs showing the color of Fe(II)–Mt before (a) and after (b) interacting with 1 mM K2 Cr2 O7 solution at pH 5.5. (The spent clay mineral was washed thoroughly.)

M. Vinuth et al. / Applied Surface Science 357 (2015) 1244–1250

1249

Fig. 8. FESEM with EDX of (a) fresh Fe(II)–Mt and (b) oxidized/spent Fe(II)–Mt.

oxyhydroxides (FeOOH or CrOOH) which require repeated treatment preferably at elevated temperature for complete extraction.
Although some Fe-containing clay minerals like ferrous saponite
are found in the natural environment [38], the availability as well as
the amount and/or access of structural Fe(II) is very limited for large
scale real system applications. On the other hand, few reduction
processes developed seemed to be restricted only for laboratory use
[58] and not for real system applications. If ferrous sulfate or sodium
sulfite is used as reductants for Cr(VI) → Cr(III), ferric hydroxide and
sulfur dioxide (toxic and highly volatile) will be formed, respectively, as byproducts which are difficult to handle [59]. Hence it is
essential to propose a suitable reductant/adsorbent for field application, especially to treat water and soil. We believe that Fe(II)–Mt
could be the efficient and suitable material to augment the real
filed applications. For this, a slurry of Fe(II)–Mt filled in dialysis
bags could be suspended in the contaminated bodies such as flowing or stagnant waters and/or industrial effluents. In the case of
soil contamination, it must be sufficiently moisturized (wet) before
dispersing the Fe(II)–Mt.
4. Conclusions
We have confirmed one of the potential applications of Fe(II)–Mt
to treat the hazardous Cr(VI) contamination in aqueous solution.

The anion exclusion impact of [Cr2 O7 ]2− ion with the negatively
charged clay mineral here was considered in all our estimations
of Cr(VI) reductions. In general, [Cr2 O7 ]2− reduction by Fe(II)–Mt
is a rapid process, especially under stoichiometric conditions. For
instance, a complete reduction occurred in about 5 min at pH 3–5.
At neutral pH and above, the reduction was relatively slow. The
Cr(VI) reduction increased with temperature up to 40 ◦ C. The time
taken for complete reduction at 0 ◦ C, RT (30 ◦ C), 40 ◦ C are 12 min,
8 min and 5 min, respectively. The immobilization of the reduced
Cr(III) was confirmed from the EDX spectra of spent clay mineral
and chemical extractions, especially in Na2 EDTA.
Acknowledgements
The lead author M. Vinuth wishes to thank Mr. K. Chandrasekhar
for his help with XRD analysis, Pavan S. at Tuv-Sud South Asia Pvt.
Ltd. Bangalore for assistance with ICP-OES analysis and Prof. G.U.
Kulkarni at JNCASR for providing FESEM facility.
References
[1] M. Barkat, D. Nibou, S. Chegrouche, A. Mellah, Chem. Eng. Prog. 48 (2009)
38–47.
[2] X. Xiang-Rong, L. Hua-Bin, G. Ji-Dong, Chemosphere 57 (2004) 609–613.

1250

M. Vinuth et al. / Applied Surface Science 357 (2015) 1244–1250

[3] M.E. Losi, C. Amrhein, W.T.J. Frankenberger, Environ. Contam. Toxicol. 136
(1994) 91–121.
[4] S.L. Friess, Sci. Total Environ. 86 (1989) 109–112.
[5] I.J. Buerge, S.J. Hug, Environ. Sci. Technol. 32 (1998) 2092–2099.
[6] J.M. Chen, O.J. Hao, Environ. Sci. Technol. 28 (1998) 219–251.
[7] A.R. Shakoor, S. Tahseen, R.U. Haq, Folia Biol. 44 (1999) 50–54.
[8] B. Wielinga, M.M. Mizuba, C.M. Hansel, S. Fendorf, Environ. Sci. Technol. 35
(2001) 522–527.
[9] J.L. Nyman, A.B. Cunningham, F.J. Caccavo, R. Gerlach, Biorem. J. 6 (2002)
39–55.
[10] E.I. Onstott, W.S. Gregory, Environ. Sci. Technol. 7 (1973) 333–337.
[11] L.E. Eary, D. Rai, Environ. Sci. Technol. 22 (1988) 972–977.
[12] C.D. Palmer, P.R. Wittbrodt, Environ. Health Perspect. 92 (1991) 25–40.
[13] S.E. Fendorf, G.C. Li, Environ. Sci. Technol. 30 (1996) 1614–1617.
[14] J.C. Seaman, P.M. Bertsch, L. Schwallie, Environ. Sci. Technol. 33 (1999)
938–944.
[15] C. Kim, Q.H. Zhou, B.L. Deng, E.C. Thornton, H.F. Xu, Environ. Sci. Technol. 35
(2001) 2219–2225.
[16] E.C. Thornton, J.E. Amonette, Environ. Sci. Technol. 33 (1999) 4096–4101.
[17] M. Pettine, L. Campanella, F.J. Millero, Environ. Sci. Technol. 36 (2002)
901–907.
[18] D.L. Sedlak, P.G. Chan, Geochim. Cosmochim. Acta 61 (1997) 2185–2192.
[19] B.V. Babu, S. Gupta, Adsorption 14 (2008) 85–92.
[20] L.-H. Zhang, Q. Sun, D.-H. Liu, A.-H. Lu, J. Mater. Chem. A 1 (2013) 9477–9483.
[21] T. Kendelewicz, P. Liu, C.S. Doyle, G.E. Brown Jr., Surf. Sci. 469 (2000) 144–163.
[22] D.L. Bond, S. Fendorf, Environ. Sci. Technol. 37 (2003) 2750–2757.
[23] L.E. Legrand, A. Figuigui, F. Mercier, A. Chausse, Environ. Sci. Technol. 38
(2004) 4587–4595.
[24] R.D. Ludwig, C. Su, T.R. Lee, R.T. Wilkin, S.D. Acree, R.R. Ross, A. Keeley,
Environ. Sci. Technol. 41 (2007) 5299–5305.
[25] M.J. Alowitz, M.M. Scherer, Environ. Sci. Technol. 36 (2002) 299–306.
[26] S.S. Chen, C.Y. Cheng, C.W. Li, P.H. Chai, Y.M. Chang, J. Hazard. Mater. 142
(2007) 362–367.
[27] S.W. Jeen, D.W. Blowes, R.W. Gillham, J. Contam. Hydrol. 95 (2008) 76–91.
[28] B. Flury, J. Frommer, U. Eggenberger, U. Mader, M. Nachtegaal, R. Kretzschmar,
Environ. Sci. Technol. 43 (2009) 6786–6792.
[29] Y. Xu, D. Zhao, Water Res. 41 (2007) 2101–2108.
[30] X.Q. Li, J. Cao, W.X. Zhang, Ind. Eng. Chem. Res. 47 (2008) 2131–2139.
[31] J.-H. Kim, J-H. Kim, K.V. Bokare, E.-J. Kim, Y.-Y. Chan, Y.-S. Chang, J. Nanopart.
Res. 14 (2012) 1010–1022.
[32] T. Grygar, D. Hradil, P. Bezdicka, B. Dousova, L. Capek, O. Schneeweiss, Clays
Clay Miner. 2 (2007) 165–176.

[33] C. Leungo, V. Puccia, M. Avena, J. Hazard. Mater. 2 (2011) 1713–1719.
[34] J.T. Kloprogge, J. Porous Mater. 5 (1998) 5–41.
[35] M. Eloussaief, N. Fakhfakh, A. Sdiri, S. Bouaziz, M. Benzina, Trans. Ind. Ceram.
Soc. 71 (2012) 195–202.
[36] J.W. Stucki, Iron redox processes in smectites, in: F. Bergaya, B.K.G. Theng, G.
Lagaly (Eds.), Handbook of Clay Science, Elsevier, Amsterdam, 2006 (Chapter
8).
[37] M.F. Brigatti, C. Lugli, G. Cibin, A. Marcelli, G. Giuli, E. Paris, A. Mottana, Z. Wu,
Clays Clay Miner. 2 (2000) 272–281.
[38] G. Parthasarathy, B.M. Choudary, B. Sreedhar, A.C. Kunwar, R. Srinivasan, Am.
Mineral. 88 (2003) 1983–1988.
[39] M. Eloussaief, N. Kallel, A. Yaacoubi, M. Benzina, Chem. Eng. J. 168 (2011)
1024–1031.
[40] L.-N. Shi, X. Zhang, Z.-L. Chen, Water Res. 45 (2011) 886–892.
[41] Y. Wu, J. Zhang, Y. Tong, X. Xu, J. Hazard. Mater. 172 (2009) 1640–1645.
[42] M. Eloussaief, W. Hamza, N. Kallel, M. Benzina, Environ. Prog. Sustain. Energy
32 (2013) 229–238.
[43] M. Eloussaief, S. Bouaziz, N. Kallel, M. Benzina, Desalin. Water Treat. (2013)
1–5.
[44] M. Eloussaief, I. Jarraya, M. Benzina, Appl. Clay Sci. 46 (2009) 409–413.
[45] M. Eloussaief, A. Sdiri, M. Benzina, Environ. Sci. Pollut. Res. 20 (2012) 469–479.
[46] M. Eloussaief, M. Benzina, J. Hazard. Mater. 178 (2010) 753–757.
[47] J. Manjanna, Appl. Clay Sci. 42 (2008) 32–38.
[48] B.M. Vinoda, J. Manjanna, Appl. Clay Sci. 97 (2014) 78–83.
[49] J. Manjanna, T. Kozaki, N. Kozai, S. Sato, J. Nucl. Sci. Technol. 44 (2007)
929–932.
[50] N. Kozai, Y. Adachi, S. Kawamura, K. Inada, T. Kozai, S. Sato, H. Ohashi, T.
Ohnuki, T. Banba, J. Nucl. Sci. Technol. 38 (2001) 1141–1143.
[51] L.R. Van loon, M.A. Glaus, W. Muller, Appl. Geochem. 229 (2007) 2536–2552.
[52] Y.H. Thomas, J.T. Samuel, Environ. Sci. Technol. 39 (2005) 4499–4504.
[53] Y. Panga, G. Zeng, L. Tang, Y. Zhang, Y. Liu, X. Lei, Z. Li, J. Zhang, Z. Liu, Y. Xiong,
Chem. Eng. J. 175 (2011) 222–227.
[54] E. Petala, K. Dimos, A. Douvalis, T. Bakas, J. Tucek, R. Zboril, M.A. Karakassides,
J. Hazard. Mater. 261 (2013) 295–306.
[55] B.S. Kadu, Y.D. Sathe, A.B. Ingle, R.C. Chikate, K.R. Patil, C.V. Rode, Appl. Catal. B
104 (2011) 407–414.
[56] P. Venkateswaran, K. Palanivelu, Sep. Purif. Technol. 40 (2004) 279–284.
[57] K.S. Sumit, S. Rakshapal, A. Ashutosh, S. Mangal, K. Alok, Environ. Sci. Pollut.
Res. 20 (2013) 1661–1674.
[58] E.J. Yurkow, J. Hong, S. Min, S. Wang, Environ. Pollut. 117 (2002) 1–3.
[59] E.B.-D. Carlos, L.-L. Violeta, B. Bryan, J. Hazard. Mater. 1 (2012) 223–224.