1
Submitted to Journal of Hazardous Materials – 15/02/2016
Extraction of Radioactive Cesium Using Innovative Functionalized Porous
Materials
Carole Delchet 1,2, Alexei Tokarev2, Xavier Dumail2, Guillaume Toquer1, Yves Barre3,
Yannick Guari,2* Joulia Larionova,2 Agnès Grandjean.1*
1
Institut de Chimie Séparative de Marcoule, UMR5257 CEA-CNRS-UM2-ENSCM, BP17171, 30207
Bagnols sur Cèze, France. Tel.: +33(0)4 66 79 66 22 ; e-mail: agnes.grandjean@cea.fr
2
Institut Charles Gerhardt Montpellier, UMR 5253 CNRS-UM2-ENSCM-UM1, Chimie Moléculaire et
Organisation du Solide, Université Montpellier II, Place E. Bataillon, 34095 Montpellier cedex 5, France.
Tel.: +33(0)4674805; e-mail:joulia.larionova@univ-montp2.fr
3
CEA/DEN/DTCD/SPDE/ Laboratoire Des Procédés Avancés de Décontamination, Centre de Marcoule,
BP17171, 30207 Bagnols sur Cèze, France
Abstract
A new approach to an efficient and a selective extraction of Cs+ ion from water and a
radioactive solution simulating the effluents of Fukushima reactors was developed by
using porous silica or glass –based nanocomposites containing Prussian blue type
nanoparticles Co2+/[Fe(CN)6]3- (CoFC) of 5.7 and 2.8 nm. A particular emphasis is given
on the kinetics of cesium sorption fitted by using the classical reaction order model as
well as a diffusion model in order to better understand the sorption mechanism.
Compared to the amount of CoFC, the sorption capacities of nanaocomposites studied
are higher than the one obtained for bulk materials. These innovative decontaminated
materials have been tested on a effluent simulating Fukushima radioactive seawater.
First experiments were done in sea water enriched with 10-4 mol/L of CsCl (inactive).
Then, materials have been used to remove
137Cs
(30kBq/L) ions from sea water. In all
cases, the distribution factors obtained are higher than 105 (mL/g). Thess results
demonstrate also clearly the strong effect of the particle size on both the sorption
capacity and the distribution coefficient This work demonstrates the potentiality of these
solid adsorbents for cartridge-type applications in high salinity effluents in order to
minimize wastes.
2
Key-Words: Cesium decontamination, Nanoparticles, Prussian Blue, Radioactive effluents
Sea water, Fukushima, porous glass
…
Introduction
Numerous processes from nuclear facilities (fuel processing, power plants,
laboratories, remediation or removal and others) generate important volume of radioactive
effluents which should be treated in order to minimize their impact on environment. Among
those, radioactive cesium isotopes are ones of the most abundant fission products of uranium
which are dangerous for health. Gamma-emitter
134
Cs and
137
Cs with a half-life of 2 and 30
years, respectively, are mainly presented in these fission products. They accumulate in the
food chain and persist in the environment for hundreds of years. Both isotopes are radiotoxic
because they are analogue of potassium and thus may be quickly assimilated in the body. For
these reasons, the problem of selective cesium effluents decontamination attracts a great deal
of attention in the recent years. However, the use of classical inorganic sorbents (such as
manganese oxide, zeolites, iron hydroxide or barium sulfate) usually employed for extraction
of various radioactive elements such as Sr, Co, Ni, or actinides
Adsorption, accepted; (Sepehrian, Yavari et al. 2008)]
[Goettmann, JNM submitted; Merceille,
is inefficient in the case of cesium due to their low
affinity. In the case of Fukushima disaster, one of the major problems now is to rapidly clean
up areas that have been heavily contaminated by radioactivity. Cesium is one of the major
radioactive element present in water (essentially sea water) used for cooling down the
damaged reactor in the first days of the disaster. Using innovative materials able to remove
radioactive Cesium in continuous process (such as column process) and leading to a waste
volume as small as possible, which match well with the classical waste confinement matrix,
such as cement or glass, is then a challenge for the cleanup of the Fukushima site.
The cyano-bridged coordination polymers based on hexacyanometallates and
transition metal ions called also Prussian Blue analogous present a high affinity for a capture
of cesium ion over a wide range of pH and a salinity of solutions due to a selective insertion
3
of Cs+ into the crystalline structures of cyanometallates (Haas 1993; Lehto and Szirtes 1994;
Mimura, Lehto et al. 1997). The pioneering employment of cyano-bridged coordination
polymers for cesium decontamination as a column process has been proposed around ten
years ago by Lehto’s group by using granular potassium cobalt hexacyanoferrate (CsTreat)
similar to K2[CoFe(CN)6] (Harjula, Lehto et al. 2001; Harjula, Lehto et al. 2004). This
coordination polymer is selective to the cesium ion extraction, however due to its low
mechanical hardness and its fine powder shape leading to a slow filtration rate and also a
clogging problem, only a small volume of effluent may be treated. For this reasons, the use of
this bulk compound in column process decontamination is limited. In order to avoid these
problems, the use of several composite materials with mainly silica supports loaded with
cyano-bridged coordination polymers particles has been proposed (Mimura, Kimura et al.
1999; Mardan and Ajaz 2002; Ambashta, Wattal et al. 2003; Chang, Chau et al. 2008;
Sangvanich, Sukwarotwat et al. 2010). The first method concerns the preparation of
K2M[Fe(CN)6]/SiO2 composites (M=Ni, Cu, Co) by successive impregnation of porous silica
(Mimura, Kimura et al. 1999) or polymer modified silica (Loos-Neskovic, Vidal-Madjar et al.
1999; Milonji, Bispo et al. 2002) by bivalent transition metal ions and hexacyanoferrate
precursor. The second one consists in elaboration of composite materials by direct
incorporation of coordination polymer particles (K2(CoFe(CN)6 or K2(NiFe(CN)6)) into the
silica gel(Mardan and Ajaz 2002) or into hydrated zirconia (Sharygin, Muromskiy et al. 2007)
as matrices during the synthesis. However, all these composite materials suffer several
drawbacks: the final composition is not good controlled and thus it is relatively poorly
reproducible. In addition, the cyano-bridged coordination polymer particles are only weakly
linked to the inorganic support and may be removed during the stage of the cesium extraction.
Alternative promising synthetic route developed in the recent few years concerns the
covalent grafting of specific functions on porous silica in order to anchor the cyano-bridged
complexes or cyano-bridged metallic particles into the silica pores or on the silica surface.
Two composite materials containing self-assembled monolayers on mesoporous silica
functionalized with ethylenediammine covalently coordinated to Cu[Fe(CN)6]2- complex in
the first caseSukwarotwat
et al. 2010
and silica (as powder or as films) functionalized with
ethylenediammine tetracetic acide derivative coordinated to Ni[Fe(CN)6] particles have
recently been reported.(Chang, Chau et al. 2008; Sangvanich, Sukwarotwat et al. 2010). In
both cases an increase of “selective” sorption of Cs+ has been obtained in comparison with the
bulk Prussian Blue.
4
In this line of thought, some of us recently use mesoporous silica matrix functionalized
with pyridine groups in order to synthesized cyano-bridged coordination polymer
nanoparticles of various compositions and various sizes ranging from 3 – 6 nm covalently
attached into the silica pores (Folch, Guari et al. 2008). In this system a good control of the
nanoparticles’ composition, nanoparticles’ size and nanoparticles’ quantity inserted into the
silica has been obtained that give an excellent opportunity to investigate the cesium ion
extraction by using this nanocomposite materials. Recently we performed the preliminary
study of a cesium extraction by using such nanocomposites which show a good and a
selective Cs+ sorption in pure water (Grandjean, BARRE et al. déposé). These results
encourage us to extend this approach to the synthesis of cyano-bridged coordination polymer
nanoparticles inserted into the porous glass matrix instead of silica. Even if porous glasses
have smaller surface area than silica, they present better thermal and chemical stability,
mechanical hardness and irradiation damage resistance (Tkachev, Antropova et al. 2004). In
addition, they are available in very flexible forms, for instance as pearls, that permits the
utilization of the obtained nanocomposites for a cesium extraction in column or cartriges
processes. Today, at industrial scale, selective extraction of Cs+ from contaminated effluents
is performed by using co-precipitation of Prussian Blue analogous but, the maximum
extraction capacity is never reached. In addition, this extraction generates a large quantity of
sludge which should be confined and stored thereafter. In the present work we propose an
innovative materials consisting of coordination polymer nanoparticles covalently grafted to a
matrixes specifically designed for column process decontamination. This allows achieving the
maximum extraction capacity and significantly reducing the amount of wastes.
The present manuscript describes the synthesis of both, silica- and glass-based
nanocomposites containing small sized coordination polymer nanoparticles covalently linked
to pores walls of support and a cesium extraction from pure water solution and from a
radioactive solution simulating the effluents of Fukushima reactors. A particular emphasis is
given on the kinetics of cesium sorption fitted by using the classical reaction order model as
well as a diffusion model in order to better understand the sorption mechanism. Adsorption
capacities and distribution coefficient for these materials in pure water and in radioactive sea
water solutions have also been studied to evaluate the potentially of these nanocomposite for
an industrial process.
1. Experimental Part
5
All the chemical reagents are of analytical grade.
Syntheses.
The synthesis of the bulk Prussian Blue analogous Co3[Fe(CN)6]2.H2O, used for comparison
of Cs+ absorption with the nanoparticles inserted into the nanocomposites has been performed
by using previously described procedures from Co(BF4)2 and TBA3[Fe(CN)6] (Folch, Guari et
al. 2008).
Synthesis of functionalized matrices:
Silica matrices: SBA-15 type silica with pore diameters of 10 nm were synthesised by using a
triblock copolymer (P123) as surfactant as previously described (Zhao, Feng et al. 1998),
(Zhao, Huo et al. 1998). Before grafting, silica matrices have been activated at 150 °C in
vacuo over night. The grafting of the organic functionality -(CH2)2C5H4N into the silica pores
was performed by refluxing in toluene under Argon of the pristine silica in the presence of
(CH3O)3Si(CH2)2C5H4N over night (Mercier and Pinnavaia 1997) (Cauvel, Renard et al.
1997; Corriu, Lancelle-Beltran et al. 2002). The resulting powder was washed in toluene and
dried at 90 °C in vacuo. As a result, the hybrid silicas NC5H4(CH2)2SiO1.5/SiO2 (Silica-Py)
were obtained.
Porous glasses: Pristine mesoporous glasses coming from Vitrabio® with pore size of
approximately 30 nm, specific area of 150 m2g-1 and grain sizes from 200 to 500 µm were
used.
The grafting of the organic functionality -(CH2)2C5H4N into the glasses pores was performed
by using the same procedure than in the case of SBA-15 type silica matrix. The hybrid glass
NC5H4(CH2)2SiO1.5/SiO2 (PG-Py) were obtained.
Synthesis of nanocomposite materials Co2+/[Fe(CN)6]3-/matrices-Py:
The intrapore growth of cyano-bridged coordination polymer nanoparticles Co2+/[Fe(CN)6]3into the silica or glass pores was performed by using simplified previously described
procedure (Folch, Guari et al. 2008). Firstly, a pristine matrice as powder (silica or glass) (1g)
was added to 150 mL of a 10-2 mol L-1 methanolic solution of Co(BF4)2 under air. The
mixture was stirred 2 h at room temperature. After filtration, the powder was thoroughly
washed 3 times with methanol. Secondly, the so-obtained powder was added to a 10-2 mol L-1
methanolic solution of [N(C4H9)4]3[Fe(CN)6]; the mixture was stirred again 2 h, the powder
6
was filtered off, thoroughly washed with methanol. Such consecutive treatments with metal
salts and cyanometallate precursors were repeated 5 times. The resulting powder was then
dried in air over night. The elemental analyses of the obtained nanocomposites
Co2+/[Fe(CN)6]3-/Silica-Py (CoFC/Silica-Py) and Co2+/[Fe(CN)6]3-/PG-Py (CoFC/PG-Py)
are given in Table 1.
Cesium sorption experiments:
Cs+ sorption experiments on both nanocomposites were performed by using first nonradioactive Cs+ containing aqueous solutions in order to investigate the sorption kinetic and
non-radioactive and radioactive Cs+ containing aqueous saline solutions (in real conditions).
The saline solution (sea water) contains Na (9.6 g.L-1); Mg (1.28 g.L-1); Ca (0.4 g.L-1); K (0.5
g.L-1) and Sr (0.008 mg.L-1). In order to compare the sorption capacity of nanocomposites, the
analogous bulk compound Co3[Fe(CN)6]2.H2O was used in the same conditions (pure water
and sea water).
A sorption isotherm has to be constructed from data obtained at reaction equilibrium. So
equilibrium studies in pure water were performed before isotherm sorption experiments also
in pure water. All Cs extraction experiments were performed in batch solution under shaking
at room temperature.
Equilibrium studies:
For the equilibrium studies, 10 mg of powder nanocomposite was shaken in a 20 mL of 0.5
mmol.L-1 of CsNO3 solution in dionized water. Experiments were performed at different time
from 5 min to 2 days. Then the powder was separated from the liquid phase phase by filtration
through a 0.2 µm cellulose acetate membrane and the remaining Cs+ concentration of the
supernatant was analysed using ionic chromatography. . To determine the time to reach
equilibrium, the quantity of cesium Qt (mmol.g-1) adsorbed is plot as a function of time. The
quantity adsorbed is defined as following:
Q t  (C 0  C t ).
V
m
(E1)
Where C0 (mmol.L-1) is the initial concentration of caesium and Ct (mmol.L-1) is the remaining
concentration of cesium in solution after the specified time t (min), V (L) is the volume of the
solution, and m (g) is the mass of adsorbent.
Isotherm studies:
7
For adsorption isotherm studies, a first parent solution of 2 mmol.L-1 of Cs+ was prepared by
dissolving of CsNO3 in deionised water. Solutions with different Cs+ concentration, from 0.01
mmol.L-1 to 2 mmol.L-1 were prepared by corresponding diluting the parent solutions with
deionised water. In each experiment 10 mg of sorbent powder was equilibrated in 20 mL of
various solutions for 24 hours, where equilibrium is always reached. Then powder was
separated from liquid phase by filtration through a 0.2 µm cellulose acetate membrane. The
remaining Cs+ concentration of the liquid phase was then analyzed by ionic chromatography.
To establish the adsorption isotherm, the remaining solute concentration of the compound at
equilibrium Ce (mmol.L-1) is compared with the concentration of this compound retained on
solid particles Qe (mmol.g-1). The relation ship Q=f(C) is known as “sorption isotherm”. The
solid concentration at equilibrium is given by this equation:
Qe 
V
Ce  Ci 
m
(E2)
With V being the volume of solution (L), m is the solid mass of the powder used (g), Ci is the
initial concentration of compound in solution (mmol.L-1) and Ce is the equilibrium
concentration of the compound in solution (mmol.L-1).
Sorption experiments with a simulated radioactive sea water:
The third kind of sorption experiment concerns the effect of the salt on the sorption properties
and then a test in a water sea with radioactive 137Cs.
In the one hand, 10-4 mol.L-1 of CsCl is dissolved in sea water. Then 10 mg of each sorbent
powder was equilibrated in 20 mL of this Cs+ enriched sea water for 24 hours. Then powder
was separated from liquid phase by filtration through a 0.2 µm cellulose acetate membrane
and the remaining Cs+ concentration of the liquid phase was then analyzed by ionic
chromatography.
On the other hand, one liter of radioactive solution is realized from sea water and
radioactive source. From this
137
Cs
137
Cs enriched sea water (29 kBq.L-1), 10 mg of each hybrid
materials was shaken with 20 mL of this solution during 24 hours. After that, the solution was
filtered and analysed by scintillation.
Physical Measurements
Thermal analysis was performed with a Setaram Instrument under nitrogen flow with a
heating rate of 5°C/min. IR spectra were recorded on a Perkin Elmer 1600 spectrometer with
a 4 cm-1 resolution. UV-Vis spectra were recorded in KBr disks on a Cary 5E spectrometer.
8
Elemental analyses were performed by the Service Central d'Analyse (CNRS, Vernaison,
France). Nitrogen and carbon content in the samples was determined using LECO instrument.
The samples were heated at 3000°C under oxygen. Nitrogen and carbon were transformed
respectively in NOx and CO and detected by using an IR detector. Powder X-ray diffraction
patterns and Small Angle X-Ray Scan (SAXS) were measured on a Bruker® D8 advanced
Diffractometer in Bragg-Bentano geometry with Ni-filtered Cu-K radiation. The
measurement parameters are: stepsize, 0.02008; counting time, 15 sec. Samples for
Transmission Electron Microscopy (TEM) measurements were prepared using extractive
replicas or ultramicrotomy techniques. Extractive replica preparation technique consists in
depositing on a freshly cleaved mica plate a suspension of the nanocomposite in ethanol.
After evaporation of the ethanol, a carbon film is deposited onto the mica plate. Then, the
latter is immersed in a diluted HF solution. The carbon film is then detached from the mica
plate and floats on the surface of the solution which allows dissolving the silica part of the
nanocomposite keeping the cyano-bridged coordination polymer nanoparticles sticking on the
carbon film. After washing twice the carbon film, it is deposited onto copper grids for TEM
observations. Thus, this technique allows visualizing the cyano-bridged coordination polymer
nanoparticles after the removal of silica. However, TEM pictures obtained using this
technique doesn’t allow to give an idea of the nanoparticles’ organization or their degree of
aggregation in the matrix. Ultramicrotomy technique consists in suspending the material in a
resin which is polymerized at low temperature (i.e. 70 °C), then slices of ca. 60 to 100 nm are
cut with an ultramicrotome apparatus equipped with a diamond knife. This technique allows
visualizing both the cyano-bridged coordination polymer nanoparticles and the matrix. TEM
measurements were carried out at 200kV with a microscope JEOL CX200. The nanoparticles
size distribution histograms were determined using enlarged TEM micrographs taken at
magnification of x50K. A large number of nanoparticles (400-600) were counted in order to
obtain a size distribution with good statistics. Scanning Electron Microscopy measurements
were performed with a Philips Quant 200 operating at 15 kV equipped with a Bruker detector.
System software for EDX analysis was developed by Bruker.
Surface area was obtained using nitrogen adsorption isotherms on an ASAP2020 analyser
from Micromeritics. Samples were outgassed under vacuum at 60 °C over night prior to
analysis. Surface area was determined using BET method. Cesium species concentrations in
solution for sorption experiment in inactive conditions were measured using ionic
chromatography (from Metrohm). Solutions were injected in a column as mobile phase. Cs
species are retained on the stationary phase and are then eluted after 20 min and analyzed by
9
conductivity. Cobalt and Iron species concentrations in solution for sorption experiment were
measured using coupled plasma atomic emission spectroscopy from SPECTRO.
2. Results and discussion
2. 1. Synthesis of the nanocomposite materials.
For the synthesis of materials containing cyano-bridged coordination polymer
nanoparticles, we used first two different matrices: (i) SBA-15 mesostructured silica with 10
nm pore size as powder. This material was used as a model for investigations of different
synthetic steps and sorption mechanisms; and (ii) porous glass pearls presenting higher
chemical and thermal stability than silica which can be used for a cesium extraction in column
process. On the other hand, the Co3[Fe(CN)6]2 nanoparticles were choose for their high
cesium absorption capacity in the bulk state. [Delchet, unpublished work] Note that cobalt
ferrocyanides are most often used in low radioactive waste treatment (Lehto, Paajanen et al.
1992) (Ca and Cox 2004) and also mentioned in the literature as the most selective
compounds for Cs+ extraction stable under  irradiation (Lehto and Szirtes 1994).
Figure 1 schematically represents the method that we used in order to form the
nanocomposite materials containing Co2+/[Fe(CN)6]3- coordination polymer nanoparticles into
the porous silica or glass matrices functionalized by -(CH2)2C5H4N. It consists of the intrapore
growth of cyano-bridged networks at specific sites of the matrices performed by consecutive
coordination of Co2+ and [Fe(CN)6]3- using the following typical procedure described
previously.
The elemental analysis of the obtained nanocomposites Co2+/[Fe(CN)6]3-/Silica-Py
(CoFC/Silica-Py) and Co2+/[Fe(CN)6]3-/PG-Py (CoFC/PG-Py) are given in Table 1 as well
as the Co/Fe atomic ratio. For all the as-obtained nanocomposites the latter is closed to 1.2.
Consequently, the content of Co2+/[Fe(CN)6]3- particles inserted to the pores of Silica-Py and
Glass-Py matrixes can be estimated to about 10 wt % and 3 wt %, respectively. Compared to
the specific surface of each support, the amount of Co-ferrocyanide based nanoparticles is
proportional to the available surface of the porous materials (Table 2).
10
Figure 1 : Schematic representation of the intrapore growth of cyano-bridged coordination polymer
nanoparticles Co2+/[Fe(CN)6] 3- by using mesostructured silica or porous glass as matrix.
Table 1: Elemental analysis of obtained nanocomposites and IR spectra data.
Sample
Co/Si
Fe/Si
Co/Fe
Wt % CoCF
IR, cm-1
Bulk
-
-
1.64
-
2087, 2159
CoFC@Silica-Py
0.027
0.021
1.28
10
2117, 2159
CoFC@PG-Py
0.006
0.005
1.17
3
2121, 2159
Co3[Fe(CN)6]2.H2O
2.2. Textural characterizations of nanocomposites.
The porous structure of the samples was characterized by using nitrogen adsorption
isotherm and SAXS measurements. The nitrogen adsorption isotherms allow determining the
specific surface by Brunauer-Emmett-Teller (BET) method, and the total pore volume Vp
(cm3.g-1). The nitrogen physisorption isotherms of the pristine mesoporous matrices Silica and
PG as well as the corresponding nanocomposites CoFC@Silica-Py and CoFC@PG-Py are
shown Figure 2a. Mesostructured silica matrix exhibits a typical adsorption-desorption
isotherm of a mesoporous structure of type IV with an H1 hysteresis loop. Similar type of
isotherm and mesoporosity is obtained for the correcponding nanocomposite CoFC@SilicaPy, that proves the preservation of the cylindrical pore system after intrapore growth of the
cyano-bridged metallic coordination polymer. The amount of adsorbed nitrogen as well as the
BET surface was reduced after formation of the nanoparticles. The decrease of the
mesoporous volume and the decrease of the pore diameter after formation of the
nanoparticles, clearly demonstrate a filling of the pores with the guest species leading to the
conclusion that the nanoparticles are formed inside the pores. The total pore volume
calculated at p/p0 = 0.9 is 0.83 cm3.g-1 for the pristine mesoporous silica and 0.41 cm3.g-1 for
11
CoFC@Silica-Py (Table 2). This indicates a degree of filling of the pores of ca. 50 %. The
nitrogen physisorption isotherms and pore-size distributions at the adsorption branch of the
pristine mesoporous glass PG-Py and its respective nanocomposite CoFC@PG-Py (Figure
2a, Table 2) also present a filling of the pores with the cyano-bridged polymer nanoparticles
with degrees of filling of ca. 10 % .
SAXS measurements have been performed on samples Silica-Py and CoFC@SilicaPy. As shown Figure 2b, the initial functionalized silica has a hexagonal porous structure with
the characteristic peaks of SBA-15: a first peak at 2=0.9° corresponding to the 100 plane, a
second peak at 2=1.5° for the 110 plane and the third one at 2=1.8° for the 200 plane. After
pore filling by nanoparticles, the identical to the hexagonal type with the characteristics of a
SBA-15 porous texture was obtained (Table 2). Combining SAXS and adsorption experiment
gives an easy and accurate method for the determination of the pore diameter ((nm))
(Table 2) thanks to this following equation (Kruk, Jaroniec et al. 1997):
 V p
  1.2d 100 
 1  V
p





1
2
where =2.2 is the density of the wall in case of silica.
Note that pristine hybrid glass PG-Py and its respective nanocomposite CoFC@PG-Py are
amorphous and cannot be characterized by SAXS. Here, the pore diameter was evaluated
from BJH method on the nitrogen adsorption curve..
Table 2: Nitrogen adsorption isotherms and SAXS data for the pristine functionalized matrices and
corresponding nanocomposites.
Samples
SBET
Porous
Pore
d100
Pore
(m2.g-1)
Volume
filling, %
(nm)
diameter
(nm)
(cm3.g-1)
Silica-Py
540
0.83
CoFC@Silica-Py
231
0.41
PG
113
1.1
CoFC@PG-Py
94
1.0
* nanoparticles diameter obtained from TEM measurements
d* (nm)
50
10
9.72
9.4
-
9.83
8.1
5.8 ± 1.4
-
30
-
28
2.7 ± 0.8
12
1.2
Silica
CoFC@Silica-Py
PG
CoFC@PG-Py
3
Vads (cm /g)
0.9
0.6
0.3
0.0
0.0
0.2
0.4
0.6
0.8
1.0
Intensity (u.a.)
P / P0
Silica
CoFC@Silica-Py
0
1
2
3
4
2 (°)
Figure 2 : (a) Nitrogen adsorption isotherms for the mesoporous Silica and glass PG as well as for the related
nanocomposites CoFC@Silica-Py and CoFC@PG-Py; (b) SAXS experiment of the samples after each steps of
grafting.
The results of the nitrogen physisorption and SAXS were substantiated by Transmission
Electronic Microscopy (TEM). The TEM images of the pristine silica clearly show the
hexagonal ordering of the pores (Figure 3a). The TEM measurements performed for the
nanocomposite CoFC@Silica-Py indicate that the hexagonal structure of the host materials is
still retained (Figure S1a, ESI). No visible particles separated out of the surface of pores
13
were observed. As expected, no aggregates can be observed into the pores that is indicative of
a homogeneous dispersion of the CoFC cyano-bridged polymer network in the silica matrix.
The nanoparticles of CoFC can be clearly seen after removal of silica from the
nanocomposite material using an extractive replica technique (see experimental section for
details) (Figure 3b). Sample presents relatively narrow size distribution of nanoparticles
centered at 5.8 ± 1.4 nm (Table 2). This value is slightly smaller than the pore channel mean
diameter of the respective host silica of 10 nm obtained from BET and SAXS measurements
(Table 2).
The same conclusions may be obtained from TEM observations of pristine glass PG-Py
(Figure 3c) and its nanocomposite CoFC@PG-Py (Figures S1b, ESI), and after removal of
the glass from the nanocomposite material using extractive replica (Figure 3d). Nanoparticles
with a size distribution centered at 2.7±0.8 nm are obtained, which is even smaller than when
using a SBA-15 matrix (Table 2).
14
Figure 3: TEM pictures for (a) the pristine silica (SBA-15)( scale bar = 200 nm); (b) the nanoparticles of CoFC
after removal of the silica from the nanocomposite ( scale bar = 100 nm); (c) the pristine glass beads (scale bar
= 200 nm) and (d) for the nanoparticles of CoFC after removal of the glass from the nanocomposite (scale bar =
50 nm). Inset shows magnification of 3d.
2.3. Structural characterisation of the nanocomposites.
The IR spectroscopy was performed on the functionalized silica and glass matrices before and
after nanoparticles formation especially in the spectral window 2000 – 2300 cm-1, i.e. in the
vicinity of the CN stretching mode, which is a fingerprint of structural and electronic changes
occurring in Prussian Blues analogous. The CN stretching frequency of a free CN - ion is 2080
cm-1, whereas upon coordination to a metal ion, it shifts to higher frequencies.NJC article The IR
spectra of both obtained nanocomposites (Table 1, Figure S2, ESI) show two characteristic
absorption bands in the CN stretching region at 2159 cm-1, 2117 cm-1 for CoFC@Silica-Py
and at 2159cm-1, 2121 cm-1 for CoFC@PG-Py. The high frequency bands can be attributed to
the stretching of the CN ligand bridged between Co2+ and Fe3+ in Co2+-CN-Fe3+ mode and the
low frequency band can be attributed to the linkage isomer with Co2+-NC-Fe3+ coordination
mode, as it was reported for the bulk cyano-bridged coordination polymer (Table 2). [E. Reguera,
J.F. Bertran, C. Diaz, J. Bianco, S. Rondon, Hyperfine Interact., 1990, 53, 391, D.H.M. Buchold and C. Feldmann, Chem. Mater., 2007, 19,
3376.
]. As expected, the IR spectra of the nanocomposites also present SiO2 vibration bands at
15
1080, 948, 798 and 459 cm-1.
XRD powder patterns of the nanocomposites materials containing cyano-bridged coordination
polymer nanoparticles (silica and glass) in the range 10 – 60 2 were compared the XRD
powder pattern of the bulk materials (Figure 4). Both nanocomposites materials show phases
very similar to those of the bulk CoFC, meaning the presence of the crystalline cyano-bridged
coordination polymer in these materials, presumably in the pore size. . The presence of these
characteristic peaks in the diffraction patterns of nanocomposites indicates that the amount of
Intensity (u.a)
nanoparticles in these materials is significant.
CoFC@Silica-Py
CoFC@Glass-Py
Bulk CoFC
20
30
40
50
60
2
Figure 4 : XRD powder pattern for the nanocomposite materials CoFC@Silica-Py (blue) and CoFC@Glass-Py
(red) compared of the XRD powder pattern of bulk CoFC (black).
2.4. Cesium Sorption experiments.
In general way, a sorption of an element in solution on a solid support depends on both, the
characteristics of the support material (ionic exchange capacity, porosity, composition …) and
the parameters of the solution (ionic concentration, pH, mixing …). There are two types of
adsorption: a physisorption in which the ions in solution are adsorbed on the solid by
electrostatic force, and a chemisorption for which the ions form a chemical bond with the
solid. The mechanism of Cs+ sorption by the bulk hexacyanoferrates depends on the type of
this latter. In the case of the anionic cyanometallate complexes with potassium cation,
KM[Fe(CN)6], it is assumed in the literature that there is a true exchange between potassium
and cesium ions (Haas 1993) (Lehto, Harjula et al. 1987) without modifying the crystal
16
structure (Loos-Neskovic, Ayrault et al. 2004). It is also shown that not all the potassium ions
can be exchanged by the cesium and the exchange degree depends on the transition metal
used (Loos-Neskovic, Ayrault et al. 2004) (Avramenko, Bratskaya et al. 2011). On the
contrary, in the case of neutral hexacyannoferrates, M3[Fe(CN)6]2, the mechanism involved in
the sorption of Cs+ is far to be well described and understood. Some authors assume an ionic
exchange between the M2+ ion of the structure and the cesium (Ayrault, Jimenez et al. 1998)
(Haas 1993; Valsala, Joseph et al. 2009; Avramenko, Bratskaya et al. 2011), while other
authors assume that Cs+ ions are incorporated into the cage structure of the metal
hexacyannoferrate as ion pair with nitrate (Ca and Cox 2004), or explain that their sorption
results by an exchange process involving only the surface layer of the crystallites (Lehto,
Harjula et al. 1987) (Ramaswamy 1999). Thus the mechanism of the Cs+ sorption of by
neutral cyanometallates seems to be complex and depends on the experimental conditions,
which are different in each publication making some difficulty in comparison or prediction of
the sorption capacities (Avramenko, Bratskaya et al. 2011).
2.4.1. Adsorption kinetics
Monitoring the kinetics of sorption allows the determination of the time required to
reach equilibrium but also the empirical order of the reaction and the experimental exchange
capacity, thanks to a kinetics reaction model. Moreover, from these tests the diffusive model
allows to highlight the steps limiting sorption kinetics (Crini, Peindy et al. 2007).
The effect of the contact time on the amount of adsorbed cesium performed for both hybrid
material, CoFC@Silica-Py and CoFC@Glass-Py, as well as for the bulk CoFC was
analyzed by using two different models, a kinetics reaction model and a diffusive model
(Figure 5). These results demonstrate that the kinetics of adsorption of both hybrid material is
more rapid than the one of the bulk CoHF: the process is quite rapid and equilibrium is
reached at about 1 h in the case of the hybrid material and at about 10 h in the case of the bulk
solid. On the other hand, the adsorption capacity at equilibrium is four or twenty times smaller
in the case of hybrid material (silica and glass, respectively) compared to the bulk one. This
result is not surprising taking into account that the adsorption capacities have been calculated
as mmol of adsorbed Cs+ by gram of whole materials and that the amount of CoHF
nanoparticles into the matrices is 10 and 3 wt % for CoFC@Silica-Py and CoFC@Glass-Py,
respectively. This point will be discussed in the adsorption isotherm section.
17
Qads (mmol/g)
0.5
0.4
0.3
0.2
0.1
0.0
0
300
600 900 1200 1500
Time (min)
Figure 5a
0.5
Qads (mmol/g)
0.4
0.3
0.2
0.1
0.0
0
10
20
30
0.5
Sqrt(time) (min )
40
Figure 5b
Figure 5: Effect of contact time on the amount of adsorbed cesium on the hybrid material, CoFC@Silica-Py (■)
and CoFC@Glass-Py (○) compared to the same experiment on bulk CoFC (●) (a) kinetics reaction model (lines
correspond to the fit of experimental data with the kinetics reaction model) and (b) diffusive model, quantity of
cesium Qt (mmol.g-1) adsorbed versus square root of time.
Kinetic reaction model.
Generally, the reaction rate of a chemical reaction is expressed as a function of the reactants
concentration. When this reaction involves a solid, the reaction order called also as a pseudo
reaction order because it is based on the sorption capacity of the solid. Usually it may be
18
determined by the linearization of the reaction rate. In the most cases of the sorption on a
solid, there is a reaction of pseudo second order. The kinetic reaction model assumes that the
reaction rate is limited by only one process or the sorption mechanism occurs on a single class
of sorbing sites. The linearization of the reaction rate gives the following expression:
t
1
1


t
2
Qt k 0 Qe Qe
(E4)
where k0 (g/mmol/min) is the rate constant of pseudo second order adsorption; t (min) is the
time; Qt and Qe (meq/g) are the adsorption capacity at the time t and at the equilibrium,
respectively. The plot of (t/Qt) as a function of time gives a linear relationship. .
If it is possible to determine a reaction order for the sorption, so the sorption is a
chemisorption mechanism and not a physisorption one.
The analysis of the experimental data using this pseudo-second order kinetic model has been
conducted. This plot of t/Qt versus t is linear in both cases, hybrid materials and CoHF bulk,
and the correlation coefficient is good (larger than 0.9) suggesting that the sorption follows
pseudo second order kinetics.
Diffusive model.
The chemisorption of an ion by a solid involves several limiting steps:

Diffusion of ions in solution;

Diffusion of ions in transition layer between liquid and solid called the boundary
layer;

Adsorption (or chemisorption) of ions inside the solid;

Diffusion of these adsorbed ions through the solid (intra particle diffusion).
Except the adsorption step, all the other steps are diffusion limited. The determination of the
limiting step of the adsorption process may be achieved by plotting the amount of adsorbed
ions as a function of the square root of time. Note that we neglect the diffusion in the liquid
phase because the stirring of solutions. Figure 5b shows the plot of the quantity of adsorbed
cesium, Qt (mmol.g-1), vs the square root of time for both nanocomposite materials as well as
for the bulk CoHF.
The curve performed for the bulk CoHF shows two linear zones but the saturation is not
totally reached indicating that the equilibrium stage is not obtained. According to the
literature, the presence of two linear zones can be explained by the presence of two sorption
processes: an external mass transfer, such as boundary layer diffusion, and intra-particle
19
diffusion. The initial curve zone is attributed to the external diffusion, and the second one is
defined as the intraparticle diffusion (Crini, Peindy et al. 2007; Metwally, Rahman et al.
2007). Note also that the absence of saturation has also be noted for
bulk cupper
ferrocyanides, and has been attributed to a small reorganization of the solid accompanied by a
release of copper in the solution (Ayrault, Jimenez et al. 1998).
The curves plotted for the nanocomposite samples, CoFC@Silica-Py and CoFC@Glass-Py
don’t show clear linear portion (Figure 5b). Only a plateau corresponding to the equilibrium
stage seems to be present on these curves. In these cases, any diffusion process seems to be a
rate-controlled step. This difference could be attributed to the presence of nano-sized CoHF
particles inside the porous silica support presenting a large surface for cesium sorption
without rate limiting steps, whereas in the bulk sample, the diffusion limited process would be
due to the micron-size of sorbent. This observation can also be linked with a sorption
mechanism involving mainly the surface layer of the crystallites (Lehto, Harjula et al. 1987;
Ramaswamy 1999).
2.4.2. Adsorption isotherm.
The measurements of sorption isotherms describe here the quantity of the adsorbed
cesium on the solid at the equilibrium (Qe, mmol.g-1) vs concentration of cesium in solution
(Ce,mmol/L) (Crini, Peindy et al. 2007). Qe is defined as:
Qe  C0  Ce 
V
m
(E5)
where C0 (mmol.L-1) is the initial Cs+ concentration in solution; V is the volume of solution
(mL) and m the mass of sorbent used (mg).
Figure 6 presents the adsorption isotherms for all sorbents studied. All these curves are
positive and concave to the concentration axis which reflects the efficiency of these materials
for the sorption of cesium ions in a wide range of concentration. For these three materials, the
plots present similar shape with a high slope for the small Cs+ concentrations in solution
indicating that these materials have a good affinity to cesium ions, as expected for
hexacyanoferrates, and a plateau for the large concentrations. This later indicates that the
saturation of powders is achieved when all the sites available for cesium were used.
In order to evaluate the efficiency of a sorbent to remove a desired ion from a liquid phase we
should take into account two main parameters:
20

The maximum adsorption capacity (Qmax) which indicates the efficiency of the
materials to remove Cs+ at (or near) saturation. It can be estimated from the isotherm’s
plateau and more precisely obtained by using the Langmuir model (see above);

The distribution coefficient (Kd) which indicates the uptake of the cesium ions in trace
concentration. Kd is a useful parameter to evaluate the selectivity of the sorbent for Cs+
especially in solution containing competitive ions (such as Na+, K+, Ca2+) in high
concentrations,which can be determined from the slope of the isotherm curve even the
Cs+ concentration in solution is very dilute.
Note that the sorbent with highest Kd does not necessary offer the highest sorption capacity
and vice versa. Consequently, The main parameters required for a sorbent in decontamination
process are first a high distribution coefficient and then a high sorption capacity.
Adsorption capacity
The Langmuir model is often used to describe the isotherms of equilibrium adsorption and to
calculate the adsorption capacity. It assumes that the sorption occurs on structurally
homogeneous sites and all sorption sites are energetically identical. The linear form of the
Langmuir equation is:
Qe
LCe

Qmax 1  LCe
(E6)
where Ce(meq.L-1) is the cesium concentration in solution at equilibrium, Qe(meq.g-1) is the
cesium concentration in solid, Qmax (meq.g-1) is the monolayer adsorption capacity and L is
the Langmuir constant related to the affinity of binding sites, linked to the free energy of
adsorption.
This non-linear form of Langmuir isotherm equation and the least square method were applied
to determine the isotherm parameters and especially the monolayer adsorption capacity. The
high correlation coefficient confirms a chemisorption isotherm mechanism with a monolayer
adsorption. These fits are useful for the determination of the numerical value of Qmax, which
corresponds to the maximum sorption capacity (Table 3). As expected, the sorption capacity
of bulk materials is higher than the one of nanocomposite materials. These values are in
agreement with the adsorption capacities obtained by kinetics experiments.
In order to compare the bulk CoFC and the nanocomposite materials, the adsorption
capacities were calculated from the isotherm experiments in mmol per gram of CoFC (Table
3). First off, the sorption capacities of both nanocomposites (CoFC@Silica-Py
and
21
CoFC@Glass-Py) of 1.3 mmol.g-1 are higher than the one obtained for the bulk CoFC of 0.4
mmol.g-1. This fact may be explained by the presence of a large surface in the case of
nanoparticles in comparison with the bulk analogous. These results are in agreement with a
hypothesis involving surface layer of the crystallite to explain Cs+ sorption by Prussian blue
analogous (Lehto, Harjula et al. 1987; Ramaswamy 1999).
Previously, sorption capacities of bulk cobalt hexacyanoferrates and potassium cobalt
ferrocyanide have been found respectively between 0.4 to 1.2 mmol.g-1 (Mardan, Ajaz et al.
1999; Valsala, Joseph et al. 2009; Avramenko, Bratskaya et al. 2011) and between 0.3 to 1.5
mmol.g-1 (Lehto, Harjula et al. 1987; Ramaswamy 1997; Liu, Li et al. 2009; Avramenko,
Bratskaya et al. 2011). Although as we mentioned above it is difficult to compare the sorption
capacities obtained from different works due to the difference in sorption conditions, these
values are in agreement with these literature values.
Qe (mmol/g)
0.4
0.3
0.2
0.1
0.0
0.0
0.5
1.0
Ce (mmol/L)
Figure 6 : Cs+ adsorption isotherm from pure water of (●) CoFC; (■) CoFC@Silica-Py and (○) CoFC@GlassPy. Dashed lines represent the Langmuir isotherm model.
Table 3: Maximum adsorption capacity and correlation factor of Langmuir fit for each sorbent studied
Sorbent
.
Qmax
Qmax* (mmol.g-1
Kd (sea water at
Kd (radioactive sea
(mmol.g-1
of CoFC)
10 ppm)
water at 29kBq.L-1)
mL.g-1
mL.g-1
composite)
22
CoFC
0.38
0.4
>104 (*)
6 105
CoFC@Silica-
0.13
1.3
>104 (*)
8 105
0.04
1.3
103
3 105
Py
CoFC@GlassPy
(*) The Cs equilibrium concentration in sea water here was out of detection limit (lower than
0.5ppm).
Distribution coefficient
Distribution coefficient, Kd ( L.g-1) is an useful parameter which evaluates a selectivity of the
sorbent to uptake
cesium ions at very low concentrations in the presence of another
competitive ions such as Na+, K+, Ca2+ at high concentrations (Lin, Fryxell et al. 2001). In
another words, high Kd values represent high selectivity for Cs+. Graphically, Kd may be
obtained as the slope of the isotherm curves at low concentrations and may be determined at
equilibrium by the equation:
Kd 
C0  Ce  V
Ce
EX
M
where C0 and Ce are the initial and equilibrium concentrations; V is the volume of the solution
and M is the mass of sorbent used.
First off, the distribution coefficient of the bulk CoFC and the nanocomposite materials
CoFC@Silica-Py, CoFC@Glass-Py were determined in sea water containing 10 ppm of
natural Cs+ with a concentration of 0.5g.L-1 of sorbent (V/M = 2 L.g-1) (Table 3). Sea water
was used to simulate inactive solution with a composition close to the Fukushima site. It is
well known that the affinity of Prussian blue analogous for alkaline ions follows Na<<K<Cs
(Sinha, Humphrey et al. 1984; Schneemeyer, Spengler et al. 1985; Haas 1993). In the case of
sea water, even if the sodium concentration is very high (9.6 g.L-1 of Na), these materials
uptake very low concentrations of Cs+. Therefore, these inorganic sorbents are the materials
of choice for decontamination of
137
Cs enriched water. Secondly, the sorbents have been
tested with the radioactive sea water simulated as the one from the Fukushima site, with
initially 29 kBq.L-1 of 137Cs. The obtained distribution coefficients of 830 000 and 290 000 for
CoFC@Silica-Py and CoFC@Glass-Py, respectively, are higher than 105, close to the ones
found in the literature for other composite materials (Table 3) (Harjula, Lehto et al. 2001).
The difference at Cs concentration of 10ppm (natural sea water) between the values obtained
23
for silica and glass supports can be attributed to the amount of CoFC inserted into the
porosity of these materials.
Moreover, the distribution coefficients obtained at trace concentration of Cs (radioative
enroiched sea water) for both nanocomposite is similar than than the bulk CoFC one,. This
result demonstrate clearly for the first time the effect of the particle size on the distribution
coefficient. That means that nanoparticles presents both higher selectivity and higher capacity.
Therefore our composites are promising materials for the decontamination of sea water
enriched with 137Cs like Fukushima site.
Furthermore, the same nanocomposites may be obtained by using functionalized glass
beads required as an excellent support for a continuous decontamination process in a column
method. After Cs+ absorption, the porosity closing by using a soft method like a treatment at
low temperature or a treatment in basic conditions may be performed that allow a Cs +
confinement for further storage. These studies are actually in run.
3. Conclusion
In summary, the concept of using nanocomposites containing Prussian blue type nanoparticles
covalently linked to the matrix appears to be a promising route to uptake cesium ion. In this
article, we prepared two types of nanocomposites based on functionalized silica or glass
matrixes containing cyano-bridged coordination polymers Co2+/[Fe(CN)6]3-. They have been
obtained by successive coordination of the cobalt ions and the ferricyanide precursors at the
specific amino groups of matrixes and contain the uniform-sized spherical NPs of 5.8 and 2.7
nm homogeneously distributed in the matrix’s pores. The as-obtained nanocomposites have
been used for the selective Cs+ extraction from different effluents: pure water, sea water and
simulated radioactive sea water from Fukushima site.
The first point to note is that the kinetics of cesium sorption is ten times faster and the
sorption capacities per gram of ferrocyanide are about 3 times higher in case of
nanocomposites compared to the bulk Prussian blue analogous CoFC. This fact was attributed
to the high surface area of the Prussian blue nanoparticles in case of composites. The amount
of inserted nanoparticles and the adsorption capacities per gram of composite materials are
directly linked to the surface are of the porous silica support used (ordered silica or glass
beads).
The second point is that these nanocomposites present a high selectivity to Cs+. In the
presence of high concentration of sodium in sea water, they uptake very low concentrations of
24
Cs+. Therefore, these inorganic sorbents are the materials of choice for decontamination of
137
Cs enriched water.
Thirdly, the experiments done with sea water enriched with
137
Cs simulating Fukushima
contaminated effluents demonstrate the high potential of these nanocomposite materials for an
efficient decontamination column process. Moreover, after the decontamination, these glassbased composites should be used as confinement matrices by closing the porosity softly.
Acknowledgements
This work was supported by the ANR Matepro, the University Sud de France, the Matinex
National Research Group and the CEA. We acknowledge Guillaume Serve, and Célia
Lepeytre for their help for radioactive experiments.
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28
Electronic Supplementary Information
Extraction of Radioactive Cesium Using Innovative Functionalized Porous
Materials
Carole Delchet 1,2, Alexei Tokarev2, Xavier Dumail2, Guillaume Toquer1, Yves Barre3,
Yannick Guari,2* Joulia Larionova,2 Agnès Grandjean.1*
1
Institut de Chimie Séparative de Marcoule, UMR5257 CEA-CNRS-UM2-ENSCM, BP17171, 30207
Bagnols sur Cèze, France. Tel.: +33(0)4 66 79 66 22 ; e-mail: agnes.grandjean@cea.fr
2
Institut Charles Gerhardt Montpellier, UMR 5253 CNRS-UM2-ENSCM-UM1, Chimie Moléculaire et
Organisation du Solide, Université Montpellier II, Place E. Bataillon, 34095 Montpellier cedex 5, France.
Tel.: +33(0)4674805; e-mail:joulia.larionova@univ-montp2.fr
3
CEA/DEN/DTCD/SPDE/ Laboratoire Des Procédés Avancés de Décontamination, Centre de Marcoule,
BP17171, 30207 Bagnols sur Cèze, France
29
Figure S1. TEM images of the nanocomposites (a) CoFC@Silica-Py and (b) CoFC@PG-
Intensity (a.u)
Py.
2300
CoFC (bulk)
CoFC@Glass-Py
CoFC@Silica-Py
2200
2100
-1
IR Shift (cm )
2000
Figure S2. IR spectra of the nanocomposites CoFC@Silica-Py and CoFC@PG-Py compared
to the bulk materials.